Earthworm ecology

SECOND EDITION

Earthworm Ecology

© 2004 by CRC Press LLC

SECOND EDITION

Earthworm Ecology Edited by

Clive A. Edwards

CRC PR E S S Boca Raton London New York Washington, D.C.

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Library of Congress Cataloging-in-Publication Data Earthworm ecology / edited by Clive A. Edwards. -- 2nd ed. p. cm. Rev. ed. of: Earthworm ecology / edited by Clive A. Edwards. 1994. Includes bibliographical references and index. ISBN 0-8493-1819-X (alk. paper) 1. Earthworms--Ecology--Congresses. I. Edwards, C. A. (Clive Arthur), 1925- II. Earthworm ecology. III. Title. QL391.A6E25 2004 592′.64—dc22 2003070024 This book contains information obtained from authentic and highly regarded sources. Reprinted material is quoted with permission, and sources are indicated. A wide variety of references are listed. Reasonable efforts have been made to publish reliable data and information, but the author and the publisher cannot assume responsibility for the validity of all materials or for the consequences of their use. Neither this book nor any part may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopying, microfilming, and recording, or by any information storage or retrieval system, without prior permission in writing from the publisher. All rights reserved. Authorization to photocopy items for internal or personal use, or the personal or internal use of specific clients, may be granted by CRC Press LLC, provided that $.50 per page photocopied is paid directly to Copyright clearance Center, 222 Rosewood Drive, Danvers, MA 01923 USA. The fee code for users of the Transactional Reporting Service is ISBN 0-8493-1819-X/04/$1.00+$.50. The fee is subject to change without notice. For organizations that have been granted a photocopy license by the CCC, a separate system of payment has been arranged. The consent of CRC Press LLC does not extend to copying for general distribution, for promotion, for creating new works, or for resale. Specific permission must be obtained in writing from CRC Press LLC for such copying. Direct all inquiries to CRC Press LLC, 2000 N.W. Corporate Blvd., Boca Raton, Florida 33431. Trademark Notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation, without intent to infringe.

Visit the CRC Press Web site at www.crcpress.com © 2004 by CRC Press LLC No claim to original U.S. Government works International Standard Book Number 0-8493-1819-X Library of Congress Card Number 2003070024 Printed in the United States of America 1 2 3 4 5 6 7 8 9 0 Printed on acid-free paper

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Preface Charles Darwin was the first scientist to bring earthworms to the attention of scientists and the general public, more than a century ago. Darwin noted the importance of earthworms in breaking down dead plant materials, recycling the nutrients they contain, and turning over soil. His book The Formation of Vegetable Mould through the Action of Worms (1881) summarized his conclusions on earthworms, which he reached after 40 years of observation and experimental work. In this book, he expressed the opinion that “earthworms have played a most important part in the history of the world.” The importance of his personal contributions to our knowledge of the roles and biology of earthworms cannot be stressed enough and led to a great upsurge in research into the morphology, histology, and taxonomy of earthworms in the late 19th and early 20th centuries. However, it was only in the last 25 years that interest in and research into the ecology and biology of earthworms has peaked. Much of this work was summarized by Edwards and his coauthors in their book The Biology and Ecology of Earthworms (first edition 1972, second edition 1977, third edition 1996) and by Lee in his book Earthworms: Their Ecology and Relationships with Soil and Land Use (1985). Interest in earthworm ecology and the importance of earthworms to soil formation and fertility has been increasing at an extremely rapid rate and so has research into the subject. This is evidenced by the increases in the number of references cited by the authors of The Biology and Ecology of Earthworms in its three editions. In 1972, they cited 565 references; in the second edition (1977), they cited 674; but in the third edition (1996), they cited more than 1500. This probably represented only a third of scientific papers published up to that time. The first edition of Earthworm Ecology (1998) owed its origin to the Fifth International Symposium on Earthworm Ecology, which was held in Columbus, Ohio, in July 1994. At this Symposium, attended by more than 220 scientists from 38 countries, 165 research presentations were made, many of which are published in a special volume of the journal Soil Biology and Biochemistry. In the eight sessions that were held at the Columbus Symposium, each opened with an invited review paper of a key topic by a distinguished earthworm scientist and concluded with a final overview of the subject and conclusions by another well-known earthworm scientist. The 16 invited papers were edited to form the eight sections in the first edition of Earthworm Ecology, which covered all the major aspects of earthworm ecology, including earthworm diversity, behavior, physiology and general ecology, and the roles of earthworms in nutrient cycling, soil maintenance, plant growth, ecotoxicology, and waste management, with two chapters summarizing research on each topic. Since the first edition of Earthworm Ecology was published in 1998, there have been two further Symposia on Earthworm Ecology, in Vigo, Spain, in 1998 and in Cardiff, Wales, in 2002; the number of publications on earthworms has continued to increase rapidly. The first edition was extremely well received by scientists, students, and the general public. In view of the rapidly expanding developments and discoveries in earthworm biology and ecology, it seemed appropriate to update, and revise extensively, the first edition of the book and add new chapters that address the most rapidly developing areas of earthworm research. This second edition includes extensive revisions of the original chapters as well as additional chapters on the history of earthworm research, mechanisms by which earthworms increase soil fertility and promote plant growth, and the importance of invasions of exotic species of earthworms in North America and other regions of the world; there is a new chapter on vermiculture and vermicomposting in Europe. These changes make this book an even more valuable addition to the

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publications that summarize the increasing importance of earthworms in natural ecosystems and crop production. It also addresses key issues in earthworm biology and ecology and is an essential key reference work for soil scientists and agronomists as well as those people with a great interest in earthworms. Clive A. Edwards The Ohio State University, Columbus

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About the Editor Clive Edwards, Ph.D., is recognized as a world authority on earthworms, and his book Ecology and Biology of Earthworms is now in its third edition. After graduating from Bristol University and then earning a M.S. and Ph.D. at the University of Wisconsin, U.S.A., Dr. Edwards was appointed to the U.K. Ministry of Agriculture. In 1960 he joined Rothamsted Experimental Station as a Senior Principal Scientific Officer where his work focused on research into the effects of agricultural chemicals on the soil environment. From 1966 to 1968 he was visiting professor at Purdue University, U.S.A. He was appointed as Chair of the Department of Entomology at The Ohio State University, U.S.A. in 1985. Dr. Edwards has published extensively on soil ecology, environmental toxicology, and sustainable agriculture, and he is currently recognized as a world authority on earthworms. His book Ecology and Biology of Earthworms is the first comprehensive book on earthworms since Charles Darwin’s The Formation of Vegetable Mould Through the Action of Worms, which was published in 1881. In 1996, Professor Edwards’ book Ecology of Earthworms won a Presidential Citation from the U.S. Soil & Water Conservation Society. In 2001, Dr. Edwards presented The Ohio State University Distinguished Lecture The Future of Human Populations; Energy, Food and Water Availability in the 21st Century — one of the university’s highest honors for a faculty member. His involvement with the British Crop Protection Council has been an outstanding contribution to all the Pests & Diseases Conferences.

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Contributors Jean Andre Université de Savoie Chambery, France Norman Q. Arancon Soil Ecology Laboratory The Ohio State University Columbus, Ohio, U.S.A. Geoff H. Baker CSIRO Entomology Canberra, Australia Nicolas Bernier Université de Savoie Chambery, France John M. Blair Division of Biology Kansas State University Manhattan, Kansas, U.S.A. Patrick J. Bohlen Archbold Biological Station Lake Placid, Florida, U.S.A. George G. Brown Embrapa Soya Londrina, Brazil Lijbert Brussaard Soil Quality Section Wageningen University Wageningen, The Netherlands Fabienne Charpentier Laboratoire d’Ecologie des Sols Tropicaux Bondy, France

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James P. Curry Department of Environmental Resource Management University College, Belfield Dublin, Ireland Laurent Derouard Laboratoire d’Ecologie des Sols Tropicaux Bondy, France Jorge Domínguez Departamento de Ecoloxía e Bioloxía Animal Universidade de Vigo Vigo, Spain Bernard M. Doube Wood Duck Cellars Bridgewater, South Australia, Australia Clive A. Edwards Soil Ecology Laboratory The Ohio State University Columbus, Ohio, U.S.A. Herman Eijsackers Alterra, Wageningen University and Research Centre Institute of Ecological Sciences Vrije Universiteit Amsterdam, The Netherlands Cécile Gilot Yurimaguas, Loreto, Peru Paul F. Hendrix Institute of Ecology University of Georgia Athens, Georgia, U.S.A.

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Samuel W. James Department of Life Sciences Maharishi University of Management Fairfield, Iowa, U.S.A. Radha D. Kale Department of Zoology University of Agricultural Sciences Bangalore, India André Kretzschmar INRA-Biometrié Avignon, France

Adriana Antonia Pop Institute of Biological Research Cluj-Napoca, Romania Victor V. Pop Institute of Biological Research Cluj-Napoca, Romania Adriaan J. Reinecke Department of Zoology University of Stellenbosch Stellenbosch, South Africa

Patrick Lavelle Laboratoire d’Ecologie des Sols Tropicaux Bondy, France

Sophié A. Reinecke Department of Zoology University of Stellenbosch Stellenbosch, South Africa

Renée-Claire Le Bayon Department of Plant Ecology Neuchâtel University Neuchâtel, Switzerland

John W. Reynolds Oligochaetology Laboratory Kitchener, Ontario, Canada

Mary Ann McLean Department of Biology Indiana State University Terre Haute, Indiana, U.S.A.

Jean-Pierre Rossi Laboratoire d’Ecologie des Sols Tropicaux Bondy, France

Dennis Parkinson Department of Biological Sciences University of Calgary Calgary, Alberta, Canada

Stefan Scheu Institute of Zoology Darmstadt University of Technology Darmstadt, Germany

Robert W. Parmelee Yucca Valley, California, U.S.A.

Martin J. Shipitalo North Appalachian Experimental Watershed U.S. Department of Agriculture Agricultural Research Service Coshocton, Ohio, U.S.A.

Beto Pashanasi Estacion Experimental San Ramon INIAA Yurimaguas, Loreto, Peru Jean-François Ponge MNHN Brunoy, France

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Cécile Villenave Laboratoire d’Ecologie des Sols Tropicaux Bondy, France

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Table of Contents Part I Introduction ......................................................................................................................................1 Chapter 1 The Importance of Earthworms as Key Representatives of the Soil Fauna.....................................3 Clive A. Edwards Chapter 2 How Earthworms Affect Plant Growth: Burrowing into the Mechanisms.....................................13 George G. Brown, Clive A. Edwards, and Lijbert Brussaard Part II Earthworm Taxonomy, Diversity, and Biogeography.................................................................51 Chapter 3 Planetary Processes and Their Interactions with Earthworm Distributions and Ecology......................................................................................................................................53 Samuel W. James Chapter 4 The Status of Earthworm Biogeography, Diversity, and Taxonomy in North America Revisited with Glimpses into the Future ...............................................................63 John W. Reynolds Chapter 5 Invasion of Exotic Earthworms into North America and Other Regions.......................................75 Samuel W. James and Paul F. Hendrix Part III Earthworm Biology, Ecology, Behavior, and Physiology............................................................89 Chapter 6 Factors Affecting the Abundance of Earthworms in Soils..............................................................91 James P. Curry Chapter 7 A Comprehensive Study of the Taxonomy and Ecology of the Lumbricid Earthworm Genus Octodrilus from the Carpathians.....................................................................115 Victor V. Pop and Adriana Antonia Pop

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Part IV Influence of Earthworms on Soil Organic Matter Dynamics, Nutrient Dynamics, and Microbial Ecology .................................................................................................................143 Chapter 8 Effects of Earthworms on Soil Organic Matter and Nutrient Dynamics at a Landscape Scale over Decades......................................................................................................145 Patrick Lavelle, Beto Pashanasi, Fabienne Charpentier, Cécile Gilot, Jean-Pierre Rossi, Laurent Derouard, Jean Andre, Jean-François Ponge, and Nicolas Bernier Chapter 9 Integrating the Effects of Earthworms on Nutrient Cycling across Spatial and Temporal Scales..........................................................................................................161 Patrick J. Bohlen, Robert W. Parmelee, and John M. Blair Part V Effects of Earthworms on Soil Physical Properties and Function..........................................181 Chapter 10 Quantifying the Effects of Earthworms on Soil Aggregation and Porosity .................................183 Martin J. Shipitalo and Reneé-Claire Le Bayon Chapter 11 Effects of Earthworms on Soil Organization ................................................................................201 André Kretzschmar Part VI Interactions of Earthworms with Microorganisms, Invertebrates, and Plants......................211 Chapter 12 Functional Interactions between Earthworms, Microorganisms, Organic Matter, and Plants .......................................................................................................................................213 George G. Brown and Bernard M. Doube Chapter 13 Impacts of Earthworms on Other Biota in Forest Soils, with Some Emphasis on Cool Temperate Montane Forests ..................................................................................................241 Dennis Parkinson, Mary Ann McLean, and Stefan Scheu Part VII Earthworms in Agroecosystems..................................................................................................261 Chapter 14 Managing Earthworms as a Resource in Australian Pastures.......................................................263 Geoff H. Baker

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Chapter 15 Earthworms in Agroecosystems: Research Approaches ...............................................................287 Paul F. Hendrix and Clive A. Edwards Part VIII Earthworms and Environmental Pollution ...............................................................................297 Chapter 16 Earthworms as Test Organisms in Ecotoxicological Assessment of Toxicant Impacts on Ecosystems...................................................................................................299 Adriaan J. Reinecke and Sophié A. Reinecke Chapter 17 Earthworms in Environmental Research .......................................................................................321 Herman Eijsackers Part IX Earthworms in Waste Management ...........................................................................................343 Chapter 18 The Use of Earthworms in the Breakdown of Organic Wastes to Produce Vermicomposts and Animal Feed Protein .....................................................................................345 Clive A. Edwards and Norman Q. Arancon Chapter 19 The Use of Earthworms: Nature’s Gift for Utilization of Organic Wastes in Asia .....................381 Radha D. Kale Chapter 20 State-of-the-Art and New Perspectives on Vermicomposting Research .......................................401 Jorge Domínguez

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Part I Introduction

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Importance of 1 The Earthworms as Key Representatives of the Soil Fauna Clive A. Edwards Soil Ecology Laboratory, The Ohio State University, Columbus, Ohio, U.S.A.

CONTENTS History ................................................................................................................................................3 Earthworm Taxonomy........................................................................................................................4 Earthworm Ecology ...........................................................................................................................4 Earthworms and Soil Fertility............................................................................................................5 Soil Formation ............................................................................................................................5 Turnover of Soil..........................................................................................................................6 Soil Aeration and Drainage ........................................................................................................6 Organic Matter Breakdown and Incorporation into Soil ...........................................................6 Nutrient Availability....................................................................................................................7 Effects of Agriculture on Earthworms...............................................................................................7 Earthworms as Indicators of Soil Quality and Health ......................................................................8 Earthworms and Soil Pollution..........................................................................................................8 Earthworm Immigrations ...................................................................................................................9 Need for Earthworm Research ..........................................................................................................9 Conclusions ........................................................................................................................................9 References ..........................................................................................................................................9

HISTORY The great importance of the soil biota in soil pedogenesis and in the maintenance of structure and fertility is not always fully appreciated by physical and chemical soil scientists. Earthworms are arguably the most important components of the soil biota in terms of soil formation and maintenance of soil structure and fertility. Although not numerically dominant, their large size makes them one of the major contributors to invertebrate biomass in soils. Their activities are important for maintaining soil fertility in a variety of ways in forests, grasslands, and agroecosystems. Aristotle was one of the first people to draw attention to the role of earthworms in turning over the soil; he aptly called them “the Intestines of the Earth.” However, it was not until the late 1800s that Charles Darwin, in his definitive work The Formation of Vegetable Mould through the Action of Worms (1881), really brought attention to the extreme importance of earthworms in the breakdown of dead plant and animal matter that reaches soils and in the continued turnover and maintenance of 3 © 2004 by CRC Press LLC

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soil structure, aeration, drainage, and fertility. Before Darwin’s book was published, earthworms were commonly considered soil-inhabiting crop pests. His views on the beneficial aspects of earthworms were supported and expanded subsequently by other contemporary scientists such as Muller (1878), Urquhart (1887), and many others. The observations Darwin described were so advanced that it was half a century before many of them were confirmed (see Chapter 2, this volume).

EARTHWORM TAXONOMY Earthworms belong to the order Oligochaeta, which includes more than 8000 species from about 800 genera. Earthworms are common all over the world in natural forests and grasslands as well as agroecosystems. However, many oligochaetes have an aquatic habit, and there is considerable controversy over earthworm systematics (see Chapters 3 to 5, this volume). Earthworms are found in most regions of the world, except those with extreme climates, such as deserts and areas that are under constant snow and ice. Some genera and species of earthworms, particularly those belonging to the Lumbricidae, are extremely widely distributed and are termed peregrine; often, when these species are introduced to new areas, they become dominant over the endemic species. This situation applies to parts of the northern United States and Canada, particularly those areas close to major waterways (see Chapters 3 and 5, this volume). However, the indigenous earthworm fauna of North America has not been well studied other than by Gates and Reynolds and earlier workers (Chapter 4). Endemic species include those in the Acanthodrilidae, with its most abundant genus Diplocardia; members of the Sparganophilidae; and species in the Megascolecidae, of which the most common genus is Pheretima. There are very few earthworm taxonomists, which has an impact on earthworm research the world over (see Chapter 4, this volume).

EARTHWORM ECOLOGY The size of earthworms ranges from a few millimeters to as much as 2 m in length, from 10 mg to nearly a kilogram in weight, and up to 40 mm in diameter. The record was a specimen believed to be a Microchaetus sp. that was 7 m long and 75 mm in diameter (Lungström and Reinecke 1969). The larger earthworms are usually found in southern latitudes, such as South America, South Africa, Southeast Asia, Australia, and New Zealand. No other terrestrial invertebrate has such a wide range of sizes between the smallest and the largest individuals (Lee 1985) Populations of earthworms vary greatly in terms of numbers or biomass and diversity. Populations range from only a few individuals per square meter to more than 1000 per square meter (Lee 1985; Edwards and Bohlen 1996; Lavelle et al. 1999). The size of populations depends on a wide range of factors, including soil type, pH, moisture-holding capacity of the soil, rainfall, and ambient temperatures, but most importantly, on the ready availability of organic matter. This is because interactions between organic matter and microorganisms provide food for earthworms. Earthworm populations in cultivated land usually do not exceed 100 per square meter or 400 per square meter in grassland, and similar populations to those in grassland are usually found in woodlands, where the availability of organic matter is seldom limiting. Numbers as high as 2000 per square meter have sometimes been recorded, although relatively few earthworms occur in the more acidic mor soils under coniferous forests. Usually, the largest earthworm populations are lumbricids, which seem to be able to survive adverse soil and litter conditions much better than species belonging to many of the other families. The earthworm biomass in most soils exceeds the biomass of all other soil-inhabiting invertebrates. It has been stated that earthworm biomass in a pasture may be ten times that of stock animals that graze on it (see Chapters 6 and 14, this volume). The diversity of species of earthworms varies greatly between sites and habitats, and there often tend to be species associations in different soil types and habitats. Earthworm communities

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in soils in temperate countries are dominated by lumbricids and tend to be considerably less diverse than in soils with other earthworm families in warmer latitudes (Lavelle et al. 1999). However, even in the most complex soil systems, the diversity of earthworm species does not seem to be very great, rarely exceeding ten, and there are usually only three to five species in any particular site. There is some evidence that species that fill the same ecological niche do not normally occur in the same degree of abundance at a particular site (Edwards and Lofty 1982a,b; Edwards and Bohlen 1996). The activity of earthworms differs greatly between seasons in temperate regions, where earthworms are active mainly in the spring and autumn. During the winter, they penetrate deeper into soil, where they are much more protected from the adverse winter cold temperatures. In dry summer periods, they also burrow deeper into soil and sometimes construct cells lined with mucus in which they estivate in a coiled position until environmental conditions become favorable again. Although cocoons may be produced at almost any time of the year, cocoon production is usually seasonal. In temperate regions, the most cocoons are produced in spring or early summer, with a second, much smaller peak in autumn. Numbers of cocoons range from 1 to 20 per mating, depending on species. The life cycles of many species of earthworms have not been well studied. There probably is adequate information on about 12 species of temperate lumbricid earthworms, 7 species from Africa (Lavelle et al. 1999), and 20 species of earthworms common in tropical agroecosystems (Barois et al. 1999). Earthworms have potential for very long life cycles of up to 10 to 12 years, although in the field, many species may live only 1 or 2 seasons because of their susceptibility to a wide range of predators (Edwards and Bohlen 1996). Indeed, their potential longevity, combined with their fecundity, means that very large populations could build up rapidly in the absence of predation or adverse environmental conditions. In addition, some species can produce cocoons parthenogenetically without mating, which increases their potential to spread to new sites. Their moisture and temperature relationships have major effects on their ability to populate new sites. Earthworms lose moisture through their cuticles, so they are very dependent on soil moisture, and their activities are linked closely with rainfall patterns in both temperate and tropical environments. However, for some reason, in periods of intense precipitation, some species may emerge from their burrows, and they are often found in large numbers on the soil surface, where they may die. Cocoon production and the growth of earthworms are correlated positively with temperature, but the cocoon incubation period, percentage hatching, and number of hatchlings produced per cocoon are correlated negatively with temperature (Edwards 1998). Many species cannot survive below 0°C, and most species cannot survive above 30 to 35°C (Edwards 1983). Nevertheless, they have behavioral patterns and resistant cocoons that enable them to survive adverse climate conditions.

EARTHWORMS AND SOIL FERTILITY SOIL FORMATION Earthworms are extremely important in soil formation, principally through activities in consuming organic matter, fragmenting it, and mixing it intimately with soil mineral particles to form waterstable aggregates. During feeding, earthworms promote microbial activity by an order of magnitude, which in turn also accelerates the rates of breakdown and stabilization of humic fractions of organic matter. Different species of earthworms do not affect soil formation in the same way because of very different behavior patterns. Some species consume mainly inorganic fractions of soil, whereas others feed almost exclusively on decaying organic matter (see Chapters 8 and 9, this volume). They can deposit their feces as casts either on the soil surface or in their burrows, depending on the species concerned, but all earthworm species contribute in different degrees to the comminution

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and mixing of the organic and inorganic components of soil and decrease the size of not only organic particles, but also mineral particles (Shrickhande and Pathak 1951; Joshi and Kelkar 1952). During passage through the earthworm gut, the different kinds of mineral particles become mixed intimately with organic matter and form aggregates, which improve both the drainage and moisture-loading capacity of the soil. These aggregates are usually very water stable and improve many of the desirable characteristics of soils. There have been various suggestions as to the possible ways in which earthworms form aggregates, such as by production of gums (Swaby 1950) or calcium humate (Meyer 1943), by plant residues (Ponomareva 1953), or by means of polysaccharide molecules (Parle 1963). Various authors have estimated that up to 50% of the aggregates in the surface layers of soil are formed by earthworms (Kubiena 1953). Earthworms also contribute in many ways to soil formation, structure, and physical characteristics (see Chapters 10 and 11, this volume).

TURNOVER

OF

SOIL

As Darwin first noted, earthworms move large amounts of soil from the deeper strata to the surface. The amounts moved in this way range from 2 to 250 tons per hectare per annum, equivalent to bringing a layer of soil between 1 mm and 5 cm thick to the surface every year, creating a stonefree layer on the soil surface. In temperate climates, all the upper 15 cm of soil may be turned over every 10 to 20 years (Edwards and Bohlen 1996). However, much larger turnovers have been reported from tropical agroecosystems (Lavelle et al. 1999).

SOIL AERATION

AND

DRAINAGE

Earthworms also affect soil structure in other ways. Some species make permanent burrows, whereas others move randomly through the soil, leaving cracks and crevices of different sizes. Both sorts of burrows are important in maintaining soil aeration, drainage, and porosity. Moreover, earthworm burrows are usually lined with a protein-based mucus that helps stabilize these channels, and many of the species with permanent burrows cast their feces around the lining of the burrows, with the cast material usually containing more plant nutrients in a readily available form than the surrounding soil. There is good evidence that earthworm activity increases both the porosity and the air-to-soil volume (Wollny 1890; Hopp 1974; Edwards and Lofty 1977). Burrows are also important in improving soil drainage, particularly because those of some species, such as Lumbricus terrestris L., penetrate deep into soil in permanent burrows (Edwards and Lofty 1978, 1982a,b) and can even pass through layers of clay. The burrows and pores also increase the infiltration rate greatly (Slater and Hopp 1947; Teotia et al. 1950; Carter et al. 1982), and there are numerous reports of water penetrating the surface soil between two and ten times faster when earthworms were present than when they were not (Stockdill 1966; Wilkinson 1975; Tisdall 1978). These effects on infiltration can be of two kinds. The first is the presence of large surface-opening holes that are not usually taken into account by soil scientists when conventional models of infiltration are developed (Edwards and Lofty 1982a). Second, the crevices also created by earthworms, but which are smaller, not only increase infiltration, but also aid in water retention (see Chapters 10 and 11, this volume). Finally, earthworm activity makes a significant contribution to soil aeration (Stockli 1928; Kretzschmar 1978) by creating channels, particularly in heavy soils, that allow air to penetrate into deeper layers of soil, minimizing the incidence of anaerobic layers.

ORGANIC MATTER BREAKDOWN

AND INCORPORATION INTO

SOIL

Although all species of earthworms contribute to the breakdown of plant-derived organic matter, they differ greatly in the ways in which they break down organic matter and incorporate it into the soil. Their activities can be of three kinds, each associated with a different group of species. Some © 2004 by CRC Press LLC

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species are limited mainly to the plant-litter layer on the soil surface, decaying organic matter or wood, and seldom penetrate soil more than superficially. The main role of these species seems to be comminution of the organic matter into fine particles, which facilitates microbial activity. Other species live just below the soil surface most of the year, except when the weather is very cold or very dry; do not have permanent burrows; and ingest both organic matter and inorganic materials. These species produce organically enriched soil materials in the form of casts, which they deposit either randomly in the surface layers of soil or as distinct casts on the soil surface. Finally, there are the truly soil-inhabiting species with permanent burrows that penetrate deep into the soil. These species feed primarily on organic matter but also ingest considerable quantities of inorganic materials and mix these thoroughly through the soil profile. These last species are of primary importance in pedogenesis. All species depend on consuming organic matter in some form and play an important role in the final stages of organic matter decomposition, which is humification into complex amorphous colloids containing phenolic materials, probably by promoting microbial activity. There is little doubt that, in many ecosystems, earthworms are the key organisms in the breakdown of plant organic matter. Populations of earthworms usually expand in relation to the availability of organic matter; in many temperate and even tropical forests, it seems that earthworms have the capacity to consume the total annual litter fall. Such a total turnover has been calculated for an English mixed woodland (Satchell 1967), an English apple orchard (Raw 1962), a tropical forest in Nigeria (Madge 1965), and an oak forest in Japan (Sugi and Tanaka 1978); it seems likely that similar calculations would be valid for other sites (Edwards and Bohlen 1996). There is current speculation that invasions of lumbricids into North American forests are changing them dramatically and having an impact on rates of organic matter turnover and soil cover (see Chapters 5, 8, 9, and 13, this volume).

NUTRIENT AVAILABILITY During feeding by earthworms, the carbon:nitrogen ratio in the organic matter falls progressively; moreover, most of the nitrogen is converted into the ammonium or nitrate form. At the same time, the other nutrients, phosphorus and potassium, are converted into a form available to plants. Soils that have poor populations of earthworms often develop a structure with a mat of decomposed organic matter at the soil surface (Kubiena 1953); this can also occur in grassland and is common on poor upland grasslands in temperate countries and in New Zealand in areas where earthworms have not yet been introduced (Stockdill 1966) (see Chapters 6 and 14, this volume).

EFFECTS OF AGRICULTURE ON EARTHWORMS Earthworm populations are affected greatly by many of the main agricultural practices; in particular, cultivations, fertilizers, pesticides, and crop rotations exert major effects on earthworm activities and communities. Cultivations have considerable effects on earthworm communities, particularly those species with deep burrows. A single cultivation does not have any drastic effects on earthworm populations other than by mechanical damage, destruction of permanent burrows, and exposure to bird predators. However, repeated heavy cultivations progressively diminish earthworm populations. No till (direct drill) and a variety of conservation tillage practices, such as ridge tillage and shallow plow, favor the buildup of larger earthworm populations that are limited only by the availability of food (Edwards and Lofty 1982a; Edwards and Bohlen 1996). Fertilizers can be either organic or inorganic, including a broad range of organic manures from sources such as cattle, pigs, poultry, sewage wastes, and wastes from industries such as those involving a brewery, paper pulp, or frozen potatoes. These materials are major factors in the buildup of large field earthworm populations; when such organic wastes are added to agricultural land, earthworm populations may double or triple in a single season. Some liquid manures that have not © 2004 by CRC Press LLC

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aged or composted can have temporary adverse effects on earthworm populations when applied to soils as slurries because of their ammonia and salt contents, but these effects are usually short term. Many inorganic fertilizers also contribute indirectly to the buildup of earthworm populations because of increased crop yields and hence increased amounts of crop residues added to the soil. However, earthworms are very sensitive to ammonia, and ammonia-based fertilizers often have adverse effects on earthworm populations, especially when these fertilizers are applied annually over several seasons (Edwards and Lofty 1982b). Pesticides, which include insecticides, herbicides, fungicides, and nematicides, are used extensively on agricultural land in developed countries. It is often assumed that many pesticides are toxic to earthworms or have harmful effects on them. However, most herbicides have few direct effects on earthworms, although the triazine herbicides are slightly toxic. However, herbicides have drastic indirect effects on earthworms through their influence on the availability of organic matter (Edwards and Thompson 1973). Most fungicides have few effects on earthworms, with the exception of the carbamate-based fungicides, such as benomyl, which are very toxic. Of the insecticides in current use, only the organophosphate, phorate, and most carbamate-based compounds such as carbaryl, carbofuran, and methiocarb, and the avermectins are toxic to earthworms (Edwards 1984a, b). Of more than 200 pesticides reviewed by Edwards and Bohlen (1992), fewer than 20 were seriously toxic to earthworms (see Chapters 16 and 17, this volume). Crop rotations have been progressively decreasing in industrialized agriculture. There has been relatively little work on the effects of crop rotations on earthworm problems. In general, the inclusion of crops such as cereals that leave considerable organic residues encourage the buildup of earthworm populations more than do legumes, which decompose quite rapidly. Root crops, for which most of the crop is removed, discourage the buildup of earthworm populations (Edwards and Bohlen 1996).

EARTHWORMS AS INDICATORS OF SOIL QUALITY AND HEALTH There has been considerable interest in the concept of maintaining soil quality and health. There has been considerable discussion on defining these terms and on identifying appropriate physical, chemical, and biological indicators of soil quality. One definition is “the ability of a soil to sustain biological productivity, maintain environmental quality and promote plant, animal and human health” (Doran and Parkin 1996). Soil is a heterogeneous mixture of abiotic and living components, including a very complex range of soil-inhabiting organisms. The basic functions of soils depend on their structural and functional integrity and the impacts of disturbances on management on these functions. A wide range of indicators of soil quality and health criteria has been suggested, but it is becoming increasingly clear that it is essential that the indicators must include biological components because soil is a dynamic entity (Blair et al. 1996). It is difficult to use microbial indicators of soil quality and health as much as desired because of a lack of simple methodologics that can be used in the field by relatively untrained workers. Soil microinvertebrates have been suggested as possible indicators of quality and health (Linden et al. 1994), but sampling microarthropod or nematode populations is difficult, so their identification and utility as suitable indicators is a complex problem. There is a consensus among soil ecologists and most farmers that earthworms may be one of the best indicators available of soil quality (Doube and Schmidt 1997). They are easy to sample and identify and, as the discussions in this book illustrate, are important indicators of both soil health and soil quality (see Chapters 2 and 6, this volume).

EARTHWORMS AND SOIL POLLUTION There has been increasing interest in the use of earthworms as organisms to assess the environmental effects of soil pollution. Three Conferences on Earthworm Ecotoxicology (1991, U.K.; 1997, the

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Netherlands; 2001, Denmark) were each attended by more than 100 scientists and provide good evidence of this interest. Standardized testing protocols have been developed by such national and international organizations as the Organization for Economic Cooperation and Development and the European Union (Edwards 1983, 1984b). Many aspects of earthworm ecotoxicology are reviewed in Chapters 16 and 17 of this volume.

EARTHWORM IMMIGRATIONS Interest has increased greatly in migrations of earthworms across regions and continents. Peregrine earthworms, especially lumbricids, are invading soils across the world, particularly into agricultural soils, but more recently into forest soils. These issues are discussed extensively in Chapters 5 and 13 of this volume.

NEED FOR EARTHWORM RESEARCH Although the number of publications on earthworm biology and ecology is increasing rapidly, there still seems an urgent need for greatly expanded research, particularly on some aspects of earthworm activity. There still is inadequate knowledge of the basic biology and ecology of even some of the more common species of lumbricoids. Very few studies have addressed the problems of the detailed interrelationships among earthworms, microorganisms, and decaying organic matter and its incorporation into soil (see Chapters 2 and 12, this volume). There is good empirical evidence that introduction of earthworms together with organic matter into impoverished soil, with addition of organic matter and adjustment of pH, can increase soil fertility greatly, but there is little knowledge of the mechanism of such increases or even the best ways of introducing earthworms. Most important is the worldwide lack of knowledge of the geographic distribution of earthworms and populations of the different species. Until more is known of the fundamental biology and ecology and the activities of the many different species and their role in maintaining soil structure and fertility, it is impossible to assess their potential role in soil improvement. These problems are particularly acute in North America, where there are few earthworm specialists, and taxonomic research is extremely sparse.

CONCLUSIONS This second edition of Earthworm Ecology appears only 5 years after the first edition; it has been revised extensively, and four new chapters on important issues have been added. The reasons for creating a second edition so soon were partially because of rapid developments in earthworm biology and ecology and, to some extent, because of the great reception of the first edition by scientists and the public. It is hoped that this new edition will find a ready audience, and that it will encourage further interest in earthworms.

REFERENCES Barois, I., P. Lavelle, M. Brossand, L. Tondal, M. Martinez, J.P. Rossi, B.K. Senapati, A. Angeles, C. Fragoso, J.J. Jimienez, T. Decaens, C. Lattand, J. Kamyono, E. Blanchart, L. Chapius, G.E. Brown, and A. Monerno. 1999. Ecology of earthworms with large environmental tolerance and extended distribution, in Earthworm Management in Tropical Ecosystems, Lowell, P., L. Brussaard, and P. Hendrix, Eds., CABI Wallingford, Oxford, U.K., pp. 57–86.

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Blair, J.M., P.J. Bohlen, and D.W. Freckman. 1996. Soil invertebrates as indicators of soil quality, in Methods for Assessing Soil Quality, Doran, J.W. and Jones, A.J., Eds., Soil Science Society of America Special Publication 49, Madison, WI, pp. 273–291. Carter, A., J. Heinonen, and J. deVries. 1982. Earthworms and water movement, Pedobiologia, 23, 395–397. Darwin, C.R. 1881. The Formation of Vegetable Mould through the Action of Worms, with Observations on Their Habitats, Murray, London. Doran, J.W. and T.B. Parkin. 1996. Quantitative indicators of soil quality: a minimum data set, in Methods for Assessing Soil Quality, Soil Society of America Special Publication 49, Madison, WI, 25–38. Doube, B.M. and O. Schmidt. 1997. Can the abundance or activity of soil macrofauna be used to indicate the biological health of soils, in Biological Indicators of Soil Health, Pankhurst, C.E., B.M. Doube, and Gupta, Eds., USSR, CAB International, Wallingford, Oxford, U.K., pp. 265–296. Edwards, C.A. 1983. Development of a Standardized Laboratory Method Assessing the Toxicity of Chemical Substances to Earthworms, Report EUR 8714 EN, Environment and Quality of Life, Commission of the European Communities, Brussels, Belgium. Edwards, C.A. 1984a. Changes in agricultural practice and their impact upon soil organisms, in Proceedings of Symposium No. 13, The Impact of Agriculture on Wildlife, Agriculture and the Environment, Jenkins, D., Ed., N.E.R.C. U.K. pp. 46–65. Edwards, C.A. 1984b. Report of the Second Stage of a Standardized Laboratory Method Assessing the Toxicity of Chemical Substances to Earthworms, Report EUR 8714 EN, Environment and Quality of Life, Commission of the European Communities, Brussels, Belgium. Edwards, C.A. 1998. The use of earthworms in processing organic wastes into plant growth media and animal feed protein, in Earthworm Ecology, Edwards, C.A., Ed., CRC Press, Boca Raton, FL, pp. 327–354. Edwards, C.A. and P.J. Bohlen. 1992. The effects of toxic chemicals on earthworms, Rev. Environ. Contamination Toxicol., 125, 23–99. Edwards, C.A. and P.J. Bohlen. 1996. Earthworm Ecology and Biology, Chapman & Hall, London. Edwards, C.A. and J.R. Lofty. 1977. Biology of Earthworms, 2nd ed., Chapman & Hall, London. Edwards, C.A. and J.R. Lofty. 1978. The influence of arthropods and earthworms upon root growth of direct drilled cereals, J. Appl. Ecol., 15, 789–795. Edwards, C.A. and J.R. Lofty. 1982a. The effect of direct drilling and minimal cultivation on earthworm populations, J. Appl. Ecol., 19, 723–724. Edwards, C.A. and J.R. Lofty. 1982b. Nitrogenous fertilizers and earthworm populations in agricultural soils, Soil Biol. Biochem., 14, 515–521. Edwards, C.A. and A.R. Thompson. 1973. Pesticides and the soil fauna, Residue Rev., 45, 1–79. Edwards, W.M., R.R. Van der Ploeg, and W. Ehlers. 1979. A numerical study of noncapillary sized pores upon infiltration, J. Soil Sci. Soc. Am., 43, 851–856. Hopp, H. 1974. What Every Gardener Should Know About Earthworms, Garden Way Publishing, Charlotte, VT. Joshi, N.V. and B.V. Kelker. 1952. The role of earthworms in soil fertility, Indian J. Agric. Sci., 22, 189–196. Kretzchmar, A. 1978. Quantification ecologique des gaeeries de lombriciens. Techniques et premieres estimations, Pedobiologia, 18, 31–38. Kubiena, W.L. 1953. The Soils of Europe, Murray, London. Lavelle, P., L. Brussaard, and P. Hendrix. 1999. Earthworm Management in Tropical Agroecosystems, CABI Wallingford, Oxford, U.K. Lee, K.E. 1985. Earthworms: Their Ecology and Relationships with Soils and Land Use, Academic Press, Sydney, Australia. Linden, D.R., P.F. Hendrix, D.C. Coleman, and P.C.J. Van Vliet. 1994. Faunal Indicators of Soil Quality for a Sustainable Environment, Doran, J.W., D.C. Coleman, D.F. Bezolicek, and B.A. Stewart, Eds., Soil Science Society of America Special Publication 35, Madison, WI, pp. 91–10. Lungström, P.O. and Reinecke, A.J. 1969. Ecology and natural history of the microchaelid earthworms of South Africa 4. Studies on the influence of earthworms upon the soil and the parabiological question, Pedobiologia, 9(1–2), 152. Madge, D.S. 1965. Leaf fall and disappearance in a tropical forest, Pedobiologia, 5, 273–288. Meyer, L. 1943. Experimenteller Beiträge zu makrobiologischen Wirkungen auf Humus and Boden Bildung, Arch. Pflanzenerahrung Dungung Bodenkunde, 29, 119–140. Muller, P.E. 1878. Studier over Skovjord I. Om Bogemuld od Bogemor paa Sand og Ler, Tidsskrift Skogbruk, 3, 1–124.

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Parle, J.N. 1963. A microbiological study of earthworm casts, J. Gen. Microbiol., 31, 1–3. Ponomareva, S.I. 1953. The influence of the activity of earthworms on the creation of a stable structure in a sod-podzolised soil, Trudy Pochvenie Institut Dokuehaeve, 41, 304–318. Raw, F. 1962. Studies of earthworm populations in orchards. I. Leaf burial in apple orchards, Ann. Appl. Biol., 50, 389–404. Satchell, J.E. 1967. Lumbricidae, in Soil Biology, Burgess, A. and F. Raw, Eds., Academic Press, London, pp. 259–322. Shrickhande, J.E. and A.N. Pathak. 1951. A comparative study of the physico-chemical characters of the castings of different insects, Indian J. Agric. Sci., 21, 401–407. Slater, C.S. and H. Hopp. 1947. Relation of fall protection to earthworm populations and soil physical conditions, Proc. Soil Sci. Soc. Am., 12, 508–511. Stockdill, S.M.J. 1966. The effect of earthworms on pastures, Proc. N.Z. Ecol. Soc., 13, 68–75. Stockli, A. 1928. Studien über den Einfluss der Regenwurmer auf die Beschaffenheit des Bodens, Landwirtschaft Jahrbuch Schweiz, 42, I. Sugi, Y. and M. Tanaka. 1978. Number and biomass of earthworm populations, in Biological Production in a Warm Temperature Evergreen Oak Forest of Japan, Kira, T., Y. Ono, and T. Hosokawa, Eds., J.I.B.P. Synthesis 18, University of Tokyo Press, pp. 171–178. Swaby, R.J. 1950. The influence of earthworms on soil aggregation, J. Soil Sci., 1, 195–197. Teotia, S.P., F.L. Duley, and T.M. McCalla. 1950. Effect of stubble mulching on number and activity of earthworms, Neb. Agric. Exp. Stn. Bull., 165, 20. Tisdall, J.M. 1978. Ecology of earthworms in irrigated orchards, in Modification of Soil Structure, Emerson, W.W., R.R. Bond, and A.R. Dexter, Eds., Wiley, Chichester, U.K., pp. 297–303. Urquhart, A.T. 1887. On the work of earthworms in New Zealand, Trans. N.Z. Inst., 19, 119–123. Wilkinson, G.E. 1975. Effect of grass fallow rotations on the infiltration of water into a savanna zone soil of northern Nigeria, Trop. Agric. (Trinidad), 52, 97–103. Wollny, E. 1890. Untersuchungen über Beeinflussung der Fruchtbarkeit e der Ackerkrume durch die Tätigdeit der Regenwurmer, Forschungen Gebeit Agrik Physik Bodenkunde, 13, 381–395.

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Earthworms Affect Plant 2 How Growth: Burrowing into the Mechanisms George G. Brown Embrapa Soja, Londrina, Brazil

Clive A. Edwards Soil Ecology Laboratory, The Ohio State University, Columbus, OH, U.S.A.

Lijbert Brussaard Soil Quality Section, Wageningen University, Wageningen, The Netherlands

CONTENTS Effects of Earthworm on Plants: The History.................................................................................14 Earthworms and Plant Production in the Tropics ....................................................................15 The Mechanisms by Which Earthworms Affect Plant Growth: A Conceptual Background ...........................................................................................................17 Types and Modes of Interaction ...............................................................................................17 Spatial and Temporal Scales of Earthworm Action .................................................................17 Why Focus on Effects of Earthworms on Plant Roots? ..........................................................18 The Seven Main Mechanisms by Which Earthworms Affect Plants..............................................18 1. Dispersal and Changes in Populations and Activities of Beneficial Microorganisms........19 2. Changes in Populations and Impacts of Plant Pests, Parasites, and Pathogens..................23 Potential Role of Earthworms in the Reduction of Plant Disease and Pest Problems ..............................................................................................................23 Potential Role of Earthworms in Increasing Plant Disease or Pest Problems ..............25 3. Earthworms and Plant Growth-Regulating and Growth-Influencing Substances ...............25 4. Root Abrasion and Ingestion of Living Plant Parts by Earthworms...................................26 5. Interactions of Earthworms with Seeds ...............................................................................27 6. Changes in Soil Structure Caused by Earthworms..............................................................27 Earthworm Casts .............................................................................................................28 Earthworm Burrows ........................................................................................................30 7. Changes in Nutrient Spatiotemporal Availability Caused by Earthworms .........................31 Nutrients from Earthworms (Death, Excretion) .............................................................34 Crawling Forward: The Challenge of Identifying and Quantifying the Potential of Earthworms to Increase Plant Growth .........................................................................................34 “All-Minus-One” Tests and Field Trials...................................................................................35 The Earthworm Threshold Concept .........................................................................................36 Future Needs in Earthworm Research......................................................................................36

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Acknowledgments ............................................................................................................................37 References ........................................................................................................................................37

EFFECTS OF EARTHWORM ON PLANTS: THE HISTORY The importance of earthworms for soils, plant growth, and society has undergone various phases, from profound recognition to utter ignorance and disdain. They were highly regarded as promoters of soil fertility during the Egyptian empire (Minnich 1977), and early philosophers such as Aristotle considered them beneficial animals, calling them the “earth’s entrails” (or intestines) (Kevan 1985). From antiquity to Darwin’s time, however, not much information is available on earthworms (see review by Kevan 1985); throughout much of the 17th up to the beginning of the 20th century, earthworms were considered garden pests that needed elimination from soils (Minnich 1977; Brown et al. 2004). Probably the earliest and best-known report of the potential benefits of earthworms to soils is the much-quoted letter of Rev. Gilbert White to the Hon. Daines Barrington, written on May 20, 1777. This letter also provided some first hints of the mechanisms by which earthworms affect plant growth. White (1789) wrote: Dear Sir — … Earthworms, though in appearance a small and despicable link in the chain of Nature, yet, if lost, would make a lamentable chasm. For to say nothing of half the birds, and some quadrupeds which are almost entirely supported by them, worms seem to be the great promoters of vegetation, which would proceed but lamely without them, by boring, perforating, and loosening the soil, and rendering it pervious to rains and the fibers of plants, by drawing straws and stalks of leaves and twigs into it; and most of all, by throwing up such infinite numbers of lumps of earth called worm-casts, which, being their excrement, is a fine manure for grain and grass … Gardeners and farmers express their detestation of worms; the former because they render their walks unsightly, and make them much work; and the latter because, as they think, worms eat their green corn. But these men would find that the earth without worms would soon become cold, hard-bound, and void of fermentation, and consequently sterile; and, besides, in favour of worms, it should be hinted that green corn, plants, and flowers, are not so much injured by them as by many species of coleoptera (scarabs), and Tipulidae (long-legs) in their larva, or grub-state; and by unnoticed myriads of small and shell-less snails, called slugs, which silently and imperceptibly make amazing havoc in the field and garden.

It was not until almost a century later that Darwin (1881), in his book The Formation of Vegetable Mold Through the Action of Worms, firmly established the benefits of earthworms to soils. Other authors (Hensen 1877, 1882; Müller 1878, 1884; Wollny 1890) supported the positive role of earthworms in soil processes and plant growth, and Wollny (1890) was the first actually to quantify this relationship. Despite initial skepticism about the reports of Darwin and Hensen (Wollny 1882a), he became convinced that earthworms were important for plant production when his experiment showed increased yields of 12 species of plants, ranging from negligible amounts up to 733% (rape), by adding earthworms (Wollny 1890). However, he continued to warn about the generalization of these results to field situations. From the early 20th century to the present, the number of experiments increased, and the intervals between them decreased, so that there are presently more than 120 papers published on the effects of earthworms on plant production. The aim of most of these investigations was to answer the following questions: • • •

Do earthworms affect plant growth (positively or negatively), and if so, by how much? Which plants are affected most (positively or negatively)? Which earthworm species are most efficient at promoting plant growth?

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However, despite the abundant literature on the responses of plants to earthworms and the identification of a number of soil, environmental, or earthworm factors associated with particular plant responses, rarely has the question of how these effects occur (i.e., what the mechanisms behind the observed effects are) been addressed properly (Blakemore and Temple-Smith 1995; Edwards and Bohlen 1996; Brussaard 1999). In most papers, mechanisms were alluded to only briefly, and in several instances, the possible reasons for the observed effects of earthworms were not even mentioned. Furthermore, the proposed mechanism often cannot be confirmed or validated. The reason for this apparent lack of focus on the mechanisms behind the effects of earthworms on plants may be partly because of the following: •

•

• •

The predominant paradigms driving agricultural development from Liebig (1840) up to the “green revolution” period (ending in the 1970s), with research focusing mainly on alleviating physical and chemical constraints to plant production through the use of artificial inorganic inputs and improved (often hybrid) crop varieties (Sánchez 1994) Production (yield-oriented) research that has concentrated mainly on aboveground plant responses and rarely has studied changes in root growth, morphology, distribution, and the belowground interactions (e.g., of earthworms with microorganisms) Inadequate experimental designs or insufficient criteria on parameters measured to assess the possible mechanisms involved The very complex nature of indirect and direct biological interactions that occur in soils, particularly between earthworms, soil properties and processes, and other organisms in soils

EARTHWORMS

AND

PLANT PRODUCTION

IN THE

TROPICS

Many aspects of the effects and management of earthworms in tropical agroecosystems were reviewed by Lavelle et al. (1999). In particular, Brown et al. (1999) summarized the results of 28 experiments in the greenhouse and at the field level that identified the various soil properties and processes affected by earthworm activities and their impacts on plant production. The experiments were done in 8 tropical countries and involved at least 34 earthworm species and 19 plant species and were tested in 23 soil types belonging to 8 soil groups. An analysis of 246 studies of the effects of earthworms on plant shoot production (Figure 2.1) and 88 studies of the effects of earthworms on grain yields demonstrated clearly that earthworms usually have positive effects on plant growth (75% of all studies resulted in plant growth increases) and biomass. A mean 57% increase was observed in plant shoot mass, and a 36% increase was found for grain yields. Important negative effects occurred only rarely, usually because of some dysfunction in the soil created or induced by earthworm activities. They also observed that root production, contrary to that of the aboveground parts, was usually affected less by earthworm activity, possibly because of difficulties in studying this parameter or because plants growing in more healthy soils (presumably the case in earthwormworked soils) tend to invest more energy in growing the aboveground plant parts, producing fewer roots per unit shoot biomass, resulting in higher shoot:root ratios. The factors that seemed to affect the ultimate responses of plants to earthworms were the following: • •

•

The part of the plant harvested, with greater effects of earthworms on biomass (positive) of shoots than grains and with the smallest effects on root growth. The species of plant involved, with greater effects of earthworms on the shoot growth of perennial plants (trees and bushes) and larger effects on yields of gramineous grain crops compared with legumes. The species of earthworms involved, with the pantropical endogeic species Pontoscolex corethrurus producing the greatest yield increases and the widespread Indian Dichogastrini

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In 43% of cases, increase was >20%

100 to 200% (9%)

>300% (2%) –100 to – 60% 200 to 300% (0.4%) (5%) –60 to –20% (4%) –20 to 0% (20%)

60 to 100% (6%)

20 to 60% (21%) 0 to 20% (33%) In 75% of cases, earthworms affected plant biomass positively

Region of generally nonsignificant effects (with little short-term importance but with possible cumulative importance)

FIGURE 2.1 Effects of tropical earthworm species on plant shoot production. Each slice of the pie indicates a range of shoot biomass increase due to earthworms (e.g., 0 to 20%), and the percentage of cases where that range of increase was observed (values in parentheses). The chart was built using 246 data points (cases) taken from a total of 28 experiments involving at least 34 earthworm and 19 plant species tested in 23 soil types belonging to 8 great groups. (Modified from Brown et al., 1999.)

species Drawida willsii and West African species Millsonia anomala and various small eudrilids all having a good potential for introduction and management into soils. • The earthworm biomass introduced or present in the soil, with higher yields usually occurring in response to greater earthworm biomass in a curvilinear relationship (moderate yield increases of 20 to 40% occurred with earthworm biomass values above 17 g m−2 and over 40% of grain production increases occurred with earthworm biomasses above 32 g m−2). • Εarthworm survival. In both pot and field trials, the mortality of introduced earthworms was often high, particularly when the species was not adapted properly to the soil used or when few or no organic residues were applied (survival was greater when organic residues were present). • The presence of organic residues on the soil surface, with greater effects on plant yields when such residues were present. • The timescale of the measurements (i.e., the duration of the experiment), usually with positive cumulative increases in plant biomass because of earthworm activities with time, although occasionally (depending on the soil type or earthworm and plant species) the cumulative effects on plant biomass observed were negative. • Τhe spatial scale of the experiment (i.e., pot vs. field experiments), with effects on yield usually greater at the pot scale for any given plant and earthworm combination. • The natural richness of the soil used in the experiment, with greater benefits on productivity in poorer soils (low percentage carbon content, coarser textures) than in richer soils (more carbon, clayey texture) with earthworms producing higher yields in moderately acid soils (pH between 5.6 and 7.0) than in strongly acid soils (pH < 5.6) or alkaline soils (pH > 7.0). Different combinations of earthworm species, soil types and conditions, plant species, and various imposed human or environmental constraints may alter the potential effects of the earthworms on soil properties and plant growth. Thus, pinpointing the exact reasons for mechanisms

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of a specific plant response to earthworms in an experiment is not easy; more often than not, several mechanisms rather than a single mechanism are probably operating simultaneously. The main aim of this chapter is to seek and identify possible mechanisms by which earthworms can promote or suppress plant growth. Furthermore, we dig further into some of these mechanisms and provide both conceptual diagrams of how they may be functioning and a few case studies dealing with each of the seven main mechanisms we propose. Finally, we end with some suggestions on how advancement will occur in this biologically complex area of research. Our basic premise is that through better understanding of the ways by which earthworms affect plant growth and production, plant and soil management techniques and practices can be adapted, improved, or implemented to prevent the occurrence of negative effects of earthworms on soils and plants and to maximize their positive effects on crops for the benefit of farmers, gardeners, ranchers, foresters, and other land users.

THE MECHANISMS BY WHICH EARTHWORMS AFFECT PLANT GROWTH: A CONCEPTUAL BACKGROUND TYPES

AND

MODES

OF INTERACTION

The effects of earthworms on soils can take three main forms: effects on biological, physical, or chemical soil properties and processes. Furthermore, because earthworms share the soil environment with roots, their effects on plant growth and root development can be either direct or indirect. Thus, the mechanisms of how earthworms influence plant productivity can be divided into three main types: physical, chemical, and biological. These can operate either directly or indirectly. Indirect effects mean that the plant is affected by earthworm activities through changes in the physical, chemical, or biological soil or rooting environment produced by earthworms; the direct mode of action means that the earthworms or their activities lead to direct changes in root growth and productivity.

SPATIAL

AND

TEMPORAL SCALES

OF

EARTHWORM ACTION

The soil volume affected by earthworm activities has been termed the drilosphere (Lavelle 1988); it constitutes one of the main soil functional domains (Beare et al. 1995; Lavelle 2002) that have significance in regulating major soil processes and functions, such as structure, organic matter (OM) decomposition, nutrient cycling, microbial and invertebrate populations, and plant growth. Because earthworm burrows and casts may outlive the earthworms themselves, and regulate the soil as an environment for other organisms (including plant roots) by controlling its physical structure, nutrient fluxes, and energetic status (resource availability), they have been termed ecosystem engineers (Jones et al. 1994; Lavelle et al. 1997). It is important to note that the drilosphere and the engineering effects of earthworms are very variable and depend on biological factors such as the type of vegetation and the characteristics and composition of the earthworm community at a particular location (species, abundance, biomass, age structure, ecological strategy) and abiotic regulating factors, including climate, soil type, and imposed anthropic (management) factors. Furthermore, the earthworm drilosphere is a dynamic zone of action that is constantly changing in both space and time as the earthworms ingest and reingest soil, burrow, and cast at different rates and in different locations in the soil. Therefore, the drilosphere can affect soil functions (including plant productivity) at different spatiotemporal scales, manifesting its effects at levels that range from the earthworm gut up to the soil profile (Lavelle 1997); these ideas are explored in Chapter 12. The effects of earthworms on plants in a given situation and the mechanisms involved are difficult to assess because, although earthworms and their structures (burrows, casts) are often easily identifiable or separable from the edaphosphere and their sphere of influence on the soil © 2004 by CRC Press LLC

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(drilosphere) can be measured and quantified physically, chemically, and biologically under controlled conditions, the drilosphere is connected with the rest of the soil system. This means that it can interact profoundly with other soil organisms and functional domains (e.g., rhizosphere, porosphere, aggregatusphere, detritusphere, mermycosphere, termitosphere) (Brown et al. 2000). This interconnectedness becomes even more evident as in attempts to separate the mechanisms responsible for plant responses to earthworms in any given situation, soil type, or area.

WHY FOCUS

ON

EFFECTS

OF

EARTHWORMS

ON

PLANT ROOTS?

Roots, as sensitive sensors of the soil environment and the producers of many signals that ultimately control plant shoot growth (Aiken and Smucker 1996), are the primary and immediate receivers of the contributions of earthworms to soil functions. By controlling nutrient and water supply to the shoots, it is the biomass, density, distribution, and activity (growth rate and longevity) of roots within the soil profile that will largely determine plant productivity (Brown and Scott 1984). Thus, it is the response of roots to earthworm activity that usually controls the overall plant response. A simple conceptual model connecting the physical, chemical, and biological effects of earthworms on soils with their potential effects on plant root or shoot growth and nutrition is provided in Figure 2.2. The interdependence of earthworm physical activities (production of casts and burrows) and earthworm physiological activities (excretions, secretions, and tissue death) in interactions with soil properties such as organic matter (soil OM, root and residue inputs), microbial populations, and plant production is evident. The effects of chemical substances on soil properties and processes are based on the selection by earthworms of particular soil particles and organic matter, the different nutrient compositions of their feces compared with uningested soil, cutaneous mucus secretion, and excretion of metabolic products. Biological effects on soils are caused primarily by interactions of earthworms with the rhizosphere and soil microorganisms, depending especially on feeding and digestive habits of the earthworms; the physical effects are associated mainly with the structural properties of the drilosphere. The following sections in this chapter explore the various ways in which earthworms can directly and indirectly affect plant growth, and we propose seven main mechanisms by which this is achieved. The focus is mainly on roots, although we recognize that indirect interactions with the aboveground plant parts and other organisms (both above- and belowground) may also be important (Wurst and Jones 2003). Given that the latter subject is a very recent field of study and that few results are available, we will limit the discussion primarily to belowground interactions and processes.

THE SEVEN MAIN MECHANISMS BY WHICH EARTHWORMS AFFECT PLANTS We define the seven main mechanisms by which earthworms affect plant growth as follows (see details in Table 2.1): 1. 2. 3. 4. 5. 6. 7.

Dispersal and changes in populations and activity of beneficial microorganisms Changes in populations and impacts of plant pests, parasites, and pathogens Production of plant growth-regulating (PGR) and plant growth-influencing (PGI) substances Root abrasion and ingestion of living plant parts by earthworms Interactions of earthworms with seeds Changes in soil structure caused by earthworms Changes in nutrient spatiotemporal availability caused by earthworms

Mechanisms 1 to 5 are mainly biological, operating indirectly (1 and 2) or directly (3 to 5); the last two (6 and 7) are indirect physical (6) or chemical (7) mechanisms. © 2004 by CRC Press LLC

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PLANT

OTHER ORGANIC MATTER SOURCES

Nutrient absorption

ORGANIC MATTER

Rhizodeposition

Excretions, secretions, dead tissue

MICROBIAL POPULATIONS and ACTIVITY H2O and air circulation

EARTHWORMS CASTINGS

BURROWS Better root penetration SOIL FIGURE 2.2 Simplified conceptual model connecting the physical, chemical, and biological earthworm effects on soils with their potential effects on plant growth and nutrition. (Modified from Cuendet and Bieri, 1999; Syers and Springett, 1983.)

1. DISPERSAL AND CHANGES MICROORGANISMS

IN

POPULATIONS

AND

ACTIVITIES

OF

BENEFICIAL

Large populations of beneficial (plant growth promoting [PGP]) microorganisms such as saphrophytic and mycorrhizal fungi, actinomycetes (e.g., Frankia), bacteria, and microinvertebrates, such as protozoa and microbivore (fungivorous, bacteriophagous, predatory omnivorous and entomophathogenic) nematodes inhabit the soil. Nevertheless, because of their limited ability to disperse within the soil and the soil environmental and nutritional limitations to their activities, a large proportion of soil microorganisms are inactive at any given time, waiting for suitable conditions to promote higher levels of activity (Lavelle 1997). Invertebrate activities, such as earthworm burrowing and casting, promote soil mixing and bring microorganisms into contact with inaccessible soil resources, stimulating both their populations and their activity. The earthworm gut also provides an ideal environment for enhanced activity levels or multiplication of some microorganisms; others may be digested or their activity levels reduced by passage through the earthworm gut (Brown et al. 2000). The complex resulting effects of earthworms on microbial communities in soils (activity, populations, diversity) depend on the reactions of microorganisms to passage through the earthworm gut and the ability of microorganisms to utilize the drilosphere. Thus, earthworms may affect microbial populations (beneficial, facultatively pathogenic, and adverse species) directly, by feeding and digestive processes or indirectly by burrowing and casting activities, which change root growth and development and the soil environment, thereby making it more or less favorable to the development of microorganisms (Figure 2.3). Furthermore, as earthworms move through the soil matrix, they may disperse microorganisms, both superficially (on the earthworm body) or via ingestion-egestion (in casts). The ability of earthworms to disperse microorganisms or stimulate microbial activity and increase microbial populations depends greatly on the earthworm’s spatial range of activity, food requirements and sources, and behavior. Epigeic, litter feeding, and dwelling species of earthworms are much more © 2004 by CRC Press LLC

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TABLE 2.1 The Seven Main Mechanisms by Which Earthworms Affect Plant (Mostly Root) Growth either Directly or Indirectly through Physicochemical or Biological Changes to the Soil Environment Mechanism Category (Type) Mechanism Mode Indirect (mediated through changes in the rooting environment, or via interactions with organisms that affect root growth and production)

Direct (earthworm activities that influence root growth/production in a direct manner)

Biological

Physical

Chemical

1. Dispersal or changes in populations 6. Changes in soil 7. Changes in nutrient and activity of beneficial structure caused by spatiotemperal microorganisms (plant growth earthworms (pore and availability caused by promoting rhizobacteria, N2 fixing root aggregate size distribution earthworms (release or symbionts, saprophytic and mycorrhizal and associated processes, immobilization of fungi, microbial biocontrol agents, including aeration, water different plant nutrients, microbivorous and entomopathogenic retention, hydraulic leaching, denitrification, nematodes, protozoa) conductivity, infiltration, volatilization, OM 2. Effects of earthworms on populations erosion, runnoff, aggregate mineralization, protection and crust formation and and/or humification, of plant pests, parasites, and breakdown, chelation of metals, pH pathogens (increase or decrease in compaction/soil slumping changes) populations and incidence of plantand decompaction/soil parasitic nematodes, phytopathogenic fungi and bacteria, plant viruses?, shoot- loosening) and root-feeding insects) 3. Production of plant growth promoting/regulating substances (hormones, vitamins, humic matter, auxins, cytokinins, gibberellins, ethylene, microbially induced and/or excreted by earthworms. 4. Root abbrasion and ingestion of living plant parts by earthworms (feeding and/or ingestion by earthworms of living roots or plant shoots, and direct damage to growing roots) 5. Interactions between earthworms and seeds (ingestion, digestion, burial, dispersal, changes in germination rates and potential)

likely to affect microorganisms in the litter layer and the roots growing through the organic matter/humus (O/H) horizons and the soil surface-litter interface compared with the endogeic (soil-dwelling) geophagous (soil-feeding) earthworm species, which tend to have a greater effect on microorganisms living within the soil. Anecic, litter-burying species of earthworms, which create deep vertical burrows and surface middens (small mounds of leaves blocking the entrance of vertically oriented burrows connected to the soil surface) can have a major influence on microorganisms (fungi, bacteria, actinomycetes) and micro-, meso-, and macroinvertebrates (protozoa, nematodes, mites, springtails, enchytraeids, millipedes, isopods, other earthworms) in surface litter communities (Brown 1995; Anderson and Bohlen 1998; Maraun et al. 1999). However, their effects on the microbial communities living within the soil are probably less than those of endogeic species because of their decreased soil-burrowing activities as they tend to build more permanent burrow systems. Nevertheless, anecic earthworm species (and large endogeic species) often have burrows that reach depths of more than 2 m, which can represent important pathways of microbial dispersal and hot spots of microbial and root growth activity compared with that in the surrounding soil matrix (Bhatnagar 1975; Ehlers et al. 1983). © 2004 by CRC Press LLC

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Beneficial plant-growth promoting (PGP), facultative, or obligatory pathogenic rhizosphere microorganisms Earthworm feeding Time

Root growth and development

Soil/rooting environment

Earthworm burrowing and casting

FIGURE 2.3 Interactions among beneficial, facultative and obligatory plant-pathogenic rhizosphere microorganisms, earthworms, plant roots and the abiotic root environment, determining plant root growth and development (note: this is a modified version of the classic “plant disease triangle” taught in plant pathology).

It has often been suggested that earthworms tend to promote changes in the microbial community toward a bacterial-based trophic chain. Actually, phospholipid fatty acid (PLFA) methyl esters analyses of earthworm-worked soils indicated that Gram-negative bacteria seem to be favored compared with Gram-positive bacteria (Clapperton et al. 2001; Enami et al. 2001). Lumbricid earthworms also increase bacterial-to-fungal ratios (Clapperton et al. 2001), although when a plantpathogenic fungus was inoculated into the soils, earthworms decreased this ratio, implying that they may also increase the soil fungal biomass. Nevertheless, several species of fungi have been shown to be ingested preferentially by the earthworm Lumbricus terrestris (Moody et al. 1995; Cooke 1983; Moody et al. 1995; Bonkowski et al. 2000), and Edwards and Fletcher (1988) reported that fungi were a major food source for earthworms. This implies that earthworms (particularly the litter-burying or fragmenting anecic and epigeic species) may impose some selection pressures on fungal populations in both litter and soils. Bacterial-to-fungal ratios in soils are also often greater in earthworm-worked soils because bioturbation tends to affect fungal populations negatively more than those of bacteria (Hendrix et al. 1986). The rhizosphere, a less-than-0.5-mm soil layer surrounding plant roots, is rich in microorganisms, with species that are beneficial or adverse to root growth. Several earthworm species (especially some endogeics) seem to feed mainly in the rhizosphere (James and Seastedt 1986; Rovira et al. 1987; Robertson et al. 1994; Hirth et al. 1998). Activity of lumbricid earthworms has been reported in the rhizosphere of a temperate pasture (Carpenter 1985) and of wheat (Doube and Brown 1998), and feeding in the rhizosphere was inferred from radio- (14C) or stable isotope (15N, 13C) analyses of the tissues of earthworms (L. terrestris and P. corethrurus) living in soils under various plants (wheat, maize, Brachiaria decumbens, and sugarcane) (Spain et al. 1990; Spain and Le Feuvre 1997; Cortez and Bouché 1992; Brown 1999). There are also records of earthworms feeding on living and dead root tissues (see mechanism 4), but the role of root tissues and their derivatives (rhizodeposition) in earthworm diets remains little understood (Brown et al. 2000). Earthworm feeding or movement in or around the rhizosphere can have important consequences for associated microbial and faunal communities (activity, populations, diversity) and thus, indirectly, on plant productivity (Figure 2.3).

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T3

T2

T1

700 mg (10 earthworms) 350 mg (5 earthworms) 0 mg (no earthworms) 31% MICORRIZAE

30% MICORRIZAE

6.6% MICORRIZAE

FIGURE 2.4 Stimulation of Eugenia stipitata (arazá) growth and root mycorrhizal colonization 120 days after inoculating tree nursery bags (filled with 2 parts soil and 1 part composted sawdust) with five (0.35 g total wet weight) or ten (0.7 g) individuals of the pantropical geophagous endogeic earthworm species P. corethrurus. (Ydrago 1994; Photograph P. Lavelle.)

Dispersal of mycorrhizal propagules (hyphae, infected root fragments, spores) has been reported by various authors (McIlveen and Cole 1976; Rabatin and Stinner 1988; Ponge 1991; Reddell and Spain 1991a; Gange 1993; Lee et al. 1996; Cavenden et al. 2003), and although some hyphae and spores may be digested, many are still infective after passage through the earthworm gut (Reddell and Spain 1991a; Gange 1993). Mycorrhizal dispersal and deposition of earthworm casts in the rhizosphere may benefit root colonization by fungi, aid plant establishment in early successional stages, and contribute to the heterogeneous nature of mycorrhizal distribution in soil communities (Gange 1993). For example, the pantropical geophagous endogeic earthworm species P. corethrurus increased colonization of roots by arbuscular mycorrhizae in various tropical tree seedlings (Ydrogo et al. 1994; Figure 2.4) and a pasture grass (Brown et al. 2000), also increasing plant biomass on several occasions. The actinomycete Frankia and ectomycorrhizae were also shown to be dispersed by P. corethrurus (Reddell and Spain 1991b; Reddell et al. unpublished), although the effects of this on plant productivity are little known. Nevertheless, soil bioturbation and feeding in the rhizosphere by earthworms may break up extramatrical hyphae and the hartig net, thereby reducing root colonization by these root symbionts, hence providing potential benefits to the plants (Pattinson et al. 1997; Brown et al. 2000; Tuffen et al. 2002). Plant growth-promoting rhizobacteria (PGPR) such as Enterobacter cloacae, Azotobacter, Azospirillum, Acinetobacter, Bacillus, and Pseudomonas spp. may also be dispersed and their populations or activity increased in the drilosphere (Bhat et al. 1960; Kozlovskaya and Zdhannikhova 1961; Kozlovskaya and Zaguralskaya 1966; Bhatnagar 1975; Loquet et al. 1977; Hand and Hayes 1983; Savalgi and Savalgi 1991; Pederson and Hendriksen 1993). The metabolites released by these microorganisms may be particularly important to the potential plant responses (mechanism 3). Dispersal of these and other microorganisms such as biocontrol bacteria (e.g., Pseudomonas corrugata) and fungi (e.g., Gliocladium virens, Trichoderma harzianum) that colonize the rhizosphere and prevent root diseases needs further investigation. The dispersal of various symbiotic N2-fixing rhizobacteria that nodulate legume roots (e.g., Rhizobium trifolii in clover; Doube et al. 1994a) also needs further research (Stephens et al. 1994e; Stephens and Davoren 1994; Singer et al. 1999). These microorganisms all have an inability to spread actively and rapidly through the

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soil and colonize plant roots extensively, so earthworms may act as important dispersal vectors for them (Rouelle 1983; Stephens et al. 1993b; Doube et al. 1994b). Populations and activity of several groups of beneficial soil organisms important in plant litter decomposition and nutrient mineralization processes in soils (e.g., microfauna, mesofauna, and macroinvertebrates) may be affected by earthworms (Brown 1995). For instance, protozoa may be part of earthworm nutrition (Miles 1963; Flack and Hartenstein 1984; Bonkowski and Schaefer 1997), but many protozoan cysts can survive passage through the earthworm gut and can hatch, become more active, and reproduce rapidly in earthworm casts and earthworm-worked soils (Shaw and Pawluk 1986; Barois 1987; Winding et al. 1997; Binet et al. 1998). Earthworm casts may benefit bacteriophagic nematode populations preferentially over those of other nematode trophic groups (Roessner 1981, 1986; Senapati 1992), but the total numbers of freeliving nematodes in earthworm-worked soils may be reduced (e.g., Alphei et al. 1996; Dominguez et al. 2003) or increased (Winding et al. 1997), depending on the situation. Populations of other organisms, such as enchytraeids and various micro- and macroarthropods, may also be increased (e.g., in anecic earthworm middens) or decreased because of changes in microbial populations and food resources in earthworm-worked soils (Brown 1995). However, most of the consequences to plant growth of changes in the populations and activity of micro and macroinvertebrates in earthworm middens, castings, and earthworm-worked soils are unknown and deserve much more attention.

2. CHANGES PATHOGENS

IN

POPULATIONS

AND IMPACTS OF

PLANT PESTS, PARASITES,

AND

As with beneficial microorganisms, earthworm feeding, burrowing, casting, and dispersing activities can alter the distribution of populations of plant pathogens such as viruses, bacteria, fungi, parasitic nematodes, or insect pests in soils. Furthermore, by making plants more or less susceptible to these pests, parasites, and pathogens, earthworms can affect root health (Brown 1995). These relationships are illustrated in a modified version of the classic “plant-disease triangle” (Figure 2.3) in which plant root growth and development are shown as a function of the interactions between a favorable environment for both roots and pathogens and the presence or activity of “virulent” of “infective” plant pathogens. The result of these interactions (i.e., plant health status) may therefore be influenced directly or indirectly by earthworm activities. Earthworms are known to transport and consume a wide variety of plant pathogenic fungi and bacteria and plant-parasitic nematodes (Brown 1995). If populations of these organisms are reduced either directly by transit through the earthworm gut or indirectly via changes in the soil environment, then the indirect consequences to plant growth may be important, particularly when disease or nematode pressure is reducing crop yields. The role of earthworms as vectors of plant diseases, parasites, and pests depends on the type of organism and species ingested, the amount of soil and inoculum ingested, the extent of beneficial or antagonistic intestinal secretions, the number of organisms digested in the earthworm gut, the amount of organisms deposited in casts, the infectivity of surviving organisms deposited in casts, the feeding and casting behavior of the earthworms (dependent on the earthworm species and ecological category), and the mobility and behavior of the earthworm. Potential Role of Earthworms in the Reduction of Plant Disease and Pest Problems Several reports of beneficial results to plants of earthworm-induced reductions of plant pathogens are known. For instance, work in the Soil Ecology Laboratory at The Ohio State University has shown that vermicomposts can suppress plant diseases such as Pythium and Rhizoctonia (Chaoui et al. 2002, 2003) in the greenhouse and Verticillium in the field. When cabbages were grown in the presence of the earthworm Pheretima hilgendorfi, Nakamura et al. (1995) observed lower incidence of club-root disease (Plasmodiophora brassicae) damage in the seedlings. They attributed this decrease to the © 2004 by CRC Press LLC

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physical, chemical, and biological changes in the soil environment because of earthworm activity, reinforced by possible consumption of the pathogens by the earthworms (Nakamura 1996). In Australia, several complementary studies (Stephens et al. 1994a,b,c,e,f,g; Stephens and Davoren 1997) reported that the earthworms Aporrectodea rosea and Aporrectodea trapezoides could increase yields of wheat, ryegrass, and subterranean clover under greenhouse and field conditions by decreasing the incidence of Rhizoctonia solani (bare patch disease). Furthermore, wheat yields were also increased by these earthworms through a reduction in incidence of Gauemannomyces graminis var. tritici (takeall disease); A. trapezoides appeared to be more effective in disease suppression, probably because of higher feeding and casting levels compared with A. rosea. Although the exact mechanisms by which earthworms influence root diseases (including takeall) remain unknown, Clapperton et al. (2001) suggested that they are most probably mediated through changes in the soil microbial community, possibly via stimulation of biocontrol agents, antagonists, or microbial competition with the pathogens. Various other indirect mechanisms have also been proposed, such as acceleration of residue decomposition, burial of infected litter, increased soil porosity, and greater availability of plant nutrients in earthworm-worked soils. For instance, in various fruit-tree orchards, the burial of 12 fungal pathogens overwintering in the surface leaf litter (including Venturia inaequalis, the causal agent of apple scab) by the anecic earthworm species L. terrestris (Raw 1962) reduced their survival and ability to disperse, colonize, and infect the apple trees the following spring (Hirst and Stedman 1962; Niklas and Kennel 1981; Laing et al. 1986; Kennel 1990). Decreases in plant parasitic nematode populations by earthworm activity have also been documented for various (tropical and temperate) earthworm and nematode species combinations ( Dash et al. 1980; Roessner 1981, 1986; Senapati 1992; Boyer 1998). For instance, Boyer (1998) observed a reduction of Pratylenchus zeae populations in small pots (200 g soil) sown with rice and containing the earthworm species P. corethrurus. However, the effects of the earthworms on plant shoot and root growth was negative. Conversely, Boyer et al. (1999) observed significantly greater maize productivity and decreased Pratylenchus vulnus populations on maize roots when maize was undersown with the legume birdsfoot trefoil, and earthworms (Amynthas corticis) were introduced into the field. Yeates (1980, 1981) also reported greater plant productivity and lower populations of nematodes, including some plant parasitic species in pastures inoculated with lumbricid earthworms in New Zealand. Reduction of plant parasitic nematode populations in the field have also been observed after application of vermicomposts (Arancon et al. 2002, 2004a,b). Earthworm-induced decreases in nematode populations may be caused by direct ingestion and digestion of nematodes (Dash et al. 1980; Boyer 1998; Dominguez et al. 2003) or the release of fluids (enzymes, etc.), which affect the fertility, viability, and germination of cysts present in earthworm-worked soils and casts (Ellenby 1945; Roessner 1981; Boyer 1998), or they may be caused indirectly through modifications by earthworms of soil structure, water regimes, and nutrient cycling processes (Yeates 1981). Edwards and Fletcher (1988) and Manku (1980) have also suggested that earthworms may spread nematode-trapping fungi and nematode cyst pathogens of major importance in controlling nematode populations. Nematodes that pass unharmed through the earthworm gut or are able to take advantage of or adapt to earthworm-induced changes in soil properties and processes may be dispersed by earthworms. In the case of plant parasitic species, this could lead to potential problems, but for entomopathogenic nematodes commonly used in insect pest biocontrol, this may be beneficial (Shapiro et al. 1993). Several studies have demonstrated the potential effects of earthworms in reducing plant-pest incidence and damage. Boyer et al. (1999) reported fewer maize plants infested with the stalk borer Sesamia calamistis when the earthworm A. corticis was inoculated into field soils. The percentages of fertile maize plants infested by the borer were 75% without earthworms and 55% in soils with earthworms, although the total aboveground biomass of the two treatments did not differ significantly. In another study, L. terrestris was shown to reduce the numbers of leaf miners (Phyllonorycta blancardella) and leaf suckers (Psylla piri) overwintering in leaf litter of fruit-tree orchards by © 2004 by CRC Press LLC

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promoting leaf burial and decomposition (Laing et al. 1986; Kennel 1990), thus reducing their potential to damage the orchard trees. However, leaf-litter burial also reduces populations of the natural biocontrol agents (brachonid wasps) of these insects (Laing et al. 1986). Potential Role of Earthworms in Increasing Plant Disease or Pest Problems Several species of plant pathogenic fungi have been found in earthworm casts (Hutchinson and Kamel 1956; Hoffmann and Purdy 1964; Thornton 1970; Melouk and Horner 1976; Toyota and Kimura 1994), and plant parasitic nematodes may survive passage through the earthworm gut (Ellenby 1945; Russom et al. 1993). However, there are relatively few data available on the potential negative effects that earthworm-induced microbial dispersal may have on incidence of plant diseases (of fungal or bacterial origin) or nematode damage. Increased dispersal of a plant pathogenic fungus, Syntrichium endobioticum, the causal agent of wart disease of potato, by L. terrestris and various other (probably lumbricid) earthworms was reported by Hampson and Coombs (1989), resulting in increased infection of several potato plants. Similarly, Melouk and Horner (1976) reported infection of mint seedlings by verticillium wilt (Verticillium dahliae) when the plants were grown with earthworm casts that contained viable spores of these pathogens. Dispersal of plant parasitic nematodes by earthworms was reported by Ellenby (1945) and Russom et al. (1993), but the potential of this for increased damage to plant roots was not evaluated. Casts of the Nigerian earthworm species Agrotoreutus nyongii had larger and more diverse populations of parasitic nematodes than did the surrounding soil (Russom et al. 1993). Casts of Aporrectodea longa contained nematode cysts with greater fertility, viability, and germination potential than those in surrounding soil (Ellenby 1945). Ilieva-Makulec and Makulec (2002) reported an increase in plant parasitic nematode populations in soil cores inoculated with Lumbricus rubellus after 60 and 90 days, but no negative effects on growth of grass roots were observed. The interactions between earthworms and plant insect pests still remain poorly explored. Kirk (1981) reported large numbers of the northern maize rootworm (Diabrotica: Coleoptera) eggs in earthworm burrows and suggested that this may contribute to the spottiness of rootworm distribution and damage often observed in maize fields. More recently, Wurst and Jones (2003) and Scheu et al. (1999) showed effects of lumbricid earthworms (Aporrectodea sp.) on increased numbers of leaf sap sucking aphids (Myzus persicae) and their offspring.

3. EARTHWORMS SUBSTANCES

AND

PLANT GROWTH-REGULATING

AND

GROWTH-INFLUENCING

The first suggestion that earthworms might produce plant growth regulators (PGRs) was by Gavrilov (1963). This was supported by the first report of the presence of PGR substances in the tissues of Aporrectodea caliginosa, L. rubellus, and Eisenia fetida by Nielson (1965), who extracted indole substances from earthworms and reported increases in the growth of peas because of them. He also extracted a substance that stimulated plant growth from A. longa, L. terrestris, and Dendrobaena rubidus, but his experiments did not exclude the possibility that the PGR substances he obtained came from microorganisms living in the earthworm guts and tissues. The presence of PGR substances in the tissues of A. caliginosa, L. rubellus, and E. fetida was confirmed by Nielson (1965), who isolated indole substances from whole earthworm tissues. This was confirmed for A. rosea and A. caliginosa by Nardi et al. (1988). More recently, El Harti et al. (2001a,b) isolated indole acetic acid (IAA)-like substances from gross extracts of tissues and feces of L. terrestris. These substances stimulated rhizogenesis and enhanced root growth of Phaseolus vulgaris (common beans) in a manner very similar to that of IAA. Graff and Makeschin (1980) tested the effects of substances produced by L. terrestris, A. caliginosa, and E. fetida on the dry matter production of ryegrass. They added liquid eluates from pots

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containing earthworms to pots containing no earthworms and concluded that PGI substances were released into the soil by all three species, but the authors did not speculate further on the nature of these substances. Earthworms may liberate PGRs or PGIs themselves (Atlavinyte and Daciulyte 1969; El Harti et al. 2001a,b), or their production may be mediated by interactions with microorganisms in the drilosphere in a process that is not fully understood. It is clear that microorganisms are capable of producing PGR and PGI substances such as hormones, auxins, gibberellins, cytokinins, ethylene, and abscisic acid (Arshad and Frankenberger 1993; Frankenberger and Arshad 1995). Many microorganisms commonly found in the rhizosphere can produce PGR substances. Krishnamoorthy and Vajranabhaiah (1986) showed, in field experiments involving large earthworm populations, that seven species of earthworms could promote the production of cytokinins and auxins in soils. They also demonstrated significant positive correlations (r = 0.97) between earthworm populations and the levels of cytokinins and auxins present in ten different field soils and concluded that earthworm activity was linked strongly with PGR production. They reported that auxins and cytokinins produced through earthworm activity could persist in soils for up to 10 weeks although degraded in a few days if exposed to sunlight. For a more in-depth discussion of the role of earthworms in producing PGR substances through promoting populations and activity of microorganisms, see Chapter 18 this volume.

4. ROOT ABRASION

AND INGESTION OF

LIVING PLANT PARTS

BY

EARTHWORMS

Because earthworms burrow and cast near or within the rhizosphere, the soil disturbance and abrasion may affect plant roots negatively, particularly the small, fine roots or the root tips, which have not yet produced a protective cortex and are more susceptible to physical disturbance. This abrasion may also break up the mycorrhizal hyphal network (mechanism 1), decreasing root colonization and the many potential benefits of these fungi to plants. Several authors have reported damage by earthworms to rice crops in Southeast Asia (Stephenson 1930; Otanes and Sison 1947; Chen and Liu 1963; Inoue and Kondo 1962, cited in Lee 1985; Pradhan 1986; Barrion and Litsinger 1996), which may be caused by root abrasion if the earthworm population is large, although other factors such as excessive casting on the rice tillers, soil loosening, water drainage, and increased water turbidity have been proposed as the main factors responsible for the damage (Kale et al. 1989; Stevens and Warren 2000). Some authors have proposed that earthworms (mainly lumbricid species) can feed on living plant roots (Stephenson 1930; Carpenter 1985; Baylis et al. 1986; Sackville-Hamilton and Cherret 1991; Cortez and Bouché 1992; Gunn and Cherrett 1993; Hameed et al. 1993), although only in a few instances was this associated with decreased plant productivity. This phenomenon does not seem to be widespread because studies on the crop, gizzard, or gut contents of over 30 earthworm species revealed that roots form a very minor component of the ingested materials in most species (see Brown et al. 1999). The extent of root feeding by earthworms, the identification of the species involved, the conditions encouraging this to happen, and its possible damage to plant productivity still need further evaluation. Other negative effects, probably mostly caused by anecic earthworm species, involve the burial of living plant leaves (Darwin 1881; Zicsi 1954) or damage to germinating seedlings (Walton 1928; Olson 1929; Trifonov 1957; Patel and Patel 1959; Lee 1985; Shumway and Koide 1994). For instance, Darwin (1881) noted that the end of a Triticum repens leaf, still attached to the plant, had been pulled into the burrow of an anecic earthworm species and had dried and turned dark brown; although the rest of the leaf remained fresh and green. He attributed this to the fluids secreted by the earthworm mouth, which rapidly stained the plant tissues, causing cortical cell discoloration and disintegration. Edwards and Bohlen (1996) reported that L. terrestris destroyed a large part of a lettuce crop when soil containing large numbers of the earthworms was taken into a greenhouse. © 2004 by CRC Press LLC

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Summarizing the available results on earthworms as pests of crops, Lee (1985) and Edwards and Bohlen (1996) stated that, although earthworms occasionally damage healthy plants, more commonly they attack very tender or moribund plants already damaged by some other mechanism, and that there is no reason to regard earthworms as serious pests of plants. However, there are clearly some instances when earthworms can damage plants either directly or indirectly (Edwards and Bohlen, 1996; Brown et al. 1999). Care should be taken to prevent these situations from occurring whenever possible.

5. INTERACTIONS

OF

EARTHWORMS

WITH

SEEDS

From the moment a seed germinates, it comes into contact with the soil, a physicochemical environment and a wide range of soil organisms, all of which may have variable degrees of influence on its growth and success as a plant. Moreover, even before a seed germinates, some of these factors may already be influencing its fate. For example, some earthworm species (e.g., L. terrestris) appear to show a preference for ingesting the seeds of certain plant species, depending on their size, shape, texture, and taste (Piearce et al. 1994; Shumway and Koide 1994). Observations made more than a century ago by Hensen (1877) and Darwin (1881) demonstrated the potential importance of surface-feeding anecic and endogeic earthworms in ingesting, transporting, and distributing seeds in the soil. Moreover, seed germination may be slower or more rapid in egested earthworm castings than in surrounding soils (McRill 1974; Atlavinyté and Zimkuviene 1985; Piearce et al. 1994). For example, Grant (1983) and Decaëns et al. (2001) observed lower germination rates and slower germination of the seeds of several weed species in earthworm casts. Furthermore, many seeds are damaged by passage through the earthworm gut, often affecting their germination success or vigor (Grant 1983). In view of the selective consumption and the digestive processes of earthworms, the preferential germination of different seed species in earthworm-linked structures, the dispersal of seeds through the soil, and the physical-chemical effects of earthworms on the soil environment, it has been suggested that earthworms may influence plant recruitment and the composition of plant communities considerably (Piearce et al. 1994; Willems and Huijsmans 1994). Some authors have suggested that earthworms seem to favor the proportion, and often biomass, of clover in pastures (Stebler et al. 1904; Bates 1933; Hopp and Slater 1948; Nielson 1953; Satchell 1955; Thompson et al. 1994; Nuutinen et al. 1998). Positive associations of earthworm casts with the frequency and distribution of the weeds Plantago spp., Trifolium, and Ranunculus were also observed in meadows in the U.K. (Bates 1933; Piearce et al. 1994). The effect of earthworms on the soil weed seed bank, particularly the influence of anecic species that preferentially ingest seeds, should not be underrated. Decäens et al. (2001) estimated that 1 to 13% of the total germinatable soil seed bank of a native savanna and two pastures were deposited in the surface casts of the anecic earthworm species Martiodrilus carimaguensis from the Colombian Eastern Plains. However, if there is preferential ingestion of weed seeds and differential growth of weed seedlings in earthworm casts or earthworm-worked soils (Piearce et al. 1994), this may eventually increase the level of weed infestations of crop fields or grasslands, potentially increasing competition of weeds with the crops or desired plants (Edwards and Bohlen 1996; Stinner et al. 1997).

6. CHANGES

IN

SOIL STRUCTURE CAUSED

BY

EARTHWORMS

The activities of earthworms in the physical “engineering” of soils can modify a wide range of chemical and biological properties and processes influenced by soil structure (see Chapters 10 and 11 this volume). Earthworm pedoturbation of soils can change soil structure by affecting aggregation (mostly by casting) and porosity to water and air (by burrowing and casting), thereby affecting soil physical functions important in root growth and penetration, such as aeration, gaseous exchange, water infiltration. and water-holding capacity (Figure 2.5). Earthworm burrowing creates mostly macropores (pores larger than 30 µm), and casting affects mainly the meso- and microporosity in

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Earthworms Casting Micropores < 0.2 µm

Burrowing

Mesopores 0.2 to 30 µm

Water-holding capacity

Macropores >30 µm

Infiltrability, aeration

Soil structure (pore size distribution and aggregate stability)

Soil physical functions

Root penetration and growth FIGURE 2.5 Diagrammatic representation of ways by which earthworms can affect plant growth via physical changes in the soil environment by burrowing and casting. (Expanded from Syers and Springett, 1983.)

soils (pores smaller than 30 µm) and the stability of soil aggregates. However, earthworm species differ greatly in their ability to modify soil structure, depending on their ecological strategies and behavior. Plants also differ tremendously in their nutrient and water requirements and rooting strategies. The minimum pore size for effective penetration of the roots of most crop species is approximately 200 µm (Wiersum 1957), so many roots become concentrated in macropores, although some root hairs may penetrate mesopores 5 to 20 µm wide (Hofer 1996). Earthworm Casts Earthworms produce basically four types of casts (Lee 1985; Lavelle 1988; Edwards and Bohlen 1996): 1. Globular, consisting of coalescent round or flattened units, generally produced by the larger earthworm species (anecic and endogeic species). 2. Pastelike slurries, mainly produced by endogeic or anecic species and excreted as single masses of soil without a distinct shape, but that take on irregular shapes once dried. 3. Tall vertical heaps or columns of variable shapes, usually deposited on the soil surface where they are most visible by endogeic or anecic species. These are usually created by the sequential deposition of globular casts and, when in tower form, often have a hole in the middle (Darwin 1881; Edwards and Bohlen 1996). 4. Granular, typically in the form of pellets, produced mainly by smaller earthworm species (epigeic, small endogeic, and some anecic species) and distributed on or beneath the soil surface. Casts from different earthworm species can have very different effects on soil structure. The first three types of casts tend to be larger, heavier, and more compact and are usually produced by “compacting” earthworm species; the granular casts are normally smaller, lighter, and looser and break down more easily, and are mostly produced by “decompacting” earthworm species (Blanchart et al. 1997, 1999). Often, the casts of compacting species are consumed by decompacting species, a process that breaks up the larger aggregates into smaller ones, helping regulate overall soil aggregation (Blanchart et al. 1997; Decäens and Rossi 2001) and liberate nutrients that were protected in the casts for plant roots (see Figure 2.6).

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FIGURE 2.6 Pasture root growth into burrows and casting of earthworms from native savannas and pastures planted on highly weathered soils of the Colombian Eastern Plains. Note the two different types of structures: globular “compact” castings created by M. carimaguensis and their breakdown by smaller polyhumic endogeic “decompacting” earthworm species and mesofauna. (Photo P. Lavelle.)

The inner porosity of earthworm casts is also very variable depending on the earthworm species producing them, particularly the earthworm’s anterior and posterior internal morphology and musculature (Lapied and Rossi 2000). A predominance of mesopores (10 to 20 µm) was reported in the casts of M. anomala (Blanchart et al. 1999), whereas pores in the casts of the compacting species P. corethrurus were all smaller than 1 µm (Chauvel et al. 1997). Thus, casts are much more important for retaining plant-available water (fresh casts of many species have water contents above 70%) (Blanchart et al. 1999) and nutrients, whereas earthworm burrows are more important for water by-pass flow, infiltration rates, gaseous exchanges, and root penetration and elongation. Subsequently, several authors (Doube et al. 1997; Stockdill 1966; van Rhee 1969) reported increased water use efficiency by crops in soils inoculated with earthworms in both pot and field experiments. Earthworm casts, once they have undergone a stabilization process still not well understood (Edwards and Shipitalo 1998), become water-stable aggregates, although their stability is very dependent on the soil type, earthworm species, and earthworm feeding habits (Blanchart et al. 1999). Often, an important part (5% or more) of the surface (A) horizon of soils passes annually through earthworm intestines, particularly in tropical regions that are dominated by endogeic species (Lavelle 1988). Under some circumstances, most of the topsoil may be composed of earthworm castings of different ages, sometimes remaining long after the earthworms have disappeared (Buntley and Papendick 1960; Graff 1971b; Pop and Postolache 1987; Lavelle 1988). Thus, because interaggregate spaces are important in soil macroporosity, the physical arrangements of casts, particularly the larger casts containing mostly water-stable macroaggregates (>2 mm diameter), can also have an important effect © 2004 by CRC Press LLC

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on the total number of macropores in soils (see Chapter 10 this volume). Furthermore, casting on the soil surface may open new pores in the soil and can even break surface crusts, thereby helping germinating seedlings reach the soil surface (Kladivko et al. 1986). Compacted soils may also benefit from the activity of decompacting earthworm species (Blanchart et al. 1997, 1999), the incorporation of OM (aggregating agent) by anecic species, and the burrowing strength and stable aggregate formation by endogeic species (Zund et al. 1997; Larink and Schrader 2000). For example, the introduction of various endogeic and deeper-burrowing anecic species of lumbricid earthworms into New Zealand pastures aided the rates of decomposition of accumulated thatch and physical incorporation of lime, fertilizers, and pesticides into the soil, reducing physical, chemical, and biological limitations to root growth and pasture productivity (Stockdill 1982; Springett 1985). However, excessively loose soils or soils with greater proportions of sand that are prone to water stress, may actually benefit from the aggregating action of compacting earthworms. Not all the effects of earthworms on soil structure help plants to grow better. First, the deposition of fresh earthworm casts on the soil surface and the burial of protective surface litter by anecic earthworm species can expose soil particles to splash erosion ( Darwin 1881; Sharpley and Syers 1976; Sharpley et al. 1979; van Hoof 1983; Binet and Le Bayon 1999), promoting their downhill soil movement if the area is sloping. In particular situations and over long time periods, this could reduce the topsoil layer upslope considerably and increase its downslope, as well as change its texture (Nooren et al. 1995) and suitability for plants. In addition, when soils are prone to compaction and a single earthworm species of the compacting type dominates the community, reaching large populations, biomass, and activity levels, the ultimate effect of the earthworms on plant growth may be negative. Hence, Puttarudriah and Shivashankara-Sastry (1961), Blackemore (1994), Barros et al. (1996, 1998), Chauvel et al. (1999) and Ester and Rozen (2002) all observed increased soil compaction and “clodding” caused by earthworm (P. corethrurus and various other species) activities and related the lower soil porosity and water infiltration rates that occurred with decreased plant (radish, carrot, bean, pasture, sorghum, and potato) productivity. Excessive casting on the soil surface and base of plants by lumbricid earthworms in England caused difficulties in harvesting cereals and hay (Stephenson 1957; Edwards and Bohlen 1996), and large amounts of casts on the soil surface of grazed pastures led to “poaching” from cattle trampling, decreasing grass growth in the Netherlands (Hoogerkamp 1984) and New Zealand (Lee 1959). Earthworm Burrows Macropores usually represent only a very small part of the total soil porosity (particularly in clayey soils), yet they are very important in hydraulic conductivity and water infiltration rates when connected with the soil surface and in increasing aeration (Kretzschmar 1998, see Chapter 11, this volume). The positive effects of earthworms on water infiltration may help decrease runoff rates (Roth and Joschko 1991), thereby allowing more water to enter the soil and reducing overall erosion (Hopp 1946, 1973; Sharpley et al. 1979), as well as increasing the potential for water storage in the soil. Thus, the effect of earthworms on soil porosity and infiltration, as well as on organic matter breakdown, has been associated consistently with increased yields in New Zealand pastures (Stockdill 1959, 1982) and reclaimed Dutch polders (e.g., van de Westeringh 1972; Hoogerkamp 1984) and with greater hay and bean yields in large container experiments (Hopp and Slater 1948, 1949), although the interactions with incorporated or surface OM (another aggregating agent) are also likely to be implicated (Cogle et al. 1994) in some responses observed by these authors. Earthworm burrows can serve as preferential pathways for root elongation (Ehlers 1975; Edwards and Lofty 1980; Kirkham 1981; Ehlers et al. 1983; Wang et al. 1986; Kladivko and Timmenga 1990; Hirth et al. 1997; Jiménez 1999), especially in compacted zones found typically in deeper soil layers. In open, abandoned earthworm burrows, the greater aeration and the small © 2004 by CRC Press LLC

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amounts of nutrients associated with the earthworm burrow walls can benefit root growth (Graff 1971a), and cast-filled earthworm burrows usually have large quantities of plant-available nutrients stored in the casts (mechanism 7). The distribution of roots in soil is often related closely to the zones of earthworm activity (Edwards and Lofty 1978, 1980), and root densities can be increased significantly by earthworm activities. In newly reclaimed polders inoculated with earthworms and planted with fruit trees in the Netherlands, van Rhee (1977) reported significantly greater root densities in the earthworm-inoculated sites but no effects on fruit production. Conversely, in a pot experiment in Mexico that compared pots inoculated with earthworms those with no earthworms, Brown (1999) observed significantly greater root densities, as well as more root and shoot biomass, but no increase in productivity of beans in the presence of Polypheretima elongata. The earthworm burrows were commonly filled with roots, and the root distribution throughout these pots showed a much more even (homogeneous) distribution, a factor considered to confer greater plant resistance to environmental stresses (Smucker 1993). The proportion of roots found in deep earthworm burrows (e.g., in the B horizons) compared with those in the soil matrix can be very high (Kirkham 1981; Logsdon and Linden 1992), and these roots may be important in maintaining plant water dynamics. However, estimates of the proportion of roots in earthworm burrows may be exaggerated because roots in earthworm burrows are more easily observed, whereas the rest of the root system may be concealed in the soil matrix (Logsdon and Linden 1992; Kretzschmar 1998). A three-dimensional estimation of interactions between roots and earthworm burrows is still not available (Kretzschmar 1998), and considerable efforts need to be made to understand these interactions and the mechanisms that control them (Tisdall and McKenzie 1995). Thus, it is a combination of the composition (ecological category, species) of the earthworm community present at a given location, the placement of their casts (surface, belowground, deep in soil, near roots, etc.), the quantities of casts deposited and their age, and the amount, type, depth, and openness of the earthworm burrows produced, the interaction of microorganisms with earthworm structures, the physicochemical soil environment, and land management that determine the ultimate effects of earthworms on soil structure and the rooting environment.

7. CHANGES IN NUTRIENT SPATIOTEMPORAL AVAILABILITY CAUSED BY EARTHWORMS The availability of many essential plant nutrients has been shown to increase in structures produced by various earthworm species, especially in their casts (e.g., Mulongoy and Bedoret 1989; Barois et al. 1999) and burrow walls. This greater nutrient availability is mainly a result of the selective feeding of earthworms on regions of the soil rich in organic matter, clay, and nutrients (Barois et al. 1999; Cortez and Hameed 2001), gut-associated processes, and cast-associated processes (Figure 2.7), together with some earthworm burrow-associated processes (especially with anecic earthworm species; Devliegher and Verstraete 1997; Brussaard 1999). Such processes include the grinding action of the gizzard, the priming of microbial activity in the gut, and the greater populations and activity of microorganisms in the earthworm casts and burrows (Figure 2.7), that induce chemical changes in earthworm-worked soil (e.g., Lee 1985; Edwards and Bohlen 1996). These nutrient enrichment processes (Devliegher and Verstraete 1995; Brussaard 1999) differ greatly according to the earthworm species involved, their ecological categories, and the feeding habits, particularly the amounts of plant litter they ingest. The type and placement of the earthworm casts are also important, affecting the spatiotemporal availability of the nutrients they contain (Figure 2.7). Surface earthworm casts dry out much more quickly, harden, and, if compact, are likely to limit root penetration, thereby reducing the ability of plant roots to obtain the nutrients stored inside the casts (nutrient protection) until they are broken down (Figure 2.6 and Figure 2.7). Belowground earthworm casts remain fresh and moist for much longer periods of time and, if they are of the decompact types (with more meso- and macropores and macroaggregates), allow roots to penetrate more easily (Figure 2.6) and profit from the greater nutrient contents available to plants. © 2004 by CRC Press LLC

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Mineral soil + OM (food) inputs

Physical breakdown

comminution

(Bio)chemical breakdown

Mucus, enzyme production, pH changes

Fresh casts Compact

Decompact

Aging casts Nutrient “protection”

Breakdown

GAPs (grinding in gizzard, microbial activity)

CAPs (microbial activity and soil chemical changes)

Nutrient release

Plant nutrient spatiotemporal availability

Root growth and development FIGURE 2.7 Diagrammatic representation of ways by which earthworms can positively affect plant growth via chemical changes in the soil environment induced by gut-associated processes (GAPs) and cast-associated processes (CAPs).

Most of the reported increases in uptake of nutrients (especially N and P) by plants in response to earthworms has been related to increased P and N mineralization rates and their availability in castings, earthworm burrow linings, and earthworm-worked soils (e.g., Graff 1967; 1970; Aldag and Graff 1975; Lee 1985; Lavelle et al. 1992; López-Hernández et al. 1993; Brossard et al. 1995; Chapuis-Lardy et al. 1998; Barois et al. 1999; Rangel et al. 1999). This is particularly important because N and P are commonly the most limiting nutrients in soils for optimum plant productivity. Because cast production rates can reach large quantities, ranging from a few tonnes per hectare in temperate arable land up to more than 1000 t ha−1 in tropical savannas with a predominance of geophagous endogeic earthworm species (Lavelle 1988), the amount of nutrients cycled and made available to plants by earthworm activity can be enormous, ranging from a few up to several hundred kilograms of mineral N per hectare and tens of kilograms per hectare of plant-available P (Lee 1985, Hauser 1993). Because of the enhanced rates of nutrient release and availability in the drilosphere, plants grown in the presence of earthworms often have more nutrients, particularly N and P (e.g., Atlavinyte and Vanagas 1982; McColl 1982; Graff and Makeschin 1983; Spain et al. 1992; Blakemore 1994; Baker et al. 1997; Stephens et al. 1994a; Tomati et al. 1996, etc.). Greater availability and rates of

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uptake of nutrients by plants in response to earthworm activity have often led to greater transfer of C and N from soils to shoots and more shoot biomass relative to that of roots (Wolters and Stickan 1991; Klebsch et al. 1995). The actual benefits derived by the plants from earthworms depend on the ability of the plant to extract these nutrients from the drilosphere and the soil solution. If nutrients are released and made available to plants during drier or colder periods when plants are dormant or not growing as actively or when the field is bare, then they may be lost (e.g., by leaching) and will provide very few benefits to the plants. Thus, the synchrony of nutrient availability (especially N) with the needs of the plants is critical, and earthworms may play an important role in this process (Fragoso et al. 1997). The increased amounts of plant-available nutrients in earthworm casts and burrows can promote root growth considerably (Darwin 1881; Ehlers 1975; Spiers et al. 1986; see Figure 2.6), and their importance to plant nutrition increases proportionally to the differences in nutrient status between the earthworm casts the surrounding soils, the quantities of casts produced and their synchronization and synlocalization with root growth needs. Thus, in deeper and possibly poorer soil zones, earthworm casts and burrows may serve as hot spots of nutrient availability to plant roots (Mouat and Keogh 1987) and promote fine root growth. Conversely, when the plants are growing in nutrientrich soils, the relative nutrient bioavailability stimulation in response to earthworm activities may be less than in poor soils, and expected plant growth increases may also be less (e.g., Atlavinyte and Vanagas 1973; Brown et al. 1999; Buse 1990; Doube et al. 1997) because the plants can obtain most of their required nutrients without the earthworms. Many experiments have demonstrated the importance of plant-available nutrients and PGR substances in earthworm casts to plant responses (e.g., Dash and Das 1989; Kang and Ojo 1996; Kang et al. 1994; Nijhawan and Kanwar 1952; Reddy et al. 1994; Tomati et al. 1987; Norgrove and Hauser 1999; Kollmannsperger 1980). The plant response is usually proportional and related positively to the quantity of earthworm casts applied or to the ratio of casts to soil or other substrates used. However, experiments of this nature have the disadvantage of unrealistic experimental conditions compared with the field and the vastly different chemical, physical, and biological properties of the casts, depending on their source and age. The effects of earthworms on nutrient mineralization are especially evident in sites newly invaded by earthworms. For example, lumbricid or pheretimoid earthworm invasions into the forests of North America have sometimes resulted in dramatic changes in the chemical status of soil, transforming humus types from mor to mull (Langmaid 1964; Nielson and Hole 1964; see Chapter 5 this volume). Corresponding C and N losses and increased nutrient turnover rates resulting from earthworm invasion of new sites may be on the order tens to hundreds of kilograms per hectare (O’Brien and Stout 1978; Alban and Berry 1994; Scheu and Parkinson 1994a; Burtelow et al. 1998). Some plants may benefit from the increased amounts of available nutrients (Scheu and Parkinson 1994b), but in general little is known of the effect of such large nutrient fluxes on overall plant growth and their potential effects on plant communities (e.g., species composition and biodiversity). Eutrophic and opportunistic plants may profit preferentially from earthworm presence during the nutrient release phase (which may last several years) until a new “equilibrium” is reached, when the lower total nutrient stocks and more rapid rates of turnover may exert relatively larger and different selection pressures on the plant communities present. Litter burial, ingestion, and digestion by earthworms, particularly anecic and epigeic earthworm species, accelerates rates of decomposition, whereas endogeic earthworm species tend to promote mineralization of the particular light and coarse organic matter fractions in soils (McCartney et al. 1997; Lavelle et al. 1998; Parmelee et al. 1998). Thus, the incorporation of organic matter by an anecic earthworm species such as M. carimaguensis in the savannas of Colombia has been associated with reduced Al saturation (through binding with organic matter) and reduced Al limitation to grass growth (Decaëns et al. 1999). In New Zealand pastures, earthworms have been shown to accelerate the incorporation of lime, fertilizers, and insecticides such as DDT (for grass grub control) (Stockdill 1966, 1982; MacKay et al. 1982; Springett 1985), thereby promoting grass productivity. © 2004 by CRC Press LLC

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The soil-mixing activities of earthworms have been shown to promote the recovery of N and P from organic residues, inorganic fertilizers, and rock phosphate in various pot and field trials using the plant species B. decumbens, Panicum maximum, ryegrass, and maize (MacKay et al. 1982; Mansell et al. 1981; Spain et al. 1992; Hameed et al. 1993, 1994a,b; Hu and Wu 1994; GilotVillenave et al. 1996; G.G. Brown et al. 2000; Cortez et al. 2000). However, earthworm activities can also sometimes lead to decreases in the availability of nutrients to plant roots. Fresh earthworm casts rich in nutrients (especially N and P) that are deposited on the soil surface may be eroded easily and are also hot spots of denitrification and NH3 volatilization (Elliott et al. 1990; Lensi et al. 1992; Karsten and Drake 1997). Furthermore, earthworm burrows, especially those of anecic species, connected to the soil surface may promote water bypass or mass flow, causing increased leaching of soluble nutrients (Anderson et al. 1983; Hoogerkamp 1984; Knight et al. 1989; Haimi and Boucelham 1991; Edwards and Shipitalo 1998). Nutrients from Earthworms (Death, Excretion) Some authors have proposed that most N excretions and mucus secretions from earthworms may be utilized rapidly by plants (e.g., Bouché and Ferrière 1986; Bouché et al. 1987; Hameed et al. 1994a,b; Whalen et al. 1999, 2000). However, this contribution is probably not very large unless the earthworm biomass is high and most of their activities are concentrated in the root zone. Further field research on this topic, particularly using homogeneously labeled (15N, 32P) earthworms, is needed. The release of nutrients from dead earthworm tissues has often been believed to play important role in plant productivity (Russell 1910; Satchell 1958; Barley 1961; Callaham and Hendrix 1998; Whalen et al. 1999). However, although visual observations show that earthworm bodies decompose very rapidly in soil, only a few reports have been published on the quantities of nutrients made available from dead earthworm biomass (e.g., Satchell 1967; Christensen 1988; Martin 1990; Whalen et al. 1999; Hodge et al. 2000). Furthermore, earthworm biomass is probably an important and significant source of plant nutrients only in field and pot experiments, in which the inoculation rates of earthworm biomass into soils is very high and when earthworm mortality or turnover rates are high. For instance, the small soil volumes often used in pot experiments may be insufficient to maintain earthworm populations inoculated into them to levels above the carrying capacity of the soil. Under such conditions, many earthworms may die, liberating nutrients that, although in small amounts relative to typical soil nutrient supplies, are enough to influence plant growth because of the low soil:earthworm biomass ratio. This has occurred in many experiments, even in those that based the rates of addition of earthworms to field populations but not to field biomass, thus adding greater biomass than would normally occur in the field (e.g., Satchell 1958; Doube et al. 1994c; Baker et al. 1996, 1997; Callaham and Hendrix 1998; Whalen et al. 1999). When larger soil volumes were used, or more realistic earthworm populations and biomass were added, nutrients from dead earthworms played a much smaller role in plant nutrition. This is probably the case even in field situations with large earthworm biomass turnovers of up to 600 kg ha−1 year−1 (fresh mass), which nevertheless can supply only a few kilograms per hectare per year of mineral N from decomposing earthworm tissues (Brown et al., unpublished data).

CRAWLING FORWARD: THE CHALLENGE OF IDENTIFYING AND QUANTIFYING THE POTENTIAL OF EARTHWORMS TO INCREASE PLANT GROWTH It is clear that a wide range of direct and indirect mechanisms of earthworm activities on plant growth can be identified. However, to benefit from the potential of plant growth stimulation by earthworms, promotion and decrease mechanisms in plant growth and their modes of action must be understood within a given practical soil-plant-earthworm species and population context. There

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are many possible combinations of soil types and associated plant and earthworm species. Because several PGI mechanisms may be operating simultaneously, identifying the most important modes of action on the plant in a given situation is a difficult task. However, there are guiding principles that can be followed to help predict potential benefits of earthworm activities to plant growth and identify some of the mechanisms involved.

“ALL-MINUS-ONE” TESTS

AND

FIELD TRIALS

First, existing limitations to plant growth in a given site, situation, soil type, and abiotic and biotic environmental conditions should be assessed and classified hiearchically. Then, the potential of earthworms to ameliorate any of the limiting factors must be evaluated. Such potentials can be assessed using pot and laboratory methodologies, not necessarily including plants. To be able to deal adequately with the multiple limiting factors, commonly present under field conditions, and identify the interactions between the limiting factors and effects of earthworms on the soil and plants, a series of “all-minus-one” trials could be designed in a greenhouse. In such experiments, all the environmental, nutrient, and physical requirements would be provided to the plant except one. Each requirement can then be altered to test how earthworms affect it and to determine any associated plant responses. Such experiments could begin with proper identification of all the soil-associated chemical limitations to plant growth, which can be assessed by soil analyses, plant analyses, and limitingnutrient pot experiments. Fertilization with all but one nutrient and inclusion or exclusion of earthworms could permit the assessment of the role of earthworms in facilitating nutrient availability and uptake by the plants. Soil sterilization and inoculation with a range of different microorganisms is also possible to help assess the role of their interactions with earthworms on plant productivity, but this could be time consuming and laborious. In shorter-term experiments, the physical effects of earthworms on soil studies are probably less important than the chemical and biological effects, although as the length of the experimental periods increase, physical effects (especially limitations when only one earthworm species is included) may become more important. To separate the chemical, physical, and biological mechanisms, simultaneous experiments using different methodologies and experimental designs, such as pots with and without earthworms and pots with soil previously processed by earthworms and various other treatment combinations (e.g., ± residues, ± fertilizers, ± sterilized or unsterilized), or with one or more earthworm species in combination could be developed. The approaches described are far from ideal because some interactions may occur when only a single factor is manipulated, and such experiments do not take into account the complementary or adverse roles of other organisms present in soils. Furthermore, such an experimental design will require a very large number of replicates and treatments, making it very complex and somewhat impractical. Therefore, experiments should be mainly in the field, while recognizing both their drawbacks and benefits (Brown et al. 1999). In field experiments, a great many more variables, many of them uncontrollable, can occur, although the results obtained should be much more realistic and useful to practicing farmers, foresters, or other environmental managers. In such instances, proper assessment of the soil physical limitations to plant growth is essential because, under such conditions, physical effects of earthworms on soils may be of much greater importance (Alegre et al. 1996; Barros 1999). Obviously, the characterization of changes in earthworm communities by periodic population sampling, the quantification of earthworm casting and burrowing activities (e.g., by physical description of soil cores or x-ray commuted tomography), and their chemical and biological analyses are important to assess the extent of earthworm changes to soil structure, biology, and fertility. In such studies, controls (without introduced or native earthworms) must be used and maintained, and if earthworms are to be introduced to them, realistic biomass, populations, and species assemblages should be used.

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THE EARTHWORM THRESHOLD CONCEPT Probably the most important effects of earthworms on crop yields will occur when earthworms are able to significantly modify factors that also happen to be those most limiting to plant growth. Almost all the soil physical, biological, and chemical needs of plants can become limiting when they fall below a certain lower threshold or become excessively above it. The range between these two thresholds is the zone ideal for plant growth. Hence, possible effects of earthworms on plants can be regulated by their ability to increase some of these factors that are limiting plant growth above the lower thresholds and decrease factors adverse to growth below the upper thresholds. The combination of effects of the drilosphere and the threshold level at which each of the biological, chemical, or physical factors are limiting plant growth will ultimately determine the effects of earthworm activities on the plants. For example, if a population of a parasitic nematode or an infestation by a particular fungal or bacterial pathogen has reached disease proportions and become a primary limiting factor in plant production, and if earthworms are able to reduce the population, it seems likely that this would be a dominant mechanism influencing plant growth, although other factors, such as the influence of earthworms on soil structure and fertility and biological interactions with soil microorganisms and invertebrates, will also be important and operate simultaneously. Similarly, if availability of N or P are limiting factors, earthworm-induced increases in the availability of the nutrients or changes in the mycorrhizal colonization of roots may become important controlling mechanisms. Finally, if soils are compacted or prone to compaction and associated hydrological limitations are complicating plant growth, earthworm bioturbation and associated soil structural changes may be the most important mechanisms that enhance plant productivity.

FUTURE NEEDS

IN

EARTHWORM RESEARCH

An ultimate goal in the process of assessing the potential effects of earthworms on plant productivity would be the development of effective simulation models in which the soil environmental constraints are matched with the earthworm and plant species (or communities) present at a site to predict any potential direct and indirect influences of earthworm activities on soil physical, chemical, and biological limitations and plant productivity. Unfortunately, there are still many gaps in knowledge of these processes that need to be filled to be able to develop such models and predict accurately whether a particular earthworm species, population, or community will enhance or suppress plant productivity. Thus, future research should strive to use more holistic approaches using detailed experiments that address potential mechanisms individually as well as in combination, but do not attempt to “kill all birds with one stone.” On the positive side, the study of increased production of PGR substances; promotion of beneficial rhizobacteria; reductions in plant pests, pathogens, or parasites; the interactions of earthworms with plant seeds; and the potential attraction of earthworms to roots and of roots to earthworm structures such as casts deserve more attention. On the negative side, the assessment of plant growth suppression by direct and indirect effects of earthworms also requires more research to verify any potential incompatibilities between existing earthworm communities and established plant species and populations, soils, cropping systems, and crops. There may be situations for which it is better to plant gramineous plant species rather than legumes or vegetables (Puttarudriah and Sastry, 1961) or to manage the soil in a way to decrease potentially negative effects of earthworm activity on plant production. Finally, many earthworm species with potential benefits for plant growth have never been studied experimentally and deserve more attention. Further studies should focus especially on the potential benefits of earthworms to plant productivity in crop fields, pastures, and managed forests, where site-specific management practices and technological innovations (e.g., precision agriculture) could be used to manage the earthworm community for the benefit of plant growth.

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ACKNOWLEDGMENTS We would like to thank the Instituto de Ecología, A.C., IRD and CNPg-Prefix for their support during the writing of this chapter. We also thank the many colleagues who provided insightful ideas and comments for its development.

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Scheu, S., Heenhausen, A., and Jones, T.H. 1999. Links between the detrivore and the herbivore system: Effects of earthworms and Collembola on plant growth and aphid development, Oecologia, 119, 541–555. Senapati, B.K. 1992. Biotic interactions between soil nematodes and earthworms, Soil Biol. Biochem., 25, 1441–1444. Shapiro, D.I., Berry, E.C., and Lewis, L.C. 1993. Interactions between nematodes and earthworms: enhanced dispersal of Steinernema carpocapsae, J. Nematol., 25, 189–192. Sharpley, A.N. and Syers, J.K. 1976. Potential role of earthworm casts for the phosphorus enrichment of runoff waters, Soil Biol. Biochem., 8, 341–346. Sharpley, A.N., Syers, J.K., and Springett, J.A. 1979. Effect of surface-casting earthworms on the transport of phosphorus and nitrogen in surface runoff from pasture, Soil Biol. Biochem., 11, 459–462. Shaw, C. and Pawluk, S. 1986. Faecal microbiology of Octolasion tyrtaeum, Aporrectodea turgida and Lumbricus terrestris and its relation to the carbon budgets of three artificial soils, Pedobiologia, 29, 377–389. Shumway, D.L. and Koide, R.T. 1994. Seed preferences of Lumbricus terrestris, Appl. Soil Ecol., 1, 11–15. Singer, A.C., Crowley, D.E., and Menge, J.A. 1999. Use of an anecic earthworm, Pheretima hawayana, as a means for delivery of fungal biocontrol agents, Pedobiologia, 43, 771–775. Smucker, A.J.M. 1993. Soil environmental modifications of root dynamics and measurement, Annu. Rev. Phytopathol., 31, 191–216. Spain, A.V., Lavelle, P., and Mariotti, A. 1992. Stimulation of plant growth by tropical earthworms, Soil Biol. Biochem., 24, 1629–1633. Spain, A.V. and Le Feuvre, R. 1997. Stable C and N isotope values of selected components of a tropical Australian sugarcane ecosystem, Biol. Fertil. Soils, 24, 118–122. Spain, A.V., Safigna, P.G., and Wood, A.W. 1990. Tissue carbon sources for Pontoscolex corethrurus (Oligochaeta, Glossoscolecidae) in a sugarcane ecosystem, Soil Biol. Biochem., 22, 703–706. Spiers, G.A., Gagnon, D., Nason, G.E., Packee, E.C., and Lousier, J.D. 1986. Effects and importance of indigenous earthworms on decomposition and nutrient cycling in coastal forest systems, Can. J. For. Res., 16, 983–989. Springett, J.A. 1985. Effects of introducing Allolobophora longa Ude on root distribution and some soil properties in New Zealand pastures, in Fitter, A.H., Atkinson, D., Read, D.J., and Usher, M.B., Eds., Ecological Interactions in Soil, Blackwell, London, pp. 399–407. Stebler, F.G., Thielé, E., Volkart, U., and Grisch, U. 1904. Sechsundzwanzigster Jahresbericht der Schweizer, Sammeluntersuchungs und Versuchsanstalt in Zurich, Landwirtschaftliche Jahrbucher der Schweiz, 18, 45–100. Stephens, P.M. and Davoren, C.W. 1994. Increased growth of subterranean clover and increased nodulating competitiveness by a strain of Rhizobium leguminosarum bv. trifolii, associated with the presence of the earthworm Aporrectodea caliginosa, in Pankhurst, C.E., Ed., Soil Biota: Management in Sustainable Farming Systems, Poster Papers, CSIRO, East Melbourne, Australia, pp. 15–17. Stephens, P.M. and Davoren, C.W. 1995. Effect of the lumbricid earthworm Aporrectodea trapezoides on wheat grain yield in the field, in the presence or absence of Rhizoctonia solani and Gaeumannomyces graminis var. tritici, Soil Biol. Biochem., 28, 561–567. Stephens, P.M. and Davoren, C.W. 1997. Influence of the earthworms Aporrectodea trapezoides and A. rosea on the disease severity of Rhizoctonia solani on subterranean clover and ryegrass, Soil Biol. Biochem., 29, 511–516. Stephens, P.M., Davoren, C.W., Doube, B.M., and Ryder, M.H. 1994a. Ability of the earthworms Aporrectodea rosea and Aporrectodea trapezoides to increase plant growth and the foliar concentrations of elements in wheat (Triticum aestivum cv. Spear) in a sandy loam soil, Biol. Fertil. Soils, 18, 150–154. Stephens, P.M., Davoren, C.W., Doube, B.M., Ryder, M.H. 1994b. Ability of the lumbricid earthworms Aporrectodea rosea and Aporrectodea trapezoides to reduce the severity of take-all under greenhouse and field conditions, Soil Biol. Biochem., 26, 1291–1297. Stephens, P.M., Davoren, C.W., Doube, B.M., and Ryder, M.H. 1994c. Field experiments demonstrating the ability of earthworms to reduce the disease severity of Rhizoctonia on wheat, in Pankhurst, C.E., Ed., Soil Biota: Management in Sustainable Farming Systems, CSIRO, East Melbourne, Australia, pp. 19–20.

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Stephens, P.M., Davoren, C.W., Ryder, M.H., and Doube, B.M. 1994d. Greenhouse and field experiments demonstrating the ability of earthworms to reduce the severity of take-all on wheat, in Pankhurst, C.E., Ed., Soil Biota: Management in Sustainable Farming Systems, CSIRO, East Melbourne, Australia, pp. 21–23. Stephens, P.M., Davoren, C.W., Ryder, M.H., and Doube, B.M. 1993. Influence of the lumbricid earthworm Aporrectodea trapezoides on the colonization of wheat roots by Pseudomonas corrugata strain 2140R in soil, Soil Biol. Biochem., 25, 1719–1724. Stephens, P.M., Davoren, C.W., Ryder, M.H., and Doube, B.M. 1994e. Influence of the earthworm Aporrectodea trapezoides (Lumbricidae) on the colonization of alfalfa (Medicago sativa L.) roots by Rhizobium meliloti L5-30R and the survival of R. meliloti in soil, Biol. Fertil. Soils, 18, 63–70 Stephens, P.M., Davoren, C.W., Ryder, M.H., and Doube, B.M. 1994f. Influence of the earthworms Aporrectodea rosea and Aporrectodea trapezoides on Rhizoctonia solani disease of wheat seedlings and the interaction with a surface mulch of cereal-pea straw, Soil Biol. Biochem., 26, 1285–1287. Stephens, P.M., Davoren, C.W., Ryder, M.H., Doube, B.M., and Correll, R.L. 1994g. Field evidence for reduced severity of Rhizoctonia bare-patch disease of wheat, due to the presence of the earthworms Aporrectodea rosea and Aporrectodea trapezoides, Soil Biol. Biochem., 26, 1495–1500. Stephenson, J. 1930. The Oligochaeta, Oxford University Press, Oxford, U.K. Stephenson, J.W. 1957. Excessive Casting by Earthworms, Rothamsted Experimental Station, Report for 1956, pp. 155. Stevens, M.M. and Warren, G.N. 2000. Laboratory studies on the influence of the earthworms Eukerria saltensis (Beddard) (Oligochaeta: Ocnerodrilidae) on overlying water quality and rice plant establishment, Int. J. Pest Manage., 46, 303–310. Stinner, B.R., McCartney, D.A., Blair, J.M., Parmelle, R.W., and Allen, M.F. 1997. Earthworm effects on crop and weed biomass, and N content in organic and inorganic fertilized agroecosystems, Soil Biol. Biochem., 29, 423–426. Stockdill, S.M.J. 1959. Earthworms improve pasture growth, N.Z. J. Agric., 98, 227–223. Stockdill, S.M.J. 1966. The effect of earthworms on pastures, Proc. N.Z. Ecol. Soc., 13, 68–75. Stockdill, S.M.J. 1982. Effects of introduced earthworms on the productivity of New Zealand pastures, Pedobiologia, 24, 29–35. Syers, J. K. and Springett, J.A. 1983. Earthworm Ecology in Grassland Soils, in Satchell, J.E., Ed., Earthworm Ecology: From Darwin to Vermiculture, Chapman & Hall, New York, pp. 67–83. Thompson, K., Green, A., and Jewelsh, A.M. 1994. Seeds in soil and worm casts from a neutral grassland, Funct. Ecol., 8, 29–35. Thornton, M.L. 1970. Transport of soil-dwelling aquatic Phycomycetes by earthworms, Trans. Br. Mycol. Soc., 3, 391–397. Tisdall, J.M. and McKenzie, B.M. 1995. Impact of earthworms on soil physical properties, in Temple-Smith, M. and Pinkard, T., Eds., The Role of Earthworms in Agriculture and Land Management, Department of Primary Industry and Fisheries, Tasmania, Technical Report 1/96, pp. 59–66. Tomati, U. and Galli, E. 1995. Earthworms, soil fertility and plant productivity, Acta Zool. Fenn., 196, 11–14. Tomati, U., Galli, E., and Pasetti, L. 1996. Effect of earthworms on molybdenum-depending activities, Biol. Fertil. Soils, 23, 359–361. Toyota, K. and Kimura, M. 1994. Earthworms disseminate a soil-borne plant pathogen, Fusarium oxysporum f. sp. raphani, Biol. Fertil. Soils, 18, 32–36. Trifonov, D. 1957. Über die Bekämpfiung der Maulwurfurgrille und des Regenwurms mit dem Präparat, Alon Kombi Bulgar. Tiutium, 2, 114–115. Tuffen, F., Eason, W.R., and Scullion, J. 2002. The effect of earthworms and arbuscular mycorrhizal fungi on growth of and 32P transfer between Allium porrum plants, Soil Biol. Biochem., 34, 1027–1036. van de Westeringh, W. 1972. Deterioration of soil structure in worm free orchard soils, Pedobiologia, 12, 6–15. van der Rees, P.J. and Rogaar, H. 1988. The effect of earthworm activity on the vertical distribution of plant seeds in newly reclaimed polder soils in the Netherlands, Pedobiologia, 31, 211–218. van Hoof, P. 1983. Earthworm activity as a cause of splash erosion in a Luxembourg forest, Geoderma, 31, 195–204. van Rhee, J.A. 1969. Development of earthworm populations in polder soils, Pedobiologia, 9, 133–140. van Rhee, J.A. 1977. A study of the effect of earthworms on orchard productivity, Pedobiologia, 17, 107–114.

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Part II Earthworm Taxonomy, Diversity, and Biogeography

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Processes and Their 3 Planetary Interactions with Earthworm Distributions and Ecology Samuel W. James Department of Life Sciences, Maharishi University of Management, Fairfield, Iowa, U.S.A.

CONTENTS Plate Tectonics and Earthworm Phylogeny .....................................................................................54 Applications of Earthworm Biogeography to the Earth’s History .................................................56 How the Earth’s History Affects Earthworm Distributions ............................................................58 References ........................................................................................................................................61

Studies of the history of earth are traditionally the domain of geologists and those in various paleodisciplines (paleontology, paleoecology, paleoclimatology, paleobotany, etc.). However, the history of biology has many examples of research into the connections between present-day biogeography and ecology on the one hand and earth history on the other. Such research may be considered in three broad categories. First, there is application of data derived from life forms of the present to questions of earth history. For example, the use of phylogenetic trees in combination with distributional data can provide valuable insights into the history of land-area relationships, as in the field of vicariance biogeography, which studies processes that split distributions (vicariating events) in relation to phylogeny (Wiley 1988). It may be claimed that earthworms were of some importance in the early development of plate tectonic theory and, therefore, of vicariance biogeography, because the former discipline has been very important to the ascendance of the latter. Plate tectonics theory attempts to account for the movement of parts of the Earth’s crust and provides a means of recovering the ancient positions of those crustal pieces. The Oligochaete systematist Michaelsen (1933) named an earthworm genus after Wegener in honor of Wegener’s work toward understanding the distributions of earthworms. It is possible that the two men (who were office neighbors at the University of Hamburg) discussed earthworm distributions, providing Wegener with more evidence for his theory, which was radical at the time. Many other groups of organisms have been used in arguments supporting or refuting various hypotheses about the history of events (Rosen 1985; Humphries and Parenti 1999; Liebherr 1988). One can also find a significant thread of research to trace the contributions of living organisms to global processes. In its boldest version, this area is termed the Gaia hypothesis (Lovelock 1988), but numerous more mundane examples can be found in the literature on biogeochemical cycles, trace gas fluxes, and many other topics relevant to global climate models. For example, earthworms may influence landforms. Charles Darwin (1881) described a process by which earthworms could increase rates of soil erosion and thereby the rates of change of land 53 © 2004 by CRC Press LLC

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topography. This subject has not been thoroughly researched, and there is some evidence that suggests that earthworms may actually reduce rates of soil erosion (Hopp 1946; Sharpley and Syers 1977) in agricultural land. Nooren et al. (1995) suggested that earthworms could increase rates of clay loss from an African soil, thereby creating nutrient-poor sandy topsoils. Lavelle and Martin (1992) hypothesized that, by protecting soil organic matter from oxidation, earthworms could have an important influence on atmospheric carbon dioxide levels. Because rates of carbon loss from soils make an important contribution to the elevation of atmospheric carbon dioxide levels, any organism that is capable of contributing to reversing or moderating that trend should be investigated closely (see Chapter 10 this volume). A third way in which earth history can play a significant role in field biology is in the understanding of ecological interactions and of organism taxon distributions. For example, the presence of a particular taxon and its unique ecological contributions to the biological community of one site, but its absence from another, could be interpreted as caused by differences in history. Perhaps the taxon could have existed in the latter site but did not get there. Thus, it may not be some ecological or evolutionary necessity that determines certain characteristics of the system but instead a historical accident. This is an important distinction because some simplistic interpretations of community dynamics assume a long history of community optimization. It seems probable that nature has not tried all possible combinations but has instead relied on the interactions and evolution of what is present combined intermittently with unpredictable arrivals of new organisms to the site. An analogy can be seen in communities of earthworms in which most or all the species are introduced or exotic to a location. The organismal content of such communities is clearly accidental, and any further interpretation of that content assumes that other potential invaders (i.e., those species not currently present) have had an opportunity to invade, but only those ecologically compatible with the community qualities have succeeded. Of these three connections of earth history to modern biology, in terms of earthworms, I focus on two: the first and the last. These may be stated briefly as using modern organisms to learn more about the history of the Earth and viewing modern organismal distributions and ecology as the outcomes of an interaction between evolutionary and large-scale abiotic processes. My purpose in this chapter is to outline some of the ways that research into earthworm systematics and biogeography can contribute significantly to the broader subject of earth history and vice versa. As I develop these discussions, I suggest that researchers into earthworm systematics and biogeography can profit from close attention to some of the recent developments in organizing their ideas and analyzing their data.

PLATE TECTONICS AND EARTHWORM PHYLOGENY The first subject is mutual enrichment between geology and systematics, from either geological data informing on probable phylogeny or vice versa. An example of the first case is provided by ongoing debates about the higher-level classification of the Clitellata. The earliest application of earthworm systematics to earth history was in the connection with Wegener’s theory that continents move, which has since been transformed into plate tectonics theory. The vicariance events (splitting of land masses) are quite ancient, and the resulting taxonomic divisions within earthworms are generally at the family level, although some are within genera. Various family-level classifications have been superimposed on paleoreconstructions of the past 250 million years of crustal movements (Jamieson 1981; Bouché 1983; Omodeo 2000), but all those classification systems were contradicted by the results of Jamieson et al. (2002), who were the first to apply deoxyribonucleic acid data to studies of earthworm phylogeny (Figure 3.1). For instance, the Octochaetidae were shown to be polyphyletic, the various concepts of the Acanthodrilidae were paraphyletic, and the Crassiclitellata (earthworms as commonly understood minus the Moniligastridae) were clearly monophyletic.

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*

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Glossoscolecidae

Eudrilidae

Microchaetidae

Lumbricidae

Hormogastridae

other Lumbricoidae

Ocnerodrilidae

Acanthodrilidae

Megascolecidae

Planetary Processes and Their Interactions with Earthworm Distributions and Ecology

other Clitellata

FIGURE 3.1 Simplified phylogenetic tree of major groups of earthworms. The node with the asterisk is the polytomy discussed in detail in the text. (After Jamieson et al. 2002.)

To provide an example with biogeographic relevance, the Glossoscolecidae (South America) are clearly a sister group of the Eudrilidae (Africa) rather than allied to the Microchaetidae (Africa), as presented by Bouché (1983) and Omodeo (1998). The Lumbricina of Omodeo and similar concepts of other groups emerged relatively intact but with significant modifications, particularly the possible inclusion of the Eudrilidae. This remains unclear because a probable trichotomy of Megascolecoidea (M), Lumbricoidea (L), and Glossoscolecidae plus Eudrilidae (G) was unresolved by Jamieson et al. (2002) and could be resolved in three different ways: (M, (L, G)); ((M, L),G); or ((M, G), L). Like Omodeo (2000), one could resort to paleogeography to support a phylogenetic theory, choosing the last of the three because this leaves the Gondwanan taxa sharing a more recent common ancestor than with the primarily Laurasian Lumbricoidea. However, this was not the conclusion reached by Omodeo (1998), who placed the Glossoscolecidae with the Lumbricoidea, nor is it consistent with the work of Bouché (1983), who placed the Glossoscolecidae with the Microchaetidae, Kynotidae, and Almidae as the Glossoscolecoidea, leaving the Eudrilidae with the Megascolecoidea. What is needed is to expand the taxon sampling (see Zwickl and Hillis 2002) of Jamieson et al. (2002) to include more Lumbricoidea, Glossoscolecidae, Ocnerdrilidae, and Eudrilidae to resolve the polytomy and to address more definitively Omodeo’s (1998) polyphyletic model of earthworm evolution. This should also include some species of Alluroididae, because that family is proposed as the source of two independent ancestors of major trunks of the Crassiclitellata tree (Omodeo 1998). Another problem is that some versions of the distribution of Lumbricoidea have members (the Microchaetidae) located in sub-Saharan Africa. Can this indicate that earthworms have a significant pre-Pangaea history, such that their biogeography can be understood only with reference to two cycles of continental fragmentation? Clearly, this issue cannot be settled until the proposed phylogeny can be stabilized, but some pre-Pangaea reconstructions link continental geologic units that are now far apart, such as northeastern North America and South Africa. Although geological models of earth evolution can provide some corroboration of phylogenetic hypotheses, the main burden of gathering evidence lies with biologists. I now consider the contributions of biological data to understanding geological events.

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APPLICATIONS OF EARTHWORM BIOGEOGRAPHY TO THE EARTH’S HISTORY There are good reasons to apply earthworm systematics and biogeography to the study of the geological history of the Earth, with particular reference to movements of the Earth’s crust. Such an idea began from two pieces of biogeographical data. First, earthworms (except anthropochorous peregrines) are absent from midoceanic volcanic islands (Gates 1969; Nakamura 1990; Talavera 1990) and uplifted carbonate platforms, indicating that earthworms may experience great difficulty crossing saltwater (Stephenson 1930). Second, there are endemic earthworm species on some oceanic islands. The possibility of over-water dispersal has been debated extensively in the past, particularly in relation to the endemic species of the subantarctic islands (Michaelsen 1911; Stephenson 1930; Lee 1959). There is an opportunity to test the hypothesis of no over-water dispersal of earthworms in a general way and to use it, if supported, to argue that earthworms are biogeographical model organisms. By the term model organism in this context, I mean an organism with dispersal that has been so poor that its distribution can be viewed as determined entirely by past land connections and vicariance events that introduced saltwater barriers between land areas. If earthworm transoceanic dispersal has been negligible in history, then earthworm phylogenies can be used to unlock many earth historical riddles, such as the geological evolution of complex areas like the Caribbean Basin (Maury et al. 1990; Pindell and Barrett 1990) and the archipelagoes of Southeast Asia (Hall 1996, 1998). This application of phylogeny to earth history follows from the simplest form of allopatric speciation. The biota of the separated areas will have evolutionary histories that mirror the fragmentation history of the land. However, postfragmentation dispersal muddles the land area cladogram that can be derived from the hypothesized phylogenetic tree of the organisms. Therefore, a low- or no-dispersal taxon is preferable to one able to cross the barrier (Noonan 1988; Sober 1988). In spite of this rather obvious conclusion, most of the work to date in the field of historical biogeography has focused on relatively vagile organisms, such as reptiles, birds, and insects. Biogeography can be considered under many formats, and it is useful to define which kinds of biogeography are concerned. Ball (1975) described three phases of the science of biogeography. The first is the empirical or descriptive phase, in which basic data are collected. At this point, it is known where the various taxa are located, and there may be some synthesis, such as the definition of the classical biogeographic provinces (e.g., Nearctic, Ethiopian). Then, an attempt is made to explain the distributions, and a narrative phase is the result. A plausible story is constructed, one that seems to fit the evidence fairly well. Work to date on the subject of oceanic island earthworm distributions has achieved this much, leaving room for an analytical approach. As mentioned earlier, it appears that earthworms have rarely or never crossed saltwater (other than the saltwater-tolerant species inhabiting seashores). This could lead to the statement of an hypothesis that earthworms cannot survive, or do not have natural means of, transport across bodies of saltwater. This is testable, although it would be difficult to release test earthworms or cocoons by all possible means of conveyance (rafting vegetation, logs, on debris in violent cyclonic storms, and so on) and even harder to track their fates. It may be easier to test the predictions of the hypothesis. The hypothesis that earthworms cannot survive natural means of transport across bodies of saltwater predicts that land areas arising from midoceanic uplift of submerged rock or from volcanic eruption should not have earthworms. Obviously, this hypothesis arose from the observation that such islands either lack earthworms or harbor only earthworms that have been distributed widely by human activity. It can be tested by examining islands conforming to the modes of origin just mentioned and that have earthworm fauna about which little or nothing is known. The Lesser Antilles fit this description. These islands constitute a fairly strong test because they are close to one another and to potential sources of colonization, unlike the Hawaiian archipelago and the © 2004 by CRC Press LLC

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islands of the Mid-Atlantic Ridge. There is ample evidence that other fauna have colonized the Lesser Antilles from South America and from the Greater Antilles and other islands to the north (Savage 1982; Roughgarden 1990; Humphries and Parenti 1999), so it would not be particularly surprising to find another group of organisms doing the same. Another region with many islands, a mix of geological histories, and partly known or unknown earthworm faunas is Southeast Asia. There, the possibility of dispersal of earthworms from the Australian region or East Asia into the extensive archipelagoes is found. The general pattern of distribution of earthworms in the Lesser Antilles has been presented elsewhere (Fragoso et al. 1995). No earthworms other than peregrine species were found in the southern Lesser Antilles or the small volcanic islands north of Guadeloupe. Many new species endemic to their islands were found on Guadeloupe, Dominica, Martinique, and St. Lucia (James, unpublished data). Based on the number of genera and morphologically homogeneous groups of species within genera, at least nine successful dispersal events would be required to establish the current earthworm fauna of those four islands. If so, this is a very dense cluster of dispersal events in the middle of the archipelago, farthest from sources of colonization. However, earthworms have failed to colonize the Lesser Antilles by over-water dispersal from nearby land masses populated with indigenous earthworms. There is no evidence that the South American earthworm fauna has spread northward into the Lesser Antilles, or that the elements of the earthworm fauna of the Greater Antilles, particularly Puerto Rico and the Virgin Islands, have dispersed to the east and south. The Guadeloupe–St. Lucia axis poses a challenge to the conclusions that otherwise could be reached easily from the data from other islands. Without recourse to an analytical approach, the challenge cannot be met because hypotheses of dispersal are logically unfalsifiable. Any distribution could arise by postfragmentation dispersal if sufficient complexity of dispersal history is allowed. In this case, progress will come only from testing the hypothesis that the distributions are the result of vicariance because it is logically conceivable that such a hypothesis could be rejected. If, for the time being, it can be accepted that over-water dispersal can be ignored as a factor in earthworm distribution to islands, or between any two land areas separated by saltwater, then earthworms provide nearly ideal indicators of past land-area connections. Their evolutionary history should mirror these past connections because a severing of the land connection isolates populations of species. From the perspective of vicariance biogeography, these isolating or vicariating events are the primary factors of interest in the history of the flora and fauna of land areas. Dispersal is seen as a source of noise in the data, analogous to homoplasious character evolution (Sober 1988). Vicariance biogeography represents a system of formulating hypotheses that predict organismal distribution patterns based on underlying historical models (Wiley 1988). The underlying historical models have to do with creation of barriers to genetic exchange. Any species, earthworms included, could be affected by the fragmentation of a range because of climatic, geologic, or other processes. Thus, it should be possible to make a very good map relating the branch points on an earthworm cladogram and the separations of the land masses on which earthworm taxa occur. In the jargon of vicariance biogeography, one replaces the terminal taxa on the phylogenetic tree with the names of collection locations to make an area cladogram. A geological model provides a predicted area cladogram, which can then be compared with a biologically derived one. The success of this approach depends heavily on a very good phylogenetic analysis, on having the data set that is as free as possible from dispersal-induced noise, and on a proper choice of area units. Hausdorf (2002) argued for the use of biotic elements, rather than the more traditional areas of endemism, as the units of biogeographical analysis. An area of endemism is a region characterized by a number of taxa unique to the area. Hausdorf defined a biotic element as “a group of taxa whose ranges are significantly more similar to each other than to those of taxa of other such groups.” The difficulty in defining areas of endemism arises from dispersal: “The delimitation of areas of endemism is not problematic when species originate by vicariance and there is no dispersal” (Hausdorf 2002). The biotic element concept is a general rule, encompassing a range of natural possibilities, including sharply delimited areas of endemism inhabited by nearly dispersal-free taxa. © 2004 by CRC Press LLC

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Again, it is seen that earthworms potentially simplify the job of biogeographical analysis if their biotic elements have very little overlap. It should be clear that any analytical biogeography of this sort absolutely requires a prior phylogenetic analysis (Wiley 1988). Ball (1975) remarked pointedly that earthworms had nothing to offer to biogeographical analyses because the treatment of the data available then was too superficial. With two exceptions (Jamieson 1988; Jamieson et al. 2002), all the higher-level taxonomic research on earthworms is based on classical methodology (e.g., Sims 1980; Omodeo 1998, 2000). This is not to say that the classical method was bad, but that “new” methods, particularly cladistic analysis (Hennig 1966; Sudhaus and Rehfeld 1992), have emerged since the older generation of earthworm taxonomists was trained. Unfortunately, very few of those specialists have taken advantage of the revolution in systematics that has occurred. Techniques of analyzing data for phylogenetic reconstruction are evolving and diversifying rapidly. Although maximum parsimony remains very important, other tools are coming into more general use. Maximum likelihood and Bayesian methods have been extended from molecular to morphological data sets and to combined morphological and molecular data (Lewis 2001). These tools allow certain flexibilities in data treatment and coping with homoplasy (independent evolution of characteristics by lineages with common ancestor that did not have the characteristics in question) that will be useful in redressing the phylogenetic deficiencies mentioned here. Both Bayesian and maximum likelihood methods use underlying mathematical models of evolution and compare the results of the model with the actual data obtained (molecular sequences or morphological data) on the way to deriving an estimate of phylogenetic relationships. Jamieson et al. (2002) provided an example of the application of several methods of analysis to a phylogenetic problem, but it is just the beginning of what needs to be done. The field of earthworm systematics is long overdue for a reinvigoration and an influx of ideas and techniques that are new to many of us. Good classical systematics will never lose its fundamental value and is the foundation for the standards of publication that are often ignored by modern authors. However, a century of work by a small but dedicated group has failed to resolve such fundamentals as the definitions of families (e.g., Sims 1980). To date, only Jamieson (1988) and Jamieson et al. (2002) have offered a rigorous analysis of this problem. The higher-level taxonomics of earthworms seems to be a very good area in which to apply molecular data. Biogeographical contributions from earthworm phylogeny need not be confined to questions involving taxa separated by oceanic barriers. In general, earthworms are slow to disperse, mating is highly localized, and populations may be easily isolated. If the earthworm fauna of New Zealand is an example, it can be seen that, in mountainous regions, there can be a tremendous overall species diversity, although the number of species present at any one site may be small (Lee 1959). Thus, the possibility exists to investigate the phylogeny of the earthworms of a single land area in relation to many potential fragmentations and reannealings of ranges because of geological, hydrological, or climatological processes. Although the noise in the data may be greater than in the case of transoceanic questions, it should still be as good as or better than the data derived from other taxa widely cited in the biogeographical literature.

HOW THE EARTH’S HISTORY AFFECTS EARTHWORM DISTRIBUTIONS The question of how to learn about the history of the Earth from earthworm phylogeny and biogeography is worth asking only because there is good reason to believe that the history of life is linked in many ways to the history of the Earth. On the other hand, have earthworms been affected by historical processes and on what scales of time and space? Certainly, much of the distributions of higher earthworm taxa can be illuminated by a study of the history of the Earth’s crust. What appears in the present as anomalous, disjunct distributions can be understood as the

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results of the fragmentation of a continuous distribution. At other time scales, the impacts of other events of earth history on earthworm distributions can be examined. Glaciation is one example of such an event. The second subject I use as an illustration of biogeographical research looks at the distributions of Nearctic earthworms in relation to the maximal extent of the Wisconsin (Würm) glaciation and how the process of postglacial recovery affects the composition of earthworm fauna at different points from south to north. This may also provide some basis for figuring out the larger modern ecological consequences of the processes affecting earthworm distributions. Work to date on the subject of glaciation and earthworm distributions has resulted in limited distributional data (Gates 1977; Reynolds 1994, 1995) and a plausible explanation for the distribution patterns (Gates 1977; Reynolds 1994), still leaving a need for an analytical approach. Analytical biogeography requires that hypotheses be formulated and their predictions tested. For example, it could be stated as a hypothesis that earthworms cannot survive beneath glaciers. This is testable, although it would be difficult to release test earthworms at the underside of a glacier and even harder to go back and look for them. It may be preferable to examine the predictions of the hypothesis: There should be no native earthworms in areas recently uncovered by receding glaciers, regardless of preglacial history of earthworm presence on that site. However, even if the observed distributions are in accord with the above historical model (the glaciers made it impossible for earthworms to be there), this does not prove a causal link. In the present case, there are conflicting data and alternative explanations to consider. For instance, Schwert (1979) found a fossil earthworm cocoon in postglacial sediments of southern Ontario, Canada. This is well north of the extent of maximum glaciation, and the discovery is old enough to indicate that colonization must have been more rapid than conventional estimates allow (about 6 to 10 m per year from points of introduction). It also worth noting that periglacial conditions (permafrost, for example) also affected earthworm distributions, such that simply marking the locations of terminal moraines is not sufficient to tell the whole picture (Schwert 1977, 1990). Finally, semiaquatic species have such as Eisenoides lönnbergi natural distributions in glaciated areas in the eastern United States (Schwert and James, unpublished data). It inhabits wetlands and so may have different propagule dispersal mechanisms than terrestrial worms. Returning to the alternative hypotheses, it is conceivable that modern-day earthworm species in North America were not capable of surviving the climate any farther north than they presently occur. Second, it is conceivable that North American earthworms were present farther north before the colonization of North America by settlers from Europe, but that European earthworms (or some combination of European earthworm invasion and habitat destruction) eliminated the native species by competition. An additional hypothesis needs to be made that climate affects the outcome of the process of competition, rendering it possible for native species to persist in the southerly areas. To meet these complaints against what seems a simple and obvious narrative explanation, I conducted experiments and made several altitudinal transects to test the hypothesis of climate limitation. Transects are comparative, not experimental, and so do not constitute a strong class of evidence. To address the climate question, I transplanted Eisenoides carolinensis and several Diplocardia species from Pennsylvania and southern Iowa, respectively, to forest and grassland sites in northern Minnesota in 1990. All the species survived the two winters of the experimental period and were still present on the site in 1994. Juveniles were found in both Minnesota sites, indicating that reproduction had taken place. Similarly, on altitudinal transects in the Appalachian Mountains of Virginia, West Virginia, and North Carolina, native earthworms were found in the highest elevations attainable, in vegetation zones characteristic of low elevation areas much farther north and presently not occupied by native species in those northern locations. Thus, climate limitation (and vegetation type) has been eliminated for the species in question and is unlikely to be a significant determinant of the northern boundary of native North American earthworm distributions. The evidence favors abiotic historical factors to explain modern North American earthworm distributions: Glaciation removed earthworms from areas once covered by ice or underlain by © 2004 by CRC Press LLC

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permafrost in periglacial areas, and the earthworms have been very slow to diffuse northward again. I hesitate to use the word “disperse” because there appear to be few events in the normal course of the earthworm life cycle that promote dispersal, other than coming out to crawl over the soil surface after heavy rains. Using estimates of rates of spread obtained from the Netherlands (Marinissen and van den Bosch 1992), 10,000 years is enough time for some peregrine earthworms to advance 60 to 100 km. Thus, range expansions of 100 to 200 km would be expected because of the maximal extent of the icecap in the Northern Hemisphere. Bouché (1983) reported on the earthworm recolonization of glaciated France along the Rhône River, with some ecological interpretations of species interactions. Range expansion rates may differ among earthworm species for various reasons, including modes of reproduction and ecological niche. For example, epigeic species may have a higher natural rate of diffusion than do oligohumic endogeic species. If this is true, this would result in something analogous to a chromatographic fractionation of the earthworm fauna. Should such a pattern emerge, earthworm ecological functions would be represented differently in the natural vegetation of locations along north-south transects crossing the limits of native earthworms. This should have some observable impact on ecosystem processes. The extreme case would be in the northern regions, where no earthworms are present and forest litter layers are quite deep. The contribution that I have tried to make to the question of how glaciation (or any other historical factor important to a species distribution) affects earthworm distributions has been to remove the question from the narrative domain and to formulate and test hypotheses. Those hypotheses tested were either the predictions of the central hypothesis (in the present case, that glaciation made impossible the occupation of ice-covered lands) or the alternatives to the central hypothesis. Although negative data are never as satisfying as positive data and can be overturned by a single positive datum, broad-scale events are often amenable to investigation only by a process of elimination of alternatives. Much earthworm biogeography (e.g., Reynolds 1995) (see Chapter 4 this volume) has been directed to the distributions of peregrine species. Completely different historical agents, such as patterns of human migration and horticultural and agricultural trade routes, would seem to be involved. Totally different ecological concerns emerge: Are absences of earthworm species because of historical factors alone or because of incompatibilities between sites and species? Are the present distributions of the peregrine earthworm species determined by their transport history or by ecological factors? To what extent is habitat disturbance involved in successful establishment of peregrines? Does it make sense to talk of species associations when the species found together may have nothing more in common than a collection of travelers in a train station? Are there forms of data, such as molecular data, that will allow the origins of populations of peregrine species to be traced to their areas of origin? The task of bringing rigorous scientific methodology to this topic will require much insight and may yield further insights into the connections between ecology and human history. Earthworm systematics has been making primary contributions to the work of people whose interests lie in ecology, agriculture, and other fields in need of a coherent classification of earthworm species. This is caused in part by the service functions of the field and a shortage of earthworm taxonomic specialists. However, using the new types of data and analytical methods discussed here, it should soon be possible to move toward providing the robust phylogenies required for good biogeography. The potential for contributions to the understanding of the Earth’s history and its interactions with the history of the biota is great. There may be few other terrestrial taxa so widely distributed, so ancient, and so amenable to collection and study as earthworms. When the high degree of endemicity of earthworm species also is considered, it is surprising that scientists working on biogeographical questions have not already taken more advantage of the marvelous research opportunities provided in earthworm systematics and biogeography.

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REFERENCES Ball, I. 1975. Nature and formulation of biogeographic hypotheses, Syst. Zool., 24:407–430. Bouché, M.B. 1983. The establishment of earthworm communities, in J.E. Satchell, Ed., Earthworm Ecology, Chapman & Hall, London, U.K. Darwin, C. 1881. The Formation of Vegetable Mould Through the Action of Worms with Observations on Their Habits, Murray, London. Fragoso, C., S.W. James, and S. Borges. 1995. Native earthworms of the north neotropical region: current status and controversies, in P. Hendrix, Ed., Earthworm Ecology and Biogeography in North America, CRC Press, Boca Raton, FL, pp. 67–116. Gates, G.E. 1969. On the earthworms of Ascension and Juan Fernandez Islands, Breviora, 323:1–4. Gates, G.E. 1977. More on the earthworm genus diplocardia, Megadriogica, 3:1–48. Hall, R. 1996. Reconstructing Cenozoic Southeast Asia, in R. Hall and D.J. Blundell, Eds., Tectonic Evolution of Southeast Asia, Geological Society of London Special Publication, London, pp. 106, 153–184. Hall, R. 1998. The plate tectonics of Cenozoic Southeast Asia and the distribution of land and sea, in R. Hall and J.D. Holloway, Eds., Biogeography and Geological Evolution of Southeast Asia, Backhuys Publishers, Leiden, the Netherlands, pp. 99–131. Hausdorf, B. 2002. Units in biogeography, Syst. Biol., 51:648–652. Hennig, W. 1966. Phylogenetic Systematics, University of Illinois Press, Urbana, IL. Hopp, H. 1946. Earthworms fight erosion, too, Soil Cons., 11:252–255. Humphries, C.J. and L.R. Parenti. 1999. Cladistic Biogeography, Oxford University Press, Oxford, U.K. James, S.W. 1996. Nine new species of dichogaster (Oligochaeta: Megascolecidae) from Guadeloupe (French West Indies), Zool. Scripta, 25(l):21–34. Jamieson, B.G.M. 1981. Historical biogeography of Australian Oligochaeta, in A. Keast, Ed., Ecological Biogeography of Australia, W. Junk, The Hague, the Netherlands, pp. 887–921. Jamieson, B.G.M. 1988. On the phylogeny and higher classification of the Oligochaeta, Cladistics, 4:367–401. Jamieson, B.G.M., S. Tillier, A. Tillier, J.-L. Justine, E. Ling, S. James, K. McDonald, and A.F. Hugall. 2002. Phylogeny of the Megascolecidae and Crassiclitellata (Annelida, Oligochaeta): combined vs. partitioned analysis using nuclear (28S) and mitochondrial (12S, 16S) rDNA, Zoosystema, 24(4):707–734. Lavelle, P. and A. Martin. 1992. Small-scale and large-scale effects of endogeic earthworms on soil organic matter dynamics in the humid tropics, Soil Biol. Biochem., 24:1491–1498. Lee, K.B. 1959. The Earthworm Fauna of New Zealand, New Zealand Department of Scientific and Industrial Research Bulletin 130, Wellington, New Zealand. Lewis, P.O. 2001. A likelihood approach to estimating phylogeny from discrete morphological character data, Syst. Biol., 50(6):913–925. Liebherr, J.K. 1988. General patterns in West Indian insects, and graphical biogeographic analysis of some circum-Caribbean Platynus beetles (Carabidae), Syst. Zool., 37:385–409. Lovelock, J. 1988. The Ages of Gaia: A Biography of Our Living Earth, Norton and Company, New York. Marinissen, J.C.Y. and F. van den Bosch. 1992. Colonization of new habitats by earthworms, Oecologia, 91:371–376. Maury, R.C., G.K. Westbrook, P.E. Baker, Ph. Bouysse, and D. Westercamp. 1990. Geology of the Lesser Antilles, in G. Dengo and J.E. Case, Eds., The Caribbean Region, Vol. H, The Geology of North America, Geological Society of America, Boulder, CO, pp. 141–165. Michaelsen, W. 1911. Zur Kenntnis der Eodrilaceen und ihrer Verbreitungsverhältnisse Zool, Jahrb. Abt. f. Syst., 30:527–572. Michaelsen, W. 1933. Die Oligochaetenfauna Surinames mit Erörterung der verwandtschaftlichen Beziehungen der Octochatinen, Tijdschr. Ned. Dierk. Vereen., 3:112–131 Nakamura, M. 1990. How to identify Hawaiian earthworms, Chuo Univ. Res. Notes, 11:101–110. Noonan, G.R. 1988. Biogeography of North American and Mexican insects, and a critique of vicariance biogeography, Syst. Zool., 37:366–384. Nooren, C.A.M., N. van Breemen, J.J. Stoorvogel, and A.G. Jongmans. 1995. The role of earthworms in the formation of sandy surface soils in a tropical forest in Ivory Coast, Geoderma, 65:135–148. Omodeo, P. 1998. History of Clitellata, Ital. J. Zool., 65:51–73. Omodeo, P. 2000. Evolution and biogeography of megadriles (Annelida, Clitellata), Ital. J. Zool., 67:179–201.

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Pindell, J.L. and S.F. Barrett. 1990. Geological evolution of the Caribbean region; a plate-tectonic perspective, in G. Dengo and J.E. Case, Eds., The Caribbean Region, Vol. H, The Geology of North America, Geological Society of America, Boulder, CO, pp. 405–432. Reynolds, J.W. 1994. The distribution of the earthworms (Oligochaeta) of Indiana: a case for the postQuaternary introduction theory for megadrile migration in North America, Megadrilogica, 5:13–32. Reynolds, J.W. 1995. Status of exotic earthworm systematics and biogeography in North America, in P. Hendrix, Ed., Earthworm Ecology and Biogeography in North America, CRC Press, Boca Raton, FL, pp. 1–27. Rosen, D.E. 1985. Geological hierarchies and biogeographical congruence in the Caribbean, Ann. Mo. Bot. Gard., 72:636–659. Roughgarden, J. 1990. Origin of the eastern Caribbean: data from reptiles and amphibians, in D. LaRue and G. Draper, Eds., Trans. 12th Caribbean Conf. St. Croix, USVI Miami Geological Society, Miami, FL, pp. 10–26. Savage, J.M. 1982. The enigma of the Central American herpetofauna: dispersals or vicariance? Ann. Missouri Bot. Gard., 69:464–547. Schwert, D.P. 1977. The first North American record of Aporrectodea icterica (Savigny 1826) (Oligochata: Lumbricidae), with observations on the colonization of exotic earthworm species in Canada, Can. J. Zool., 55(1):245–248. Schwert, D.P. 1979. Description and significance of a fossil earthworm cocoon (Oligochata: Lumbricidae) from postglacial sediments in southern Ontario, Can. J. Zool., 57(7):1402–1405. Schwert, D.P. 1990. Lumbricidae, in D. L. Dindal, Ed., Soil Biology Guide, John Wiley & Sons, New York, pp. 341–356. Sharpley, A.N. and J.K. Syers. 1977. Seasonal variation in casting activity and in the amounts and release to solution of phosphorus forms in earthworm casts, Soil Biol. Biochem., 9:227–231. Sims, R.W. 1980. A classification and the distribution of earthworms, suborder Lumbricina (Haplotaxida Oligochaeta), Bull. Br. Mus. (Nat. Hist.) Zool. Ser., 39(2):103–124. Sober, E. 1988. The conceptual relationship of cladistic phylogenetics and vicariance biogeography, Syst. Zool., 37:245–270. Stephenson, J. 1930. The Oligochaeta, Clarendon Press, Oxford, U.K. Sudhaus, W. and K. Rehfeld. 1992. Einführung in die Phylogenetik und Systematik, Gustav Fischer, New York. Talavera, J.A. 1990. Claves de identificacion de las lombrices de tierra (Annelida Oligochaeta) de Canarias, Vieraea, 18:113–119. Wiley, E.O. 1981. Phylogenetics, Wiley Interscience, New York. Wiley, E.O. 1988. Parsimony analysis and vicariance biogeography, Syst. Zool., 37:271–290. Zwickl, D.J. and D.M. Hillis. 2002. Increased taxon sampling greatly reduces phylogenetic error, Syst. Biol., 51(4):588–598.

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Status of Earthworm 4 The Biogeography, Diversity, and Taxonomy in North America Revisited with Glimpses into the Future John W. Reynolds Oligochaetology Laboratory, Kitchener, Ontario, Canada

CONTENTS Introduction ......................................................................................................................................63 Earthworm Biogeography, Diversity, and Taxonomy .....................................................................65 Biogeography ............................................................................................................................65 North America .................................................................................................................65 Other Countries ...............................................................................................................67 Diversity ....................................................................................................................................67 Taxonomy..................................................................................................................................67 Presentations at the International Earthworm Ecology Symposia..................................................68 Future Trends and Research Imperatives in Earthworm Taxonomy...............................................69 Training of Earthworm Taxonomists........................................................................................69 Earthworm Parthenogenesis and Effects on Taxonomy...........................................................69 Earthworm Surveys...................................................................................................................69 Earthworm Life Histories .........................................................................................................70 Modern Earthworm Techniques................................................................................................70 Use of Earthworms for Waste Management ............................................................................70 Earthworms for Environmental Monitoring .............................................................................70 Plain Language and Less Esotery ............................................................................................70 References ........................................................................................................................................71

INTRODUCTION A discussion of the distribution of earthworms in North America Reynolds (1994a) included the three steps of biogeographical theory. The first step is descriptive or faunistics, the gathering of facts and identification and enumeration of animals in an area, including research and literature surveys. This step, although seldom fully achieved, seeks to present a clear distributional picture of all the animals in all areas. The second step is the classification of data, grouping of the distributional data according to as many different points of view as a particular investigation may

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find necessary, including comparison of the distributional ranges of phylogenetically related groups. This step also includes analyzing the fauna of a geographically, ecologically, or historically uniform area. The third and final step, causal analysis, tries to explain the reasons for the present distribution as identified. It is from this step that the principles of biogeography emerge. At present, in North America, with megadrile earthworms the situation is probably somewhere between the first two stages (Reynolds, 1975a, 1976b, 1994a). Studies of the status quo of animal distributions and of the grouping of these distributions according to different principles are investigations of momentary phenomena. Faunistic and regional zoogeography is basically static zoogeography because present distributions are the result of processes that have moved the animals in space, resulting in change of the distributional picture with the passage of time. Thus, causal zoogeography can be identical to dynamic zoogeography because causations of distributions are dynamic processes. As long as the inquiry probes the reasons for the arrival and colonization of a species in a certain area, the investigation is within the field of dynamic zoogeography. For example, the lack of an animal species in a certain area may be caused by ecological (it cannot exist there) or zoogeographical (it has not arrived there) reasons. The answers pertaining to dynamic zoogeography are often ecological, but the ecological cause is followed by movement of individuals or populations and of the subject organism. When the ecological cause is followed by a positive or negative response of the organism or population, and when a spatial shift is not involved, the whole phenomenon falls within the sphere of ecology. This, however, is usually assigned to the taxonomist because evolutionary changes are most easily discerned by noting morphological changes. Lindroth (1962) expressed the relationship of these two disciplines as follows: Zoogeography always depends on taxonomy to know what to study, but taxonomy also depends on zoogeography because geographic speciation is the accepted norm of the formation of its basic working unit. Every animal species originated from a few ancestors in a limited area; if a particular species is now widespread, it must of necessity have reached parts of its present range at an earlier period. The first aspect of dynamic zoogeography pertains to dispersal. If the details of dispersal processes are known, much about the presence or absence of animals can be explained. Dispersal may result as a by-product of other important phenomena, belonging to the biological habits of the animal, or it can result from distinct, adaptive characteristics of the species that directly assist dissemination into wider areas. Although every animal species has a capability to migrate, dispersing individuals must find suitable areas in which to settle and reproduce through many generations. When the process of settling or colonization is studied, the ecological factors that make the existence of a species possible in a given area must be scrutinized, as well as the adaptations and limitations of a species, such as structural, physiological, behavioral, or population dynamic properties that enable it to initiate a new population and survive (successfully) in the newly colonized area. Factors of dispersal as well as factors of existence in an area can influence the size, extent, and dynamism of the distributional range of the animal. James (1995) (see Chapter 3 this volume) uses a coincidently similar approach when he elaborates on the concepts of Ball (1975). Ball described three phases of biogeographic methodology: (1) empirical or descriptive, (2) narrative, and (3) analytical. James concentrates his discussion on two main points: (1) deoxyribonucleic acid (DNA) sequence information and (2) biogeographical research related to glaciation. I have very limited personal experience with DNA research, but for more than 2 decades (Reynolds 1973a) I have advocated this approach but have never had the opportunity for follow-up. It now appears that James, his colleagues, and their students have the means to do so, and earthworm ecology and biology should benefit greatly from the results of their work. I have devoted considerable time and activity to the second topic over the years. My earthworm research has concentrated on presenting wide-ranging formal surveys in North America and

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elsewhere (Table 4.1). James and others are going back to some of these areas and, with knowledge of the species that should be present, are conducting more detailed studies (e.g., using transects to assess distributions).

EARTHWORM BIOGEOGRAPHY, DIVERSITY, AND TAXONOMY In this section, I discuss some of the advances and innovations in oligochaetology during the past 2 years, particularly in those areas in which I have participated either as a journal referee or an editor, and for which I know, sometimes more than a year or two in advance, what will be appear in the literature regarding earthworm biogeography, diversity, and taxonomy. I summarize some of my relevant impressions on these developments.

BIOGEOGRAPHY North America Canada Considerable advances have been made in the distributional reports of earthworm taxa for many regions of North America. The first complete picture of the megadrile earthworm distributional patterns in the Province of Quebec was presented in the early 1990s (Reynolds and Reynolds 1992). Several earlier reports (Reynolds 1975a,b,c,d; 1976a,c,d) presented segments of the Quebec earthworm distribution patterns, but it was not until later (Reynolds and Reynolds 1992) that distribution over the whole province was considered. In this report, 19 species were recorded for the province, and Sparganophilus eiseni was reported for the first time from two widely separated collections more than 700 km east of any previously known records. Scheu and McLean (1993) presented the first report of the earthworm distributions in wideranging areas of southern Alberta. In their article, they reported eight species, with Lumbricus rubellus reported for the first time from the province. They also included considerable discussion on the ecological aspects of the species in the area. One of the more interesting and surprising reports was by two Russian scientists working in Siberia and the Yukon (Berman and Marusik 1994). Until their paper appeared, there were no published reports on earthworms from these Canadian territories. They found Bimastos parvus, Dendrobaena octaedra, and Dendrodrilus rubidus in these far north areas. They presented their views in great depth on earthworm migrations, isolated populations, and earthworm introductions in the far north to explain why the Siberian megadrile earthworm fauna is lacking in the Yukon. Subsequent papers (Reynolds and Moore 1996; Reynolds 2003b,c) have added to the earthworm species list of the Northwest Territories and the new territory of Nunavut in northern Canada. United States There has been considerable expansion in the knowledge of earthworm distributions in the United States. There were 37 species recorded (Reynolds 1994b) in the state of Indiana, 10 of which were reported for the first time. In North Carolina, Reynolds (1994f) reported 42 species, of which 10 were reported for the first time from the state. In Virginia (Reynolds 1994h), another Atlantic coastal state, 37 species were reported, with 6 species recorded for the first time for the state. Later in that year, the earthworm distributions in two more southern states, Florida (Reynolds 1994e) and Mississippi (Reynolds 1994g), were published. In the state of Florida, there were 51 species, and 3 subspecies were reported from 8 families. Of these species, 7 were reported for the first time from Florida. In the case of Mississippi, 27 species were recorded, and 10 of these were reported for the first time for the state. The earthworm distributional patterns of several other states and provinces are currently in preparation (Reynolds and Wetzel 2004b). The scientific and common names plus the distributional ranges for all the groups in the Clitellata and Aphanoneura have been described (Coates et al. 1995). © 2004 by CRC Press LLC

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TABLE 4.1 Regional Earthworm Surveys in North America Number of Species

Number of Units (%) Surveyed

Alabama Alberta

28 12

75 35

Arkansas Bermuda Colorado Connecticut Delaware Florida Georgia Illinois Indiana Louisiana Manitoba Maryland Massachusetts Michigan Minnesota Mississippi Missouri Montana New Brunswick

21 9+ 14 21 14 54 42 32 37 17 12 26 21 20 15 27 21 8 15

29 100 51 100 100 85 75 45 100 100 n/a 100 100 64 59 77 27 14 100

New York Newfoundland and Labrador North Carolina North Dakota Northwest Territories Nova Scotia Nunavut Ohio Ontario Oregon Prince Edward Island Quebec (south shore) Quebec (north shore) Rhode Island Saskatchewan South Carolina South Dakota Tennessee Virginia Washington

20 11 42 5 2 15 2 22 19 25 12 15 19 14 6 26 3 41 37 22

47 n/a 79 15 n/a 100 n/a 63 96 70 100 100 86 100 n/a 100 3 100 77 40

Wyoming Yukon

12 3

71 100

Region

n/a, not applicable

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Ref. Reynolds 1994c Scheu and McLean 1993; Reynolds and Clapperton 1996 Causey 1952, 1953 Reynolds and Fragoso 2004 Reynolds and Reeves 2004 Reynolds 1973b Reynolds 1973c Reynolds 1994e Reynolds in prep. Harman 1960 Reynolds 1994b Harman 1952; Gates 1965, 1967 Reynolds 2000a Reynolds 1974a Reynolds 1977b Snider 1991 Reynolds et al. 2002 Reynolds 1994g Olson 1936 Reynolds 1972 Reynolds 1976d, 2001b; Reynolds and Christie 1977; McAlpine et al. 2001 Olson 1940; Eaton 1942 Reynolds 2000b Reynolds 1994f Reynolds 1978a Reynolds and Moore 1996; Reynolds 2003b Reynolds 1976b Reynolds 2003b,c Olson 1928, 1932 Reynolds 1977a MacNab and McKey-Fender 1947; Fender 1985 Reynolds 1975d Reynolds 1975c,d,e, 1976a Reynolds and Reynolds 1992 Reynolds 1973d, 2002 Reynolds and Khan 1999 Reynolds 2001b; Reynolds and Reeves 2004 Gates 1979 Reynolds et al. 1974; Reynolds 1977c,d, 1978b Reynolds 1994h Altman 1936; Fender 1985; MacNab and McKey-Fender 1947 Reynolds 2004 Berman and Marusik 1994

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The first attempt to present a regional review of the earthworm distributions in continental North America was in the 1970s (Reynolds 1975a, 1976b). During the past 20 years, there have been many advances and additional available data for many regions (states and provinces) in North America. The first summary of regional earthworm surveys in North America was made (Reynolds et al. 1974). Since that time, there have been wide-ranging earthworm collections in North America, resulting in many publications on the distribution of various species of earthworms in North America and Mexico (Fragoso et al. 1994). An updated version of that table appeared in 1994 and 1995 (Reynolds 1994c, 1995), and the most recent revision is given in Table 4.1. Other Countries There has been considerable expansion of knowledge of megadrile earthworm distributions on other continents as well. For instance, de Mischis (1992, 1993) expanded the distributional records for Argentina. The first survey of the earthworms of Swaziland appeared (Reynolds 1993). This survey reported seven species from that small African country. Earthworm surveys undertaken during 1992 and 1993 contributed to the first survey of the earthworms of Bangladesh (Reynolds 1994d), where 14 species were reported, and there were speculations on the possible presence of an additional 28 species. The first survey and report of earthworms from Belize indicated three species: Dichogaster bolaui, Pontoscolex corethrurus, and a new species, Eodrilus jenniferae (Reynolds and Righi 1994). Additional earthworm samples were obtained subsequently (Reynolds and Guerra 1994). Two additional examples of original surveys were in San Andres and Contadora Island (Reynolds and Reynolds 2002a,b) and Gough Island, South Atlantic Ocean (Reynolds et al. 2002).

DIVERSITY There have been considerable data published on earthworm diversity in the Americas. The current status of various aspects of earthworm research on this continent has been summarized (Fragoso et al. 1994; Fender 1995; Hendrix 1995; James 1995; Reynolds 1995). Bøgh (1992) published an interesting and innovative article, “Identification of Earthworms (Lumbricidae): Choice of Method and Distinction Criteria.” In this article, he discussed the use of electrophoretic techniques in the identification of earthworm species. One of the immediate benefits of this approach is the potentially more accurate determination of juveniles and fragments of the caliginosa complex that the late Dr. Gordon Gates and I have long advocated. The morphological criteria developed by Gates (1972a) that were used in many of my surveys throughout North America were supported by Bøgh (1992). I (1994a) prepared a summary in Earthworms of the World. This included discussions of global distributions, barriers to migration, habitat requirements, and functions of earthworms in the soil.

TAXONOMY For a field of science as limited as oligochaetology, it is fortunate there are books that combine all the description citations and type depositions of earthworms: the Nomenclatura Oligochaetologica and its three supplements (Reynolds and Cook 1976, 1981, 1989, 1993). The third supplement (Supplementum Tertium), which recorded new taxa found up to December 31, 1992, suggested that 739 earthworm genera, 40 subgenera, and 7254 species have been described. A fourth supplement (Reynolds and Wetzel 2004a) described more than 1048 new species. Some of the most exciting discoveries regard the presence of nearctic species in the far reaches of North America, areas where they were not previously recorded. These include Fender’s (1985) and McNab and McKey-Fender’s (1947) work on Bimastos and Arctiostrotus in the northwestern United States and southwestern British Columbia (Canada) as well as James’s (1995) discoveries of Argilophilus and Diplocardia in the southwestern United States.

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The status of exotic earthworm systematics and biogeography in North America was reviewed (Reynolds 1995). This review dealt mainly with the Lumbricidae and excluded the nearctic genera Bimastos and Eisenoides. The section on historical perspectives traced the associations and classiù ´ (1991) and discussed the lumping of fications of authors from Linnaeus (1758) through Mrsic certain species inter- and intragenerically. These frequent realignments by taxonomists have created many nomenclatural and taxonomic problems for ecological researchers for many years. The review also included discussions on earthworm contributions and activity in North America, which is often lacking in some European publications. Some portions of the biogeographical section are summarized here and in Table 4.1. James (1995) addressed the issue of systematics, biogeography, and ecology of nearctic earthworms from the eastern, central, southern, and southwestern United States. He dealt with five families, two of which are monospecific: Lutodrilidae (Lutodrilus multivesiculatus) and Komarekionidae (Komarekiona eatoni). The family Sparganophilidae is monogeneric with 12 species. The Lumbricidae were included mainly in the two nearctic genera Bimastos and Eisenoides, with nine and two species, respectively. James’s version of the Megascolecidae (Acanthodrilidae sensu Gates and Reynolds) was restricted to the genus Diplocardia and its 42 species. James’s ecological section included discussions of earthworm population studies and emphasized the lack of earthworm community studies and economic applications. Fender (1985, 1995) provided an ecological overview of native earthworms of the Pacific Northwest. He dealt with a group of earthworms (Megascolecidae, Argilophilini) unfamiliar to many earthworm scientists. Fender stated that the Pacific Northwest possesses a “rich, varied, and interesting, but highly underreported” earthworm fauna. There are a vast number of taxa to be described and data that have been collected and are waiting analysis. Fender discussed the historical biogeography, ecology, and variation of this group of little-known oligochaetes. Fragoso et al. (1994) reviewed the native earthworms of the north neotropical region and their current status and controversies. Their lists of earthworm species and the authors cited may not be familiar to many North American earthworm scientists, although the early works cited in their historical perspective included researchers such as Beddard, Benham, Cognetti, Eisen, Gates, and Michaelsen, who are well-known earthworm taxonomists. They also reviewed species of earthworms from Mexico (Fragoso et al. 1994). The authors explained in great detail the biogeography, ecology, and taxonomy of the earthworms from their region. In their list of the earthworm fauna of the north neotropical region, Fragoso et al. (1995) indicated the absence of any reports on earthworms from Belize (formerly British Honduras). I had the opportunity to work in this small Central American country and collect earthworms. The results of these collections were published in a paper by Gilberto, and two well-known species (Dichogaster bolaui and Pontoscolex corethrurus) are reported together with a new species, Eodrilus jenniferae. In addition to the description of a new species and the first record of earthworms from Belize, this article includes a discussion on the retention of eodrilus and diplotrema as separate and distinct genera based on four criteria. We also continued with suggestions for distinguishing four closely related genera: eodrilus, notiodrilus, acanthodrilus, and microscolex. Our solution for distinguishing these genera will undoubtedly spark discussion in the future, and as a result of these debates, earthworm taxonomists should come closer to agreement.

PRESENTATIONS AT THE INTERNATIONAL EARTHWORM ECOLOGY SYMPOSIA There have been seven International Earthworm Ecology Symposia, at Grange-over-Sonids, England (1981); Bologna, Italy (1985); Hamburg, Germany (1987); Avignon, France (1990); Columbus, OH, United States (1994); Vigo, Spain (1998); and Cardiff, Wales (2002). Each was

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attended by 200 to 300 earthworm scientists, including many international systematists. All the symposia included sections on taxonomy, distributions, and biogeography of earthworms.

FUTURE TRENDS AND RESEARCH IMPERATIVES IN EARTHWORM TAXONOMY One of the major problems for earthworm taxonomy in North America and elsewhere has been the paucity of earthworm systematists and their geographical isolation. Gordon Gates, one of the most productive taxonomists, died in 1987. Although he began publishing on earthworm systematics in 1926, it was not until the latter part of his career that he devoted much time to describing the Lumbricidae and other earthworms of North America. Before 1950, taxonomists included Frank Smith (publication period 1985 to 1937) and Henry Olson (publication period 1928 to 1940), who made significant contributions to earthworm taxonomy and distributions, respectively. In the past few decades, Dorothy McKey-Fender and William Fender have concentrated on the taxonomy and distribution of the native and exotic earthworm fauna of the west coast of North America, and Sam James has investigated the endemic species of the southeastern and plains areas of the United States. Since 1972, I have collected earthworms widely throughout North America, and the results of these collections have been published primarily as distribution data, with relatively minimal contributions to systematics sensu stricto. The Nomenclatura Oligochaetologica series put together in a single source the essential basic reference data for anyone involved in the taxonomy and nomenclature of earthworms or who need a ready, up-to-date reference list of species authorities have described. The fourth supplement (Reynolds and Wetzel 2004a) should be available at the end of 2004. I suggest the following priorities for future earthworm research.

TRAINING

OF

EARTHWORM TAXONOMISTS

For more than 2 decades, it has been obvious that the scarcity of competent earthworm systematists and taxonomists was detrimental to progress in research by ecologists and others (Reynolds 1973a; Reynolds et al. 1974a,b). Institutions that normally employ taxonomists and encourage their development (e.g., museums, departments of agriculture, and universities) have not done so in North America. There is a need for a concerted effort to support this type of research before there are few or no earthworm taxonomists remaining. However, the large number of scientists actively working in various aspects of earthworm biogeography and taxonomy elsewhere in the world is most gratifying and bodes well for the future of the science.

EARTHWORM PARTHENOGENESIS

AND

EFFECTS

ON

TAXONOMY

A major exotic group of earthworms in North America (Megascolecidae and pheretimoid groups) has long been plagued with taxonomic problems, which have resulted from the incidence of widespread parthenogenesis among its species (Gates 1972). Parthenogenesis also occurs within the Lumbricidae. One study showed that localized populations of Octolasion tyrtaeum (Jaenike et al. 1980, 1982; Jaenike and Selander 1985) have exhibited parthenogenesis. Previously, taxonomic problems with some morphotypes of what is now recognized as Dendrodrilus rubidus may be attributed to parthenogenesis. The issue of parthenogenesis in earthworms has only recently received more attention and may have major impacts on earthworm taxonomy.

EARTHWORM SURVEYS It is obvious from the data in Table 4.1 that, in spite of what has already been achieved, there are major areas of North America in which there have been no earthworm surveys. In certain areas

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where native species still exist, there is still very considerable potential for discovery of new species. Any new species reported in the lumbricidae will most probably come from the two native genera bimastos and isenoides or even from a new genus. Additional new species might also be expected to occur in other nearctic genera, such as arctiostrotus, diplocardia, and komarekiona.

EARTHWORM LIFE HISTORIES It is amazing that, of the nearly 8000 oligochaete species described (Reynolds and Cook 1993), ecological and life history studies have been made on relatively few species, probably fewer than 20. Lee (1992) suggested that only about six lumbricid and six tropical species have been studied in sufficient detail to provide adequate life history information. However, Barois et al. (1999) provided basic ecological and life history data on 58 tropical earthworm species, but not all described in detail. Some of the information gathered on the ecology of common lumbricidae was obtained from a time when species lumping occurred; that is, descriptions were attributed to Allolobophora caliginosa, which now includes several species, so further studies may be needed for clarification.

MODERN EARTHWORM TECHNIQUES One technique that has been considered for years, but only recently had any evidence to support its potential in taxonomy, is electrophoresis. Using this technique, Bøgh (1992) illustrated that certain earthworm species were different (i.e., Aporrectodea tuberculata and Aporrectodea turgida were distinct species), and he demonstrated how to identify species from fragments. There is also considerable scope for use of DNA analyses in taxonomy (see Chapter 3, this volume).

USE

OF

EARTHWORMS

FOR

WASTE MANAGEMENT

The acceptance of organic waste recycling and vermicomposting using earthworms has gained increasing acceptance over the past 2 decades and is increasing rapidly in both industrialized and nonindustrialized countries (see Chapter 18, this volume). In North America, the research is restricted to only a few species (Eisenia foetida, Eudrilus eugeniae, Perionyx excavatus), but with almost 4000 megadrile species available, the search should be for additional species that may be harnessed to assist the decomposition and transformation of waste products into useful materials (see Chapters 18, 19, and 20 this volume).

EARTHWORMS

FOR

ENVIRONMENTAL MONITORING

The ability of many species of earthworms to accumulate heavy metals and various pesticides into their tissues offers opportunities to trace the movement of these materials in the soil. One aquatic microdrile species, Tubifex tubifex, has been used for decades as a biological indicator in polluted waters. Eisenia andrei is used in a standardized terrestrial assay using artificial soil to assess the toxicity of pesticides and other chemicals (Edwards and Bohlen 1992). Research is needed on the toxicities uptake mechanisms, distribution, and concentration of chemicals in various types of earthworm tissues. In particular, interpretation of the findings on chemicals and earthworms as they relate to our daily lives is needed.

PLAIN LANGUAGE

AND

LESS ESOTERY

I have long advocated the necessity for scientific information to be more accessible to the general nonscientific community. Entomologists and ornithologists have advanced more rapidly because of the joint contributions of “amateurs” and general collectors. In the early part of this century, earthworm biology had one such person, the Hilderic Friend, who was turned aside and dismissed

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by the many earthworm specialists of the day. He was a good naturalist, but many of his contributions were ridiculed and discounted. If he had been encouraged, his work might have had a greater impact on oligochaetology as we know it today.

REFERENCES Altman, L.C. 1936. Oligochaeta of Washington, Univ. Wash. Publ. Biol., 4(1), 1–137. Ball, I.R. 1975. Nature and formulation of biogeographical hypotheses, Syst. Zool., 24(4), 407–430. Barois, I., Lavelle, D., Brossard, M., Tondoh, J., Martinez, M.A., Rossi, J.P., Senapati, B.K., Angeles, A., Fragoso, C., Jimenez, J.J., Decaens, T., Lattand, C., Kanyono, J., Blanchart, E., Chapuis, L., Brown, G., and Moreno, A. 1999. Ecology of earthworm species with large environmental tolerances and/or extended distributions, in Earthworms Management in Tropical Agroecosystems, Lavelle, P., Brussaard L., and Hendrix, P. Eds., CABI Publishing, Wallingford, U.K., pp. 57–86. Berman, D.I. and Marusik, Y.M. 1994. On Bimastos parvus (Oligochaeta: Lumbricidae) from Yukon Territory (Canada) with notes on distribution of the earthworms in northwest North America and northeast Siberia, Megadrilogica, 5(10), 113–116. Bøgh, P.S. 1992. Identification of earthworms (Lumbricidae): choice of method and distinction criteria, Megadrilogica, 4(10), 163–174. Causey, D. 1952. The earthworms of Arkansas, Proc. Ark. Acad. Sci., 5, 31–42. Causey, D. 1953. Additional records of Arkansas earthworms, Proc. Ark. Acad. Sci., 6, 47–48 de Mischis, C.C. 1992. The first record of the species Amynthas diffringens (Baird, 1869) (Oligochaeta: Megascolecidae) in the province of Cordoba (Argentina), Megadrilogica, 4(8), 143–144. de Mischis, C.C. 1993. A contribution to the knowledge of megascolecid fauna (Annelida, Oligochaeta) from the province of Cordoba, Argentina, Megadrilogica, 5(2), 9–12. Eaton, T.H. 1942. Earthworms of the northeastern United States: a key, with distribution records, J. Wash. Acad. Sci., 32(8), 242–249. Edwards, C.A., Ed. 1997. Proceedings of Fifth International Symposium on Earthworm Ecology, Soil Biol. Biochem., 29(3/4), 217–766. Edwards, C.A. and Bohlen, P.J. 1992. The effects of toxic chemicals on earthworms, Rev. Environ. Contam. Toxicol., 125, 23–99. Fender, W.B. 1985. Earthworms of the western United States. Part. I. Lumbricidae, Megadrilogica, 4(5), 93–129. Fender, W.B. 1995. Native earthworms of the Pacific Northwest: an ecological overview, in Ecology and Biogeography of Earthworms in North America, Hendrix, P.F., Ed., Lewis Publishers, Boca Raton, FL, pp. 53–66. Fragoso, C. and Fernández, P.R. 1994. Earthworms from southwestern Mexico. New Acanthodriline genera and species (Megascolecidae, Oligochaeta), Megadrilogica, 6(1), 1–12. Fragoso, C., James, S.W., and Borges, S. 1995. Native earthworms of the north neotropical region: current status and controversies, in Ecology and Biogeography of Earthworms in North America, Hendrix, P.F., Ed., Lewis Publishers, Boca Raton, FL, pp. 67–114. Gates, G.E. 1965. Louisiana earthworms. I. A preliminary survey, La. Acad. Sci., 28(1), 12–20. Gates, G.E. 1967. On the earthworm fauna of the Great American desert and adjacent areas, Gt. Basin Nat., 27(3), 142–176. Gates, G.E. 1972. Burmese earthworms. An introduction to the systematics and biology of megadrile oligochaetes with special reference to southeast Asia, Trans. Am. Philos. Soc., 62(7), 1–326. Gates, G.E. 1979. South Dakota does have earthworms! Megadrilogica, 3(9), 165–166. Harman, W.J. 1952. A taxonomic survey of the earthworms of Lincoln Parish, Louisiana, Proc. La. Acad. Sci., 15, 19–23. Harman, W.J. 1960. Studies on the Taxonomy and Musculature of the Earthworms of Central Illinois, Ph.D. dissertation, University of Illinois, Champaign, IL. Hendrix, P.F. 1995. Ecology and Biogeography of Earthworms in North America, Lewis Publishers, Boca Raton, FL. Jaenike, J. and Selander, R.K. 1985. On the co-existence of ecologically similar clones of parthenogenetic earthworms, Oikos, 44(3), 512–514. © 2004 by CRC Press LLC

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Jaenike, J., Ausubel, S., and Grimaldi, D.A. 1982. On the evolution of clonal diversity in parthenogenetic earthworms, Pedobiologia, 23(3–4), 304–310. Jaenike, J., Parker, E.D., and Selander, R.K. 1980. Clonal niche structure in the parthenogenetic earthworm Octolasion tyrtaeum, Am. Nat., 116, 196–205. James, S.W. 1995. Systematics, biogeography and ecology of earthworms from eastern, central, southern and southwestern USA, in Ecology and Biogeography of Earthworms in North America, Hendrix, P.F., Ed., Lewis Publishers, Boca Raton, FL, pp. 29–51. Kretzschmar, A., Ed. 1992. Proceedings of Fourth International Symposium on Earthworm Ecology, Soil Biol. Biochem., 24(12), 1193–1771. Lee, K.E. 1992. Some trends and opportunities in earthworm research or: Darwin’s children — the future of our discipline, Soil Biol. Biochem., 24(12), 1765–1771. Lindroth, C.H. 1962. Foreword, in Taxonomy and Geography, Nichols, D., Ed., Systemics Assoc. Publ., London, U.K., pp. 3–5. Linnaeus, C. 1758. Systema Naturae. Regnum Animale, 10th ed. MacNab, J.A. and McKey-Fender, D. 1947. An introduction to Oregon earthworms with additions to the Washington list, Northwest Sci., 21(2), 69–75. McAlpine, D.F., Reynolds, J.W., Fletcher, T.J., Trecartin, J.L., and Sabine, D.L. 2001. Megadrilogica, 8(10), 53–56. ÿ ´ N. 1991. Monograph on Earthworms (Lumbricidae) of the Balkans, Slovenska Akademija Znanosti Mrsic, Umetnosti, Ljubljana, Slovenia. Olson, H.W. 1928. The Earthworms of Ohio, with a study of their distribution in relation to hydrogen-ion concentration, moisture and organic content of the soil, Bull. Ohio Biol. Surv., 4(2), Bull. 17, 47–90. Olson, H.W. 1932. Two new species of earthworms for Ohio, Ohio J. Sci., 32, 192–193. Olson, H.W. 1936. Earthworms of Missouri, Ohio J. Sci., 36(2), 102–193. Olson, H.W. 1940. Earthworms of New York State, Am. Mus. Nov., 1090, 9. Reynolds, J.W. 1972. A contribution to the earthworm fauna of Montana, Proc. Mont. Acad. Sci., 32, 6–13. Reynolds, J.W. 1973a. Earthworm (Annelida, Oligochaeta) ecology and systematics, in Proc., Dindal, D.L., Ed., Proc. 1st Soil Microcommunities Conf., U.S. Atomic Energy Commission, National Tech. Inform. Serv., Springfield, pp. 95–120. Reynolds, J.W. 1973b. The earthworms of Connecticut (Oligochaeta: Lumbricidae, Megascolecidae and Sparganophilidae), Megadrilogica, 1(7), 1–6. Reynolds, J.W. 1973c. The earthworms of Delaware (Oligochaeta: Acanthodrilidae and Lumbricidae), Megadrilogica, 1(5), 1–4. Reynolds, J.W. 1973d. The earthworms of Rhode Island (Oligochaeta: Lumbricidae), Megadrilogica, 1(6), 1–4. Reynolds, J.W. 1974a. The earthworms of Maryland (Oligochaeta: Acanthodrilidae, Lumbricidae, Megascolecidae and Sparganophilidae), Megadrilogica, 1(11), 1–12. Reynolds, J.W. 1974b. Are oligochaetes really hermaphroditic amphimictic organisms? Biologist, 56(2), 90–99. Reynolds, J.W. 1975a. Die biogeographie van Noorde-Amerikaanse (Oligochaeta) noorde van Meksike — I, Indikator, 7(4), 11–20. Reynolds, J.W. 1975b. The earthworms of Prince Edward Island (Oligochaeta: Lumbricidae), Megadrilogica, 2(7), 4–10. Reynolds, J.W. 1975c. Les Lombricidés (Oligochaeta) de la Gaspésie, Québec, Megadrilogica, 2(4), 4–9. Reynolds, J.W. 1975d. Les Lombricidés (Oligochaeta) des Îles-de-la-Madeleine, Megadrilogica, 2(3), 1–8. Reynolds, J.W. 1975e. Les Lombricidés (Oligochaeta) de Î’Ile d’Orléans, Québec, Megadrilogica, 2(5), 8–11. Reynolds, J.W. 1976a. Die biogeographie van Noorde-Amerikaanse (Oligochaeta) noorde van Meksike — II, Indikator, 8(1), 6–20. Reynolds, J.W. 1976b. A preliminary checklist and distribution of the earthworms of New Brunswick, New Brunswick Nat., 7(2), 16–17. Reynolds, J.W. 1977a. The Earthworms (Lumbricidae and Sparganophilidae) of Ontario, Life Sci. Misc. Publ., Roy. Ont. Mus., 141 pp. Reynolds, J.W. 1977b. The earthworms of Massachusetts (Oligochaeta: Lumbricidae, Megascolecidae and Sparganophilidae), Megadrilogica, 3(2), 49–54 Reynolds, J.W. 1977c. The earthworms of Tennessee (Oligochaeta). II. Sparganophilidae, with the description of a new species, Megadrilogica, 3(3), 61–64.

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Reynolds, J.W. 1977d. The earthworms of Tennessee (Oligochaeta). III. Komarekionidae, with notes on distribution and biology, Megadrilogica, 3(4), 65–69. Reynolds, J.W. 1978a. A contribution to our knowledge of the earthworm fauna of North Dakota, Megadrilogica, 3(8), 148–149. Reynolds, J.W. 1978b. The earthworms of Tennessee (Oligochaeta). IV. Megascolecidae, with notes on distribution, biology and a key to the species in the state, Megadrilogica, 3(7), 117–129. Reynolds, J.W. 1993. On some earthworms from Swaziland (Oligochaeta: Glossoscolecidae, Megascolecidae, Microchaetidae and Octochaetidae), Megadrilogica, 5(1), 1–8. Reynolds, J.W. 1994a. The distribution of earthworms (Annelida, Oligochaeta) in North America, in Advances in Ecology and Environmental Science, Mishra, P.C., Behera, N., Senapati, B.K., and Guru, B.C., Eds., Ashish Publication, New Delhi, India, pp. 133–153. Reynolds, J.W. 1994b. The distribution of the earthworms (Oligochaeta) of Indiana: a case for the post quaternary introduction theory for megadrile migration in North America, Megadrilogica, 5(3), 13–32. Reynolds, J.W. 1994c. Earthworms of Alabama (Oligochaeta: Acanthodrilidae, Eudrilidae, Lumbricidae, Megascolecidae, Ocnerodrilidae and Sparganophilidae), Megadrilogica, 6(4), 35–46. Reynolds, J.W. 1994d. The earthworms of Bangladesh (Oligochaeta: Megascolecidae, Moniligasteridae and Octochaetidae), Megadrilogica, 5(4), 33–44. Reynolds, J.W. 1994e. Earthworms of Florida (Oligochaeta: Acanthodrilidae, Eudrilidae, Glossoscolecidae, Lumbricidae, Megascolecidae, Ocnerodrilidae, Octochaetidae and Sparganophilidae), Megadrilogica, 5(12), 125–141. Reynolds, J.W. 1994f. Earthworms of North Carolina (Oligochaeta: Acanthodrilidae, Komarekionidae, Lumbricidae, Megascolecidae, Ocnerodrilidae and Sparganophilidae), Megadrilogica, 5(6), 53–72. Reynolds, J.W. 1994g. Earthworms of Mississippi (Oligochaeta: Acanthodrilidae, Lumbricidae, Megascolecidae, Ocnerodrilidae and Sparganophilidae), Megadrilogica, 6(3), 17–29. Reynolds, J.W. 1994h. Earthworms of Virginia (Oligochaeta: Acanthodrilidae, Komarekionidae, Lumbricidae, Megascolecidae and Sparganophilidae), Megadrilogica, 5(8), 77–94. Reynolds, J.W. 1995. The status of exotic earthworm systematics and biogeography in North America, in Ecology and Biogeography of Earthworms in North America, Hendrix, P.F., Ed., Lewis Publishers, Boca Raton, FL, pp. 1–28. Reynolds, J.W. 2000a. A contribution to our knowledge of the earthworm fauna of Manitoba, Canada (Oligochaeta, Lumbricidae), Megadrilogica, 8(3), 9–12. Reynolds, J.W. 2000b. A contribution to our knowledge of the earthworm fauna of Newfoundland and Labrador, Canada (Oligochaeta, Lumbricidae), Megadrilogica, 8(2), 5–8. Reynolds, J.W. 2001a. The earthworms of New Brunswick (Oligochaeta: Lumbricidae), Megadrilogica, 8(8), 37–47. Reynolds, J.W. 2001b. The earthworms of South Carolina (Oligochaeta: Acanthodrilidae, Lumbricidae, Megascolecidae, Ocnerodrilidae and Sparganophilidae), Megadrilogica, 8(7), 25–36. Reynolds, J.W. 2002. Additional earthworm (Oligochaeta: Lumbricidae and Megascolecidae) records from Rhode Island, USA, Megadrilogica, 9(4), 21–27. Reynolds, J.W. 2003a. The earthworms (Oligochaeta: Lumbricidae) of Wyoming, USA, Megadrilogica, 9(6), 33–39. Reynolds, J.W. 2003b. First earthworm record from Nunavut, Canada and the second from the Northwest Territories, Megadrilogica, 9(6), 40. Reynolds, J.W. 2003c. A second earthworm species (Lumbricidae) from Nunavut, Canada, Megadrilogica, 9(8), 52. Reynolds, J.W. and Christie, D.S. 1977. Additional records of New Brunswick earthworms, New Brunswick Nat., 8(3), 25. Reynolds, J.W. and Clapperton, M.J. 1996. New earthworm records for Alberta (Oligochaeta: Lumbricidae) including the description of a new Canadian species, Megadrilogica, 6(8), 73–82. Reynolds, J.W. and Cook, D.G. 1976. Nomenclatura Oligochaetologica, a Catalogue of Names, Descriptions and Type Specimens of the Oligochaeta, University of New Brunswick, Fredericton, Canada. Reynolds, J.W. and Cook, D.G. 1981. Nomenclatura Oligochaetologica Supplementum Primum, Catalogue of Names, Descriptions and Type Specimens of the Oligochaeta, University of New Brunswick, Fredericton, Canada.

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Reynolds, J.W. and Cook, D.G. 1989. Nomenclatura Oligochaetologica Supplementum, A Catalogue of Names, Descriptions and Type Specimens of the Oligochaeta, New Brunswick Museum Monograph Series (Natural History), No. 8, Fredericton, New Brunswick, Canada, 37 pp. Reynolds, J.W. and Cook, D.G. 1993. Nomenclatura Oligochaetologica Supplementum Tertium, A Catalogue of Names, Descriptions and Type Specimens of the Oligochaeta, New Brunswick Museum Monograph Series (Natural History), No. 9, New Brunswick, Canada, 93 pp. Reynolds, J.W. and Fragoso, C. 2004. The earthworms (Oligochaeta) of Bermuda, Megadrilogica, 10, in press. Reynolds, J.W. and Guerra, C.A. 1994 Two species of earthworms newly reported from Belize, C.A. (Oligochaeta: Glossoscolecidae and Megascolecidae), Megadrilogica, 5(10), 122–124. Reynolds, J.W. and Khan, M.N. 1999. A contribution to our knowledge of the earthworm fauna of Saskatchewan, Canada, Megadrilogica, 7(12), 81–82. Reynolds, J.W. and Moore, S.M. 1996. Note on the first earthworm record from the Northwest Territories, Canada, Megadrilogica, 6(10), 96. Reynolds, J.W. and Reeves, W.K. 2004. Additional earthworm records (Oligochaeta: Acanthodrilidae, Lumbricidae, Megascolecidae, Ocnerodrilidae and Sparganophilidae) from South Carolina, USA, Megadrilogica, 9(12), 100–111. Reynolds, J.W. and Reynolds, K.W. 1992. Les vers de terre (Oligochaeta: Lumbricidae et Sparganophilidae) sur la rive nord du Saint-Laurent (Québec), Megadrilogica, 4(9), 145–161. Reynolds, J.W. and Reynolds, D.W. 2002a. Primeros datos de lombrices de tierra (Oligochaeta) de la Isla de San Andrés, Colombia, Megadrilogica, 8(6), 21–24. Reynolds, J.W. and Reynolds, D.W. 2002b. Primeros datos de lombrices de tierra (Oligochaeta) de la Isla de Contadora, Panamá, Megadrilogica, 9(1), 1–4. Reynolds, J.W. and Righi, G. 1994. On some earthworms from the Belize, C.A. with the description of a new species (Oligochaeta: Acanthodrilidae, Glossoscolecidae and Octochaetidae), Megadrilogica, 5(9), 97–106. Reynolds, J.W. and Wetzel, M.J. 2004a. Nomenclatura Oligochaetologica Supplementum Quartum, a Catalogue of Names, Descriptions and Type Specimens of the Oligochaeta, Illinois Natural History Survey Special Publication, in press. Reynolds, J.W. and Wetzel, M.J. 2004b. Terrestrial Oligochaeta (Annelida: Clitellata) in North America north of Mexico, Megadrilogica, 9(11), 71–99. Reynolds, J.W., Clebsch, E.E.C., and Reynolds, W.M. 1974. The Earthworms of Tennessee (Oligochaeta). I. Lumbricidae, Bull. Tall Timbers Res. Stn., 17, 1–133. Reynolds, J.W., Jones, A.G., Gaston, K.J., and Chown, S.L. 2002. The earthworms (Oligochaeta: Lumbricidae) of Gough Island, South Atlantic Ocean, Megadrilogica, 9(2), 5–15. Reynolds, J.W., Linden, D.R., and Hale, C.M. 2002. The earthworms of Minnesota (Oligochaeta: Acanthodrilidae, Lumbricidae and Megascolecidae), Megadrilogica, 8(12), 85–99. Scheu, S. and McLean, M.A. 1993. The earthworm (Lumbricidae) distribution in Alberta (Canada), Megadrilogica, 4(11), 175–180. Snider, R.M. 1991. Checklist and distribution of Michigan earthworms, Mich. Academician, 24, 105–114.

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of Exotic Earthworms 5 Invasion into North America and Other Regions Samuel W. James Department of Life Sciences, Maharishi University of Management, Fairfield, Iowa, U.S.A.

Paul F. Hendrix Institute of Ecology and Department of Crop and Soil Sciences, University of Georgia, Athens, Georgia, U.S.A.

CONTENTS Introduction ......................................................................................................................................75 Criteria of Exotic Earthworm ..........................................................................................................77 Characteristics of Earthworms That Make Them Invasive .............................................................78 Mechanisms of Earthworm Invasion and the Dynamics of Invading Populations.........................79 How Do Invasive Earthworms Interact with Native Earthworms?.................................................80 How Do Invasive Earthworms Interact with Other Organisms?.....................................................81 Effects of Exotic Earthworms Invasions on Ecosystem Processes.................................................82 What Can be Done about Exotic Earthworm Invasions? ...............................................................83 Acknowledgments ............................................................................................................................86 References ........................................................................................................................................86

INTRODUCTION Invasions of earthworms and their redistribution around the globe involve only a small proportion of the overall diversity of earthworms, but the phenomenon is very widespread. It is safe to say that few places occupied by humans are free of introduced exotic earthworms. Those few areas would be environments that are inhospitable to earthworms because of either climate or soil characteristics; otherwise, it is almost certain that there will be foreign earthworms in most soils affected by human activity. In this chapter, we discuss the timing and extent of earthworm movements, how and when they came to the attention of scientists, and how to be sure an earthworm is exotic in a particular location. Then, we explore several ecological aspects of exotic earthworm species and their invasions into new habitats and regions. The ecological aspect includes the general biological features contributing to the ability of an earthworm species to be invasive or simply to be transported inadvertently by humans; the dynamics of invasions; the biotic interactions of exotic earthworms with other groups of organisms, including native earthworms; and the potential for ecosystem-level effects of these invasions.

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It is certain that humans have been aware of earthworms for a very long time, yet only in the past several millennia have people had the motivation or wherewithal to alter global distributions of earthworms in any ecologically significant way. This required the birth of agriculture and thus an incentive to transport live plant materials, perhaps with or in soil, from one place to another. It may be speculated that the Polynesian peoples, or other seafaring groups with some reliance on agriculture, were among the first to cause the long-distance transport of earthworms, although there is no reason to suspect them of doing so intentionally. As transportation technology advanced and trade routes, road networks, and convenient means of moving heavy loads over long distances became readily available in Europe and Asia, earthworms were likely to move as well. To be sure or suspect that a species of earthworm or any other organism is outside its natural habitat range, there has to be a concept of natural range and some knowledge of the diversity and phylogenetic relationships of the organisms concerned. Otherwise, there is no basis for inferring that an occurrence at a particular site is unnatural; after all, until the late 19th century A.D., any such situation could be explained as the will of the Creator. Even well into the post-Darwin age, describers of new earthworm species found outside what is now believed to be the natural ranges of the species often used the foreign place names as specific epithets. There are numerous cases of earthworm species that were either first discovered out of their indigenous ranges or were given synonyms based on specimens collected outside their natural range (Reynolds and Cook 1976). For example, Amynthas gracilis Kinberg 1867 is an East Asian species that is found commonly well out of its normal range and in Hawaii was given the junior synonym A. hawayanus Rosa 1891. On the other hand, certain more familiar species of European Lumbricidae (virtually all earthworm taxonomists were European until the 20th century) kept turning up in odd places, such as Australia (Blakemore 1999) and New Zealand (Lee 1959), such that by 1900, it was generally believed that certain earthworm species had been widely distributed by human activity (e.g., Michaelsen 1900); these are collectively referred to as peregrine species. Once some basic outlines of earthworm taxa, including families and their generic distributions, were worked out (1895 to 1930), the extent of artificial earthworm distributions could be much better appreciated. Because certain earthworm species turn up in unexpected places, early observations often considered that native indigenous earthworms of many areas were declining in numbers and species, and that exotic earthworms were increasing in abundance. Eisen (1900) observed the phenomenon in California, and in the central United States, Smith (1928) noted a transition from an abundance of the native Diplocardia communis Garmann in 1888 to domination by Lumbricus terrestris L. over the period 1900 to 1925. Significantly, this occurred in urban lawns and gardens, so the Diplocardia had survived a habitat conversion. Much later, Ljungstrom (1972) noted replacement of the South African native earthworm fauna by exotic species. The mechanisms of replacement of native earthworm species by exotic species are complex, and there has been little experimentation to identify the key processes involved. These mechanisms could include the intolerance of the native earthworm to altered habitats, a loss of key biotic relationships present in intact ecosystems, competition pressures from exotic species, and an inability to reestablish populations after partial recovery of ecosystems (Stebbings 1962; Kalisz and Wood 1995). Earthworms are still traveling, although recent increased stringency of border controls may have reduced international traffic to some extent. Gates (1982) received thousands of specimens that had been intercepted at the U.S. borders over several decades in the 1900s. The result of the many years of earthworm transport is global homogenization of earthworm communities, in agricultural and urban lands, modified by climate conditions. The U.K. is now populated by the same species of Lumbricidae that were present in the rest of glaciated Europe. The same set of species, for the most part, occupy soils in the temperate zone of North America and the cooler regions of Australia, New Zealand, North Africa, South America, and South Africa and are part of a mixed earthworm fauna in temperate East Asia. Aporrectodea spp. have been seen in irrigated highway rest areas in the Humboldt Sink region of Nevada, a cold desert with salt alkaline soils, where they © 2004 by CRC Press LLC

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must have arrived with nursery stock or other plant materials used in landscaping. Species of Lumbricidae are present in Costa Rican montane forests at an elevation of 2000 m near Heredia and in similar habitats in Venezuela. Significantly, the peregrine Lumbricidae are now the only earthworms currently present in large areas of glaciated North America and Eurasia, where they are modifying soils and litter decomposition dynamics, but this subject is discussed below. A later wave of immigrant earthworms from temperate Asia (probably Japan) is now progressing through North American soils. The most important of these is Amynthas hilgendorfi Michaelsen 1892, which may be a range of parthenogenetic morphs. In tropical climates, a different set of exotic species is now distributed globally, dominated by the ubiquitous Pontoscolex corethrurus Müller 1856, a member of the South American family Glossoscolecidae. This species is believed to have originated in northeast South America, where the rest of its congenitors are found (Righi 1984). This earthworm is now abundant in all tropical areas where rainfall is adequate to support earthworm activity for at least a few months each year. It lives in lowland evergreen tropical forests up to cloud forests along elevational transects in high rainfall areas and can maintain populations in seasonally tropical dry forests, including some with xerophytic vegetation. It also prospers under agricultural conditions, including pastures, tree plantations, row crops, and paddies (along the margins, not in flooded soils). Reforestation programs using root-ball planting stock provide another means of dispersal of these species, and in many places in Southeast Asia and Fiji, only the disturbance of logging and the probable movement of P. corethrurus on heavy equipment was required for virtual elimination of the native species and introduction and establishment of this exotic species. Some other widespread tropical peregrine earthworm species are Dichogaster bolaui, Dichogaster affinis, Eudrilus eugeniae, Drawida barwelli, Pithemera bicincta, Perionyx excavatus, Amynthas corticis, and Polypheretima elongata (Fragoso et al. 1999). Of these, the most important is probably the last because it is capable of surviving in a broad range of environmental conditions and has been blamed for degradation of soil structure in some agricultural settings (Shah and Patel 1978). Some of the species in this list are used in vermicomposting and so are maintained in cultures even where they cannot survive ambient winter temperatures.

CRITERIA OF EXOTIC EARTHWORM We have mentioned briefly some criteria for determining whether an earthworm is an exotic species. Essentially, it requires broad knowledge of the global or regional distribution of the larger taxon to which the species belongs. Thus, if outside South America and some neighboring areas (Central America and a few Caribbean islands), a glossoscolecid such as P. corethrurus is clearly an exotic species because that family is otherwise confined to South America and its environs (Righi 1972; Fragoso et al. 1995). Similarly, a species of the exclusively African family Eudrilidae would be exotic on any other continent, and even within Africa it is exotic north of the Sahara and south of the Kalahari deserts. The problem is harder when the larger taxon has a wider global distribution, such as the Lumbricidae, with species native to Europe and North America and a known direct land connection prior to the opening of the Atlantic Ocean. In this case, the debate has raged longer (Omodeo 1963) but was settled ultimately by the observation that the two genera occurring in North America are completely absent from Europe, and there are no endemic North American species of any, otherwise European, genera (Gates 1982). Furthermore, the same members of European lumbricid genera that are found in North America are also found in temperate zones of other continents. Looking in more detail at earthworm distributions in Europe, it is clear that the peregrine species are the only earthworms in the northern regions of the continent, whereas in southern Europe, a very diverse earthworm fauna exists. Thus, the peregrine species are only a small subset of the total, a pattern that is repeated in all higher earthworm taxa containing peregrine species.

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Some of the earthworm distributions in East Asia are more difficult to assess because numerous Amynthas and Metaphire species have been transported, perhaps for hundreds of years, and concomitant habitat alterations have all but eliminated any chance of locating the ancestral homes of these peregrine species. Some species are often found mainly in anthropogenic habitats, and others are found only in undisturbed or relatively undisturbed soil systems. Without any other data, it can be determined that certain earthworm species are at least potentially peregrines, but it may not be possible now to determine their original ranges (see Chapter 3, this volume).

CHARACTERISTICS OF EARTHWORMS THAT MAKE THEM INVASIVE Four primary features of earthworms are suggested as important for their ability to travel and establish populations in new areas: a tolerance of environmental variability; suitable ecological niches; high reproductive potential; and an appropriate reproductive system, which may be biparental or uniparental (Barois et al. 1999; Fragoso et al. 1999). To be transported successfully to other regions, an earthworm species or its cocoons must be able to survive some degree of disturbance, perhaps including fluctuations in temperature and moisture levels, and be able to survive in a new habitat with different soil conditions, novel sets of soil organisms, and potentially unaccustomed seasonal patterns of temperature and moisture. The simplest case is the transport of a body of soil containing the earthworms or cocoons; the larger the body of soil is, the greater its degree of insulation from environmental changes during the movement. It would be expected that large amounts of soil would be least challenging to an earthworm species that lives in it but also less likely to occur than the movement of smaller quantities of soil. Therefore, the most successful earthworm travelers will be those with the maximum degree of tolerance for adverse soil conditions. Among the known tolerance mechanisms of earthworms is their ability to enter dormancy in response to higher temperatures, low soil moisture, or both (Lee 1985). Many peregrine species of Lumbricidae, such as the Aporrectodea species, which are common throughout the world, have this ability. Similarly, the ubiquitous tropical peregrine species P. corethrurus can become dormant in drought conditions. On the other hand, Diplocardia species in the central United States can enter dormancy during drought periods, but none of these species is known to be peregrine, although several do well in moderately disturbed anthropogenic habitats such as lawns and gardens. The only obvious difference suggesting why the Diplocardia species may not have traveled with human aid is the physical robustness of construction among the Lumbricidae, which appears to be greater than that of Diplocardia spp. Yet, robustness is hard to quantify and perhaps even harder to put to a meaningful test. On the other hand, some Diplocardia species commonly collected for fishing bait can survive soil temperatures of 35 °C in small containers, which is invariably fatal to most peregrine Lumbricidae. Bouché (1977) outlined several major aspects of earthworm behavior and morphology correlated with ecology and defined the ecological categories now in wide use, that is, epigeic, endogeic, and anecic species. Epigeic species are those inhabiting superficial organic matter accumulations and the soil-litter interface and are typically small bodied, dark colored, and capable of rapid movements. Many of the common vermicomposting species fall into this category. Their natural food sources are short lived and widely scattered, so they produce large numbers of offspring or may even be parthenogenetic and rely on dispersal to locate new resource patches. Endogeic earthworm species live and feed in the mineral soil layer, are usually lightly pigmented or unpigmented, range in size from small to very large, do not move quickly, and have lower reproductive rates. Further division of endogeic species has been made (i.e., polyhumic, mesohumic, and oligohumic) based on the relative age, concentration, and location of organic matter they utilize as food. The polyhumic species somewhat resemble epigeic species in coloration but are capable

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of living in mineral soil. Oligohumic species are those that feed mostly on greatly modified organic matter in the soil and are virtually always unpigmented. Most anecic species maintain deep burrows from which they emerge to feed on surface organic matter. They usually are large bodied, have dark anterior pigmentation, and have lower reproductive rates. From this quick overview of earthworm ecological categories, it might be guessed that the epigeic species are those most likely to be invasive, but this is not entirely true because some of their habitat requirements can be restrictive. Many invasive species are endogeic, and these are primarily of the poly- or mesohumic types. Some anecic species, such as L. terrestris, are also invasive. The reproductive potential of a species also plays some role in its potential to be invasive, although it probably is less important than its ecological relations, which are intimately related to the life history strategy of the particular species (r vs. K selection, for example). A more obviously important factor in the ability of earthworms to colonize new habitats successfully is the nature of the reproductive system. Although most species of earthworms are obligately outcrossing hermaphrodites, some use other options. Some species may self-fertilize, and others are parthenogenetic. In either of these cases, reproduction is or can be uniparental as opposed to biparental. A single uniparentally reproducing earthworm can found a new population, and it is clear that this has been important in establishing earthworms in new areas. For instance, the invasive Octolasion species are parthenogenetic, as are many of the invasive Amynthas and Microscolex species and probably P. corethrurus. Uniparental reproduction, although not necessary or sufficient to achieve successful invasiveness, is certainly a useful property for this purpose.

MECHANISMS OF EARTHWORM INVASION AND THE DYNAMICS OF INVADING POPULATIONS We mentioned some of the mechanisms of earthworm invasions, especially those directly involving human activity and the transportation of earthworms to new locations. In this section, we review the available knowledge of earthworm invasion dynamics in earthworm-free habitats (typically, higher latitudes subject to recent glaciation). There are still some unresolved questions about situations in which an indigenous earthworm fauna is still present on a site where exotic species have been introduced. Invasions of agroecosystems by exotic earthworms are virtually the general rule, particularly in North America, because native species capable of tolerating a less-buffered soil environment and the frequent soil disturbances are uncommon. It is fair to say that the diversity of agriculturetolerant earthworms is to the total diversity of earthworms as the diversity of cultivated plants is to the total diversity of plants. If it is not quite true yet, it seems likely that soon there will be two or three major earthworm groups operating in temperate, subtropical, and tropical climate zones. For this reason, and because the vast majority of earthworm research on agroecosystems has dealt with exotic species, readers are referred to reviews of earthworm ecology in agroecosystems (Edwards and Bohlen 1996; Baker 1998; Hendrix 1998; Lavelle et al. 1999). Here, we confine discussion to earthworms in nonagricultural land. In the temperate deciduous and mixed forests of North America, several research groups have recorded profound changes in the structure of soils and the forest floor following the introduction of exotic species of earthworms. The first publication on this subject was by Nielsen and Hole (1964), who noted the conversion of podzols to a mixed mineral-humus soil in New Brunswick, Canada. These authors saw nearly total destruction of forest floor litter, leaving only the most recent leaf fall on the ground, in several locations in Michigan and Iowa. Similar situations have been reported or are under investigation in Minnesota (Alban and Berry 1994; Hale et al. 2000); New Jersey (Kourtev et al. 1999); Rhode Island, Pennsylvania, and New York (Steinberg et al. 1996; Burtelow et al. 1998); Alberta, Canada (Scheu and Parkinson 1994;

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McLean and Parkinson 1997, 2000a,b); and in North American natural grasslands (Callaham and Blair 1999; Callaham et al. 2001). In Minnesota, the invasion by exotic earthworms consisted of several species arriving in successive invasion fronts. First came the epigeic litter-dwelling species, followed by polyhumic endogeic species and then slower-moving anecic and endogeic species. This phenomenon of successive waves of invasion by earthworms depends on the presence of diverse ecological types of earthworm, so it is not necessarily going to apply to all places. However, it demonstrates that an earthworm invasion is not a simple monolithic process, and that it may require several years to determine the outcome of an invasion of a single site. The species content of invasive waves are in accordance with what we might predict given the life histories, ecology, and reproductive potentials of the various ecological categories of earthworms. On the other hand, sometimes the results of the initial invasive wave, such as that of epigeic species (Dendrobaena octaedra), have proven different in places like Minnesota (Gundale 2002; Hale et al. 2000) and Alberta (Scheu and Parkinson 1994; McLean and Parkinson 1997). In the former area, little impact of the invasion on soil horizons was noted, whereas in the latter, there were significant effects on the soil strata. In Minnesota, the main effects of invasive earthworms on soil organic horizons occurred after the arrival of Lumbricus rubellus in the second wave of invasion. In some New England forests, the arrival of A. hilgendorfi is now causing total destruction of the organic horizon in spite of centuries of opportunity for European earthworms to do the same. When earthworms arrive in a forest with soils with thick organic horizons but previously devoid of earthworms, there is a large and readily available supply of food. Years later, the amount of organic matter per square meter has probably diminished significantly or, if not, it has been transformed profoundly and mixed with the mineral soil. In either case, negative consequences would be expected for the early invading species that took advantage of the mass of surface organic matter, with more favorable conditions occurring for endogeic species. Anecic species might be less affected, except as small juveniles, because the larger earthworms can pull surface organic material into their burrows. Nevertheless, populations may be limited by food availability at some point because the entire litter fall is usually consumed annually. Such long-term changes in the total food supply available should mean that populations of some earthworm species would go through peaks and declines during the course of the invasion and then reach something approximating a steady state or equilibrium, because the earthworms have to live off current food income rather than the organic capital accrued prior to their arrival.

HOW DO INVASIVE EARTHWORMS INTERACT WITH NATIVE EARTHWORMS? When exotic earthworm species arrive at a site already occupied by a native earthworm fauna, there could be several other things occurring. We have noted the replacement of native earthworm species by exotic species in Illinois, California, and South Africa, but this was mainly in urban and agricultural areas, so it is a different situation from similar replacements in undisturbed habitats or in less-disturbed systems such as second-growth forests, where earthworm invasions are unlikely (Kalisz and Dotson 1989). In some disturbed habitats, native and exotic species coexist (Fragoso et al. 1995, 1999; Callaham and Blair 1999; Bhadauria et al. 2000); in others, the exotic species dominate. Stebbings (1962) suggested that exotic earthworm species may be out-competing the native species on some forested sites in the central United States, but Kalisz and Wood (1995) suggested that replacements might not occur in at least some minimal area of undisturbed habitats. Lavelle and Pashanasi (personal communication) reported that P. corethrurus did not invade adjacent primary Peruvian forests from agricultural land, where it was well established. The primary forest had an endemic earthworm fauna that was absent from the land that was converted to agriculture. Dalby et al.

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(1998) reported resistance to invasion by exotic earthworm species in Australian forests inhabited by native species. On the other hand, we have observed that the species Diplocardia riparia which inhabits the riparian zone was absent from a Skunk River, Iowa, riverbank habitat near public access points, and that the Asian species Amynthas hupeiensis (used as fish bait) was abundant. However, after prolonged inundation during the summer floods of 1993, populations of A. hupeiensis were apparently wiped out, and populations of D. riparia recovered. In this example, the native species was better adapted to the natural disturbance regime of the riparian zone, but under current conditions of flood control by reservoirs upstream (decreased disturbance frequency or intensity), the exotic species can dominate. A combination of factors is probably involved in the outcome of earthworm invasions when a native earthworm fauna is already present (Hendrix and Bohlen 2002). Habitat destruction or disturbance is almost always cited as a precursor to earthworm invasions, but it is difficult to say whether it is a necessary prerequisite. Appropriate experiments have not been done to determine the outcome of earthworm introductions to areas of primary vegetation with an intact native earthworm community. Parallel experimental introductions could be made to primary vegetation soils where earthworms have been removed and in sites with populations of native earthworms intact. A related question is whether the rate of earthworm expansion into a new area is slower (but greater than zero) when native species are present than when they are absent. The reduction or elimination of native earthworm species is another factor or prerequisite to invasion, but it is difficult to say if it is a necessary factor for successful earthworm invasions; most examples suggest that it certainly helps (Kalisz and Wood 1995). The only factor clearly necessary for successful establishment is the arrival of the exotic species, without which there can be no invasion.

HOW DO INVASIVE EARTHWORMS INTERACT WITH OTHER ORGANISMS? Hawaii, which had no earthworms prior to human colonization, is now occupied by a variety of exotic earthworm species (Nakamura 1990). In a massive invasional meltdown (Simberloff and von Holle 1999), the invasive nitrogen-fixing tree Myrica faya is increasing the amounts of nitrogen cycling in the ecosystems and increasing earthworm populations by a factor between 2 and 10, depending on site (Aplet 1990). In addition, feral hogs, when foraging for earthworms, are ripping up the forest floor, creating more open sites for germination of the seeds of an exotic flora. The best that can be said of the Hawaiian situation is that it creates an opportunity for humans to hunt feral pigs. There are several recorded examples of invasive earthworms altering the composition of soil microbial and faunal communities. For example, McLean and Parkinson (2000a) reported that populations of D. octaedra in pine-forest floors in Alberta significantly altered microfaunal abundance and diversity in the organic horizon and in the mineral soil, increasing the number of the faster-growing microinvertebrate taxa. Similarly, soil microarthropod communities were changed by the presence of P. elongate in pastures in Martinique (Loranger et al. 1998) and by that of D. octeadra in pine forests in Alberta (McLean and Parkinson 2000b). In both instances, there were positive and negative correlations between earthworm populations and measures of microarthropod diversity and abundance; these effects were attributed to changes in physical structure of the soil and organic layers. In Minnesota, perturbation by introduced earthworms was responsible for the reduction of populations of the endangered fern Botrychium mormo (Gundale 2002), and interactions between effects of earthworms and deer browsing dramatically reduced the forest-floor vegetation (Hale et al. 2000). We have observed a large gallery-forming beetle larva in the thick, forest organic horizons

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with no earthworms at the University of Michigan Biological Station. Adjacent earthworm-populated areas seem to have no place for this beetle to live. If earthworm introductions were the only thing that changed, there would be little cause for concern, but many interactions and ecosystem processes can be altered profoundly, sometimes with clear negative effects (Hendrix and Bohlen 2002).

EFFECTS OF EXOTIC EARTHWORMS INVASIONS ON ECOSYSTEM PROCESSES In addition to issues concerning the population and community dynamics of exotic species of earthworms, there are questions about the impacts of earthworm invasions on ecosystem processes. Of particular importance are soil processes mediated by biological activities (e.g., litter decomposition, nutrient mineralization), which may be susceptible to the intensified earthworm activities characteristic of earthworm invasions. Although exotic earthworm invasions have been reported for over a century, quantitative studies of their effects on soils have been much more recent. In North America, invasions by European lumbricids and Asian megascolecids have been reported north of Pleistocene glacial margins in Canada and the United States, that is, in areas previously devoid of earthworms. One early study by Buntley and Papendick (1960) described a large area in eastern South Dakota (approximately 3400 km2) where the soil profile, of a chernozem on Pleistocene deposits, had been disrupted by earthworms. Soil horizon boundaries were obliterated, and the physical structure of the soil consisted of casts and filled earthworm burrows to a depth of 90 cm. Organic carbon and nitrogen, CaCO3, and clay distributions were also changed because of vertical translocation of these substances. The earthworm species involved was not indicated, but the area was probably previously devoid of native earthworms. However, Gates (1967) reported that the species in question was L. terrestris, which could have been introduced less than 100 years earlier with the first European settlements in the area. He cited several examples of early settlers transporting fruit trees and other horticultural materials during the westward migrations of the 19th century. In Idaho, there was even an attempt to establish exotic earthworms into soils for use as fishing bait because no indigenous species of earthworms were found in the region. The time required for the extensive alteration of soils observed in South Dakota was probably less than 100 years. More recent studies in New Brunswick and Minnesota demonstrated that only a few years are needed for transformations of mor horizons into mull horizons in forests that are invaded by several lumbricid species (Langmaid 1964; Alban and Berry 1994). Current research in Minnesota (Hale et al. 2000) and Alberta (Migge et al. 2003) suggest that these lumbricid invasions tend to proceed across the landscape in sequential waves, as noted in the section on mechanisms of invasion. In the Minnesota instance, it began with D. octaedra, which consumed the F and H layers; followed by L. rubellus and Aporrectodea spp., which mixed the organic and mineral horizons; and finally by L. terrestris, which drew recent litter into its burrows and produced middens and casts on the soil surface. The final result of this multistage invasion was a compacted, bare mineral soil surface covered only by recent litterfall. Impacts of the loss of the O-horizon on plant and animal populations in the forest floor have been reported (e.g., Maerz et al. 2001; Gunndale 2002). Intensive experimental studies of D. octaedra and Octolasion lacteum invasions in aspen and pine forests in the Rocky Mountains of Alberta showed significant changes in structures of microbial communities, decreases in microbial biomass, and increases in nutrient leaching rates in the Ohorizon in addition to physical disruptions of the soil profile. Declines in the total organic matter content were also noted in response to the highest earthworm population densities (Scheu and Parkinson 1994; McLean and Parkinson 1997, 2000a,b). Similarly, studies of undisturbed hardwood-forest patches in New York that were invaded by a variety of lumbricids revealed the

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elimination of the forest floor litter, reductions in carbon storage in soil, changed carbon:nitrogen ratios, and reductions in fine-root biomass in the upper mineral soil (Bohlen et al. 2003a; Fisk et al. 2003). The impacts of earthworm invasions on soil carbon dynamics in previously cultivated forest patches were less, emphasizing the importance of the site history as a determinant of earthworm effects of invasions on soil processes (Bohlen et al. 2003b). Interestingly, soil phosphorus fractions were affected differentially by different earthworm species. The anecic species L. terrestris increased the total phosphorus in surface soils by transporting parent material from deeper soil layers, and the epi-endogeic species L. rubellus increased the extent of exchangeable phosphorus and phosphorus leaching by consuming organically enriched surface materials and producing phosphorus-enriched castings (Suárez et al. 2003). Such results demonstrate the difficulty in making generalizations about effects of earthworm invasions. Further research is needed under a wider range of climatic, edaphic, and land-use conditions and a larger number of invasive earthworm species. Models of earthworm effects on soil process dynamics (e.g., Chertov and Komarov 1997) may be useful in the search for broad-scale patterns of earthworm invasions. Exotic earthworm invasions also have been reported in areas inhabited by native earthworms, but their impacts on ecosystem processes appear to depend on the previous disturbance history of the site, earthworm invasion pressure, and the degree to which the native earthworm assemblage is intact (Figure 5.1). In highly disturbed soils, native earthworm populations are often reduced, and invasions by the typical anthropochore species usually have significant effects on soil processes. Observations have been made worldwide in agricultural and pastoral ecosystems, where introduced European lumbricids have modified soil structure, rates of organic matter decomposition, nutrient dynamics, and in some cases plant productivity, for example, in reclaimed polders in the Netherlands (Hoogerkamp et al. 1983); grasslands in New Zealand and California (Stockdill 1982; Winsome 2003); and cropping systems in Australia and the United States (Parmelee et al. 1990; Baker 1998). Similar impacts of invasive earthworms have also been reported in at least moderately disturbed natural areas where native earthworm species are still present but probably not at natural abundances, for example, in California chaparral (Graham and Wood 1991) and tropical forests (Fragoso et al. 1995; Liu and Zou 2002). Where native and exotic earthworm species coexist, the magnitude of their effects on soil processes may be determined by the relative abundance of the various species of earthworms, ecological strategies, and environmental fitness of the dominant species. James (1991) suggested that native earthworms were usually better adapted to local soil and climatic conditions and hence could maintain longer periods of activity and have greater effects on nutrient dynamics in tallgrass prairie soils than could invading European lumbricids. Lachnicht et al. (2002) reported that native Estherilla spp. and the exotic P. corethrurus coexisted by partitioning the soil volume physically in microcosms containing forest floor and mineral soil from tabonuco forests in Puerto Rico. Interactions between the two species reduced the impact of P. corethrurus on carbon and nitrogen mineralization rates. The dynamics and impacts of earthworm invasions in undisturbed ecosystems in which soil, vegetation, and indigenous earthworm assemblages are intact have not been well studied. In such instances, there may even be some degree of biotic resistance to earthworm invasions; hence, the impacts on ecosystem processes in these situations may be different from those in areas that were previously devoid of earthworms. This possibility needs further research.

WHAT CAN BE DONE ABOUT EXOTIC EARTHWORM INVASIONS? Many exotic earthworm species are now naturalized in areas beyond their place of origin, notable examples being European lumbricids in temperate regions worldwide and P. corethrurus throughout the tropics (Reynolds 1994; Fragoso et al. 1999). Although humans and their disturbed habitats are

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INITIAL STATE

Disturbance

A

Severe

INTERMEDIATE STATE

Exotic Invasion

Native Earthworms Eliminated

Successful

CURRENT STATE Exotic Earthworms Exclusively

Successful "Pristine" System (Native Earthworms B Exclusively)

Moderate

Native Earthworms Diminished

1 2

Successful?

Native and Exotic Earthworms Coexisting

Successful?

C

Minimal

Native Earthworms Exclusively

1 2

Unsuccessful

Native Earthworms Exclusively

FIGURE 5.1 Hypothesized pathways of exotic earthworm invasions in ecosystems inhabited by native earthworms. Pathway A is the extreme case leading to exclusively exotic assemblages, as often observed with anthropochorous earthworms in agricultural soils; the same outcome may occur under less-severe disturbance but perhaps with more aggressive exotic invaders via pathway B-1. Pathways B-2 and C-1 lead to the sometimes observed co-occurrence of native and exotic species through varying levels of habitat disturbance and invasion intensity; pathway C-1 suggests direct competitive displacement of native species by exotic species. Whether co-occurrence is a stable condition or whether native species or exotic species maintain dominance is an interesting long-term question, and hence the question marks are shown for successful invasion on these pathways. Pathway C-2 suggests that native earthworm assemblages under minimally disturbed, native conditions are resistant to invasion by exotic species. The alternative is best represented by pathway C-1, by which forest fragmentation, for example, may foster exotic invasions without direct habitat disturbance. (Modified from Hendrix et al. 2004.)

the cause, these peregrine earthworm distributions must still be considered expansions of range for these species, albeit at very rapid rates in ecological and geological time. The important issues now are how, at local-to-regional scales, the spread of invasive species into areas where they are not wanted (e.g., remote or sensitive ecosystems) can be impeded and how, at national-to-continental scales, introduction of new invasive earthworm species can be prevented. The management of earthworm populations, whether invasive or not, has received considerable attention in the context of agriculture and organic waste management, and there are numerous cases of successful outcomes of such management, for example, enhanced plant productivity or accelerated decomposition of organic wastes (Lee 1995; Edwards 1998; Lavelle et al. 1999). Indeed, there probably would be no incentive to prevent or regulate earthworm introductions for such purposes. The management of earthworm populations in the context of mitigation of invasions has only recently emerged as a topic in need of development after it was brought to light by some of the reports noted here of adverse effects in forested ecosystems. There are at least two approaches to the management of invasive exotic earthworms. The first is essentially to provide no management at all, allowing invasions to run their course and assuming that, over time, invaded ecosystems will reach a new “equilibrium” under the influence of the newly reassembled biotic community. This approach accepts any changes in soil characteristics, biogeochemical cycling rates, and above- and belowground biotic communities that are likely to occur. In fact, this approach is in effect by default, in many areas where earthworm invasions are in progress and where the changes are witnessed as they occur, as in some of the experimental studies described here. Additional chronosequence studies, which compare sites at various stages of earthworm invasion from the beginning to equilibrium, might reveal a long-term progression in such changes and whether the ultimate outcomes are acceptable in particular regions. For example, © 2004 by CRC Press LLC

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forests in central to northern European countries may represent the very long-term postglacial end points of lumbricid invasions; similar situations might occur in tropical ecosystems (e.g., Fragoso et al. 1999). Such studies should help create predictive models of earthworm invasion dynamics, which would be useful in developing improved management strategies. A second approach to managing earthworm invasions is human intervention. It is a well-known tenet in invasion biology that it is much easier to prevent an invasion than to stop one in progress. Therefore, the first step is the one most often used by governments, that is, interception and quarantine at the borders. As reviewed by Hendrix and Bohlen (2002), some countries have specific regulations concerning the importation of earthworms; others do not. Canada, for example, allows earthworm imports only from the Netherlands (L. terrestris) and the United States (i.e., only species that are known already to occur in Canada). This represents a “clean-list” or “guilty-until-proveninnocent” approach to invasive species control (Reichard and Hamilton 1997; Mack et al. 2000). The converse or “innocent-until-proven-guilty” appears to be the default approach followed by many nations, including the United States, where current regulations regarding earthworm imports are based on the Federal Plant Pest Act, under which the U.S. Department of Agriculture Animal and Plant Health Inspection Service (APHIS) controls imports of soils that might carry pathogens. In the absence of any pathogens, there are no specific considerations of earthworms as invasive organisms, although this situation may be changing as APHIS develops new guidelines (Hendrix and Bohlen 2002). There is a rich literature base on the invasion biology of plants and insects of economic importance (Simberloff 1989; Mack et al. 2000), but invasions by soil invertebrates have not been well studied (Ehrenfeld and Scott 2001). There are some precedents in the Formosan termite, fire ant, and Japanese beetle invasions in the United States and terrestrial flatworm invasions in Europe and Australia; we may at least learn what not to do from these case studies. However, invasions by more cryptic and less-mobile earthworms appear to be qualitatively and quantitatively different from those by most other invertebrates and may be more similar to plant invasions than to those of other animals (di Castri 1991). Invasions by terrestrial planarians that attack earthworms (Boag and Yeates 2001) may be the best model for understanding and controlling earthworm invasions because of similarities between these groups with respect to ecology, life history, and modes of transport. Control or mitigation of earthworm invasions after they occur has received very little attention, probably because it is a daunting proposition. As anyone knows who has sampled earthworm populations in the field, removing them from soil is often very destructive to soils and almost never 100% effective, even in small plots. Removal of invasive earthworms over large areas is probably not feasible. Therefore, the best approach may be containment of exotic earthworm populations to areas where they already exist or, at least, a reduction of rates of dispersal from such areas into surrounding ecosystems in which the earthworms may have adverse effects. Physical barriers, such as buffer zones of unsuitable habitat around agricultural sites, might impede earthworm migration. Also, simple practices, such as not dumping fishing bait on stream banks or cleaning horse hooves or off-road equipment tires before transporting them into remote areas, might reduce the likelihood of earthworm invasions. In addition to experimental studies on the effectiveness of preventive measures, public awareness campaigns, e.g., those by conservation and outreach groups such as Minnesota Worm Watch (http://www.nrri.umn.edu/worms) could help reduce the flow of propagules of earthworms as well as those of other exotic species. Ultimately, regulations at state, provincial, or regional levels may be needed to prevent the transport of exotic earthworms into remote or particularly sensitive areas. Finally, as discussed by Hendrix and Bohlen (2002), more basic knowledge is needed in terms of the natural history and ecology of invasive earthworms, both in their native habitats and in ecosystems in which they have invaded and had significant impacts. Which factors control earthworm populations under natural local conditions? Which characteristics of the organisms and of the habitats have contributed to successful invasions and to earthworm invasion failures? How © 2004 by CRC Press LLC

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important is biotic resistance, which may occur in ecosystems inhabited by native earthworms or other native competitors and predators? Are native species competitive with aggressive exotic species under native conditions? Is habitat disturbance a prerequisite to invasion? Case studies and experimental manipulations are needed to answer these and other pertinent questions. In addition, there is still a need for basic survey and taxonomic work to assess the diversity and distribution of earthworm species both in their native habitats and in new areas where they may become invasive.

ACKNOWLEDGMENTS This work was supported by grant DEB 0236276 from the National Science Foundation to the University of Georgia. We are grateful to colleagues at the Seventh International Symposium on Earthworm Ecology, Cardiff University, Wales, U.K., in 2002 for useful discussions of the topics covered in this chapter.

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McLean, M.A. and D. Parkinson. 2000b. Introduction of the epigeic earthworm Dendrobaena octaedra changes the oribatid community and microathropod abundances in a pine forest, Soil Biol. Biochem., 32:1671–1681. Michaelsen, W. 1900. Oligochaaeta, in Das Tierreich, Lief. 10, Friedlander, Berlin. Migge, S., D. Parkinson, and S. Scheu. 2004. Earthworm invasion of Canadian aspen forest soils: Impact on soil microarthropod communities, in Abstr. Ninth Biennial Mtg. Soil Ecol. Soc., Palm Springs, CA, in press. Nakamura, M. 1990. How to identify Hawaiian earthworms, Chuo Univ. Res. Notes, 11:101–110. Nielsen, G.A. and F.D. Hole. 1964. Earthworms and the development of coprogenous A1 horizons in forest soils of Wisconsin, Soil Sci. Soc. Am. Proc., 28:426–430. Omodeo, P. 1963. Distribution of the terricolous oligochaetes on the two shores of the Atlantic, in Love, D. and A. Love, Eds., North Atlantic Biota and Their History, Pergamon Press, New York, pp. 127–151. Parmelee, R.W., M.H. Beare, W. Cheng, P.F. Hendrix, S.J. Rider, D.A. Crossley, Jr., and D.C. Coleman. 1990. Earthworms and enchytraeids in conventional and no-tillage agroecosystems: a biocide approach to assess their role in organic matter breakdown, Biol. Fertil. Soil, 10:1–10. Reichard, S.H. and C.W. Hamilton. 1997. Predicting invasion of woody plants introduced into North America, Conserv. Biol., 11:193–203. Reynolds, J.W. 1994. Earthworms of the world, Global Biodiversity, 4:11–16. Reynolds, J.W. and D.G. Cook 1976. Nomenclatura Oligochaetologica: A Catalogue of Names, Descriptions and Type Specimens of the Oligochaeta, New Brunswick Museum Monographic Series No. 9, Fredericton, New Brunswick, Canada. Righi, G. 1972. Bionomic considerations on the Glossoscolecidae, Pedobiologia, 12:254–260. Righi, G. 1984. Pontoscolex (Oligochaeta, Glossoscolecidae) a new evaluation, Stud. Neotrop. Fauna, 19(3): 159–177. Scheu, S. and D. Parkinson. 1994. Effects of invasion of an aspen forest (Canada) by Dendrobaena octaedra (Lumbricidae) on plant growth, Ecology, 75:2348–2361. Shah, A.H. and C.B. Patel. 1978. Annelids as pests of paddy crop, in Edwards, C.A. and G.K. Veeresh, Eds., Soil Biology and Ecology in India, University of Agricultural Sciences, Bangalore, p. 83. Simberloff, D. and B. von Holle. 1999. Positive interactions of nonindigenous species: invasional meltdown? Biol. Invasions, 1:21–32. Smith, F. 1928. An account of changes in the earthworm fauna of Illinois and a description of one new species, Bull. Ill. Nat. Hist. Surv., 17:347–362. Stebbings, J.H. 1962. Endemic-exotic earthworm competition in the American Midwest, Nature, 196:905–906. Steinberg, D.A., R.V. Pouyat, R.W. Parmelee, and P. M. Groffman. 1996. Earthworm abundance and nitrogen mineralization rates along an urban-rural gradient, Soil Biol. Biochem., 29:427–430. Stockdill, S.M.J. 1982. Effects of introduced earthworms on the productivity of New Zealand pastures, Pedobiologia, 24:29–35. Suárez, E.R, D.M. Pelletier, T.J. Fahey, P.M. Groffman, P.J. Bohlen, and M.C. Fisk. 2004. Effects of exotic earthworms on soil phosphorus cycling in two broadleaf temperate forests, Ecosystems, in press. Winsome, T. 2003. Native and Exotic Earthworms in a California Oak Savanna Ecosystem, Ph.D. dissertation, University of Georgia, Athens.

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Part III Earthworm Biology, Ecology, Behavior, and Physiology

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Affecting the 6 Factors Abundance of Earthworms in Soils James P. Curry Department of Environmental Resource Management, University College, Belfield, Dublin, Ireland

CONTENTS Introduction ......................................................................................................................................91 Climate .............................................................................................................................................92 Soil Properties ..................................................................................................................................93 Food..................................................................................................................................................94 Competition......................................................................................................................................98 Predation...........................................................................................................................................99 Parasitism and Disease...................................................................................................................100 Land Management..........................................................................................................................100 Mining and Industrial Wastes .................................................................................................101 Deforestation ...........................................................................................................................102 Afforestation ...........................................................................................................................102 Grassland Management...........................................................................................................103 Arable Cropping .....................................................................................................................103 Manures and Fertilizers ..........................................................................................................105 Pesticides and Pollutants ........................................................................................................106 Soil Water Management..........................................................................................................106 Conclusions ....................................................................................................................................107 References ......................................................................................................................................108

INTRODUCTION Earthworm populations show a considerable amount of variability in time and space, with mean population densities and biomass ranging from fewer than 10 individuals and 1 g m–2, respectively, to more than 1000 individuals and 200 g m–2, respectively, under favorable conditions. However, within particular climatic zones, earthworm assemblages, with fairly characteristic species richness, composition, abundance, and biomass, can often be recognized in broadly different habitat types, such as coniferous forests, deciduous woodland, grassland, and arable land. There is a considerable volume of literature describing the earthworm communities of such habitats, and much of this was summarized by Lee (1985) and updated by Edwards and Bohlen (1996). There is also a considerable amount of information describing the influence of various environmental and management factors on earthworm populations, but in comparison with insects, for which the population ecology of

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many species has been subjected to quantitative analysis, earthworm population ecology is still largely at a descriptive stage. Population modeling has been used for particular purposes such as risk assessment (Baveco and De Roos 1996), but development of realistic models of field populations has been hindered by a lack of information on key life history parameters and the factors that influence them. In this chapter, the main factors that influence earthworm abundance are reviewed. These factors fall into two broad categories: external factors, which determine the habitat (climate, soil, vegetation, and litter supply and management), and biotic interactions within the communities to which earthworms belong (competition, predation, parasitism, disease, and food relationships).

CLIMATE Climate affects earthworms directly by influencing their biology and life processes and indirectly through its effects on their habitat and food supply. Temperature is a factor of primary importance because it determines individual earthworm metabolic rates, and on a global scale, it can have a major role in determining patterns of earthworm distribution and activity. The range of temperatures within which most earthworms can function is narrow, with upper lethal temperatures rather low (25 to 35°C) and optimum temperatures typically in the range 10 to 20°C for cool temperate species and 20 to 30°C for tropical and subtropical species (Lee 1985; Edwards and Bohlen 1996). Few species can tolerate temperatures below 0°C, although many species have behavioral or physiological adaptations that enable them to survive unfavorable periods in areas with strongly seasonal climates. Temperature may be a factor of primary importance in determining the composition and structures of earthworm communities (Lavelle 1983; Lavelle et al. 1989, 1999). Faster organic matter decomposition rates at higher temperatures result in decreased litter availability, and litter-feeding epigeic and anecic earthworm populations tend to be depleted in tropical soils compared with those in temperate soils (Lavelle et al. 1999). With increasing temperatures, endogeic species, which can utilize resources of increasingly lower quality through more efficient digestive processes involving mutualistic interactions with ingested soil microflora, are favored. Thus, in Mediterranean and humid tropical areas, oligohumic earthworm species that are able to feed on soils poor in organic matter are found deep in the soil profile. These are typically K-strategists, i.e., with large body size, slow growth, and low fecundity and mortality rates. However, very large endogeic species such as Octochaetus multiporus in New Zealand and Megascolides australis in Australia appear to be more common in temperate than in tropical soils (Lee personal communication). Thus, although large body size is an adaptation that facilitates earthworm feeding in nutrient-poor soils, in warmer soils body size may be constrained by increased energy demands for respiration, resulting in a severely limited energy supply for tissue production from a low-energy diet. High temperatures are often associated with moisture shortages, and seasonal earthworm mortality in temperate soils has usually been attributed to moisture stress rather than to temperature extremes (e.g., Gerard 1967; Phillipson et al. 1976). Indeed, the overwhelming importance of soil moisture in determining earthworm distributions and activity has frequently been demonstrated. Rainfall can explain more of the variance in earthworm numbers than any other variable in a range of agricultural soils, with annual rainfall varying from 230 to 1150 mm in southern Australia (Baker 1998). Optimum soil moisture content varies for different earthworm species and ecological groups, and within species, there appears to be a considerable capacity to adapt to local conditions (Lee 1985). In general, earthworms are most active at moisture tensions approaching field capacity (~10 kPa), and activity declines rapidly as the moisture tension exceeds 100 kPa and ceases for most species at moisture tensions below the permanent wilting point (1500 kPa) (Lavelle 1974; Nordström and Rundgren 1974; Nordström 1975; Baker et al. 1993).

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Although adapted populations in areas with strongly seasonal climates have the capacity to survive periods of drought, they nevertheless suffer heavy mortality, particularly among juveniles unable to escape desiccation by moving deeper into the soil and becoming inactive (e.g., Gerard 1967). The severity and duration of summer droughts impose severe constraints on the duration of earthworm activity and undoubtedly influence both the overall earthworm population density and the biomass, which can be attained when growth and development are restricted to relatively short periods in autumn and spring.

SOIL PROPERTIES Many studies have attempted to relate earthworm distributions to a range of soil physical and chemical parameters, often with inconclusive results. Other than soil moisture, the soil properties that appear to be most important include texture, depth, pH, and organic matter content. Medium-textured soils appear to be more favorable to earthworms than sandy soils or soils with high clay content (Guild 1948). Nordström and Rundgren (1974) reported a positive relationship between clay content and the abundance of Aporrectodea caliginosa, Aporrectodea longa, Aporrectodea rosea, and Lumbricus terrestris in 15 forest, pasture, and heath soils in Sweden, with clay contents ranging from 5 to 25%. A stepwise multiple regression indicated a positive relationship between clay content and the numbers and biomass of introduced Aporrectodea spp. in 113 pasture soils in South Australia (Baker et al. 1992). Decreasing population densities of A. caliginosa were linked with increasing proportions of sand and gravel in Egyptian soils (Khalaf El-Duweini and Ghabbour 1965), and low earthworm densities (max 73 m−2) occurred in sandy and silty coastal grassland sites in County Wexford, Ireland, compared with those in similarly managed loam soils (max 516 m−2) (Cotton and Curry 1980b). Although texture could have a direct effect on earthworm activity in the case of abrasive gravelly soils, more often the influence of texture may be indirect through its effect on moisture relationships. Heavy, poorly drained clay soils may become anaerobic in areas of high rainfall, and light sandy soils are prone to drought. The depth of soil is a significant factor governing earthworm distributions in temperate (Phillipson et al. 1976) and tropical (Fragoso and Lavelle 1992; Lavelle et al. 1999) forest soils. Lack of a sufficient depth of aerobic soil could be a factor limiting the establishment of deep-burrowing earthworm species in soils reclaimed after mining (Curry and Cotton 1983). Earthworms are generally absent from very acid soils (pH less than 3.5) and are scarce in soils with pH less than 4.5. Although there are considerable differences among species in their pH preference, the majority of temperate climate species are found in the pH range 5.0 to 7.4 (Satchell 1967; Bouché 1972). Other edaphic factors that have been linked with earthworm distributions include calcium, magnesium, and nitrogen content (Fragoso and Lavelle 1992), and populations can be affected adversely by high salt concentrations, which can occur, for example, in irrigated soils (Khalaf El-Duweini and Ghabbour 1965). The nature and quality of the soil organic matter are determined largely by the litter input from the vegetation. Litter from grass, herbaceous plants, and deciduous trees growing on base-rich, fertile soils is generally of high quality, with ratios of carbon to nitrogen close to or less than 20:1; the vegetation on impoverished acidic soils produces tough, unpalatable litter low in nutrients (carbon-to-nitrogen ratio more than 60:1) and unfavorable for earthworms. The organic matter that provides the food base for the earthworm community is vitally important in determining their distribution and abundance, and the soil organic matter content can sometimes be a good predictor of earthworm abundance (Edwards and Bohlen 1996). For example, Hendrix et al. (1992) reported a highly significant correlation between earthworm populations and soil organic carbon content over a range of sites in the state of Georgia, which included a wide variety of soil and vegetation types and management histories.

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FOOD There is little doubt that earthworm populations are often food limited; this is evident from the fact that populations often increase following organic amendment. The response to organic amendments can be particularly marked in disturbed habitats of low organic matter content (Edwards 1983; Lofs-Holmin 1983; Lowe and Butt 2002), but significant population increases can also occur in favorable habitats, such as permanent pasture, following the application of high-quality organic materials like animal manures (Curry 1976; Cotton and Curry 1980a,b; Edwards and Lofty 1982b). In the absence of other constraints, it is likely that the earthworm-carrying capacity of most habitats could be increased considerably by increasing the food supply. Hartenstein and Bisesi (1989) estimated from laboratory studies conducted by Hartenstein and Amico (1983) that a biomass of L. terrestris up to 0.5 kg m–2 could be sustained under conditions of unlimited food supply in soil irrigated with livestock wastes. However, it is unlikely that such a high earthworm biomass could be sustained under field conditions (Schmidt et al. 2003). The main source of the organic matter on which earthworms feed is litter from aboveground plant parts in most ecosystems, although dead roots and rhizodeposition can also be important food sources. Some species, including Allolobophora chlorotica, are found in close association with roots, and some species are known to ingest living roots (Baylis et al. 1986). Earthworm populations in woodlands can be limited by the amount and continuity of the litter supply; this was apparent, for example, in afforested coal mine dumps in Germany, where the epigeic species (Dendrobaena spp., Lumbricus rubellus) flourished when a well-developed litter layer was present but declined in importance when the litter layer was depleted by the action of anecic species such as L. terrestris (Dunger 1989). Zicsi (1983) concluded that the continuity of food supply was of cardinal importance in determining the suitability of deciduous woodlands in central Europe for the survival of largebodied, anecic earthworms. However, it appears to be the quality rather than the actual quantity of litter that most often limits earthworm populations (Satchell 1967; Swift et al. 1979; Boström and Lofs-Holmin 1986). Much of the litter input into soil is poor in nutrients, with nitrogen in particular often in short supply. Satchell (1963) calculated that the nitrogen requirement of the L. terrestris population in an English deciduous woodland (about 100 kg ha−1 year–1) was at least equivalent to, and was possibly in excess of, the nitrogen supply from litter. Nitrogen is often considered the critical factor limiting earthworm populations in many ecosystems, both temperate (Satchell 1967) and tropical (Lee 1983). Nitrogen content can be a useful indicator of food quality when comparing widely different types of litter but may be less useful as a predictor of earthworm performance on more palatable residues from agricultural crops and deciduous trees (Table 6.1). Boström and Lofs-Holmin (1986) and Boström (1987, 1988), for example, found no consistent relationships between nitrogen content and the growth rates of A. caliginosa cultured in soil amended with plant materials, with nitrogen contents ranging from 0.37 to 4%, although adult growth rates and cocoon production rates were significantly lower on unfertilized barley straw (0.35% nitrogen) than on meadow fescue (2.57% nitrogen) and lucerne (2.3% nitrogen) residues. Particle size has an important influence on the quality of plant materials as food for endogeic species, such as A. caliginosa (Boström and Lofs-Holmin 1986), but not for large anecic species such as L. terrestris. Boyle (1990) reared juvenile L. terrestris and A. caliginosa in mixed cultures (1 L. terrestris plus 1 A. caliginosa per 1-L container) in a peat/mineral soil medium with chopped (8-mm pieces) or milled (<1-mm pieces) ryegrass (Lolium perenne) provided as a food source (Table 6.2). The earthworms were kept for 18 to 19 weeks at 15°C, and grass (1 g dry mass) was either added to the soil surface or mixed into the soil every 2 weeks. Predictably, in view of its endogeic feeding habits, A. caliginosa grew faster on milled than on chopped grass, whereas for L. terrestris, which gathered the food and dragged it into the mouth of its burrow, particle size was unimportant. Also predictably, L. terrestris fared better when the food was placed on the surface, © 2004 by CRC Press LLC

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TABLE 6.1 Earthworm Tissue Production in Relation to Food (mg fresh mass g–1 dry mass)

Grass, Lolium perenne 1.5% nitrogen (Boyle 1990) Grass, Festuca pratensis Roots 0.8% nitrogen (Boström and Lofs-Holmin 1986) Shoots 1.7–2.2% nitrogen (Boström and Lofs-Holmin 1986; Boström 1987) Decayed 2.2-2.4%N (Boström 1987) Willow, Salix burjatica 1.8% nitrogen (Curry and Bolger 1984) Barley, Hordeum distichum 0.7–2.9% nitrogen (Boström and Lofs-Holmin 1986; Boström 1987) Barley, decayed 1.1–1.2% nitrogen (Boström 1987) Barley residues in field (N.C. Andersen 1983) Lucerne, Medicago sativa 2.4–4.0% nitrogen (Boström and Lofs-Holmin 1986; Boström 1987) Lucerne decayed 2.2–2.9% nitrogen (Boström 1987)

Lumbricus terrestris

Aporrectodea caliginosa

Mixed spp.

82–251

17–69

118–320

53 (17) 182 (152) to 247 187–225 180–250 74 (39) to 224 (206) 242–255 120–180 106 (76) to 132 175–249

Note: Values in parentheses are corrected to allow for growth in unamended soil.

TABLE 6.2 Earthworm Growth Rates in Peat/Mineral Soil Cultures under Different Feeding Regimes (mg ind–1 d–1± SE)

Milled grass On the surface Mixed in Chopped grass On the surface Mixed in

Lumbricus terrestris

n

Aporrectodea caliginosa

n

18.9 ± 1.2 a 6.2 ± 1.4 b

7 6

5.2 ± 1.6 a 2.7 ± 0.7 b

6 7

17.7 ± 2.3 a 8.0 ± 2.2 b

7 6

2.1 ± 0.6 b 1.2 ± 0.2 b

8 5

n = number of surviving earthworms. Note: Means in the same column followed by the same letter (a or b) do not differ significantly at p < 0.05 in ANOVA and t-tests. Source: Boyle 1990.

where it could be located easily and concentrated, but rather more surprisingly, this was also the case for A. caliginosa. It may be that A. caliginosa benefited by gaining access to concentrations of decaying grass residues in L. terrestris burrows. However, A. caliginosa growth rates were low compared with those recorded when A. caliginosa was reared individually under comparable conditions (Boström and Lofs-Holmin 1986), suggesting that A. caliginosa was at a net disadvantage as a result of its association with L. terrestris. In addition to its nutrient content, litter quality is influenced by factors such as carbohydrate content and the concentration of phenolic compounds (especially tannins), which can reduce © 2004 by CRC Press LLC

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palatability (Edwards and Heath 1963; Satchell and Lowe 1967). The presence of toxins can also have adverse effects; for example, the high mortality of juvenile A. caliginosa feeding in soil mixed with fresh lucerne residues was attributed to the glucoside saponin (Boström and Lofs-Holmin 1986; Boström 1988). Fresh litter is usually not acceptable to earthworms but must undergo a period of weathering before it is eaten. Factors involved include leaching of feeding inhibitors, softening of hard tissues, and microbial degradation (Satchell and Lowe 1967; Wright 1972). The role of the soil microflora is important not only in increasing palatability, but also in enhancing nutrient content (Waters 1951; Wright 1972; Cooke and Luxton 1980). This is particularly important in the case of lignified, nutrient-poor litter, such as straw, which undergoes a marked increase in nitrogen content when incubated in the soil (Boström 1987; Curry and Byrne 1997). The importance of microbial action in enabling endogeic earthworms to persist in tropical soils, which are low in organic matter, through mutualistic associations with ingested soil microflora has also been established (Lavelle et al. 1989). Litter ingestion rates are sometimes used as a basis for assessing limitations likely to be imposed on earthworm populations by the food supply; however, ingestion rates tend to be very variable, reflecting differences in food quality and palatability. For this reason, rates of tissue production per unit litter input may be more useful. Data from a number of studies, mostly conducted in the laboratory under favorable conditions but including some field studies (Andersen 1983; Curry and Bolger 1984), are summarized in Table 6.1. Tissue production rates ranged from 17 to 255 mg fresh mass per gram dry mass of ingested food in the case of A. caliginosa, representing production-to-consumption ratios (P/C) of 0.3 to 4.3% on a dry mass basis. The corresponding values for L. terrestris were 82 to 251 mg g−1 and P/C 1.4 to 4.2% (Table 6.2). Boström and Lofs-Holmin (1986) included data for earthworm growth rates corrected for growth in unamended soil; this was not necessary in the case of the Curry and Bolger (1984) and Boyle (1990) studies because earthworms lost weight in the unamended peat or peat-mineral soil media used. Earthworms grew faster on barley and lucerne residues that had been buried in the soil for 1 to 3 months prior to the start of the feeding trials (Boström 1987), but this pretreatment had no effects on growth performance on milled Festuca species residues, which presumably were acceptable with little or no weathering. Table 6.3 summarizes data from a 3-year study of the earthworm populations in a winter cereal field at Lyons, County Kildare, Ireland (Curry et al. 1995); the estimated food requirement of the population is compared with the food supply in Table 6.4. There were 12 earthworm species recorded; the most abundant were A. chlorotica (60 to 65% of adult numbers) and A. caliginosa (19 to 25% of adults).

TABLE 6.3 Tissue Production and Nitrogen Requirements of an Earthworm Population in a Winter Cereal Field

Mean population density (numbers m–2) Mean biomass (gm–2) Tissue production (gm–2) Nitrogen requirement for production (gm–2) Nitrogen loss via excretion and the like (gm–2) Source: Curry et al. 1995.

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1988

1989

1990

346 61 68–156 1.2–2.8 3.6

471 57 87–210 1.5–3.8 3.3

353 59 89–137 1.6–2.5 3.0

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TABLE 6.4 Food Requirement and Supply for the Earthworm Population in a Winter Cereal Field

Litter (stubble, cattle slurry) Roots and rhizodepositiona Requirement for maintenance and production a

g DM m−2

g N m−2a

484 300 340–1050

6.1 5 4.6–7.1

Estimated; see text.

Although population densities and biomass fluctuated considerably over the period of study, the mean annual values were fairly stable (Table 6.3). Annual tissue production values were 1.1 to 3.7 times the mean biomass, depending on the method of calculation. The estimates for nitrogen turnover, via excretion and mucus production, were derived from 15N laboratory studies conducted with juvenile L. terrestris. The main sources of organic matter input to the soil were postharvest crop residues, autumnapplied cattle slurry, and roots. Crop residues and slurry inputs were estimated, on the basis of six 50 × 50 cm samples collected in September 1990, as 484.2 ±199.5 (SE) g DM m−2; the mean nitrogen content of this material was 1.25%. Root inputs can be substantial but are difficult to quantify. No direct measurements were available in the present example, but tentative estimates can be made based on the results of studies that traced the fate of 14CO2-labeled photosynthate in growing cereal plants. Keith and Oades (1986) and Jensen (1994) calculated that the total amount of carbon translocated belowground during the growing season in wheat and barley was equivalent to about one third of the carbon contained in harvested grain and straw. Harvested dry matter at Lyons amounted to about 1500 g m−2, suggesting that the total input of organic matter to the soil via the roots could be on the order of 500 g DM m−2. Assuming that 40% of this is lost through root respiration (Jensen 1994), the amount available for root development and rhizodeposition (turnover of fine roots and carbohydrate exudates) could be about 300 g DM m−2. Rhizodeposition could account for about half of this, with the remainder located in macroroots. Studies on nitrogen turnover in spring barley in Sweden indicated that the quantity of nitrogen allocated to roots was equivalent to about 25% of that removed at harvest (Andrén et al. 1990); on this basis, the total nitrogen input to the soil from roots at Lyons could be about 5 g m−2. Thus, the total nitrogen input from aboveground litter and roots could be on the order of 800 g DM and 11 g N m−2 annually, with most of this becoming available to decomposer organisms in the autumn. Other potential food sources for the earthworm populations are reingested casts, microbial biomass, and soil organic matter, which could include significant quantities of nutrients recycled from dead earthworm tissues. Assuming an average earthworm tissue production rate of 200 mg fresh mass per gram dry mass organic input (Table 6.1), the inputs needed to sustain tissue productions by the earthworm population at Lyons could be in the range of 340 to 1050 g m−2 year−1; the estimated nitrogen requirement was 4.6 to 7.1 g m−2 (Table 6.3). If the actual tissue production and food requirement values fall in the lower part of this range, as seems likely in a cool temperate soil, then the figures suggest there is a reasonable match between food supply and demand. However, only a proportion of the organic matter input is likely to be available to earthworms. Litter bag studies indicated that, at most, 34% of crop residues are utilized by earthworms (Curry and Byrne 1997), and there is likely to be severe competition from other organisms for organic inputs in the form of rhizodeposition. Thus, it is unlikely that the level of earthworm population biomass recorded could be sustained for long under continuous arable cropping. © 2004 by CRC Press LLC

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COMPETITION Earthworm populations can be influenced by any animals with activities that affect their habitat conditions and food supply. These could include grazing animals in grassland and termites in tropical and subtropical habitats, both of which can drastically influence the nature and quality of organic matter available to earthworms. In a broad sense, earthworms can be regarded as in competition with such animals for food; however, the term competition in the strict ecological sense is usually reserved for interactions between and within closely related species. Interspecific competition is thought to play a major part in determining the structure of earthworm communities (Bouché 1983; Lavelle 1983). Despite the relative paucity of species (1 to 16) in most earthworm assemblages, a considerable amount of niche separation is apparent, suggesting ecological adaptation to permit the coexistence of potentially competing species. Phillipson et al. (1976), for example, found evidence for ecological separation in time and space and differences in food preferences among the 10 earthworm species present in an English beechwood; they concluded that these differences were important in reducing interspecific competition. Lavelle (1983) considered that differences in patterns of vertical distribution and in body size are most important in achieving earthworm niche separation in tropical soils, whereas horizontal and temporal separation and feeding specialization are the main factors in temperate soils. He concluded that competition is greater in temperate soils, where populations tend to be concentrated close to the surface because there is heavy dependence on litter as a food source, and the depth of soil that can be exploited is reduced by smaller body size. In tropical regions, by contrast, where higher temperature conditions favor microbial activity and enhanced levels of interaction among fauna, microflora, and organic matter, more diverse food resources can be utilized, permitting the occurrence of a wider range of earthworm sizes, greater depth distribution in the profile, and reduced competitive pressure. However, in the absence of supporting evidence, differences among them are not in themselves indications of ecological niche differentiation resulting from competition (Begon et al. 1990), and there is little direct evidence for interspecific competition among earthworms in the field. A decline in native species has been linked with the establishment of introduced lumbricid species in cultivated areas of Australia, New Zealand, and South Africa (Barley 1959; Martin and Charles 1979; Reinecke and Visser 1980), and there is evidence from pot experiments that interspecific competition could be a factor in such a decline (Dalby et al. 1998). Likewise, James (1991) reported a reduction in native Diplocardia spp. following invasion of tallgrass prairie in Kansas by A. caliginosa and Octolasion cyaneum. When the vegetation and soil have been altered radically, such trends may largely reflect the inability of native earthworm species to adapt to the new conditions and greater tolerance by lumbricid species for soil disturbance (Callaham and Blair 1999), but when no such disturbance has occurred, competitive replacement is a possibility. Interspecific competition in the field is notoriously difficult to detect, but the apparent absence of current competition does not mean that it is unimportant as a factor influencing community structure and population density. Competition may occur only occasionally, or not at all, because natural selection in response to competition in the past may have favored avoidance of competition through niche differentiation (the “ghost of competition past”; Connell 1980). Thus, unsuccessful competitors may have been eliminated already, and mature, present-day communities may comprise earthworm species able to coexist with little or no competition. Furthermore, competition need not be continuous or intense to be important because competition for scarce resources at critical stages in the life cycle could have important consequences for the species involved. Interspecific interactions may not necessarily always have negative effects on one or more of the interacting species. Temple-Smith et al. (1993) reported that A. longa gained more weight in the presence of A. caliginosa than in single-species culture; whereas decomposing leaf bundles at the entrance of L. terrestris burrows provided favorable microsites for Dendrodribus rubidus in English woodland (Phillipson et al. 1976). However, in the longer term, the litter-burying activities © 2004 by CRC Press LLC

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of anecic earthworm species such as L. terrestris are bound to affect populations of epigeic species adversely by reducing their food supply. There is more general agreement on the importance of intraspecific competition in determining earthworm populations. There may be some situations when competition is for resources other than food, as was reported when L. terrestris was cultured at increasing densities with adequate food in fixed volumes of soil (Butt et al. 1994), but in most natural habitats intraspecific competition is likely to be for food (Satchell 1967; Daniel 1992).

PREDATION Earthworms feature in the diets of hundreds of species of animals, vertebrates and invertebrates (Edwards and Bohlen 1996), but the quantitative impact of predation on populations has not been studied intensively. Most attention has been given to birds, which can be very visible and sometimes can be severe predators on earthworms. Bengtson et al. (1976) excluded golden plovers (Pluvialis apricaria) from small plots in an Icelandic hayfield in May/June and reported that the numbers and biomass of earthworms (A. caliginosa and L. rubellus) were more than twice as high after 22 days in the protected plots than in adjacent areas exposed to predation. Observations of feeding plovers showed that about 100 lumbricids m−2 could be expected to be taken over the experimental period, a figure that agreed closely with the results of the exclusion experiment. Golden plover predation was also the subject of a study by Barnard and Thompson (1985), who compared earthworm populations in areas enclosed by nets with those in unprotected control areas in English pasture at the beginning and end of a 3-month period (December to March). Earthworm populations declined by 46 and 71% in unprotected young and old pastures, respectively, and population density increased by 11 and 14% in corresponding protected enclosures. At the feeding rates observed, the authors estimated that golden plover predation could account for about 50% of the differences recorded; the other 50% was attributed to predation by other birds, foxes, and badgers, which were also affected by the nets. Earthworms were classified as a main food source for blackbirds and a regular food for song thrushes in France (Granval and Aliaga 1988); heavy predation by gulls, starlings, and magpies has been reported in New Zealand grassland (Moeed 1976). Earthworms comprised more than 90% of the food mass ingested by black-headed gulls when the earthworms were readily available on the soil surface following autumn cultivation in Switzerland (Cuendet 1983). About 10% of the total earthworm biomass was made available by cultivation, and 25 to 33% of this was taken by gulls. In a similar study in southwestern Ontario, Canada, the impact of ring-billed gull predation on earthworms in freshly plowed land was also negligible (Tomlin and Miller 1988). Many Insectivora (Mammalia) prey extensively on earthworms. Moles (Talpa europaea) consume 18 to 36 kg of prey annually per individual in Britain; at least half, and sometimes as much as 100%, of this prey consists of earthworms, depending on their availability (Mellanby 1966; Raw 1966; Funmilayo 1979). Large quantities of mutilated earthworms are stored in caches within mole fortresses as future food. Feeding on earthworms is also prevalent among shrews (Soricidae), with earthworms comprising 0 to 60% of the diet of the common shrew (Sorex araneus) in British grassland, depending on the season (Pernetta 1976b). At a daily consumption rate of 1.5 to 5 g fresh mass per individual (Pernetta 1976a), a field population of 10 to 20 shrews per hectare in mixed grassland and woodland habitat could consume 15 to 100 g fresh mass ha−1 day−1 or 0.5 to 3.7 g m−2 year, equivalent to somewhere on the order of 1 to 5% of the mean earthworm biomass in such habitats. Among the Carnivora, the European badger (Meles meles) and the red fox (Vulpes vulpes) are major earthworm feeders. The badger is highly specialized behaviorally as a predator of L. terrestris, and its population density seems to depend heavily on earthworm availability (Kruuk and Parish 1982). Adult badgers were estimated to consume 130 to 200 L. terrestris each per night in mixed deciduous woodland, pasture, and arable land in England (Kruuk 1978), but annual consumption © 2004 by CRC Press LLC

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by the badger population was estimated as less than 5% of the total earthworm biomass available, and the overall effect on the earthworm population was deemed negligible. The red fox showed similar foraging behavior in relation to L. terrestris and can have a catch rate of up to 10 earthworms per minute (Macdonald 1983). Invertebrate predators of earthworms include centipedes (Chilopoda) and ground beetles (Carabidae). On the basis of an exclusion experiment, Judas (1989) concluded that birds, shrews, and rodents did not affect earthworm abundance in a German beechwood, but chilopods and carabids, at the population densities recorded for the study area, could potentially have a significant impact on the earthworm population. Evidence available does not suggest that predation has a major long-term influence on earthworm population dynamics in most habitats, although short-term population reductions may sometimes occur following periods of intensive feeding by birds such as golden plover. A notable exception is the case of the New Zealand flatworm (Arthurdendyus triangulatus), which has recently become established in Ireland and Scotland and other areas in northwestern Europe, where it is potentially capable of causing severe reductions in earthworm populations and bringing about major changes in community structure (Blackshaw and Stewart 1992; Blackshaw 1995; Boag and Yeates 2001).

PARASITISM AND DISEASE A wide variety of parasitic and pathogenic organisms have been recorded from earthworms (Lee 1985; Edwards and Bohlen 1996); these include bacteria, fungi, protozoans, rotifers, platyhelminths, nematodes, mites, and dipterous larvae. Many are parasites of earthworm predators, and in some cases, earthworms are secondary hosts in which the parasites complete part of their life cycle. The effects of many of these organisms on their earthworm hosts are not well understood and are generally not considered significant, but some are known to be harmful. Larvae of the cluster fly (Pollenia rudis) parasitize and kill lumbricids in Europe and North America (Walton 1928 as cited by Lee 1985); Kingston (1989) attributed, at least partially, high summer mortality in A. caliginosa and L. rubellus to larvae of Calliphora dispar in irrigated pastures in Tasmania. Earthworms in nonirrigated land were not affected similarly because the earthworm population was isolated from the parasite by being quiescent 12 to 20 cm below the soil surface. An anoetid mite, Histiosoma murchiei, was reported to parasitize cocoons of A. chlorotica and to a lesser extent Eiseniella tetraedra in Michigan (Oliver 1962); the mites feed on and destroy the developing earthworms. Anoetid mites are common in wet habitats, and infection rates in A. chlorotica cocoons were consistently around 40% in a heavily forested area subject to spring floods. This parasite has also been reported from a peaty, poorly drained pasture in the Strødam nature reserve, Denmark (Gjelstrup and Hendriksen 1991), where about 20% of the cocoons of A. caliginosa and about 7% of those of L. terrestris were parasitized. The potential impact of Bacillus thuringiensis (Bt) toxin released into the soil by genetically modified plants is a cause for concern. Preliminary studies with Bt maize have not indicated adverse effects on populations of Lumbricus terrestris or other selected beneficial soil organisms (Saxena and Stotzky 2001); however, further studies are needed to evaluate the effects on a wider range of taxa and on soil biodiversity generally.

LAND MANAGEMENT Human activities can drastically alter the soil environment and influence earthworm populations directly by physical disturbances and indirectly by altering the physicochemical environment and the food supply. Some of the main ways in which activities such as mining, deforestation and afforestation, grassland management, arable cropping, pesticide use, and water management may influence earthworms are reviewed briefly.

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Most mining and industrial waste sites present conditions that are extremely hostile to biological activity (Ma and Eijsackers 1989; Hossner and Hons 1992; Logan 1992). Features of unameliorated sites likely to inhibit earthworm establishment include lack of organic matter, poor physical structure, compaction, poor drainage and unfavorable moisture conditions, excessive fluctuations in surface temperatures, and extreme acidity resulting from the reduction of sulfides and other materials in tailings ponds. Most soil-dwelling earthworms avoid low pH soils, although epigeic species such as Lumbricus eiseni and Dendrodrilus and Dendrobaena spp. may be found in the surface litter even in very acid soils. Very alkaline wastes such as pulverized fly ash (PFA) and wastes, following aluminum extraction with NaOH from bauxite, can also be toxic to the soil fauna (Satchell and Stone 1977; Southwell and Majer 1982; Eijsackers et al. 1983). High salinity is probably the main reason for PFA toxicity; this declines to harmless levels after weathering for 2 to 3 years (Townsend and Hodgson 1973). Mine tailings from low-grade copper and uranium mining can also be highly saline (Nielson and Peterson 1973). Metal toxicity can seriously impede the rehabilitation of mine spoil, adversely affecting revegetation and litter decomposition processes. Earthworms can tolerate fairly high levels of most heavy metals (Ireland 1983), although depressions in populations close to a copper refinery have been reported (Hunter and Johnson 1982). Metals in ionic form pose greater risks than organically bound forms (Malecki et al. 1982), and adverse effects on earthworm populations are more likely to occur in acidic soils (Ma 1988). Prerequisites for earthworm reestablishment in severely degraded soils include the amelioration of adverse conditions such as low pH, the stabilization of the physicochemical environment, and the provision of a suitable food supply. Liming and organic matter amendment can counteract the effects of acidity and metal toxicity and facilitate the initial stages of earthworm establishment, but long-term community development depends greatly on the nature and extent of revegetation and the litter supply. The rate and extent of earthworm population establishment varies, depending on factors such as the extent of initial disturbance and depopulation, the size and shape of the area affected, the degree and kind of restoration work carried out, and the availability of colonizers. In the case of rehabilitated coal mining dumps in the former German Democratic Republic, earthworm establishment reflected the main features of vegetation succession (Dunger 1989). Early colonizing species such as A. caliginosa and Dendrodrilus/Dendrobaena spp. made their appearance when a herb layer had developed and litter had begun to accumulate. The development of a shrub layer and well-developed litter layer was accompanied by the appearance of large populations of Dendrodrilus/Dendrobaena spp. and later L. rubellus, with activities that rapidly reduced the litter layer. By the time a closed tree canopy had developed, anecic species such as L. terrestris were well established, and during the transition to a fully woodland stage, further immigration of other species occurred. Under optimum conditions, the woodland stage with a species-rich and abundant earthworm fauna commenced 20 to 25 years after rehabilitation, but little earthworm community development occurred on acidic, infertile dumps. Earthworm population establishment can occur quite rapidly in grassland on mined sites restored to a high level of fertility. Purvis (1984) reported population densities in sites restored for 5 years or longer comparable with those in unmined sites in the English Midlands. However, earthworm biomass was low because large, deep-burrowing species were scarce: Lumbricus terrestris in particular was still largely confined to the edges of reclaimed fields after 10 years. Significant earthworm populations (9 to 10 spp., ca. 270 individuals m−2) were also found in 5- to 6-year-old productive grass leys on cutover peat in Ireland, although under less-favorable conditions, earthworm establishment occurred much more slowly (Curry and Cotton 1983). Because natural rates of earthworm dispersal into new habitats are low, at most 10 to 15 m year−1 and no more than 2 to 3 m year−1 for many species (Dunger 1969; Hoogerkamp et al. 1983; © 2004 by CRC Press LLC

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Marinissen and van den Bosch 1992), passive dispersal in water, in soil, with plants, by machinery, and by animals plays an important part in the colonization of rehabilitated land (Dunger 1989; Marinissen and van den Bosch 1992; Curry and Boyle 1995). In large tracts of land and when passive earthworm dispersal is inadequate, the process of earthworm establishment can sometimes be greatly accelerated by deliberate introductions (Vimmerstedt and Finney 1973; Stockdill 1982; Hoogerkamp et al. 1983) (see Chapter 5, this volume). Once a sustainable plant community has been established, most reclaimed soils can in time support a relatively stable earthworm population, one that is likely to be constrained to a greater or lesser degree by site limitations. Thus, earthworm population densities and biomass in grasslands on reclaimed cutover peat are lower than those in comparable grasslands on mineral soils, with L. terrestris particularly scarce (Curry and Cotton 1983; Curry and Boyle 1995). In this and perhaps in many other examples of rehabilitated sites, an important factor limiting community development may be the restricted depth of biologically active soil.

DEFORESTATION Many temperate forest species adapt well to grassland established on cleared deciduous forest sites. With the disappearance of the litter layer, there is a decline in populations of epigeic species, but anecic and endogeic species benefit from the improved soil fertility and food quality, and temperate grasslands tend to support greater earthworm populations and biomass than do temperate forest soils (Lee 1985). Tropical forest earthworms are affected more drastically by deforestation. Most tropical forests are on nutrient-poor soils and have predominantly epigeic earthworm populations, which largely disappear following forest clearance; in more nutrient-rich soils, in which endogeic and anecic species are more abundant, some may survive (Fragoso and Lavelle 1992). However, when adapted species are available for recolonization, high earthworm population densities may become established under suitable conditions. Lavelle and Pashanasi (1989) reported earthworm biomass of up to 153 g m−2, comprised almost entirely of the endogeic peregrine species Pontoscolex corethrurus, in pasture on cleared forest in Peruvian Amazonia.

AFFORESTATION The response of earthworms living on grassland to afforestation depends on the species of trees planted and the quality and quantity of litter produced. Muys et al. (1992) compared five broadleaf forest stands planted 20 years previously on meadow with a nearby meadow and two old forest stands. Significant differences in litter decomposition rates between sites could be observed in the thickness and quality of the surface organic (Ao) layer and in earthworm biomass and community structure. These effects were attributed to the quality and quantity of the annual litter produced in the various stands. Under Quercus palustris, which produces relatively unpalatable, poor-quality litter, earthworm biomass diminished, and the process of litter accumulation and moder humus formation had begun. Coniferous afforestation on fertile soils is usually accompanied by marked impoverishment of the earthworm fauna, reflecting poor litter quality, declining pH, and deteriorating soil structure. Preliminary sampling in Woodville wood, a mixed oak woodland in County Offaly, Ireland, indicated the presence of a moderate earthworm population of 40 to 60 g m−2 biomass comprising a mixture of epigeic, endogeic, and anecic species; in areas that had been cleared and planted with conifers (mainly Norway spruce, Picea abies) 40 to 50 years in the past, earthworms were scarce or absent (<10 g m−2). The earthworms recorded comprised mainly Dendrobaena and Dendrodrilus spp., with occasional specimens of Aporrectodea and Allolobophora spp.

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GRASSLAND MANAGEMENT Grassland management can alter herbage production and utilization and the dramatic return of plant residues to the soil (Andrzejewska and Gyllenberg 1980; Hutchinson and King 1980). The quality and supply of residues and the responses of soil organisms are likely to be influenced by sward composition, the extent and type of grazing, the frequency of cutting, and the quantity and type of fertilizer applied (Curry 1994). Potentially important features of grazing, as opposed to cutting, include the high proportion of ingested plant material returned to the soil in the form of dung, heterogeneity associated with selective grazing, uneven distribution of excreta by gregarious domestic animals, and negative impacts of trampling by cattle, particularly at high stocking densities (cf. Murphy et al. 1995; de Bruyn and Kingston 1997; Byers and Barker 2000). In the absence of grazing, frequent cutting and removal of herbage could result in much reduced rates of return of plant residues to the soil. A survey of the earthworm populations in 68 grassland fields in County Kilkenny, Ireland (Muldowney et al. 2003), revealed significant positive relationships between intensity of grassland utilization as measured by either stocking density or nitrogen fertilizer use and overall earthworm biomass, but not abundance (Figure 6.1). The response mainly reflected trends in populations of anecic earthworm species (mainly L. terrestris) and seemed to reflect an increase in body size rather than number. A similar overall trend of increasing earthworm biomass with increasing intensity of grassland management was reported from Switzerland by Cuendet (1996). However, no such relationship was detected by Boag et al. (1997) in Scotland, and Byers and Barker (2000) reported a negative correlation between earthworm populations and number of dairy cows per hectare in drought-prone soils in Pennsylvania. Preliminary results from ongoing studies in Ireland do not indicate any consistent relationships between management intensity and earthworm biomass in heavy-textured soils prone to poaching. Thus, it appears that the benefits for earthworms of enhanced food supply from plant residues and dung at higher fertilizer and stocking levels may be offset by negative effects of trampling in susceptible soils.

ARABLE CROPPING Earthworm populations in arable land are generally lower than those in undisturbed habitats (Chan 2001), but not always. The level of direct mortality associated with cultivations depends on the severity and frequency of soil disturbance. Ploughing per se does not appear to cause serious mortality: Cuendet (1983) estimated that 5 to 10% of the earthworm biomass was brought to the surface by plowing, with about 25% of these earthworms mortally wounded. Rotary cultivation can reduce numbers by 60 to 70% (Boström 1988), and earthworms can be virtually eliminated by very intensive soil cultivation, such as bed tilling and destoning for potato cropping (Curry et al. 2002). Populations generally recover within 1 year from less-severe forms of cultivation, provided the disturbance is not repeated. Larger, anecic earthworm species such as L. terrestris and A. longa, which require a supply of surface litter and have relatively permanent burrows, are the species most adversely affected by repeated soil disturbance; smaller endogeic species such as A. chlorotica and A. caliginosa are less affected and can benefit from plowed-in crop residues (Edwards 1983; LofsHolmin 1983). Tropical earthworms have little tolerance for cultivations (Lal, 1987). Earthworms, especially the larger, deep-burrowing species, are favored by minimum tillage and direct drilling compared with conventional methods of cultivation (Gerard and Hay 1979; Edwards and Lofty 1982a; Hendrix et al. 1986, 1992; Hutcheon et al. 2001). Indirect effects of cultivations, which are likely to affect earthworms adversely, include greater variability in surface soil temperature and moisture regimes in the absence of a permanent vegetation

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FIGURE 6.1 Relationships between earthworm biomass and grassland management intensity as measured by stocking rate (A, C) and mineral nitrogen fertilizer use (B, D). For total earthworm biomass, r = 0.32, p < 0.001 (A); r = 0.26, p = 0.007 (B). For anecic species, r = 0.33, p < 0.001 (C); r = 0.31, p = 0.001 (D). (From Muldowney et al. 2004. With permission.)

cover, reduced litter input, and more rapid oxidation of crop residues (Balesdent et al. 1988). Under these conditions, earthworms may experience severe food limitations, and earthworm population densities and biomass are likely to reflect the quality and quantity of crop residue input (Hendrix et al. 1992). Earthworm populations benefit from tillage practices that return a high proportion of crop residues to the soil, particularly when the residues remain on the soil surface, and from crops such as cereals, for which significant amounts of residues are left behind compared with root crops, for which most of the plant production is removed (Edwards 1983; Lofs-Holmin 1983; Hendrix et al. 1986; Mele and Carter 1999). Cereal-legume intercrops support much larger earthworm populations than those in conventional monocrops (Schmidt et al. 2001), probably in response to reductions in tillage (direct drilling or no till) and increased food supply in the intercrops. The relative importance of these two factors was investigated in a field plot experiment at Long Ashton Research Station, U.K., in which the earthworm populations in conventional wheat, direct-drilled wheat, and direct-drilled wheat-clover intercrops were compared (Schmidt et al. 2003). The results indicated that reduced tillage alone had only a modest effect on earthworm populations; the combination of reduced tillage and the presence of a clover understory increased earthworm populations greatly (Figure 6.2). The authors © 2004 by CRC Press LLC

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FIGURE 6.2 Earthworm populations and biomass (means ± SE) in three wheat cropping systems at Long Ashton, U.K. (From Schmidt et al. 2003. With permission.)

concluded that the earthworm populations benefited less from the reduced soil disturbance than from the enhanced quantity, nutritional quality, and continuity of food supply in the wheat-clover crop.

MANURES

AND

FERTILIZERS

Earthworm population responses to mineral fertilizers can be variable. Generally, the effects of moderate levels of application are positive, reflecting increased litter quality and quantity (Gerard and Hay 1979; Edwards and Lofty 1982b; Lofs-Holmin 1983; Boström 1988; Muldowney et al. 2003), but populations may be depressed by heavy applications of nitrogen (Nowak 1976). Adverse effects of some nitrogenous fertilizers, such as sulfate of ammonia, appear to be caused by soil acidification. Ma et al. (1990) reported a severe depression in earthworm populations in some grassland field plots treated with various types of nitrogenous fertilizers over a period of 20 years. The degree of depression reflected pH reduction, with ammonium sulfate and, to a lesser extent, sulfur-coated urea having the most marked effects. Organic manures benefit earthworms by providing additional food, by their mulching effects, and by stimulating plant growth and litter return. Farmyard manure is a particularly beneficial form of organic amendment (Edwards and Lofty 1982b; Lofs-Holmin 1983; Whalen et al. 1998). However, when populations are already very high under favorable conditions, such as those found in cereal-clover bicrops, neither the input of additional organic matter (as cattle slurry) nor mineral fertilizers may increase earthworm populations further (Schmidt et al. 2003). Heavy applications of animal wastes as semiliquid slurry containing high levels of ammonia and organic salts can be © 2004 by CRC Press LLC

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toxic (Curry 1976; Andersen 1980). However, any adverse effects of moderate slurry applications are transitory, and the long-term net population response is positive (Curry 1976; Cotton and Curry 1980a,b; Unwin and Lewis 1986).

PESTICIDES

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POLLUTANTS

Earthworms can be exposed to pesticides and other hazardous chemicals while moving through and ingesting contaminated soil or by ingesting contaminated litter. Varied responses to pesticides have been reported, ranging from little or none to very severe depressed numbers depending on the species concerned; the chemicals used; the rates, methods, and frequency of application; and other factors often not well understood (Edwards and Thompson 1973; Brown 1977; Lee 1985). Edwards and Bohlen (1992) reviewed the toxicity of nearly 200 chemicals to earthworms, including inorganic chemicals, 4 biological agents, 19 aromatic agents and organochlorine insecticides, 41 organophosphate insecticides, 13 carbamate insecticides, 6 pyrethroid insecticides, 26 fungicides, 53 herbicides, and 16 organic chemicals, and assigned a chemical toxicity index to each. Among the older pesticides, lead arsenate and mercuric chloride were highly toxic, as were soil fumigants such as DD, chloropicrin, methyl bromide, and carbon tetrachloride. Chlordane, heptachlor, and toxaphene are also very toxic and have been used as vermicides, but few of the other organochlorines affect earthworm populations to any significant degree at normal rates of application. Of the organophosphates, phorate and ethoprop are most toxic at normal rates of application; others, including fonofos, parathion, and thionazin, can be moderately toxic. Carbamate insecticides, notably aldicarb, carbaryl, carbofuran, and methiocarb, are very toxic to earthworms (Stenersen et al. 1973; Martin 1976; Stenersen 1979; Edwards 1980, 1983; Clements et al. 1986), as are benomyl and some related fungicides (Stringer and Wright 1976; Lofs-Holmin 1981). Herbicides do not appear to be directly toxic to earthworms but can have indirect effects by altering plant cover and food supply and the microclimate at the soil surface. It is unlikely that occasional applications of even the most toxic compounds have very serious consequences to earthworm populations, but their repeated use over a long period can. Long-term use of copper fungicides for disease control drastically reduced earthworm numbers in an English orchard (Raw 1962), and frequent treatments with large doses of insecticides (mainly phorate) over a period of 20 years eliminated earthworms from grassland plots (Clements et al. 1991). Other potentially hazardous chemicals include heavy metals in metal smelter emissions, in landspread sewage sludge, and in landspread pig slurry containing copper and zinc. Metals in organic wastes are not considered toxic to earthworms (Hartenstein et al. 1980; Malecki et al. 1982), but copper toxicity has been suggested as the likely reasons for low earthworm numbers in land heavily contaminated with pig slurry (van Rhee 1977; Curry and Cotton 1980; Ma 1988).

SOIL WATER MANAGEMENT As mentioned in this chapter, earthworm populations and activity are often restricted by unfavorable soil moisture conditions. Irrigation of dry soils has resulted in significant extensions of the range of lumbricid species (Barley and Kleinig 1964; Reinecke and Visser 1980) and can allow at least some species to remain active during the hot dry summer weather in South Australia (Baker 1998). However, the response to irrigation may depend on the species. Some soil-dwelling earthworm species such as A. caliginosa can suffer high summer mortality in irrigated pastures because of surface compaction and poaching by grazing animals, and dung and litter-feeding species such as L. rubellus are favored by such conditions (de Bruyn and Kingston 1997). High salinity resulting from excessive irrigation can also limit earthworm populations in some situations (Khalaf ElDuweini and Ghabbour 1965). Conversely, drainage and reclamation of wetlands such as polders and peat soils create conditions suitable for earthworm establishment (van Rhee 1969; Curry and

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Cotton 1983); drainage of water-logged soils in high-rainfall areas of southeastern Australia resulted in significantly increased abundance of A. caliginosa (Baker 1998).

CONCLUSIONS In undisturbed sites, which support relatively “stable” earthworm communities, the abundance and diversity of the earthworm fauna are determined primarily by interactions between climate and soils. These factors determine the physicochemical parameters of the soil environment, the nature of the vegetation that can be supported, and the quantity and quality of the litter it produces. On a global scale, temperature is the climatic variable of greatest significance to earthworms because it determines their metabolic rates and the diversity of food resources that can be exploited, but on a more local scale, moisture restrictions often determine local patterns of earthworm distribution and activity. Earthworms are often exposed to high rates of predation and are subject to pathogen and parasite attack; there may be times when predation especially can significantly depress population densities. However, when populations are not constrained by physicochemical environmental factors, food supply, notably the quality and quantity of the litter input, is the factor that most frequently limits earthworm abundance. Competition for food is generally believed to be important in determining earthworm populations, but there is little information on the nature of this competition and the frequency with which it occurs. Available evidence suggests that interspecific competition has been minimized through niche differentiation; when it occurs, it is probably diffuse in nature and may not have a major influence on population trends. By contrast, intraspecific competition may be common following periods of rapid population growth and may operate in a density-dependent manner to adjust population density to the available food supply. Most, if not all, earthworm species appear to undertake surface migrations to a greater or lesser degree (Mather and Christensen 1992), but the significance of this behavior for population processes is not known. If, as Mather and Christensen suggest, migration is primarily a resource-seeking activity, it could be important in enabling the population to locate more suitable habitats when conditions become unfavorable. Migration is undoubtedly significant in the colonization of new habitats, although active migration may often be less important than passive dispersal in this regard. There has been little debate about the relative importance of density-dependent and densityindependent factors in determining earthworm abundance. Although there is a lack of critical information on earthworm population dynamics, some of the conclusions that have emerged from studies of soil insects such as the garden chafer Phyllopertha horticola appear relevant (Milne 1984): Most of the times when populations are low, natural control is caused by the combined effects of weather, which is always density independent in its action, and interspecific competition of predators, parasites, and disease, which may be density independent or weakly density dependent but rarely if ever strongly density dependent in action. Under favorable conditions, when earthworms are able to “escape” from these natural control mechanisms, populations will ultimately be regulated by intraspecific competition acting in a strongly density-dependent manner. Virtually any form of human intervention will influence earthworm populations, often adversely, when the intervention is disruptive, as in mining and mechanical cultivation. One aspect of human intervention with potentially important consequences is accidental or deliberate introduction of exotic species, which could dramatically change abundance and species composition, possibly to the detriment of native species, and influence soils and plant production (see Chapter 5, this volume). However, there is considerable scope for promoting earthworm activity through management practices that remove constraints such as low pH and unfavorable moisture conditions, which minimize the adverse effects of cultivation and pesticide use and increase the food supply through organic amendment and increased crop residue return to the soil.

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Ma, W.-C., Brussard, L., and de Ridder, J.A. 1990. Long-term effects of nitrogenous fertilizers on grassland earthworms (Oligochaeta: Lumbricidae): their relation to soil acidification, Agric. Ecosyst. Environ., 30, 71–80. Ma, W.-C. and Eijsackers, H. 1989. The influence of substrate toxicity on soil macrofauna return in reclaimed land, in Animals in Primary Succession — the Role of Fauna in Reclaimed Lands, J.D. Majer, Ed., Cambridge University Press, Cambridge, U.K., pp. 223–244. Macdonald, D.W. 1983. Predation on earthworms by terrestrial vertebrates, in Earthworm Ecology — from Darwin to Vermiculture, J.E. Satchell, Ed., Chapman & Hall, London, pp. 393–414. Malecki, M.R., Neuhauser, E.F., and Loehr, R. 1982. The effect of metals on the growth and reproduction of Eisenia foetida (Oligochaeta, Lumbricidae), Pedobiologia, 24, 129–137. Marinissen, J.C.Y. and van den Bosch, F. 1992. Colonization of new habitats by earthworms, Oecologia, 91, 371–376. Martin, N.A. 1976. Effect of four insecticides on the pasture ecosystem V. Earthworms (Oligochaeta: Lumbricidae) and Arthropoda extracted by wet sieving and salt flotation, N.Z. J. Agric. Res., 19, 111–115. Martin, N.A. and Charles, J.C. 1979. Lumbricid earthworms and cattle dung in New Zealand pastures, in Proceedings of the Second Australasian Conference on Grassland Invertebrate Ecology, T.K. Crosby and R.P. Pottinger, Eds., Government Printer, Wellington, New Zealand, pp. 52–54. Mather, J.G. and Christensen, O. 1992. Surface migration of earthworms in grassland, Pedobiologia, 36, 51–57. Mele, P.M. and Carter, M.R. 1999. Impact of crop management factors in conservation tillage farming on earthworm density, age structure and species abundance in south-eastern Australia, Soil Tillage Res., 50, 1–10. Mellanby, K. 1966. Mole activity in woodlands, fens and other habitats, J. Zool. (London), 149, 35–41. Milne, A. 1984. Fluctuations and natural control of animal populations, as exemplified in the garden chafer Phylopertha horticola (L.), Proc. R. Soc. Edinburgh, 82B, 145–199. Moeed, A. 1976. Birds and their food resources at Christchurch International Airport, New Zealand. N.Z. J. Zool., 3, 378–390. Muldowney, J., Curry, J.P., O’Keefe, J., and Schmidt, O. 2003. Relationships between earthworm populations, grassland management and badger density in Co. Kilkenny, Ireland, Pedobiologia, 47, 913–919. Murphy, W.M., Mena Barreto, A.D., Silman, J.P., and Dindal, D.L. 1995. Cattle and sheep grazing effects on soil organisms, fertility and compaction in a smooth-stalked meadow grass-dominant white clover sward, Grass Forage Sci., 50, 191–194. Muys, B., Lust, N., and Granval, P. 1992. Effects of grassland afforestation with different tree species on earthworm communities, litter decomposition and nutrient status, Soil Biol. Biochem., 24, 1459–1466. Nielson, R.F. and Peterson, H.B. 1973. Establishing vegetation on mine tailings waste, in Ecology and Reclamation of Devastated Land, Vol. 2, R.J. Hutnik and G. Davis, Eds., Gordon and Breach, New York, pp. 103–115. Nordström, S. 1975. Seasonal activity of lumbricids in southern Sweden, Oikos, 26, 307–315. Nordström, S. and Rundgren, S. 1974. Environmental factors and lumbricid associations in southern Sweden, Pedobiologia, 14, 1–27. Nowak, E. 1976. The effect of fertilization on earthworms and other soil microfauna, Pol. Ecol. Stud., 2, 195–207. Oliver, J.H. 1962. A mite parasite in the cocoons of earthworms, J. Parasitol., 48, 120–123. Pernetta, J.C. 1976a. Bioenergetics of British shrews in grassland, Acta Theriol., 21, 481–497. Pernetta, J.C. 1976b. Diets of the shrews Sorex araneus L. and Sorex minutus L. in Wytham grassland, J. Anim. Ecol., 45, 899–912. Phillipson, J., Abel, R., Steel, J., and Woodell, S.R.J. 1976. Earthworms and the factors governing their distribution in an English Beachwood, Pedobiologia, 16, 258–285. Purvis, G. 1984. Earthworm Recolonization of Open Cast Coal Mine Sites Restored to Agriculture and Comparison with Populations on Unmined Agricultural Grassland, Technical Report MRPITR 1842 to the National Coal Board, U.K. Raw, F. 1962. Studies of earthworm populations in orchards. 1. Leaf burial in apple orchards, Ann. Appl. Biol., 50, 389–404. Raw, F. 1966. The soil fauna as a food source for moles, J. Zool. (London), 149, 50–54.

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Reinecke, A.J. and Visser, F.A. 1980. The influence of agricultural land use practices on the population densities of Allolobophora trapezoides and Eisenia rosea (Oligochaeta) in southern Africa, in Soil Biology as Related to Land Use Practices, D.L. Dindal, Ed., Environmental Protection Agency, Washington, D.C., pp. 310–324. Satchell, J.E. 1963. Nitrogen turnover by a woodland population of Lumbricus terrestris, in Soil Organisms, J. Doeksen and J. van der Drift, Eds., North Holland, Amsterdam, pp. 60–66. Satchell, J.E. 1967. Lumbricidae, in Soil Biology, A. Burges and F. Raw, Eds., Academic Press, London, pp. 259–322. Satchell, J.E. and Lowe, D.G. 1967. Selection of leaf litter by Lumbricus terrestris, in Progress in Soil Biology, O. Graff and J.E. Satchell, Eds., North Holland, Amsterdam, pp. 102–119. Satchell, J.E. and Stone, D.A. 1977. Colonization of pulverized fuel ash sites by earthworms, Publ. Centro Pirenaico de Biol. Exp., 9, 59–74. Saxena, D. and Stotzky, G. 2001. Bacillus thuringiensis (Bt) toxin released from root exudates and biomass of Bt corn has no apparent effect on earthworms, nematodes, protozoa, bacteria, and fungi in soil, Soil Biol. Biochem., 33, 1225–1230. Schmidt, O., Clements, R.O., and Donaldson, G. 2003. Why do cereal–legume intercrops support large earthworm populations? Appl. Soil Ecol., 22, 181–190. Schmidt, O., Curry, J.P., Hackett, R.A., Purvis, G., and Clements, R.O. 2001. Earthworm communities in conventional wheat monocropping and low-input wheat–clover intercropping systems, Ann. Appl. Biol., 138, 377–388. Southwell, L.T. and Majer, J.D. 1982. The survival and growth of the earthworm Eisenia foetida (Lumbricidae: Oligochaeta) in alkaline residues associated with the bauxite refining process, Pedobiologia, 23, 42–52. Stenersen, H. 1979. Action of pesticides on earthworms. Part. 1. The toxicity of cholinesterase-inhibiting insecticides to earthworms as evaluated by laboratory tests, Pest. Sci., 10, 66–74. Stenersen, J., Gilman, A., and Vardanis, A. 1973. Carbofuran: its toxicity to and metabolism by earthworms (Lumbricus terrestris), J. Agric. Food Chem., 21, 166–171. Stockdill, S.M.J. 1982. Effects of introduced earthworms on the productivity of New Zealand pastures, Pedobiologia, 24, 29–35. Stringer, A. and Wright, M.A. 1976. The toxicity of benomyl and some related 2-substituted benzimidazoles to the earthworm Lumbricus terrestris, Pest. Sci., 7, 459–464. Swift, M.J., Heal, O.W., and Anderson, J.M. 1979. Decomposition in Terrestrial Ecosystems, Blackwell, London. Temple-Smith, M.G., Kingston, T.J., Furlonge, T.L., and Garnsey, R.B. 1993. The effect of the introduction of the earthworms Aporrectodea caliginosa and Aporrectodea longa on pasture production in Tasmania, in Proceedings of the Seventh Australian Agronomy Conference, Adelaide, Australia, p. 373. Tomlin, A.D. and Miller, J.J. 1988. Impact of ring-billed gull (Larus delawarensis Ord.) foraging on earthworm populations of south-western Ontario agricultural soils, Agric. Ecosyst. Environ., 20, 165–173. Townsend, W.N. and Hodgson, D.R. 1973. Edaphological problems associated with deposits of pulverized fuel ash, in Ecology and Reclamation of Devastated Land, Vol. 1, R.J. Hutnick and G. Davis, Eds., Gordon and Breach, New York, pp. 45–56. Unwin, R.J. and Lewis, S. 1986. The effect upon earthworm populations of very large applications of pig slurry to grassland, Agric. Wastes, 16, 67–73. van Rhee, J.A. 1969. Inoculation of earthworms in a newly drained polder, Pedobiologia, 9, 128–132. van Rhee, J.A. 1977. Effects of soil pollution on earthworms, Pedobiologia, 17, 201–208. Vimmerstedt, J.P. and Finney, J.H. 1973. Impact of earthworm introduction on litter burial and nutrient distribution in Ohio strip-mine spoil banks, Soil Sci. Soc. Am. Proc., 37, 388–391. Waters, R.A.S. 1951. Earthworms and the fertility of pasture, Proc. N.Z. Grass. Assoc., 13, 168–175. Whalen, J.K., Parmelee., R.W., and Edwards, C.A. 1998. Population dynamics of earthworm communities in corn agroecosystems receiving organic or inorganic fertilizer amendments, Biol. Fertil. Soils, 27, 400–407. Wright, M.A. 1972. Factors governing ingestion by the earthworm Lumbricus terrestris (L.) with special reference to apple leaves, Ann. Appl. Biol., 70, 175–188. Zicsi, A. 1983. Earthworm ecology in deciduous forests in central and southeast Europe, in Earthworm Ecology — from Darwin to Vermiculture, J.E. Satchell, Ed., Chapman & Hall, London, pp. 171–177.

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Comprehensive Study of the 7 ATaxonomy and Ecology of the Lumbricid Earthworm Genus Octodrilus from the Carpathians Victor V. Pop and Adriana A. Pop Institute of Biological Research, Cluj-Napoca, Romania

CONTENTS The Carpathians .............................................................................................................................118 Taxonomy of the Lumbricid Genus Octodrilus ............................................................................118 Variability of Characters.........................................................................................................119 Taxa Discrimination and Identification of Octodrilus Species by Numerical Taxonomy..........................................................................................................120 Molecular Taxonomy and Phylogeny of the Octodrilus Species ..........................................122 Results .....................................................................................................................................122 16S rDNA Sequences....................................................................................................122 COI Gene Sequences ....................................................................................................123 Discussion......................................................................................................................124 Insularlike Accelerated Speciation in Octodrilus in the Carpathians....................................125 Ecology of the Genus Octodrilus ..................................................................................................125 Earthworm Communities Dominated by Octodrilus Species................................................125 Seasonal Dynamics of Earthworm Communities with Octodrilus frivaldszkyi ....................127 Site.................................................................................................................................128 Methods .........................................................................................................................128 Microclimatic Dynamics ...............................................................................................129 Seasonal Dynamics of the Earthworm Community .....................................................129 The Role of Giant Octodrilus Species in Building Up Vermic (Earthworm-Based) Characters in Mountain Soils ....................................................................................................130 The Earthworm Communities ................................................................................................134 Morphology and Micromorphology of Soil Profiles .............................................................135 Physical and Chemical Properties of Vermic Soils................................................................136 Experimental Study of the Influence of Calcophilous Octodrilus Species on Soil....................................................................................................................138 Material and Methods ...................................................................................................138 Results ...........................................................................................................................139

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Conservation of the Endemic Octodrilus Species in the Carpathians..........................................139 Effects of Air Pollution on Earthworms.................................................................................139 Effects of Sylvicultural Cutting on Endemic Earthworms ....................................................140 Acknowledgments ..........................................................................................................................140 References ......................................................................................................................................140

The genus Octodrilus includes a quite well-defined group of lumbricid earthworm species and has a relatively simple history. Örley (1895) established the genus Octolasion for lumbricids based on eight widely paired rows of setae, regardless of pigment. V. Pop (1941, 1948), for phylogenetic considerations, modified the diagnosis of the genus, keeping in it only species without red pigment. Omodeo (1956) divided the genus into the subgenera Octolasium and Octodrilus based mainly on the number of spermathecae; these subgenera were raised to the rank of genera by Bouché (1972). Zicsi (1986), based on the position of the male pores, divided the latter genus into two genera, Octodrilus and Octodriloides. In this concept, Omodeo (1956) includes medium-size to very large earthworms without pigment or pigmented in hues of gray or brown, with widely paired setae, four pairs of seminal vesicles, and five to eight pairs of spermathecae in the genus Octodrilus. The clitellar organs are very constant and characteristic of the species. Most Octodrilus species have relatively small distribution areas in central and eastern Europe or in North Africa. Endemic species occur in the Alps, Carpathians, and Dinara Mountains. A very few species, such as Octodrilus complanatus, are widely distributed in Europe. However, the genus Octodrilus is difficult to study because of many closely related species, often improperly described, with overlapping diagnostic characters. At least 54 Octodrilus species ÿ 1991). In the Romanian have been described to date (Zicsi and Pop 1984; Pop 1989: Mrsic Carpathians, 11 species and 7 subspecies are known (Pop 1941, 1948, 1973; Zicsi and Pop 1984; Pop 1991) (Table 7.1, Figure 7.1). Very little is known about the general biology and ecology of Octodrilus species. They occur mostly in remote mountain regions, mainly on limestone soils, in which their sampling is very difficult. Hence, studies of Octodrilus species have been neglected. During 20 years of research in the Carpathians, in different projects studying the structure of mountain ecosystems, we have collected many samples of earthworms belonging to Octodrilus species. Some of the local earthworm populations, especially those in the Apuseni Mountains, raised difficult taxonomic problems, particularly concerning the variability of certain characters. Taxonomic studies have led to more theoretical ones, such as of the patterns of speciation, followed by ecological studies, in both the field and the laboratory. The research direction that took shape covered a large range of topics concerning many aspects of this interesting group of earthworms. Results of this research have been presented at various international symposia on earthworm ecology. These results include (1) the first attempt to separate Octodrilus species by numerical taxonomy (the Darwin Centenary Symposium on Earthworm Ecology, Grange-over-Sands, U.K., 1981); (2) the theoretical background of Octodrilus species discrimination based on a hypothesis of accelerated insularlike speciation (the Michaelsen Memorial Symposium on Terrestrial Earthworms, Hamburg, 1987); (3) the structure of earthworm communities and the role of large Octodrilus species in building up certain so-called vermic characters in mountain soils (the Rosa Symposium, Bologna, Italy, 1985; Fourth International Symposium on Earthworm Ecology (ISEE 4), Avignon, France, 1990; the 11th International Colloquium on Soil Zoology, Jyväskylä, Finland, 1992; and ISEE 5, Columbus, OH, 1994); and (4) molecular taxonomy research (the ISEE 7, Cardiff, U.K., 2002). This chapter provides a summary of the main results of all this research, whether previously published or not. In addition, new laboratory research on the structure of deoxyribonucleic acid (DNA) in Octodrilus species is presented. © 2004 by CRC Press LLC

O. O. O. O. O. O. O. O. O. O. O. O. O. O. O. O. O.

aporus (Pop 1989) bihariensis (Pop 1989) b. bihariensis (Pop 1989) b. rendzinicola (Pop 1989) compromissus (Zicsi and Pop 1984) c. minimus (Pop 1989) c. compromissus (Pop 1989) exacystis (Rosa 1896) e. meziadensis (Pop 1989) e. exacysts (Pop 1989) e. oresbius (Pop 1989) frivaldszkyi (Örley 1880) lissaensis (Michaelsen 1891) ophiomorphus (Pop 1989) permagnus (Pop 1989) robustus (Pop 1973) transylvanicus (Zicsi and Pop 1984)

Male Pores

Spermathecae

Tubercula pubertatis

Clitellum

Segments

Typhlosole endig

Length (mm)

Diameter (mm)

15 15 15 15 15 15 15 15 15 15 15 15 15 15 15 15 16–19

6 6 6 6 6 6 6 6 6 6 6 6 6 6 6 5 6

30–40 29–38 29–38 29–38 29–37 29–37 29–37 30–38 30–38 30–38 30–38 29–37 29–36 30–38 30–39 30–38 29–36

29, 30–40 29–37 29–37 29–27, 38 (28), 29–36, (37) 29–36 (28), 29–36, (37) 29, 30–37, 38 30–37 29, 30–37, 38 29, 30–37, 38 28, 29–36, 37 29–36 29, 30–37, 38 29, 30–38, 39 29–38 29–36

207–273 98–256 98–195 185–256 125–206 125–161 161–206 100–271 100–187 165–271 180–259 189–262 98–150 202–263 232–262 224–253 91–156

155–163 72–158 72–120 125–158 90–136 90–111 116–136 88–162 88–118 125–152 140–162 147–170 ND 155–174 155–179 164–170 78–98

275–460 94–240 94–143 126–240 63–195 66–105 63–193 60–310 60–135 103–310 140–290 200–400 80–150 225–420 320–720 200–300 40–70

14–18 4–8 4–7 7–8 3–7 3–4 4–7 4–10 4–5 6–9 6–10 8–15 4–5 10–16 11–17 13–15 4–6

Note: Numbers in parentheses indicate differently colored, smaller clitellum segments. ND = no data.

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TABLE 7.1 The Octodrilus Species from the Southeastern and Eastern Carpathians

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FIGURE 7.1 Species of the Octodrilus genus from the Carpathians. Anterior end, lateral view: (a) O. aporus; (b) O. frivaldszkyi; (c) O. compromissus; (d) O. exacystis.

THE CARPATHIANS The southern and southeastern Carpathians, belonging to the Alpino-Carpathic orogenic system in Romania, form a 900-km long, 35- to 150-km wide mountain chain with a median altitude of 840 m and a highest peak of 2544 m (Figure 7.2). The main vegetation belts consist of oak (Quercus petraea), beech (Fagus sylvatica), spruce fir ( Picea abies) forests and subalpine-alpine grassland on cambisols, argiluvisols, spodosols, and with islands of mollisols.

TAXONOMY OF THE LUMBRICID GENUS OCTODRILUS Species separation in the genus Octodrilus is difficult because of the close similarities of many species. Many authors, when describing new species, have not indicated clearly the range of variation of diagnostic characters. Moreover, as these characters were often interpreted differently, complexes of species that are difficult or impossible to delimit on objective criteria have resulted. The synthesis of the genus Octodrilus and its separation into the two subgenera Octodrilus and Octodriloides by Zicsi (1986) clarified many aspects of species discrimination. Nevertheless, the genus Octodrilus still remains difficult.

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ÿ FIGURE 7.2 The Carpathians in the Alpine mountain system. (After V. Mihailescu 1963.)

VARIABILITY

OF

CHARACTERS

The systematic status assigned to different Octodrilus species populations depends on the range permitted in variability of characters. Thus, if wide variability in clitellar organs is assumed, some of the species described could be combined to form a single species, but by limiting the variability, they could be regarded as distinct species or, in certain cases, as subspecies of polytypic species. Because of a present lack of agreement on the taxonomic significance of some characters of lumbricid species, and even of lumbricid genera, we have studied the variability in morphology and anatomy in a range of material of Octodrilus (approximately 2000 specimens) from almost the whole range of known species. In addition to the V.V. Pop collection kept with the Institute of Biological Research in Cluj-Napoca, Romania, the collection of Prof. V. Pop (1903 to 1976) at the University of Cluj-Napoca, Prof. A. Zicsi’s collection at the Eötvös Lorand University in Budapest, Hungary, and Michaelsen’s collection in Hamburg, Germany, were studied. Statistical processing of data on the variability of characters in local populations of the Octodrilus species from the Carpathians and the Alps, when correlated with available data on other species of the genus, leads to the following ordination of their relative diagnostic value within the genus Octodrilus (Pop 1991): At (sub)genus level: Position of male pores At species level: Number and position of spermathecae; position of tubercula pubertatis; position of clitellum; presence or absence of dorsal pores; setal ratio; shape of calciferous glands At subspecies level: Extension or length of typhlosole; number of segments; body size and mass Other characters are sometimes also important but are seemingly of less-discriminatory value. Two ecological groups of Octodrilus species were recorded in the Carpathians: (1) widely distributed species, like Octodrilus exacystis and Octodrilus compromissus, found in a range of different habitats and displaying a wide variability in nonspecific characters (size, segment number, typhlosole

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ending) and seemingly a polytypic species; and (2) endemic species, like Octodrilus frivaldszkyi or Octodrilus aporus, which are confined to small and distinct habitats mainly on Mesozoic limestone and which appear to be monotypic species with a limited variability of characters.

TAXA DISCRIMINATION AND IDENTIFICATION OF OCTODRILUS SPECIES BY NUMERICAL TAXONOMY Taxa identifications have to be based on a thorough examination of the variability of the diagnostic characters in local populations. The high constancy in some sexual characters, like the number and position of the spermathecae, the position of the tubercula pubertatis, and, to a certain extent, the position of the clitellum were confirmed statistically and were considered the major diagnostic characters at the species level (Pop 1989, 1991). The position of the tubercula pubertatis seems to have a particularly good diagnostic value at the species level. We have never found any local species population with variable positions of the tubercula pubertatis, as occurs in other lumbricid genera. When differences of even one segment in the position of tubercula pubertatis occurred in the same sample, there were other differences as well, which allowed the separation of the individuals into distinct taxa. On the basis of a study of the variability of characters in more than 1500 individuals, it is justifiable to recognize as many Octodrilus species as there are different positions of tubercula pubertatis (Figure 7.3). Subspecies discrimination presents yet another difficult taxonomic task. Subspecies in lumbricids have so far been recognized on classical criteria. According to Mayr (1971), “a subspecies is an aggregate of phenotypically similar populations of a species inhabiting a geographic subdivision of the range of the species and differing taxonomically from other populations of the species.” That is, a subspecies has a character of spatial unity, a subdivision of a species in the dimensions of longitude and latitude. However, habitat differences may occur over small distances in the case of insular distributions of biotopes in mountain regions. In many areas of the Carpathians, patchy distribution patterns of biotopes are evident. Islands of soils developed on limestone are surrounded by soils developed on acid parent material, and there are grassland areas that surround forest islands. Soils, valleys, and vegetation act as barriers, so there is scope for allopatric, accelerated, insularlike speciation.

FIGURE 7.3 Pattern of Octodrilus taxa discrimination by the main diagnostic characters.

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If we admit that insular speciation may exist in these mountains, we must admit a similar process of subspecies differentiation. Widespread species, inhabiting multiple kinds of biotopes, show clear morphological differences in the framework of species diagnostic characters. Taxonomic differences among subspecies concern mostly characters related to the size of the earthworms (i.e., the length, number of segments, and ending of the typhlosole). The location of the typhlosole ending, which is a very constant character in local Octodrilus species populations, is positively related with the size and number of segments and has proved to be a convenient character for subspecies discrimination according to the “75% rule” (Mayr et al. 1953). The shape and length of the typhlosole may have an ecological significance as an adaptation to the quality and quantity of food. The size of the earthworms, which is positively related with the typhlosole ending, could also be regarded as an adaptation to soil conditions. In these mountains, the smaller species generally dwell in more compact soils under grassland; the larger ones tend to inhabit looser forest soils. Thus, statistically delimited groupings of local octodrilus populations might be regarded as also having different ecological requirements and could therefore be considered as distinct subspecies. Based on the above ordination of the diagnostic value of characters, we described, from a relatively small area in the Apuseni Mountains, five new Octodrilus species and seven new subspecies and amended the diagnoses of three species (Pop 1989; Zicsi and Pop 1984) (Table 7.1). This number of newly described taxa suggested that the Apuseni Mountains should be considered as an active speciation center of the genus Octodrilus. The validity of the species and subspecies delimitations was checked using a multiple character analysis without previous subordination of the diagnostic value of the characters. Following Sims (1969), we used the simple Sheals’ method (1964). The dendrograms, drawn according to Mountford (1962), used in classifying taxa (Figure 7.4) resemble very much the key drawn for delimiting taxa by the main diagnostic characters (Figure 7.3). 90–111 –37

116–136

29– 72–120 –39 125–158 88–118 125–152 –38

140–162

30– –39 –40

Tubercula pubertatis

FIGURE 7.4 Classification of Octodrilus taxa from the Apuseni Mountains by multiple character analyses.

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The dendrograms demonstrated the existence of two groupings of species. The first group, with O. frivaldszkyi, O. compromissus, and O. exacystis, has conspicuous dorsal pores and a somewhat shorter (nine-segment) tubercula pubertatis and clitellum. The second group of species has inconspicuous dorsal pores and longer clitellar organs (10 to 11 segments). Octodrilus aporus, with no dorsal pores, has a separate position. Even though it is only a phenotypical tree, not a phylogenetical one, the dendrograms suggest a truly divergent position of the four larger Octodrilus species, namely, O. frivaldszkyi, O. aporus, Octodrilus ophiomorphus, and Octodrilus permagnus, and the high probability that their discrimination is objective. In describing new species or in delimiting the range of characters for already described species, we adopted a rigorous statistical procedure. Thus, tables with statistically processed data are part of the description. The type series, in the majority of taxa discussed here, had tens or even hundreds of individuals. The type populations and the holotypes, as well as the neotype for O. frivaldszkyi, were selected to match as nearly as possible the arithmetic mean value of the numeric characters of the whole type series. Thus, the “nomenclatural type” or the “name bearer” of newly described or amended taxa corresponds to the “morphological” or the “biological” types. It is important to notice that such a selection avoids hazardous designation of holotypes from marginal, overlapping populations of taxa.

MOLECULAR TAXONOMY

AND

PHYLOGENY

OF THE

OCTODRILUS SPECIES

It is well known that some taxonomists do not accept, or look on without some reservation, species described by numerical taxonomy. Therefore, we had some degree of uncertainty concerning the validity of taxa we described as new in 1976 and 1989. Namely, are the four giant Octodrilus species from the Carpathians really distinct species, or assuming a higher variability of genital characters, should they have been incorporated into the old O. frivaldszkyi? Or, was it correct to distinguish Octodrilus bihariensis from O. exacystis or from O. compromissus based mainly on a difference of only one segment in the position of the clitellar organs? A. Pop studied this species by molecular taxonomy, which is expected to provide new and more precise ways of delimiting genera as well as establishing phylogenetic distances among species. New, fresh individuals of these few species were collected from the already known sites in the type localities. The molecular research was done in the laboratories of the Institute of Pharmacy and Molecular Biotechnology, Ruprecht Karls University in Heidelberg, Germany. Structure of two components of DNA, namely, the 16S ribosomal DNA (rDNA) sequences and cytochrome c oxidase (COI) genes, were established by standard or slightly modified methods (A.A. Pop et al. 2003). Six of the seven Octodrilus species occurring in the Apuseni Mountains were studied (Table 7.2). Subspecies only for O. bihariensis were resampled. For the rest of polytypical species, only the nominate subspecies were analyzed. Partial sequences for 16S rDNA of seven taxa (six species and one subspecies) and subunit I for COI for the same seven taxa were obtained. The relative nucleotide structure of analyzed genes is presented in Table 7.2, and the maximum parsimony trees are shown in Figure 7.5 and Figure 7.6. Both cladograms are rooted with a snail species, Biomphalaria tenagophila (Mollusca, Gastropoda, Planorbidae).

RESULTS 16S rDNA Sequences The parsimony analysis of molecular data (491 bp) resulted in a single optimal tree with length 314, consistency index 0.9427, homoplasy index 0.0573, and retention index 0.6087. Of the 491 nucleotide positions, 230 (46.84%) were constant, 229 (46.64%) were variable and parsimony uninformative, and 32 (6.52%) were parsimony informative.

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TABLE 7.2 Nucleotide Structure of 16S rDNA and Cytochrome c Oxidase Sequences in Octodrilus Species 16S

COI

Octodrilus Species

bp

GC%

AT%

GC/AT

bp

GC%

AT%

GC/AT

O. O. O. O. O. O. O.

497 476 476 475 474 476 500

41.05 41.18 40.97 39.79 38.82 41.18 41.40

58.15 58.82 59.03 60.21 60.13 58.61 57.80

0.70 0.70 0.69 0.66 0.64 0.70 0.71

672 652 650 651 651 650 667

46.43 46.01 46.77 45.01 47.00 44.46 46.63

52.83 53.83 53.08 54.84 53.00 53.38 52.77

0.87 0.85 0.88 0.82 0.88 0.83 0.88

aporus b. bihariensis b. rendzinicola compromissus excacystis frivaldszkyi permagus

bp = base pairs; A = adenine; C = cytosine; G = guanine; T = thymine

Biomphalaria

O. exacystis

O. compromissus

O. aporus

O. permagnus

O. frivaldszkyi

O. b. bihariensis

O. b. rendzinicola FIGURE 7.5 Maximum parsimony tree from six Octodrilus species based on 16S rDNA analysis.

COI Gene Sequences The parsimony analysis of molecular data (654 bp) resulted in an optimal tree with length 494, consistency index 0.7753, homoplasy index 0.2247, and retention index 0.4011. Of the 654 nucleotide positions, 357 (54.59%) were constant, 153 (23.39%) were parsimony uninformative, and 144 (22.01%) were parsimony informative.

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Biomphalaria O. frivaldszkyi O. compromissus

O. b. bihariensis

O. b. rendzinicola

O. exacystis

O. aporus

O. permagnus

FIGURE 7.6 Maximum parsimony tree from six Octodrilus species based on COI analysis.

Discussion The magnitude of our data, summarized in Table 7.2, agrees with that already known in lumbricids. Moreover, in the 16S sequences, our data ranged between 474 and 500 bp and seem to be more complete than others reported. In the COI sequences, our data ranged between 650 and 672. The very low homoplasy index of 0.0573 (required theoretic level under 0.5) proves that the maximum parsimony tree presented here for ribosomal fragments of 16S rDNA is correct. Nevertheless, the 6.51% parsimony informative characters are quite low. Comparatively, the 22.01% of parsimony informative characters and a homoplasy index of 0.2247 suggest a higher taxonomic discriminatory value of the fragments of mitochondrial CO (COI). Thus, at this stage of our research, the image of the cladogram for 16S (Figure 7.5) seems to be less informative than that for COI (Figure 7.6). The clustering of species in both cladograms shows the existence of the same number of taxa as differentiated by single- (Figure 7.3) and multiple-character (Figure 7.4) numerical analyses. To answer the basic questions of how many Octodrilus species occur in the Carpathians; the hypothetical and empirical trees from Figures 7.3 to Figure 7.6 are analyzed. First, the three giant species were placed on different branches, which suggests that they are really different taxa. The distance between O. aporus and O. permagnus is relatively small, in accordance with the original description, but they differ clearly from O. frivaldszkyi. Second, the position and distances among O. bihariensis, O. compromissus, and O. exacystis prove that they are also different species. The relative positions of the six Octodrilus species in the tree diagrams are different but resemble their positions in the hypothetical single-character, step-by-step tree, or in multiple-character analyses. For example, the grouping of O. frivaldszkyi with O. bihariensis and O. compromissus and of O. exacystis with O. aporus and O. permagnus fit the patterns obtained by numerical taxonomy. The subspecies of O. bihariensis show very close relatedness. In conclusion, molecular taxonomy research confirms or validates the species that were identified and described by numerical taxonomy. Species clustering and branching indicate possible phylogenetic relationships. © 2004 by CRC Press LLC

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INSULARLIKE ACCELERATED SPECIATION

IN

OCTODRILUS

IN THE

125

CARPATHIANS

The large number of endemic Octodrilus species distributed in patches in small areas of the Carpathians, as well as in the Alps and Dinara Mountains, could be explained by a process of insularlike speciation (Pop et al. 1994). The following conditions are considered to favor accelerated insularlike speciation: Stenobiontism (species that tolerate only narrow variations of environmental factors) regarding mountain habitats on Mesozoic limestone Insularity of these habitats, leading to geographical and reproductive isolation The size of populations, which is an essential factor because most of the isolated Octodrilus populations are rather small, thus enhancing the probability of quick genetic changes Low dispersal ability because earthworms are almost sedentary forms The occurrence of parthenogenesis, indicated by indirect morphologic features; this kind of reproductive isolation might also be an accelerating factor for species segregation because a mutation is more easily preserved and spread within a population in the absence of genetic recombinations through biparentalism The Apuseni Mountains are seemingly one of the most prolific speciation centers of the genus. Timing of the Alpino-Carpathian orogenesis (i.e., raising of the mountains) and phytohistorical data support the idea that endemic Octodrilus species were able to survive the Pleistocene glaciations in Carpathian refuges. Distinct Octodrilus lines are found in the limestone areas of the Carpathians, the Alps, and the Dinara Mountains.

ECOLOGY OF THE GENUS OCTODRILUS EARTHWORM COMMUNITIES DOMINATED

BY

OCTODRILUS SPECIES

As discussed in the section on variability of characters, two groups of Octodrilus species could be distinguished: a group of widely distributed, usually acid-tolerant, medium-size species and a group of endemic species, usually calcophilous or confined to limestone areas, that are medium-size to very large species of earthworms (Pop 1985, 1997). The great majority of the Octodrilus species were recorded in the vegetation belts of beech or beech-hornbeam forests and more rarely in mixed beech spruce-fir forests, as well as in corresponding grasslands (Pop 1982). Beech (F. sylvatica) is the main forest tree in Romania, occurring on both aspects of the Carpathians and forming a distinct vegetation belt, both as mixed beech-spruce-fir or beechhornbeam forests. The obvious individuality of Carpathian beech forests justifies their placement in an independent phytosociological alliance, namely, Symphyto cordato-Fagetum Vida 59, Fagion dacicum Soó 62. The earthworm fauna of the Carpathian beech forests differs correspondingly from that of similar biotopes in the Alps or Dinara Mountains. In this forest biotope are found the majority of Carpathian endemic species and, as a particular feature, some of the largest lumbricids known. The earthworm communities (Pop 1997), previously referred to as synusia (Pop 1982, 1985, 1987), comprise two main structural patterns related to the nature of the substrata (igneouscrystalline vs. limestone-dolomite rocks), which in turn reflects peculiarities of the soils and, within certain limits, some features of the phytocenoses. The structural pattern of these earthworm communities is maintained by a substitution process of the ecologically similar species, which are as a rule vicariants, that is, they do not occur together in the same place. This substitution process was found mainly at the level of the same ecologic categories, namely, epigeic species (surface living), endogeic species (living in the depth of soil), and anecic species (vertical migratory) species. Their ratios are related closely to soil properties

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B.P. 1

B.E. 2

3

4

Fir + beech Fir

Beech

Vegetation Soil Site Species Allolobophora caliginosa (Savigny, 1826) Allolobophora dacica (Pop, 1938) Allolobophora georgii Michaelsen, 1890 Allolobophora mehadiensis Rosa, 1895 Allolobophora rosea (Savigny, 1826) Dendrobaena alpina (Rosa, 1884) Dendrobaena byblica (Rosa, 1893) Dendrobaena clujensis Pop, 1938 Dendrobaena octaedra (Savigny, 1826) Dendrobaenarubida (Savigny, 1826) Fitzingeria platyura Fitzinger, 1833 Lumbricus polyphemus (Fitzinger, 1883) Lumbricus rubellus Hoffmeister, 1843 Octolasion lacteum Örley, 1885 Octodrilus c. compromissus Zicsi & V. V. Pop, 1984 Octodrilus exacystis exacystis (Rosa, 1896) Octodrilus e. oresbius V. V. Pop, 1989

Hornbeam

Beech

Earthworm Ecology, Second Edition

Oak

126

A.B. 5

6

7

B.F. B.P. 8

9 10

B.P., Brown podzolic soil; B.E., Brown earth; A.B., Acid brown soil; B.F., Brown forest soil. Dominance % 0

0–10

10.1–20

20.1–40

40.1–60

60.1–100

FIGURE 7.7 Structure of earthworm communities with O. compromissus–Dendrobaena byblica in forest ecosystems with acid soils from the Carpathians.

and the composition of the vegetation cover. The two main community types were named tentatively by characteristic species following Cassagneau’s procedure (1961) when defining Collembola synusia in the Pirinei Mountains. The earthworm community with O. compromissus-Dendrobaena byblica (Figure 7.7) is characteristic of beech and mixed beech-spruce fir forests with oligobasic, cambic, spodic, or argillic soils, with acid mulls, and developed on igneous or crystalline acid parent material. The species Dendrobaena alpina is often also present, but D. byblica could be substituted partially by Dendrobaena clujensis or Dendrobaena veneta. In places with deeper soils, the endogeic O. compromissus or Octodrilus exacystis dominate the earthworm biomass. The large species that are characteristic of the second community occur sporadically and at low population densities. The earthworm community with O. frivaldszkyi (Figure 7.8) is very characteristic for beech or mixed beech-hornbeam forests, over rendzinas or cambic eubasic of mesobasic soils (brown earth), with calcic mulls on limestone or dolomite (Pop 1985, 1987). This community type also occurs, very rarely, in neighboring acid soils. This earthworm community is dominated by giant species of Octodrilus, namely, O. frivaldszkyi, O. aporus, O. ophiomorphus, or O. permagnus, reaching 40 to 70 cm in length. These very large species, observed to be vertical burrowers or anecic earthworms, are vicariants, but one of them is always present. Lumbricus polyphemus, considered an anecid or vertical migratory species, seems to be one of the characteristic species of this type of earthworm community. However, in some places, it is replaced by a large population of the giant Octodrilus species (Pop 1980). Characteristic epigeic species are the red-pigmented D. byblica or D. clujensis. The presence of the small D. alpina is determined by the presence of spruce fir trees. Endogeic, medium-size species of earthworm are O. compromissus, O. exacystis, Allolobophora dacica, or Octolasium

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Soil Species Site Allolobophora dacica (Pop, 1938) Allolobophora georgii Michaelsen, 1895 Allolobophora rosea (Savigny, 1826) Dendrobaena alpina (Rosa, 1884) Dendrobaena byblica (Rosa, 1893) Dendrobaena clujensis Pop, 1938 Dendrobaena rubida (Savigny, 1826) Lumbricus polyphemus (Fitzinger, 1883) Lumbricus rubellus Hoffmeister, 1843 Octolasion lacteum (Örley, 1885) Octodrilus aporus V. V. Pop, 1989 Octodrilus bihariensis V. V. Pop, 1989 Octodrilus compromissus Zicsi & V. V. Pop, 1984 Octodrilus exacystis (Rosa, 1896) Octodrilus frivaldszkyi (Örley, 1880) Octodrilus ophiomorphus V. V. Pop, 1989 Octodrilus permagnus V. V. Pop, 1989

R

T.R.

1 2

3 4

B.E.

5 6

7

127

P.B.

8 9 10

FIGURE 7.8 Structure of earthworm communities with O. frivaldszkyi in forest ecosystems on limestone from the Carpathians.

lacteum. The substitution of ecologically similar species seems to follow certain zones along the Carpathian range. In the southwestern part (Mehedinti, Cerna, and Retezat Mountains), Allolobophora species often substitute for the endemic Octodrilus species. Here, the very large (40 to 80 cm long) Allolobophora robusta substitutes for the giant Octodrilus species; Allolobophora mehadiensis occurs instead of medium-size Octodrilus species. In the northern part of the eastern Carpathians, the Octodrilus species are replaced by Allolobophora carpathica. Biogeography and geochronological timing could explain the functional relationships among earthworm communities, vegetation, and soil types. Long convolution and coadaptation of the beech-earthworm association might also explain the perfection of these relationships (Pop 1982). Earthworm communities with Octodrilus species in grassland ecosystems are shown in Figure 7.9. The earthworm communities in soils developed under grassland display less-interesting structures. In areas with acid soils, only the relatively widespread O. exacystis and O. compromissus occurred. In limestone areas, endemic earthworm species such as O. bihariensis and O. exacystis meziadensis typified the local peculiarities.

SEASONAL DYNAMICS

OF

EARTHWORM COMMUNITIES

WITH

OCTODRILUS

FRIVALDSZKYI

The Padis karstic plateau, situated in the middle of the Apuseni Mountains (Figure 7.2), a quite large limestone area, is covered by beech forests with patches of fir tree forests (on acid rocks) and represents the typical habitat for O. frivaldszkyi and O. bihariensis, two of the most interesting species of this genus. These species also indicate the presence of other endemic soil-dwelling invertebrates. The scientific importance of several well-known caves and the peculiar terrestrial flora and fauna of the central part of the Apuseni Mountains have led to preliminary studies of making a nature reserve of the Padis karstic area. It is in this framework that we have studied the structure of the earthworm communities. In 1979, seasonal earthworm dynamics were studied at three representative sites, namely, in a beech forest (Padis), a mixed beech-spruce fir forest (Pârâul Ponor) and a spruce fir tree forest (Calineasa). In the beech

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Soil Site

Species

T.R. 1

2

E.B.E. C.A. 3

4

5

Allolobophora caliginosa (Savigny, 1826) Allolobophora dacica (Pop, 1938) Allolobophorarosea (Savigny, 1826) Dendrobaena byblica (Rosa, 1893) Dendrobaena clujensis Pop, 1938 Dendrobaena rubida (Savigny, 1826) Lumbricus polyphemus (Fitzinger, 1883) Lumbricus rubellus Hoffmeister, 1843 Octolasion lacteum Örley, 1885 Octodrilus b. bihariensis V. V. Pop, 1989 Octodrilus b. rendzinicola V. V. Pop, 1989 Octodrilus c. compromissus Zicsi & V. V. Pop, 1984 Octodrilus e. exacystis Rosa, 1896 Octodrilus e. oresbius V. V. Pop, 1989

T.R., Terra rossa. E.B.E., Brown earth. C.A., Colluvial–alluvial. Dominance % 0

0–10

10.1–20

20.1–40

40.1–60

60.1–100

FIGURE 7.9 Structure of earthworm communities with Octodrilus species in grassland ecosystems from the Carpathians.

forest from Padis, the peculiar, obvious vermic characters developed in the soil by the giant O. frivaldszkyi and the large O. bihariensis were first recorded (Pop and Postolache 1987). In this chapter, only the seasonal dynamics of the earthworm community that was dominated by Octodrilus species from the beech forest is presented. Mean data for the other two sites studied are given only for comparison of the earthworm biomass. Site The site was the Padis karstic plateau, Bihorului Mountain, in the Apuseni Mountains, at 1300-m altitude, south-southwest aspect, 20% slope, and in a beech forest (As. Symphyto cordati-Fagetum Vida 1959, facies with Allium ursini). The soil is a cambic rendzina with calcic mull and the following profile: O (0 to 2 cm), Am (0 to 25 cm), AmBv (25 to 36 cm), BvR (36 to 46 cm), and R (46 cm or deeper). Methods A stratified random earthworm sampling program was initiated that involved sampling earthworm populations at monthly intervals during May to October 1979 (months without snow cover). Earthworms were expelled from the soil by formalin; there were nine sample units, each with 50cm by 50-cm surface area. Biomass was estimated by weighing the earthworms preserved in 4% formalin. Comparative weighing of worms, with and without emptying the gut content, showed that the loss of weight through preservation in formalin corresponded approximately with the weight

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AIR A 0–5 cm A 5–10 cm A 10–15 cm

Temperature °C

18 16 14 12 10 8 6 4 2 0

May

June

July

Aug.

Sept.

Oct. 1979

300

OLHF A 0–5 cm A 5–10 cm A 10–15 cm

250 Moisture %

129

200 150 100 50 0 May

June

July

Aug.

Sept.

Oct.

FIGURE 7.10 Seasonal dynamics of microclimate in soil.

of the gut content. Earthworms were identiÞed to species level, and each species was divided into three age groups (adult, subadult, juvenile). Microclimatic Dynamics Seasonal dynamics of air and soil temperature (O-litter or organic stratum and the A horizon at 0 to 5, 5 to 10, and 10 to 15 cm) are shown in Figure 7.10. Air temperature ranged between 11 and 15°C. Soil temperature has two maxima, in June and in September. Soil moisture, more constant in deeper horizons, has a somewhat similar seasonal dynamics. Seasonal Dynamics of the Earthworm Community Five species were at the study site (Table 7.3). The community structure is the one described as a community with O. frivaldszkyi in the separate section above. Octodrilus frivaldszkyi is one of the giant species characteristic of the limestone areas in these mountains and is often associated with the endogeic O. bihariensis rendzinicola. Dendrobaena alpina, a small epigeic worm, is mostly characteristic of coniferous forests in the Carpathians. The study area has sparse Þr trees, which explains the presence and low density of this species. Dendrobaena byblica and D. clujensis, which are red-pigmented epi-endogeic species of earthworms, are common in the beech and Þr tree forests of the Carpathians. The biomass of the earthworm community is dominated throughout the year by the two Octodrilus species (Figure 7.11). The seasonal dynamics of mean density and biomass of individual earthworm species are summarized in Figure 7.12 and Figure 7.13.

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TABLE 7.3 Characters of the Lumbricid Species in the Beech Forest from Padis (the Apuseni Mountains) with Ecological Significance Character Species

Pigment

Length (mm)

Diameter (mm)

Fresh Weight (g)

Ecologic Category

D. D. D. O. O.

Reddish Red Red Gray Gray

40–50 45–105 60–120 126–240 200–460

2–3 3–4 4–5 4–5 8–15

0.02–0.07 0.23–0.40 1.05–2.43 4.80–9.05 19.00–36.60

Epigeic Epigeic Epi-endogeic Endogeic Anecic

alpina byblica clujensis bihariensis frivaldszkyi

It should be noted that a biomass of 256 g/m2 is the highest earthworm biomass recorded in the Carpathians. For comparison, Table 7.4 shows the earthworm biomass recorded in two neighboring forest sites investigated at the same time as the Padis beech forest. The yearly mean earthworm biomass of the community with the giant O. frivaldszkyi exceeds by more than 30 times that recorded for fir tree forests and by more than 20 times that for mixed beech-fir tree forests. The earthworm densities and biomass of these other communities are typical for the Carpathians (Pop 1987). This enormous earthworm biomass, at least for Carpathian forests, implies extremely high earthworm activity in soil and, as discussed in the next section, accounts for the characteristics of this soil.

THE ROLE OF GIANT OCTODRILUS SPECIES IN BUILDING UP VERMIC (EARTHWORM-BASED) CHARACTERS IN MOUNTAIN SOILS The very large Octodrilus species imprint conspicuous vermic (i.e., earthworm-based) characters, mainly in soils developed on limestone or dolomite but more rarely in neighboring acid soils. This earthworm activity is so intense that soils developed on different parent materials and usually classified into different classes have quite similar structures and chemical properties in the upper horizons. The term vermic (earthworm based), introduced into soil systematics by the American Seventh Approximation Survey (1960), indicates soils processed intensively by soil invertebrates, especially earthworms. According to the Romanian system of soil classification (1979), the vermic term characterizes soils that exhibit coprolites (casts of earthworms or of other soil-inhabiting invertebrates) and earthworm burrows (sometimes filled by soil material) in more than 50% of the volume of the A horizon and in more than 25% of the volume of the subsequent horizon. Using this definition, the vermic character has been considered diagnostic of only a few soil types, such as chernozems in the class of mollisols. The definition can be misleading because a normally developed soil must be vermic, that is, developed by some participation of earthworms and other soil invertebrates. Kubiena (1953), when defining mull humus, stated: “Practically all soil aggregates are earthworm casts or residues of them.” His statement should be extended because traces of the activity of soil invertebrates, especially earthworms, can be shown micromorphologically throughout the entire profile of most soil types. Hence, in the opinion of Pop and Postolache (1987), the term vermic should be used to indicate only visible, macromorphologically stable, and lasting aggregates produced by soil invertebrates. On this basis, we studied the morphology and chemical characteristics of soils inhabited by the giant Octodrilus species, and the morphological and micromorphological features of a vermic

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131

N*m–2 80 70 60 50 40 30 20 10 0 May

Total density

June

% 100 90 80 70 60 50 40 30 20 10 0 May

g*m

July

Aug.

Sept.

Oct.

O. bihariensis O. frivaldszkyi Octodrilus unident. juv. D. alpina D. byblica

D. clujensis

June

July

Aug.

Sept.

Oct.

1979

–2

300 250 200 150 100 Total biomass

50 0 May

June

July

Aug.

Sept.

Oct.

% 120

O. bihariensis

100 80

O. frivaldszkyi

60 40

Octodrilus unident. juv. D. alpina D. byblica D. clujensis

20 0 May

June

July

Aug.

Sept.

Oct.

FIGURE 7.11 Seasonal dynamics of the earthworm community with O. frivaldszkyi in the vermic cambic rendzina from Padis (the Apuseni Mountains).

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–2

N*m

100

Total density

80 60 40 20 0 8

O. bihariensis

6 4 2 0 15

O. frivaldszkyi

10 5 0 30 25 20 15 10 5 0

Octodrilus unident. juv.

D. alpina

15 10 5 0 20

D. byblica

15 10 5 0 20

D. clujensis

15 10 5 0 May

June

July

Aug.

Sept.

Oct.

1979

FIGURE 7.12 Seasonal dynamics of lumbricid density in the vermic cambic rendzina from Padis (the Apuseni Mountains).

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133

g*m –2 400

Total biomass

300 200 100 0 40

O. bihariensis

30 20 10 0 300 250 200 150 100 50 0

O. frivaldszkyi

40

Octodrilus unident. juv.

30 20 10 0 0.4

D. alpina

0.3 0.2 0.1 0 4

D. byblica

3 2 1 0 25 20 15 10 5 0

D. clujensis

May

June

July

Aug.

Sept.

Oct.

1979

FIGURE 7.13 Seasonal dynamics of lumbricid biomass in the vermic cambic rendzina from Padis (the Apuseni Mountains).

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TABLE 7.4 Yearly Mean Biomass and Monthly Mean Range of Earthworm Communities in Three Forest Ecosystems from the Apuseni Mountains (May–October 1979) (g m–2 earthworms preserved in formalin) Ecosystem Fir tree, acid brown soil (Calineasa)

Species

0.29 0.12 0.06 0.34

A. dacica

0.54

0.18 (Oct)

1.11 (Jun)

A. rosea D. alpina D. byblica D. clujensis O. compromissus

0.12 0.70 1.93 2.39 0.39 6.07

0.10 0.50 1.10 0.15 0.08 2.02

0.48 (Sep) 1.21 (Aug) 2.91 (Jul) 7.12 (Aug) 1.54 (May) 11.14 (Aug)

D. alpina D. byblica D. clujensis O. bihariensis O. frivaldszkyi Octodrilus (juv) Total

0.10 1.50 9.29 14.17 93.65 8.34 127.0

(Oct) (Oct) (Jun) (Oct)

Maximum

2.38 1.25 0.27 3.90

Total Beech forest, cambic rendzina (Padis)

Minimum

D. alpina D. clujensis D. byblica

Total Mixed beech-fir, brown forest soil (Pârâul Ponor)

Mean

(May) (May) (Oct) (Oct) (Oct) (Oct)

0.01 (Sep) 0.67 (Oct) 3.03 (Jul) 3.91 (Oct) 25.02 (May) 5.29 (Jul) 53.60 (May)

4.96 2.77 0.73 7.35

(Jun) (Aug) (Aug) (Jun)

0.23 (Aug) 3.00 (Aug) 15.79 (Aug) 23.38 (Jul) 204.09 (Aug) 19.92 (Aug) 256.97 (Aug)

cambic rendzina from the Padis karstic plateau were described (Pop and Postulache 1987). Subsequent field research confirmed the occurrence of vermic characters in other soils in which the larger Octodrilus species are present. Thus, vermic soil subtypes could be identified and described as new to science in rendzinas (mollisols); in eubasic, mesobasic, and argillic brown earth (cambisols); and even in podzolic brown soils (spodosols) (Pop and Vasu 1995). Here, the vermic characters of three different soil types (a cambic rendzina, a eubasic brown earth, and a podzolic brown soil) are discussed. They are soils developed under beech forests, but from different parent material, and are classified into three different soil classes: mollisols, cambisols, and spodosols. Site descriptions are as follows: • • •

Padis, the Apuseni Mountains, 1300-m altitude, beech forest (As. Phylitidi Fagetum), a vermic cambic rendzina on dolomitic limestone Buces Vulcan, the Apuseni Mountains, 550-m altitude, beech forest (As. Phylitidi Fagetum), a cambic brown earth on volcanic breccia with quartz, calcite, and andesite Dealul Mare, Abrud, the Apuseni Mountains, 800-m altitude, beech forest (As. Symphyto cordati Fagetum), a vermic podzolic brown soil on orthogneiss and schists

THE EARTHWORM COMMUNITIES In all three sites, similar earthworm community patterns were found but with different species combinations (Table 7.5). The three large Octodrilus species (O. aporus, O. frivaldszkyi, and O. permagnus) all occurred. All three are very large earthworms (Table 7.1). The highest biomass

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TABLE 7.5 Proportion of Total Density (D%) and Biomass (B%) of Lumbricid Species in Three Soil Types with Vermic Characters from the Carpathians Cambic Rendzina, Padis, August 16, 1979 Species Allolobophora dacica Dendrobaena alpina Dendrobaena byblica Dendrobaena clujensis Lumbricus polyphemus Octolasium lacteum Octodrilus aporus Octodrilus bihariensis Octodrilus compromissus Octodrilus exacystis Octodrilus frivaldszkyi Octodrilus permagnus Total community (g/m2)

D%

B%

12.l 22.7 2.2

0.8 1.2 6.0

15.2

28.8

66.0

Brown Earth, Buces Vulcan, June 6, 1972

Podzolic Brown, Abrud, June 12, 1992

D%

B%

D%

B%

26.5

2.2

13.9

1.7

4.4

0.4

25.0 22.2

1.5 6.6

1.5 2.9 48.5

0.3 0.6 92.5

5.6

0.4

11.8 4.4

1.7 2.3 33.3

89.8

18.0

98.0

7.0

85.0

257

34.0

200

of 256 g/m2 was recorded in the cambic rendzina, followed by that in the eubasic brown earth (200 g) and a much lower biomass of 98 g/m2 in the podzolic brown soil.

MORPHOLOGY

AND

MICROMORPHOLOGY

OF

SOIL PROFILES

All three soil types are rather shallow, with a depth of 50 to 60 cm in rendzina and podzolic brown soils and 70 to 80 cm in eubasic brown earth. Coarse rocky fragments occur sparsely in the A horizon, up to 10 to 20% in the AB, and more than 40 to 50% in the B horizons (Figure 7.14) (Pop and Vasu 1995). The most distinctive feature of these soils is the presence of a 4- to 6-cm thick surface layer of discrete, stable, and large (2 to 4 cm in diameter) earthworm casts. The earthworm cast layers, described as Am′ (mollic) or Aou (ochric-umbric) horizons, consist of old, hardened, and rather rounded earthworm casts, not linked or united together, as well as fresh casts, mostly in conical heaps. Such a cast heap can weigh as much as 200 to 300 g (oven dried). All the species of cohabiting earthworms must interact to influence soil formation, but the more evident and quite unusual traces or characters are those produced by the giant species. The size of the structural elements in the soil (burrows and casts) depends on the size of the earthworms and is species specific. Experimental studies suggested that the stability or degradation of the surface earthworm casts is conditioned by the physicochemical properties of the plasma and the texture of the soil, in which calcium plays an important role; in rendzina, the vermic structure, produced by the same species, is more stable than in acid soils (Pop et al. 1992). In the deeper horizons, the diameter of the earthworm casts diminishes, and the horizons appear more compact. The A horizon has a crumb structure; the AB and B horizons have an angular blocky structure. Earthworm burrows with a diameter up to 1.5 cm occur through the entire soil profile. Micromorphological studies showed that the earthworm cast structure does not totally disappear even in the deeper horizons. Reliable evidence of the strongly vermic character of these soils is shown in a series of structure photograms of the rendzina profile to the depth of 40 cm; the presence of hard rocks hindered us from sampling deeper (Figure 7.15).

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FIGURE 7.14 Profiles of three vermic soil subtypes from the Carpathians.

In all structural photograms, three types of earthworm casts can be distinguished; consisting of (1) mineral material brought up by earthworms from the deeper horizons; (2) predominantly organic material, mainly plant remnants in various breakdown stages, with casts or microaggregates of mesoinvertebrates; and (3) organo-mineral earthworm casts from deeper soil material mixed with organic remnants of ingested food. The series of structure photograms shows the well-aerated crumb, porous structure of the surface soil horizons, made up almost entirely of earthworm casts, with large spaces among them. In the deeper horizons, the casts are increasingly pressed together, and the pore spaces are smaller. In the upper soil horizons, a spongelike microstructure dominates because of the activity of mesoinvertebrates; the earthworm castings, which are rich in organic matter, provide a suitable medium for the development of smaller invertebrates. Large amounts of soil material from the deeper soil horizons are brought to the surface and deposited as earthworm casts, thus burying former surface cast horizons. Their secondary and tertiary comminution is achieved by a series of smaller and smaller soil organisms (animals, fungi, and bacteria) as well as by physicochemical weathering. The organic remnants brought into soil by earthworms, with an initial recognizable tissue structure, gradually turn into “fine humus.” In the deeper horizons, the diameter of earthworm casts is smaller than those at the soil surface. The spongelike microstructure caused by the mesoinvertebrates diminishes, and the horizons appear more compact. However, the earthworm cast structure does not totally disappear even in the deeper soil horizons.

PHYSICAL

AND

CHEMICAL PROPERTIES

OF

VERMIC SOILS

The physical and chemical properties of the three soils investigated are normal for their evolutionary stages and the pedoclimatic conditions. Thus, the particle size distribution, the humus content, pH, exchangeable bases, and total and mobile nutrient content are all within the usual limits of “typical” nonvermic soils (Pop and Vasu 1995).

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FIGURE 7.15 Structure photogram of the vermic cambic rendzina from Padis (the Apuseni Mountains).

It seems that the giant earthworms have affected most the morphology of the soil profile, but their biochemical influences should not be neglected, even if it is not demonstrated clearly by chemical analyses and is not entirely understood. There are some peculiarities that are common to these very different soils, suggesting a somewhat convergent evolution resulting from the unusually intense earthworm activity. The very low bulk densities, especially in the surface horizons (0.4 to 0.7 g/cm–3), low resistance to penetration (1 to 6 kgf/cm–3) associated with a high permeability (290 to 1280 mm.hr–1), and very high porosity (over 50 to 60%) provide a background to peculiar biochemical processes in these vermic soils. The thermodynamic stability conditions evidenced by redox potential (Eh) values of 400 to 600 mV (which is 100 to 200 mV higher than in typical soils) is associated with a higher humidity than usual (Wi = 53 to 65%), which indicate highly oxidative conditions because of the highly aerated soil structure.

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The humus content is very high, ranging between 4 and 11%. In all cases, the humification processes favors the formation of fulvic acids, a most unusual process in soils with high base saturation. This process is illustrated by the relatively high values of fulvic acid carbon (CF = 300 to 600 mg/100 g soil) and especially by the ratio between humic and fulvic acid carbon, which is mainly below (0.5 to 0.9) or slightly above (1.1 to 1.4) unity. The presence of ionic aluminum (2 to 80 ppm) and of ionizable SiO2 (50 to 300 ppm) in the soil solution of the saturated soil ensures suitable conditions for the formation of aluminum-fulvate and aluminum-silicate gels. As our experiments indicate, these are very efficient binding materials for soil aggregates. Differences in the base content and in aluminum and iron ion content, as well as in texture, account for the differences in stability of soil aggregates in the three investigated soils. In summary, the morphology, micromorphology, and physical and chemical properties of these soils allow the identification of new vermic subtypes in mountain rendzinas, brown eubasic earths, and podzolic brown soils that result from the high pedogenetic activity of the soil fauna, dominated by giant earthworms of the genus Octodrilus.

EXPERIMENTAL STUDY ON SOIL

OF THE INFLUENCE OF

CALCOPHILOUS OCTODRILUS SPECIES

As discussed in this chapter, several of the endemic Octodrilus species may be regarded as calcophilous, their occurrence confined to rendzinas and eubasic brown earths distributed in patches underlying beech or beech-hornbeam forests. These relatively small areas of soils developed on limestone are isolated from one another by soils developed on acid rocks. It was suggested (V.V. Pop 1994) that these acid soils could be regarded as barriers to the spreading of calcophilous earthworms, thereby promoting an accelerated insularlike speciation of the genus Octodrilus. Some of these Octodrilus species, such as O. frivaldszkyi, O. aporus, O. permagnus, or O. ophiomorphus, are very large earthworms and are believed to play an important role in developing the most evident vermic features recorded so far in the soils of the northern temperate zone. In a laboratory experiment, the relationships between Octodrilus species and calcium in the development of vermic structures in soil was investigated. Calcium is an important factor in stabilizing soil structure, and its concentration is also considered a key factor in determining the distribution of certain earthworms. The degree of stenobiotism toward calcium was also investigated by testing the tolerance of calcophilous earthworms to acid soils, presumed to act as barriers to their distribution (Pop et al. 1992). Material and Methods The activity of two calcophilous earthworm species, O. frivaldszkyi and O. bihariensis, in three different soils, either with or without supplementary calcium carbonate, was investigated in subterranean laboratory conditions. Octodrilus frivaldszkyi is a very large earthworm (length 400 to 500 mm, diameter 8 to 15 mm, weight 20 to 25 g), and O. bihariensis is a medium-size earthworm (length 94 to 113 mm, diameter 4 to 7 mm, weight 2 to 3 g). The soil types used in the experiment were (1) a vermic cambic rendzina from beech forest (1100 m altitude), Padis karstic area; (2) an acid brown soil from fir tree forest (1400-m altitude), Baisoara, the Apuseni Mountains; and (3) an albic luvisol from a mixed beech-oak forest (700-m altitude) near Cluj-Napoca. The rendzina represents the natural habitat of the two earthworm species studied; the acid soils, which separate the limestone areas in the Apuseni Mountains, can be considered a hostile environment. Experimental cages were filled with soil material (passed through 2-mm mesh sieve) from the A horizon of the three soils. For each soil type, cultures with and without a supplementary calcium carbonate layer were designed. The earthworms were fed with dried beech leaves. Experiments were in a subterranean laboratory in a deep cellar with air humidity close to saturation and a temperature of 8 to 12°C for 16 months (November 1987 to March 1988).

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Results The morphological traces of earthworm activity in soil can be explained tentatively by the physicochemical data, particularly by ionic equilibria in the soil solution. Earthworm activity in the soil was evidently species specific, but it depended on soil type. Calcophilous species can induce welldeveloped vermic characters in rendzina, which represents their natural habitat, but when placed in acid soils, especially in acid brown soil, they do not. The size of the structural elements in soil (casts and burrows) depends on the size of the earthworms, but their stability and degradation are conditioned by differences between soil types, especially by the chemical and physicochemical properties of plasma, and texture. Thus, in a rendzina, because of the existing chemical equilibria and physical properties, the vermic structure is more stable than in acid brown soil, in which the casts rapidly break down. The nature of the cement of soil aggregates depends on the soil chemical composition, but differences detected in the cement composition of the casts of the two Octodrilus species showed a specific or selective influence of the earthworms on soil structure. The CaCO3 treatment of soil leads to a marked improvement in soil structure (i.e., in the conservation and stabilization degree of both discrete earthworm casts and burrow layers). Special neoformations with calcium, iron, and aluminum were observed in different parts of the soil profile. These included ferric and humic earthworm burrow linings in variants with acid soils, which could be regarded as protecting the earthworms against adverse environmental conditions, and semicolloidal and semicrystalline gels, the function of which is unknown. The alterations induced in soil by earthworms are sometimes so great that they affect not only the mobile ionic structure of the soil, but also the clay minerals (i.e., its most stable constituents). If the bioaccumulation of potassium and the changes in clay mineral ratio detected in our experiments actually represent an illitization (alteration of the structure of clay minerals), then this could be evidence that earthworms can produce much more rapid and profound changes in soils than hitherto believed. It is generally accepted that profound alterations in soils are very slow, taking hundreds or thousands of years. Our experiment lasted only 16 months, so this issue needs further investigation.

CONSERVATION OF THE ENDEMIC OCTODRILUS SPECIES IN THE CARPATHIANS The endemic Octodrilus species, which have proved to be interesting for science and important to the stability of unique natural ecosystems in the Carpathians, are endangered by two common human activities: creation of air-soil pollution and sylvicultural cutting, especially clear-felling of the forests.

EFFECTS

OF

AIR POLLUTION

ON

EARTHWORMS

The largest part of the Romanian Carpathians is currently protected from or is not affected by destructive air pollution. Nevertheless, over a limited area, air pollution does affect earthworms, as well as whole ecosystems. The earthworm fauna from one of the most polluted areas in Romania (namely, the Ampoiu valley) in the Apuseni Mountains was investigated (Pop 1987). Here, sulfur dioxide and dusts containing heavy metals are discharged into the air from a metal-chemical works in Zlatna. The effects of these on the earthworm populations were directly related to the distance from Zlatna. In an old beech forest with well-developed tall trees in the immediate neighborhood of the pollution source, I found no earthworms. The vegetation and soil type, as well as a study of several unpolluted control forests, indicated that conditions would normally be suitable for an earthworm community such as O. frivaldszkyi, Allolobophora dugesi, A. rosea, D. byblica, and D. clujensis, with an approximate density of 15 to 20 earthworms per square meter and a biomass of 35 to 50 g per square meter. Earthworms in the most polluted area, which includes the beech forest near the chemical works, were presumably killed either by the direct action of pollutants or by starvation induced by © 2004 by CRC Press LLC

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unpalatable food. A microcosm experiment showed clearly that earthworms do not eat the polluted litter from Zlatna forest. However, other soil-inhabiting invertebrates, such as mites and springtails, are abundant in the litter. Therefore, the accumulation on the soil surface of a thick undecomposed leaf litter layer is considered evidence of the interruption of the organic matter turnover food chain at the level of earthworms, which were proved to be more affected by the pollutants than other soil-inhabiting invertebrates.

EFFECTS

OF

SYLVICULTURAL CUTTING

ON

ENDEMIC EARTHWORMS

Natural forests, with large areas of secondary grassland following forest cutting, cover most of the Carpathians in Romania. Agricultural Þelds are conÞned to areas near the sparse villages at lower altitudes. Nevertheless, substantial areas are affected by sylvicultural cutting, which also jeopardizes soil invertebrates, mostly on limestone, in which the shallow soils without the protection of natural vegetation are destroyed rapidly by erosion. The giant endemic Octodrilus species, living on small areas distributed in patches, are jeopardized most. It is considered that the disappearance of the giant endemic earthworms after clear-cutting endangers the restoration of the speciÞc Carpathian beech forests of limestone areas (Pop 1982). Only protection of their preferred environment (namely, the islands of beech forests on limestone) could ensure the conservation of endemic species of earthworms. In Romania, several projects aiming at the protection or conservation of ßoristic and faunistical elements of the Carpathians are in progress. The necessity to protect soil endemic organisms as a new aim of nature conservation in the Carpathians has been discussed previously (Pop 1983). It is considered that endemic earthworm species are good indicators of particular edaphons that can be also called endemic edaphons. The protection of the biotopes inhabited by the endemic Octodrilus species, besides the conservation of these earthworm species, may have theoretical value for the study of insular speciation in continental mountains.

ACKNOWLEDGMENTS It is our pleasant duty to express gratitude to Dr. Ken Lee, CSIRO, Division of Soils, Glen Osmond, Australia, for reading the manuscript, correcting the English, and providing valuable suggestions in preparing the chapter. We are also indebted to Dr. Clive Edwards, The Ohio State University, Columbus, for the valuable suggestions in drafting and revising the chapter. The second author expresses her gratitude to Prof. Michael Wink, Institute for Pharmaceutical Biology, RuprechtKarls University, Heidelberg, Germany, for guiding the molecular taxonomy research.

REFERENCES Bouché, M.B. 1972. Lombriciens de France. Écology et Systématique, Institut National de la Recherche Agronomique, Publication 72.2, Paris. Cassagnau, P. 1961. Ecologie du Sol dans les Pyrénées Centrales, Hermann, Paris. Kubiena, W.L. 1953. The Soils of Europe, Thomas Murby and Co., London Institute for Soil Science and Biochemistry. 1979. The Romanian System of Soil ClassiÞcation, Bucharest. Mayr, E. 1971. Populations, Species and Evolution. An Abridgment of Animal Species and Evolution, Belkamp Press, Cambridge, MA. Mayr, E., E.G. Linsley, and R.L. Usinger. 1953. Methods and Principles of Systematic Zoology, McGrawHill, New York. Mihailescu, V. 1963. The Southeastern Carpathians on the Romanian Territory, Editura StiintiÞca, Bucuresti. Mountford, M.D. 1962. An index of similarity and its application to classiÞcation problems, in Progress of Soil Zoology, Murphy, P.W., Ed., Butterworths, London, pp. 43–50. ÿ Mrsic, N. 1991. Monograph on earthworms (Lumbricidae) of the Balkans, Academia Scientiarum et Artium Slovenica, Historia Naturalis, Ljubliana, 31, I,II.

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Omodeo, P. 1956. Contributo alla revisione die Lumbricidae, Arch. Zool. Ital., 41, 129–212. Pop, A.A., M. Wink, and V.V. Pop. 2003. Use of 18S, 16S rDNA and cytochrome c oxidase sequences in earthworm taxonomy (Oligochaeta, Lumbricidae), Pedobiologia. Pop, V. 1941. Zur Phylogenie und Systematic der Lumbriciden, Zool. Jahrbücher (Syst.), 74, 487–522. Pop, V. 1948. Lumbricidele din România, Anal. Acad. Republicii Populare Romane, Sect. Stiinte Geologice, Geografice, Biol. Ser. A, Mem. 9, (Romanian, Russian, and French summaries), 1, 383–506. Pop, V. 1973. Octolasium (Octodrilus) robustum nouvelle espèce de Lumbricidae et ses affinités, Rev. Roumaine Biol., Ser. Zool., T.18, 4, 265–268. Pop, V.V. 1980. The biogeographic significance of the earthworm fauna of the future National Park area in the Cerna Valley [English summary], Ocrotirea naturii si a mediului înconjurãtor, Bucuresti, Romania, (Romanian and English summary), 24, 2, 157–164. Pop, V.V. 1982. On the earthworm fauna of the Carpathian beech forests [English summary], in Preda, V. and N. Boscaiu, Eds., Les Hétraies Carpatiques. Leur Signification Biohistorique et Ecoprotective, (Romanian and English summary), Publ. Academiei R.S. România, Filiala Cluj-Napoca, Subcomsia. Protectia Naturii, pp. 327–341. Pop, V.V. 1983. The importance of the conservation of the endemic edaphobionts from the Romanian Carpathians [English summary], Ocrotirea naturii si a mediului înconjurãtor, Bucuresti, Romania, t.21, 1, 37–39. Pop, V.V. 1985. Relationships between Lumbricide synusia and ecosystem types in the Apuseni Mountains (the Romanian Carpathians) [Romanian and English summary], in Actualitate si Perspectiva in Biologie. Structuri si Functii in Ecosisteme Terestre si Acvatice, Tipo Agronomia, Cluj-Napoca, Romania, pp. 79–86. Pop, V.V. 1987. Density and biomass of earthworm synusia in forest ecosystems of the Romanian Carpathians, in Bonvicini Pagliai, A.N. and P. Omodeo, Eds., On Earthworms. Selected Symposia and Monographs, U.Z.I., 2, Muchi, Modena, Italy, pp. 183–190. Pop, V.V. 1989. Studies on the genus Octodrilus Omodeo, 1956 (Oligochaeta, Lumbricidae) from the Apuseni Mountains (the Carpathians, Romania). I. Description of new taxa, Travaux du Museum d’Histoire naturelle “Grigore Antipa,” Bucharest, Romania, 30, 193–221. Pop, V.V. 199l. Studies on the genus Octodrilus Omodeo, 1956 (Oligochaeata, Lumbricidae) from the Apuseni Mountains (the Carpathians, Romania). II. Variability of characters, Travaux du Museum d’Histoire naturelle “Grigore Antipa,” Bucharest, Romania, 31, 397–414. Pop, V.V. 1994. On speciation in the genus Octodrilus Omodeo, 1956 (Oligochaeta, Lumbricidae), Mitteilungen aus dem Hamburgischen Zoologischen Museum und Institut, Band 89, Ergbd. 2, 37–46. Pop, V.V. 1997. Earthworm-Vegetation-Soil Relationships in the Romanian Carpathians, Soil Biol. Biochem., 29(3/4), 223–229. Pop, V.V. and T. Postolache. 1987. Giant earthworms build up vermic mountain rendzinas, in Bonvicini Pagliai, A.N. and P. Omodeo, Eds., On Earthworms. Selected Symposia and Monographs, U.Z.I., 2, Muchi, Modena, Italy, pp. 141–150. Pop, V.V., T. Postolache, A.Vasu, and C. Craciun. 1992. Calcophilous earthworm activity in soil: an experimental approach, Soil Biol. Biochem., 24(12), 483–1490. Pop, V.V. and A. Vasu. 1995. Conspicuous vermic characters in mountain soils developed by large lumbricids (Oligochaeta), Acta Zool. Fenn., 196, 83–86. Sheals, R.W. 1969. Outline of an application of computer technology to Acarine taxonomy: a preliminary examination with species of Hypoaspis-Androlaelaps complex (Acarina), Proc. Linnean Soc., London, 176, 11–121. Sims, R.W. 1969. Outline of an application of computer technique to the problem of the classification of the Megascolecoid earthworms, Pedobiologia, 9, 35–41. U.S. Department of Agriculture, Soil Classification System. A Comprehensive System, Seventh Approximation Soil Survey Staff, Soil Conservation Service, U.S. Government Printing Office, Washington, D.C., U.S.A. Zicsi, A. 1986. Über die taxonomischen Probleme der Gattung Octodrilus Omodeo, 1956 und Octodriloides gen.In. (Oligochaeta, Lumbricidae), Opuscula Zool., Budapest, Hungary, 22, 103–112. Zicsi, A. and V.V. Pop. 1984. Neue Regenwürmer aus Rumänien (Oligochaeta, Lumbricidae), Acta Zool. Hung., 30(1–2), 241–248, 1960.

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Part IV Influence of Earthworms on Soil Organic Matter Dynamics, Nutrient Dynamics, and Microbial Ecology

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of Earthworms on Soil 8 Effects Organic Matter and Nutrient Dynamics at a Landscape Scale over Decades Patrick Lavelle,1 Fabienne Charpentier,1 Cécile Villenave,1 Jean-Pierre Rossi,1 Laurent Derouard,1 Beto Pashanasi,2 Jean Andre,3 Jean-François Ponge,4 and Nicolas Bernier4 1

Laboratoire d’Ecologie des Sols Tropicaux, Bondy Cedex, France; Estacion Experimental San Ramon, INIAA, Yurimaguas, Loreto, Peru; 3Université de Savoie, France; 4MNHN, Brunoy, France 2

CONTENTS Earthworms and Soil Function: The Drilosphere Concept ...........................................................146 Earthworm Behavior ......................................................................................................................148 Selection of Soil Particles.......................................................................................................148 Spatial Patterns of Horizontal Distribution of Earthworms...................................................149 Compacting vs. Decompacting Species .................................................................................149 Medium-Term Effects: Experiments Inoculating Earthworms into Cropping Systems of the Humid Tropics ..................................................................................................151 Long-Term Effects of Earthworms: Modeling and Observation of Successional Processes ..............................................................................................................151 Modeling .................................................................................................................................152 Earthworm Activities and Successional Processes.................................................................153 Discussion ......................................................................................................................................155 References ......................................................................................................................................157

After several decades of unquestioned success, agriculture is now facing important global problems. Huge increases in productivity in developed countries have been accompanied by a severe depletion of “soil quality” in terms of resistance to erosion, organic contents, concentrations of heavy metals, and pesticide residues. Agricultural intensification in developing countries has been less successful because of various socioeconomic limitations. Nevertheless, traditional agricultural practices do not conserve the quality of soils; stocks of organic matter are rapidly becoming depleted, and erosion removes fine particles from the soil surface horizons. In a context of increasing human population pressures, particularly in developing countries, this degradation of soils results in many social and environmental problems (Eswaran 1994; FAO 2000). Features common to all kinds of soil degradation are a significant decrease in organic reserves, degradation of the soil structure, and severe depletion of soil invertebrate communities, especially earthworms (Decaëns et al. 1994; Lavelle et al. 1994).

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The contributions of earthworms to soil fertility have been described in several hundreds of articles and books (Lee 1985; Edwards and Bohlen 1996; Lavelle and Spain 2001). This has led to a growing expectation from soil users for provision of methods that protect soil fertility through the enhancement of biological processes. Earthworms may be considered a biological resource for farming systems, and the management of earthworm communities provides a promising field for innovation in agricultural practices (Lavelle et al. 1999). Demand for techniques, making optimal use of earthworms as a resource, is likely to increase, although basic research is still needed to support such developments. The relationships between earthworm activities and changes in soil properties are not thoroughly understood, especially at large timescales of years to decades. Most research results have been obtained in small-scale laboratory or field experiments that exaggerate the process under study and can by no means be extrapolated readily to larger scales of time and space. However, recent research on earthworm casts and other related biogenic structures have shown that these structures may persist for rather long periods and provide the soil with specific properties that may survive the death and elimination of earthworm populations that produced them. This property contributes significantly to the resistance and resilience of soils to disturbances (Lavelle et al. 2004). This chapter synthesizes information on the effects of earthworms on soil systems at scales longer than 1 year, and earthworm behavior that may affect these processes is detailed.

EARTHWORMS AND SOIL FUNCTION: THE DRILOSPHERE CONCEPT For example, at a real scale of a small tropical farmer’s plot, earthworm activities are only one determinant of soil fertility, and their effects are likely determined by factors operating at larger scales of time and space, such as climate, edaphic characteristics, and the quality and amount of organic inputs (Lavelle et al. 1993). Earthworms participate in soil functions through the drilosphere system, which involves earthworms, casts, and burrows and the whole microbial and invertebrate community that inhabits these structures. As a result of earthworm digestion processes and creation of soil structures, the composition, structure, and relative importance of the drilosphere system to soils is clearly determined by climate, soil parameters, and the quality of organic inputs. Earthworms in turn influence soil microbial communities and hence have effects on microbial processes related to soil organic matter (SOM) status and nutrient dynamics. They also affect the activities of other soil-inhabiting invertebrates, either by modifying their environment or through competition for feeding resources (Figure 8.1). Earthworms are not a homogeneous entity. They comprise several functional groups, each with clearly distinct ecology and impacts on the environment (Bouché 1977). Current classifications based on earthworm location in the soil profile and their feeding resources are still too general to describe the large diversity in functions. Moreover, earthworms classified into a general category may not always exhibit the behavioral traits expected to be associated to this category (Neilson et al. 2000; Mariani et al. 2001). Classifications based more on impacts on soil parameters might be more useful. The effects of earthworms on soil function thus depend on their interactions with a wide range of identified abiotic and biotic factors that operate at rather different scales of time and space. Furthermore, the effects produced will affect soil structures of different sizes and persist for highly variable periods of time, depending on the factors with which they interact. For example, it is expected that physical structures created by earthworms as a result of their interactions with other soil components will last much longer than the flush of activity of dormant microorganisms that they have activated in their guts (Figure 8.2). Most studies have described processes at smaller scales of earthworm activities, typically in microcosms, small plots or small field enclosures. Results obtained under such conditions describe existing processes but cannot be extrapolated directly to quantify and predict effects produced at the

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Interactions with microflora for digestion of complex organic molecules

ECOSYSTEM ENGINEERS

LITTER TRANSFORMERS MICROPREDATORS MICROFLORA Formation of biogenic structures FIGURE 8.1 Relationships of earthworms and other soil “ecosystem engineers” with other soil biological components. Earthworms partly influence the occurrence and activity of other soil organisms because they create large structures that may persist over significant periods and affect the environment. (From Lavelle 1997. With permission.)

PEDOGENESIS SOIL STRUCTURE

SOM DYNAMICS

AGGREGATION HUMIFICATION

CASTING BURROWING

MINERALIZATION EARTHWORM MICROBIAL ACTIVITY

DIGESTION

MYCORRHIZAL INFECTION

FIGURE 8.2 Effects of earthworm activities on soil processes operating at different scales.

scale at which SOM dynamics and nutrient cycling are generally studied. One spectacular result of this smaller-scale approach is a huge discrepancy between the great importance that pedobiologists attribute to earthworms as regulators of soil physical structure and SOM dynamics and the absence of any representation of earthworm activities in simulation models that describe SOM dynamics at

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Functional Groups

Spatial Distribution

Time Structure

HOURS GUT CONTENT

DAYS/WEEKS FRESH CASTS

MONTHS AGING CASTS

YEARS: DECADES SOIL PROFILE

Effect on SOM

ASSIMILATION COMMINUTION

NUTRIENT RELEASE

PHYSICAL PROTECTION

ACCELERATION OF TURNOVER

Activation of mineralization

Transfers

Conservation

FIGURE 8.3 Effects of earthworm activities on SOM dynamics at different scales of time and space.

scales of decades and hectares (Jenkinson and Rayner 1977; Molina et al. 1983; Parton et al. 1988; Agren et al. 1991). A few papers have already described the effects of earthworms at the scales of the different biotic and abiotic parameters with which they interact, such as (1) selection of ingested particles and digestion processes at the scale of a gut transit (0.5 to 20 hours) (Lee 1985; Barois and Lavelle 1986); (2) immobilization-reorganization of nutrients in fresh casts (1 to 20 days) (Syers et al. 1979; Lavelle et al. 1992; Lopez-Hernandez et al. 1993; Tiunov et al. 2001); and (3) evolution of SOM in aging casts (3 to 30 months) (Martin 1991; Lavelle and Martin 1992; Blair et al. 1994; McInerney et al. 2000). Long-term evolution of SOM at the scale of the whole soil profile and pedogenesis during periods of years to centuries has been identified, although little information is currently available (Figure 8.3). This chapter describes the effects of earthworms on larger-scale SOM and nutrient dynamics observed in 3-year field experiments and details three subprocesses that may determine the longterm effects of earthworms on soils: feeding behaviors, patterns of horizontal distribution, and participation of earthworm activities in successional processes. Simulations of SOM dynamics, based on the CENTURY model (Parton et al. 1988), give some insight on effects of earthworms on soils observed at a timescale of 10 to 50 years.

EARTHWORM BEHAVIOR Earthworm behavior may affect soil functions significantly. A major difference between short-scale experiments and the real world is that, in confined small experiments, earthworms have limited opportunities to choose their food and move away. This probably explains why they often lose weight or die in laboratory experiments. On the other hand, the introduction of unrealistically high earthworm populations to small enclosures in the field often creates concentrations of intense earthworm activity that would not normally have occurred in the field or that concern only microsites that are either highly dispersed in nature or infrequently visited.

SELECTION

OF

SOIL PARTICLES

Earthworms are known to select the organic and mineral soil components that they ingest. As a result, their casts often have much higher contents of SOM and nutrients than the surrounding soil (Lee 1985). This is probably because of preferential ingestion of plant residues (leaf and root litter debris and occasionally fungi) (Piearce 1978; Ferrière 1980; Kanyonyo ka Kajondo 1984; Bonkowski et al. 2000; Neilson et al. 2000), fecal pellets of other invertebrates (Mariani et al. 2001), and clay minerals. Barois and Lavelle (1986) demonstrated that the tropical peregrine species Pontoscolex corethrurus was able to select either large organic debris or small mineral particles, depending on soil type. Selection was made on aggregates rather than primary particles. However, some earthworms may selectively ingest coarser particles than the average in soils with very high clay contents. There is evidence that some endogeic species of earthworms ingest only aggregates that do not exceed the diameter of their mouths, whereas other species may feed on larger aggregates © 2004 by CRC Press LLC

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and split them into smaller aggregates (Derouard et al. 1997; Blanchart et al. 1999). The feeding behavior that allows such a selection has never been described in detail. The long-term consequences of that behavior on the dynamics of soil processes and SOM dynamics have not been addressed directly. Endogeic species of earthworms may deposit 20 to 200 t dry soil ha−1 surface casts that contain a significant proportion of SOM yearly. A much larger quantity of casts is deposited inside the soil, and a volume of soil equivalent to the whole soil of the upper horizons may be passed through earthworm guts in a few years (Darwin 1881; Lavelle 1978). Nevertheless, the higher concentration of fine particles in casts than in the surrounding bulk soil suggests that earthworms may possibly reingest the same soil several times, whereas microsites with a relatively coarser texture may be avoided by earthworms.

SPATIAL PATTERNS

OF

HORIZONTAL DISTRIBUTION

OF

EARTHWORMS

Several authors have pointed out the aggregated distribution of earthworm populations that occur in such diverse ecosystems as an arable soil from Germany (Poier and Richter 1992), a deciduous forest of England (Phillipson et al. 1976), humid African (Lavelle 1978) and Colombian savannahs (Decaëns and Rossi 2001; Jimenez et al. 2001), and an artificial pasture in southern Martinique (Rossi et al. 1997). In some places, the earthworm distribution was independent of basic soil parameters such as depth or clay or carbon contents. Furthermore, different earthworm species tended to have different horizontal distributions (Poier and Richter 1992); in an almost monospecific earthworm community of Polypheretima elongata in Martinique, Rossi (1997) observed different distribution patterns for adults and juvenile earthworms. These observations suggest that some earthworm species have patchy population distributions with an average patch diameter of 20 to 40 m. Such patches seem to be independent of soil parameters, at least to some extent. The dynamics of earthworm populations in such a patch are not synchronized with the populations of other patches. The occurrence of such distribution patterns suggests that earthworm activities concentrate on patches probably only for limited periods of time before becoming locally extinct. In the moist savanna at Lamto (Côte d’Ivoire), complementary patterns have been observed between large earthworm species that produce large casts and tend to compact the soil and smaller species that produce fine granular casts after ingesting the larger casts of the first type of species with root litter (Blanchart et al. 1999). In that case, the observed patterns suggest a succession of patches made of “compacting” and “decompacting” species with complementary effects on soils.

COMPACTING

VS.

DECOMPACTING SPECIES

The physical structure of earthworm casts is very relevant to the dynamics of SOM at intermediate scales of months to years. Two different types of casts may be distinguished in this respect: the globular casts of compacting species and the granular casts of decompacting species. Casts of the first category may be surrounded by a thin 10- to 20-m µ cortex made of clay minerals and organic particles, which seem to reduce aeration and hence inhibit microbial activity at the scale of weeks to months (Martin 1991; Blanchart et al. 1993). Soils colonized by monospecific communities of such earthworm species are prone to compaction. In an experiment in which the earthworm Millsonia anomala had been introduced into soils under a yam and a maize culture, the proportion of large aggregates bigger than 2 mm increased significantly in soils that had been sieved previously, but no significant effect was observed in a soil that had kept its original structure (Table 8.1). Soil bulk density increased significantly in both situations. Similar effects have been observed after inoculation of the peregrine, pantropical, endogeic species P. corethrurus into soils under a traditional cropping system of Peruvian Amazonia. After six successive crops, earthworms had increased the proportion of macroaggregates (>2 mm) significantly from 25.4 to 31.2% at the expense of smaller (<0.5 mm) aggregates with proportions that had decreased

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TABLE 8.1 Effects of Inoculation of the Earthworm Millsonia anomala into Cropping Systems on Bulk Density and Aggregation of Soil Bulk Density

% Aggregates Larger than 2 mm

Maize 0–10 cm

Maize 0–10 cm

Yam Mounds, Sieved Soil

Maize 0–10 cm, Undisturbed Soil

Maize 0–10 cm, Sieved Soil

3 1.31 1.24 0.06

36 1.48 1.37 <0.001

34 53.5 29.8 <0.05

36 42.5 38.8 n.s.

36 42.2 24.6 <0.01

Time (months) Inoculated Noninoculated p n.s., not significant

from 35.4 to 27.4%. Such changes in soil aggregation resulted in a slight increase of bulk density (significant during the first three cropping cycles) and a significant decrease in infiltration rates and sorptivity, the latter decreasing from 0.34 cm sec−1 in non-earthworm–inoculated soils to 0.15 cm sec−1/2 in treatments inoculated with 36 g m−2 fresh biomass of earthworms (Alegre et al. 1996). This transformation in soil physical properties resulted in eventual changes in the soil water regime because soil tended to become drier during dry periods and wetter in periods of heavy rainfall than in the non-earthworm–inoculated treatment. Other endogeic earthworm species have been reported to have opposite effects because they tend to break down large (>0.5 mm) aggregates and split them into smaller ones (Derouard et al. 1997; Blanchart et al. 1999). For example, in western African savannahs, small species of the Eudrilidae family possess such abilities, and it has been hypothesized that soil aggregation is regulated by the opposite effects of large compacting species such as Millsonia anomala and decompacting species like the common eudrilid Hyperiodrilus africanus. In the absence of such decompacting earthworms or other earthworm species, the activity of compacting earthworms may create severe soil problems. This was the case in a pasture derived from primary forest near Manaus (Brazil). At this site, the peregrine earthworm P. corethrurus had increased to rather large and active populations; at the same time, deforestation had eliminated 75% of other macroinvertebrate species, especially from the decompacting group. The accumulation of casts of P. corethrurus at the soil surface in very moist soil conditions resulted in the formation of a continuous muddy layer of earthworm casts. When droughts occurred, this layer turned into a compact 3-cm thick crust that prevented plants from growing, leaving large patches of bare soil. These results contrast with a rather broad set of other results suggesting that earthworm activities improve the aeration of soil and infiltration of water (Lee 1985; Edwards and Bohlen 1996), and P. corethrurus has been reported to repair a compacted oxisol, showing that interactions with other invertebrates and soil characteristics may be the ultimate determinants of the effects of earthworms on soils (Zund et al. 1997). Three hypotheses may explain such discrepancies: 1. Most studies on the relationships between earthworm activities and soil physical parameters have been based on Lumbricidae. This family, unlike most tropical families, comprises a large proportion of species that create semipermanent burrows that influence water infiltration significantly. 2. In natural ecosystems, the association of compacting and decompacting earthworm species may regulate soil physical properties and, in the end, favor infiltration and aeration. It is important to consider that decompacting species may belong to other taxa such as Enchytraeidae (Didden 1990; Van Vliet et al. 1993), ants (Decaëns et al. 2001), termites (Isoptera), or millipedes (Diplopoda) (Tajovsky et al. 1991). © 2004 by CRC Press LLC

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3. The effects of Lumbricidae could be a consequence of the burial of large organic particles mixed with ingested soil, because it is commonly recognized that the incorporation of litter into soil has significant effects on soil physical parameters (Springett 1983; Aina 1984; Oades 1984; Kladivko et al. 1986; Shaw and Pawluk 1986; Joschko et al. 1989; Kooistra 1991; Wolters 1991).

MEDIUM-TERM EFFECTS: EXPERIMENTS INOCULATING EARTHWORMS INTO CROPPING SYSTEMS OF THE HUMID TROPICS Three-year experiments have been conducted at two tropical sites, Lamto (Côte d’Ivoire) and Yurimaguas (Peru). Annual cropping systems were inoculated with populations of endogeic species of earthworms, and the dynamics of the system were compared with those of noninoculated systems for six successive crops over 3 years (Pashanasi et al. 1996; Gilot 1997). At Yurimaguas, the C and N contents of soil decreased significantly with time. After six cropping cycles, the C contents had decreased from 16.8 to 1.36 and 1.51 mg g−1 in systems inoculated with earthworms and in the control, respectively (Figure 8.4). Although systems inoculated with earthworms tended to have less C after the fourth cropping cycle, the observed differences were not significant at the end of the experiment. Changes in N content during the experiment showed similar trends: N concentrations increased initially in both systems as a result of N outputs after burning and incorporation of ashes in the soil. During the first three cropping cycles, N contents were higher in the earthworm-inoculated systems. From the fourth harvest, the N contents were lower in the earthworm-inoculated treatments, but differences were not significant. Earthworms did not affect soil nutrient contents for the first five cropping cycles: Ca, Mg, K, and P contents increased first after burning and incorporation of ashes and thereafter decreased steadily. By the last harvest, cation contents were slightly higher in the earthworminoculated systems, but the differences were significant only for K contents. Similar trends were observed for pH and Al saturation. At Lamto, similar results were obtained. After four cropping cycles of maize, the C contents in the upper 10 cm of soil decreased from 13.37 mg g−1 at time 0 to 9.75 and 9.64 mg/g in the control and earthworm-inoculated systems, respectively, with the differences observed between the last two treatments insignificant. In spite of these results, some differences in the quality of organic matter seemed to exist. Physical fractionation of SOM using the Feller (1979) method showed some evidence of protection of the coarse organic particles in the inoculated systems (Gilot 1997). Furthermore, laboratory respirometric tests showed a significant increase in soil respiration rates where earthworms had been active (Tsakala 1994). Experiments by Gilot (1997) showed that the effects of earthworms in protecting coarse organic fractions were significant only in soils that had been sieved previously. In this case, reaggregation of soil by earthworms had positive effects on SOM protection. In soils that had not been sieved, large aggregates, resulting from earthworm activities in the natural soil, were conserved in the no earthworm treatment conditions during the 3 years that the experiment lasted. Thus, the effects of protection linked to aggregation were retained, although no earthworms were present.

LONG-TERM EFFECTS OF EARTHWORMS: MODELING AND OBSERVATION OF SUCCESSIONAL PROCESSES In the absence of long-term experiments, evidence for effects of earthworms at scales from 10 to 100 years or more has been sought from simulation modeling and the observation of time sequences in successional processes.

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C content in the 0- to 10-cm layer

C% 1.9 1.8 1.7 1.6

a

a

1.5 a

1.4

a

a

a

1.3

a b

1.2 0

1

2

3

4

5

6

Number of Crops Earthworm + Earthworm – N content in the 0- to 10-cm layer

N% 0.14 0.13

a

0.12

a

a

a

0.11 a

0.10

a

0.09 0

1

2

3

4

5

6

Number of Crops

FIGURE 8.4 Changes in carbon and nitrogen contents in an ultisol of Peruvian Amazonia submitted to traditional slash-and-burn agriculture in the presence and absence of earthworms (Pontoscolex corethrurus). Pairs of data with different letters are significantly different. (From Pashanasi et al. 1996. With permission.)

MODELING Currently available models that describe SOM dynamics do not take into account explicitly the effects of soil invertebrates. Part of the effects of soil-inhabiting invertebrates may be implicit, either when describing initial soil conditions (e.g., through how the C:N ratio is influenced by their activities) or when assessing the decomposition rates of C pools that will actually include the overall effects of earthworms. An attempt was made to simulate the effects of earthworm activities on the three kinetically defined organic pools of the CENTURY model (Parton et al. 1988). This model, which simulates plant production and SOM dynamics in various agricultural and natural systems, considers three different SOM functional pools. A labile fraction (active SOM) has a rapid turnover and exists in the form of live microbes and microbial products. The remaining fractions comprise SOM that is stabilized, either because it is physically protected (slow pool), because it is chemically resistant (passive pool) to decomposition, or both. As a first step, the CENTURY model was used to simulate SOM dynamics and plant production in the savannah of Lamto (Martin and Parton unpublished data) and validated against observed data values. Then, the model was calibrated for this site and simulated C dynamics in earthworm casts and a control soil of the same savannah sieved at 2-mm aperture. Observations by Martin (1991) during a 450-day incubation of earthworm casts and a control surrounding soil sieved at 2mm aperture were used as a reference. For the sieved soil, the model outputs were close to the experimental results, provided slow soil C decomposition rates increased. Conversely, it was necessary to decrease the rates for both slow and active decomposition rates of soil C to simulate its dynamics in casts. Earthworm removal was simulated by replacing the active and slow soil C decomposition rates of the model with those obtained by calibration with the control soil. Under these assumptions, the CENTURY model indicated that SOM would decrease by about 10% in 30

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Earthworm removal 1500

1000

500

0 0

10

20

30

40

50

60

70

80

90

100

Time (yr)

FIGURE 8.5 Changes of carbon contents in soils of a grass savanna at Lamto (Côte d’Ivoire) in the presence and absence of endogeic earthworms; results of a simulation using the CENTURY model. (From A. Martin and W.J. Parton unpublished data, 1993.)

years, the largest proportion lost in the slow pool that includes physically protected organic matter (Figure 8.5). This suggests that slow decomposition rates of soil C may be influenced significantly by earthworm activities. This pool would comprise organic matter that binds microaggregates into macroaggregates (Elliott 1986), which is generally lost during cultivations. Earthworms may play an important role in stabilizing SOM, hence maintaining the SOM stock and soil structure in agroecosystems in the longer term.

EARTHWORM ACTIVITIES

AND

SUCCESSIONAL PROCESSES

Successional processes in vegetation dynamics such as those observed in natural forests may precede or follow significant changes in the organic status of soils. Several examples indicate that earthworm activities during these successions vary significantly (Miles 1985; Scheu 1992) and may be limited to periods when the organic matter that they can digest is available. Sampling soil invertebrate populations in a diachronic series of hevea plantations in the Côte d’Ivoire showed great changes in the soil faunal communities as plantations aged (Gilot et al. 1995). During the early stages of succession, soil invertebrate communities were dominated by termites, especially xylophagous species. After a few years, the abundance of this group of termites declined, and other groups of termites dominated the termite communities. After 20 years, earthworms became dominant; mesohumic and endogeic species categories prevailed. Finally, in a 30-year plantation, soil invertebrate populations decreased steadily, as did the productivity of the hevea. It has been suggested that these changes could reflect successions in soil invertebrate communities after changes in the quality and quantity of organic matter. When the plantations were created, woody material from the primary forest was left on the soil surface. Xylophagous termites were the first invertebrate group that used this resource. They transformed decaying wood into fecal pellets that may have been used later by other groups of termites and surface-living earthworms. Eventually, fecal pellets of humivorous termites may have been incorporated into the soil and been used as food by endogeic mesohumic and oligohumic species of earthworms. Once the organic matter from the wood had passed through this food chain and lost most of its energy, which was stored as carbon compounds, the food resource base for soil invertebrates was reduced to the plant residues currently available in the hevea plantation, and invertebrate populations decreased drastically. Interestingly, this sharp decrease in invertebrate populations coincided with a reduction in productivity of the hevea.

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Endogeic Anecic

190

160

60

50

30

Epigeic

20

120 100 80 60 40 20 0

10

n m–2

154

Years FIGURE 8.6 Changes in the abundance of earthworms of different ecological categories along a succession in an alpine spruce forest. (From Bernier et al. 1993. With permission.)

These observations suggest that soil invertebrates, especially earthworms, may at some stages use carbon sources that had been previously stored in the soil system at different stages of natural or artiÞcial successional processes. In the case of the hevea plantation, we hypothesize that the activities of soil invertebrates are sustained, at least partly, by the energy progressively released from decaying logs, resulting in positive effects on hevea productivity. Our hypothesis that earthworm activities may develop at a determined stage in plant successions is supported by data from Bernier et al. (1993) in an alpine forest of France located at a 1550-m elevation. In a succession of forest patches from 10 to 190 years old, signiÞcant changes in earthworm communities were observed (Figure 8.6). In the early stages, earthworm populations were large with a clear dominance of endogeic species populations. Population densities decreased steadily during the subsequent 20 years, and after 60 years, when the forest was mature, earthworm populations started to increase again, although their population density was low. These population changes coincided with clear changes in the amounts and quality of organic matter stored in the humus layers. The proportion of organic matter that was bound to minerals (i.e., organic matter mixed into the soil by earthworm activities) was greatest after 10 years and then decreased steadily, with bound organic matter almost absent after 60 years (Figure 8.7). In the later stages of succession, bound organic matter began to accumulate again. The pattern of changes of unbound organic matter in time was exactly opposite to that of bound organic matter. % min materials 60 50

Unbounded OM

Bounded OM

40 30 20 10 0 10

20

30

50 Years

60

160

190

FIGURE 8.7 Changes in the abundance of organic matter (OM) bounded to soil minerals and unbounded organic matter along a succession in an alpine spruce forest. (From Bernier et al. 1993. With permission.)

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The amounts of unbound organic matter decreased when earthworms were abundant and increased when earthworm populations were at low densities. This is evidence that during the cycle of growth, maturation, and senescence of the forest, the humus type changed, with maximum development of a moder after 60 years and a mull after 10 years. We hypothesize that a forest accumulates organic matter as litter and raw humus during its early phases of growth, when primary production is high. Then, earthworm populations start to develop at the expense of these organic accumulations, and they progressively incorporate the nondigested part of this raw organic matter into organomineral complexes (Figure 8.7). This process results in the release of large quantities of nutrients and the creation of physical soil structures (macroaggregates, macropores, and galleries) typical of a mull-type humus soil. This is believed to be favorable for the establishment and growth of seedlings. Processes by which earthworm populations establish and the reasons why earthworms become able to live in what was previously an acid environment have not yet been identified. Recent cases of invasion of exotic lumbricid earthworms in northeast United States and Canada provide rather similar situations. There, earthworms are able to feed on organic matter accumulated in these forests, decrease the overall organic matter content, and accelerate the turnover of C and N, increasing microbial biomass and changing microarthropod communities (see Chapter 5 this volume).

DISCUSSION The long-term effects of earthworms on SOM dynamics vary depending on the scale of time considered. When earthworms are introduced artificially into an ecosystem, they use part of the C resources for their activities. In the African savannahs of Lamto (Côte d’Ivoire), the amount of C mineralized directly through earthworm respiration was estimated as 1.2 t C ha−1 years in a grass savannah, which is equivalent to about 5% of overall primary production (Lavelle 1978). The annual average population densities and biomass of earthworms were 202 individuals and 39.7 g m−2, respectively, and this population ingested up to 1000 to 1250 t dry soil ha–1 year–1. As part of this process, nutrients were released and made available to plants or microorganisms. In the same savannah, the overall amounts of assimilable N, released as ammonium in feces, or labile organic N in dead earthworms and mucus were estimated at 21.1 to 38.6 kg ha–1 year–1 of NH4-N in a population of Millsonia anomala that comprised 60% of the overall population biomass. The overall production of mineral N for the earthworm community is therefore expected to range from 30 to 50 kg ha–1 year–1. In tropical pastures, with earthworm biomass of 1 to 3 t ha–1 and soils with higher contents of organic nitrogen, the contribution of earthworms to N mineralization may probably reach a few hundred kilograms mineral N ha−1 year−l. In temperate pastures, the flux of mineral N through earthworms may be estimated as a few hundred kilogram ha–1 year–1 (Syers et al. 1979; Hameed et al. 1994). Similar processes have been observed relating to P turnover, but no actual estimates of amounts of P released ha–1 year–1 have been produced (Sharpley and Syers 1976; Lopez-Hernandez et al. 1993; Brossard et al. 1996). There is some evidence that plants may accumulate these nutrients, but the exact proportion, especially on a yearly basis, is not known (Spain et al. 1992; Hameed et al. 1994). Increased nutrient turnover from earthworm activities usually results in increased plant growth. Most experiments on the scale of one to six successive cropping cycles showed significant positive effects of earthworm activities on plant productivity; these effects seem to be proportional to the earthworm biomass, within a limited range of biomass (Lavelle 1997). Whether this increased production is sustained in the long term is not known. On one hand, earthworms tend to feed on existing stocks of almost undecomposed organic matter and accelerate their decomposition. Once these stocks are depleted, earthworm activities may cease, and the system returns to lower levels of plant production.

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Observations made in rubber plantations of different ages in Côte d’Ivoire seem to support this hypothesis (Gilot et al. 1995). Observations of successional processes in an alpine spruce forest of France showed that earthworm activities seem to have been reduced by an accumulation of lowquality litter residues that they could not process rather than by the exhaustion of available organic resources. In that case, spruce litter may have become palatable only after long periods of maturation, during which fungal attacks progressively eliminated those toxic compounds present in fresh litter. The effect of litter quality on earthworm activities has already been stressed in studies on Finnish spruce forests, in which the input of high-quality litter allowed earthworm populations to increase significantly even in an acid environment (Huhta 1979). However, earthworms may participate in the accumulation of organic matter through (1) an overall increase of amounts of organic matter produced in an ecosystem and (2) the protection of SOM in structures of the earthworm drilosphere (burrow system) (Martin 1991). In a 3-year experiment at Yurimaguas and Lamto, the combination of C consumption by inoculated earthworms, the increased capture of C by plants, and protection of SOM in compact casts did not result in significant changes in the abundance of C. Nevertheless, there were clear indications that the quality of organic matter, estimated by either physical fractionation or respirometry, was modified. Longterm consequences of such modifications are not predictable yet. Physical effects on soils resulting from earthworm activities seem to persist for long periods. Blanchart et al. (1993) demonstrated that casts deposited by large earthworms (of the compacting category) had largely kept their structures 30 months after the earthworms had disappeared. This resistance of casts, which seems to be because of an outstanding stability of aggregates, seems largely dependent on soil type (Zhang and Schrader 1997). Stabilized earthworm casts tend to conserve organic matter because little microbial activity is possible in these compact structures (McInerney and Bolger 2000; Zhang et al. 2003). Nevertheless, in natural conditions, such macroaggregates cannot comprise more than 40 to 60% of the total soil of the Lamto savannah (Blanchart et al. 1993). Despite their continuous production, these aggregates do not accumulate beyond that limit, probably because earthworm populations that decompact soil also regulate the extent of aggregation. These may be the small species earthworms of Eudrilidae at Lamto. They may also be species of Enchytraeidae, ants, termites, myriapoda, or microarthropods. Therefore, the long-term efficiency of these processes of protection of aggregates in earthworm casts depends largely on (1) the maximum percentage of large aggregates present at a given site and (2) the lifetime of these aggregates. Spatial patterns of earthworm populations may be understood best in terms of their overall impact on soils. At Lamto, there was a significant negative correlation in some seasons among populations of compacting earthworm species that accumulate large compact casts in the upper 20 cm of soil, decompacting species that produce fine granular-shaped casts from soil macroaggregates, and the large casts produced by the former species (Rossi 1997). This may indicate that earthworm subpopulations have successions inside a patch. However, in the Colombian savannah studied by Decaëns and Rossi (2001), the spatial distribution of earthworm populations did not show much change over a 3-year monitoring period, indicating that possible changes in population distribution patterns occur only at rather long scales of time. Populations of earthworm species that compact soils can develop once those that decompact soils have produced small aggregates rich in the organic matter that populations of the compacting species may ingest. Such distribution patterns may also be determined by local availability of assimilable organic matter. Studies in both hevea plantations and spruce forests showed clearly that the impacts of earthworms become important when organic matter has been sufficiently prepared by digestion by a succession of termite and arthropods digestions in the first case and the development of fungal colonization in the second. These results and our hypotheses open several avenues for future research:

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1. Testing the effect of earthworm activities on the quality of soil organic matter (through physical fractionations and respirometry) in long-term experiments using natural 13C labeling after, for example, a change from C3 to C4 vegetation (Desjardins et al., in press). 2. Establishment of carbon budgets to quantify the amount of C converted by earthworms and identify the pools in which this C is taken. 3. Experiments on the effects of the addition of organic matter, in different physical and chemical qualities, on earthworm activities. This last research theme has clear practical implications because what is envisaged is the improvement of earthworm activities in agricultural systems with a view to increase crop production and soil sustainability.

REFERENCES Agren, G.I., R.E. McMurtrie, W.J. Parton, J. Pastor, and H.H. Shugart. 1991. State-of-the-art of models of production-decomposition linkages in conifer and grassland ecosystems, Ecol. Appl., 1, 118–138 Aina, P.O. 1984. Contribution of earthworms to porosity and water infiltration in a tropical soil under forest and long-term cultivation, Pedobiologia, 26, 131–136. Alegre, J., B. Pashanasi, and P. Lavelle. 1996. Dynamics of soil physical properties in a low input agricultural system inoculated with the earthworm Pontoscolex corethrurus in Peruvian Amazonia, Soil Sci. Soc. Am. J., 60, 1522–1529. Barois, L. and P. Lavelle. 1986. Changes in respiration rate and some physicochemical properties of a tropical soil during transit through Pontoscolex corethrurus (Glossoscolecidae, Oligochaeta), Soil Biol. Biochem., 18, 539–541. Barois, I., P. Lavelle, M. Brossard, J. Tondoh, M.A. Martinez, J.P. Rossi, B.K. Senapati, A. Angeles, C. Fragoso, J.J. Jimenez, T. Decaëns, C. Lattaud, K.K. Kanyonyo, E. Blanchart, L. Chapuis, G.G. Brown, and A. Moreno. 1999. Ecology of earthworm species with large environmental tolerance and/or extended distributions, in Earthworm Management in Tropical Agroecosystems, P. Lavelle, L. Brussaard, and P. Hendrix, Eds., CAB International, Wallingford, U.K., pp. 57–86. Bernier, N., J.F. Ponge, and J. Andre. 1993. Comparative study of soil organic layers in two bilberry spruce forest stands (Vaccinio-Picetea). Relation to forest dynamics, Geoderma, 59, 89–108. Blair, J.M., R.W. Parmelee, and P. Lavelle. 1994. Influence of earthworms on biogeochemistry in North American ecosystems, in Earthworm Ecology in Forest Rangeland and Crop Ecosystems in North America, P.F. Hendrix, Ed., Lewis Publishers, Chelsea, MI, pp. 1–44. Blanchart, E., A. Albrecht, J. Alegre, A. Duboisset, B. Pashanasi, P. Lavelle, and L. Brussaard. 1999. Effects of earthworms on soil structure and physical properties, in Earthworm Management in Tropical Agroecosystems, P. Lavelle, L. Brussaard, and P. Hendrix, Eds., CAB International, Wallingford, U.K., pp. 139–162. Blanchart, E., A. Bruand, and P. Lavelle. 1993. The physical structure of casts of Millsonia anomala (Oligochaeta: Megascolecidae) in shrub savannah soils (Côte d’Ivoire), Geoderma, 56, 119–132. Bonkowski, M., B.S. Griffiths, and K. Ritz. 2000. Food preferences of earthworms for soil fungi, Pedobiologia, 44, 666–676. Bouché, M.B. 1977. Stratégies lombriciennes, in Soil Organism as Components of Ecosystems, T. Persson and U. Lohm, Eds., Ecol. Bull. (Stockholm), 25, 122–132. Brossard, M., P. Lavelle, and J.Y. Laurent. 1996. Digestion of a vertisol by an endogeic earthworm (Polypheretima elongata, Megascolecidae) increases soil phosphate extractibility, Eur. J. Soil Biol., 32, 107–111. Chauvel, A., M. Grimaldi, E. Barros, E. Blanchart, T. Desjardins, M. Sarrazin, and P. Lavelle. 1999. Pasture degradation by an Amazonian earthworm, Nature, 389, 32–33. Darwin, C. 1881. The Formation of Vegetable Mould Through the Action of Worms, with Observations of Their Habits, Murray, London. Decaëns, T., J.H. Galvin, and E. Amezquita, E. 2001. Propriétés des structures produites par les ingénieurs écologiques à la surface du sol d’une savane colombienne, C.R. Acad. Sci. Life Sci., 324, 465–478. Decaëns, T., P. Lavelle, J.J. Jimenez Jaen, G. Escobar, and G. Rippstein. 1994. Impact of land management on soil macrofauna in the Oriental Llanos of Colombia, Eur. J. Soil Biol., 30, 157–168.

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Decaëns, T. and J. Rossi. 2001. Spatial and temporal structure of earthworm community and soil heterogeneity in a tropical pasture, Ecography, 24, 671–682. Derouard, L., J. Tondoh, Vilcosqui, and P. Lavelle. 1997. Species-specific effects in the response of tropical annual crops to the inoculation of earthworms. Short-scale experiments at Lamto (Côte d’Ivoire), Soil Biol. Biochem., 29(3–4), 541–545. Desjardins, T., F. Charpentier, B. Pashanasi, A. Pando-Bahon, P. Lavelle, and A. Mariotti. (in press). Effects of earthworm inoculation on soil organic matter dynamics of a cultivated ultisol, Pedobiologia. Devliegher, W. and W. Verstraete. 1997. Microorganisms and soil physico-chemical conditions in the drilosphere of Lumbricus terrestris, Soil Biol. Biochem., 29, 1721–1729. Didden, W.A.M. 1990. Involvement of Enchytraeidae (Oligochaeta) in soil structure evolution in agricultural fields, Biol. Fertil. Soils, 9, 152–158. Edwards, C.A. and P.J. Bohlen. 1996. Biology and Ecology of Earthworms, Chapman & Hall, London. Elliott, E.T. 1986. Aggregate structure and carbon, nitrogen, and phosphorus in native and cultivated soils, Soil Sci. Soc. Am. J., 50, 627–633. Eswaran, F. 1994. Soil resilience and sustainable land management in the context of AGENDA 21, in Soil Resilience and Sustainable Land Use, D.J. Greenland and I. Szabolcs, Eds., CAB International, Wallingford, U.K. FAO. 2000. La situation mondiale de l’alimentation et de l’agriculture. FAO (Rome), vol. 32, 497 pp. Feller, C. 1979. Une méthode de fractionnement granulométrique de la matière organique des sols: application aux sols tropicaux à texture grossière, très pauvres en humus, Cahiers de l’O.R.S.T.O.M., Sér. Pédol., 17(4), 339–346. Ferrière, G. 1980. Fonctions des Lombriciens. VII. Une méthode d’analyse de la matière organique végétale ingérée, Pedobiologia, 20, 263–273, Gilot, C. 1997. Effects of a tropical geophageous earthworm, M-anomala (Megascolecidae), on soil characteristics and production of a yam crop in Ivory Coast, Soil Biol. Biochem., 29, 353–359. Gilot, C., P. Lavelle, Ph. Kouassi, and G. Guillaume. 1995. Biological activity of soils in Hevea stands of different ages in Côte d’Ivoire, Acta Zool. Fenn., 196, 186–190. Hameed, R., M.B. Bouché, and J. Cortez. 1994. Etudes in situ des transferts d’azote d’origine lombricienne (Lumbricus terrestris L.) vers les plantes, Soil Biol. Biochem., 26, 495–501. Huhta, V. 1979. Effects of liming and deciduous litter on earthworm (Lumbricidae) populations of a spruce forest, with an inoculation experiment on Allolobophora caliginosa, Pedobiologia, 19, 340–345. Jenkinson, D.S. and J.H. Rayner. 1977. The turnover of soil organic matter in some of the Rothamsted classical experiments, Soil Sci., 123, 298–305. Jimenez, J.J., J.P. Rossi, and P. Lavelle. 2001. Spatial distribution of earthworms in acid-soil savannas of the Eastern plains of Colombia, Appl. Soil Ecol., 17(3), 514. Joschko, M., H. Diestel, and O. Larink. 1989. Assessment of earthworm burrowing efficiency in compacted soil with a combination of morphological and soil physical measurements, Biol. Fertil. Soils, 8, 191–196. Kanyonyo ka Kajondo, J.B. 1984. Ecologie alimentaire du ver de terre detritivore Millsatia lamtoiarta (Acanthdrilidae, Oligochètes) dans la savane de Lamto (Côte d’Ivoire), Ms.C. thesis, Paris VI. Kladivko, E.J., A.D. Mackay, and J.M. Bradford. 1986. Earthworms as a factor in the reduction of soil crusting, Soil Sci. Soc. Am. J., 50, 191–196. Kooistra, M.J. 1991. A micromorphological approach to the interactions between soil structure and soil biota, Agric. Ecosyst. Environ., 34, 315–328. Lavelle, P. 1978. Les Vers de Terre de la Savane de Lamto (Côte d’Ivoire): Peuplements, Populations, et Fonctions dans l’Écosystème, Publications du Laboratoire de Zoologie de l’ENS no. 12, Paris. Lavelle, P. 1997. Faunal activities and soil processes: adaptive strategies that determine ecosystem function, Adv. Ecol. Res., 27, 93–132. Lavelle, P., D. Bignell, M. Austen, Y. Brown, V. Behan-Pelletier, J. Garey, P. Giller, S. Hawkins, G. Brown, M. St. John, B. Hunt, and E. Paul. 2004. Vulnerability of ecosystem services at different scales: role of biodiversity and implications for management, in Sustaining Biodiversity and Functioning in Soils and Sediments, D. H. Wall, Ed., Island Press, New York, in press. Lavelle, P., E. Blanchart, A. Martin, S. Martin, I. Barois, F. Toutain, A. Spain, and R. Schaefer. 1993. A hierarchical model for decomposition in terrestrial ecosystems. Application to soils in the humid tropics, Biotropica, 25, 130–150.

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Lavelle, P., L. Brussaard, and P. Hendrix. 1999. Earthworm Management in Tropical Agroecosystems, CAB International, Wallingford, U.K. Lavelle, P., M. Dangerfield, C. Fragoso, V. Eschenbrenner, D. Lopez-Hernandez, B. Pashanasi, and L. Brussaard. 1994. The relationship between soil macrofauna and tropical soil fertility, in The Biological Management of Tropical Soil, M.J. Swift and P. Woomer, Eds., John Wiley-Sayce, New York, pp. 137–169. Lavelle, P. and A. Martin. 1992. Small-scale and large-scale effects, of endogeic earthworms on soil organic matter dynamics in soils of the humid tropics, Soil Biol. Biochem., 24, 1491–1498. Lavelle, P., G. Melendez, B. Pashanasi, and R. Schaefer. 1992. Nitrogen mineralization and reorganisation in casts of the geophagous tropical earthworm Pontoscolex corethrurus (Glossoscolecidae), Biol. Fertil. Soils, 14, 49–53. Lavelle, P. and A.V. Spain. 2001. Soil Ecology, Kluwer Scientific, Amsterdam. Lee, K.E. 1985. Earthworms: Their Ecology and Relationships with Soils and Land Use, Academic Press, New York. Lopez-Hernandez, D., J.C. Fardeau, and P. Lavelle. 1993. Phosphorus transformations in two P-sorption contrasting tropical soils during transit through Pontoscolex corethrurus (Glossoscolecidae, Oligochaeta), Soil Biol. Biochem., 25, 789–792. Mariani, L., N. Bernier, J.J. Jimenez, and T. Decaëns. 2001. Diet of an anecic earthworm from the Colombian savannas: a question about ecological categories, C.R. Acad. Sci. Sér. 3, 733–742. Martin, A. 1991. Short- and long-term effects of the endogeic earthworm Millsonia anomala (Omodeo) (Megascolecidae, Oligochaeta) of tropical savannas on soil organic matter, Biol. Fertil. Soils, 11, 234–238. McInerney, M. and T. Bolger. 2000. Decomposition of Quereus patracea litter: influence of burial, comminution and earthworms, Soil Biol. Biochem., 32, 1989–2000. McLean, M.A. and D. Parkinson. 1997. Soil impacts of the epigeic earthworm Dendrobaena octaedra on organic matter and microbial activity in lodgepole pine forest, Can. J. Forest Res., 27, 1907–1913. McLean, M.A. and D. Parkinson. 2000. Field evidence of the effects of the epigeic earthworm Dendrobaena octaedra the microfungal community in pine forest floor, Soil Biol. Biochem., 32, 351–360. Miles, J. 1985. Soil in the ecosystem, in Ecological Interactions in Soil: Plants, Microbes and Animals, D. Atkinson, A.H. Fitter, D.J. Read, and M.B. Usher, Eds., Blackwell Scientific, Oxford, U.K., pp. 407–427. Molina, J.A.E., C.E. Clapp, M.J. Shaffer, F.W. Chichester, and W.E. Larson. 1983. NCSOIL, a model of C and N transformations in soils: description, calibration and behaviour, Soil Sci. Soc. Am. J., 47, 85–91. Neilson, R., B. Boag, and M. Smith. 2000. Earthworm delta C-13 and delta N-15 analyses suggest that putative functional classifications of earthworms are site-specific and may also indicate habitat diversity, Soil Biol. Biochem., 32, 1053–1061. Oades, J.M. 1984. Soil organic matter and structural stability: mechanisms and implication for management, Plant Soil, 76, 319–337. Parton, W.J., J.W.B. Stewart, and C.V. Cole. 1988. Dynamics of C, N, P and S in grasslands soils: a model, Biogeochemistry, 5, 109–131. Pashanasi, B., P. Lavelle, and J. Alegre. 1996. Effect of inoculation with the endogeic earthworm Pontocolex corethrurus on soil chemical characteristics and plant growth in a low-input agricultural system of Peruvian Amazonia, Soil Biol. Biochem., 28, 801–810. Phillipson, J., R. Abel, J. Steel, and S.R. Woodell. 1976. Earthworms and the factors that govern their distribution, Oecologia, 33, 291–309. Piearce, T.G. 1978. Gut contents of some lumbricid earthworms, Pedobiologia, 18, 3–157. Poier, K.R. and J. Richter. 1992. Spatial distribution of earthworms and soil properties in an arable loess soil, Soil Biol. Biochem., 24, 1601–1608. Rossi, J.P. 1997. Rôle fonctionnel de la distribution spatiale des vers de terre dans une savane humide de Côte d’lvoire. Doctoral thesis, University of Paris VI, Paris. Rossi, J.P., P. Lavelle, and A. Albrecht. 1997. Relationships between spatial pattern of the endogeic earthworm Polypheretima elohgata and soil heterogeneity in a tropical pashue of Martinique (French West Indies), Soil Biol. Biochem., 29(3–4), 481–485. Satchell, J.E. 1983. Earthworm Ecology from Darwin to Vermiculture, Chapman & Hall, London, 495 pp.

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Scheu, S. 1992. Changes in the lumbricid coenosis during secondary succession from a wheat field to a beechwood on limestone, Soil Biol. Biochem., 24, 1641–1646. Scullion, J. and A. Malik. 2000. Earthworm activity affecting organic matter, aggregation and microbial activity in soils restored after opencast mining for coal, Soil Biol. Biochem., 32, 119–126. Sharpley, A.N. and J.K. Syers. 1976. Potential role of earthworm casts for the phosphorus enrichment of runoff waters, Soil Biol. Biochem., 8, 341–346. Shaw, C. and S. Pawluk. 1986. The development of soil structure by Octolasion tyrtaeum, Aporrectodea turgida and Lumbricus terrestris in parent materials belonging to different textural classes, Pedobiologia, 29, 327–339. Shuster, W.D., S. Subler, and E.L. McCoy. 2001. Deep-burrowing earthworm additions changed the distribution of soil organic carbon in a chisel-tilled soil, Soil Biol. Biochem., 33, 983–996. Spain, A.V., P. Lavelle, and A. Mariotti. 1992. Stimulation of plant growth by tropical earthworms, Soil Biol. Biochem., 24, 1629–1634. Springett, J.A. 1983. Effect of five species of earthworm on some soil properties, J. Appl. Ecol., 20, 865–887. Syers, J.K., A.N. Sharpley, and D.R. Keeney. 1979. Cycling of nitrogen by surface-casting earthworms in a pasture ecosystem, Soil Biol. Biochem., 11, 181–185. Tajovsky, K., G. Villemin, and F. Toutain. 1991. Microstructural and ultrastructural changes of the oak leaf litter consumed by millipede Glomeris hexasticha (Diplopoda), Rev. d’Ecol. Biol. du Sol, 28, 287–302. Tiunov, A.V., M. Bonkowski, J. Alphei, and S. Scheu. 2001. Microflora, Protozoa and Nematoda in Lumbricus terrestris burrow walls: a laboratory experiment, Pedobiologia, 45, 46–60. Tsakala, R. 1994. Evolution Spatio-Temporelle de la Minéralisation du Carbne et de l’Azote dans les Sols de Deux Parcelles à Lamto (Côte d’Ivoire), Mémoire DESS, Université Paris 12. Van Vliet, P.C.J., L.T. West, P.F. Hendrix, and D.C. Coleman. 1993. The influence of Enchytraeidae (Oligochaeta) on the soil porosity of small microcosms, Geoderma, 56, 287–299. Wolters, V. 1991. Soil invertebrates — effects on nutrient turnover and soil structure — a review, Z. Pflanzenernahr. Bodenk., 154, 389–402. Zhang, B.G., G.T. Li, T.S. Shen, J.K. Wang, and Z. Sun. 2000. Changes in microbial biomass C, N, and P and enzyme activities in soil incubated with the earthworms Metaphire guillelmi or Eisenia fetida, Soil Biol. Biochem., 32, 2055–2062. Zhang, X., J. Ang, H. Xie, J. Wang, and W. Zech. 2003. Comparison of organic compounds in the particlesize fractions of earthworm casts and surrounding soil in humid Laos, Appl. Soil Ecol., 23, 147–153. Zund, P.R., U. Pillai McGarry, D. McGarry, and S.G. Bray. 1997. Repair of a compacted oxisol by the earthworm Pontoscolex corethrurus (Glossoscolecidae, Oligochaeta), Biol. Fertil. Soils, 25, 202–208.

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the Effects of 9 Integrating Earthworms on Nutrient Cycling across Spatial and Temporal Scales Patrick J. Bohlen Archbold Biological Station, Lake Placid, Florida, U.S.A.

Robert W. Parmelee Yucca Valley, California, U.S.A.

John M. Blair Division of Biology, Kansas State University, Manhattan, Kansas, U.S.A.

CONTENTS Overview ........................................................................................................................................162 Conceptual Models ........................................................................................................................165 Mechanistic Models ................................................................................................................165 Ecosystem Budget Models .....................................................................................................168 Future Experiments.................................................................................................................170 A Hierarchical Approach ...............................................................................................................171 Individual Earthworms............................................................................................................171 Earthworm Populations...........................................................................................................173 Earthworm Communities ........................................................................................................174 Earthworms in Ecosystems.....................................................................................................174 Earthworms in Landscapes .....................................................................................................175 Conclusions ....................................................................................................................................176 Acknowledgments ..........................................................................................................................176 References ......................................................................................................................................176

There has been an exponential increase in research publications addressing the impact of earthworms on terrestrial nutrient cycling processes. This research has demonstrated that earthworms significantly affect key soil properties and processes such as microbial biomass and activity, organic matter dynamics, nutrient availability, plant uptake and production, and soil structure. There is sufficient evidence to conclude that earthworms are central to regulating nutrient cycling processes in many ecosystems. Earthworms may alter the balance between conservation and loss of nutrients in ecosystems, and their net influence at large scales defines their role in ecosystem processes. Many investigations have focused on small-scale phenomena, such as nutrient dynamics

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in earthworm casts and burrows, that are critical to a mechanistic understanding of earthworm effects but are difficult to extrapolate to the ecosystem scale. In addition to the challenge of integrating information and analysis across spatial scales, there is the challenge of extrapolating the short-term effects of earthworms through time, which, as Darwin (1881) noted more than a century ago, can lead to cumulative changes at the landscape scale. Further advances in the understanding of the role of earthworms in nutrient cycling processes will depend on experimental studies, modeling approaches, and conceptual advances that integrate across spatial and temporal scales. The greatest challenge to making these advances will continue to be integrating across the ecological hierarchy and understanding the effects of earthworms on nutrient cycling at the scale of whole ecosystems and landscapes through time.

OVERVIEW In this chapter, we focus on research that integrates soil biology, chemistry, and physics and that demonstrates how earthworms affect multiple soil nutrient cycling processes, and we highlight potentially fruitful areas for new research. Although the majority of papers presented at the Fifth International Symposium on Earthworm Ecology in 1994 (Edwards 1998) were conducted in agricultural soils and systems, there has been a substantial increase since that time in research from a variety of natural ecosystems, including tropical forests and grasslands (Fragoso et al. 1999a,b; Lachnict et al. 2002; Liu and Zou 2002; Jiménez et al. 2003; Sanchez-De Leon et al. 2003), temperate grasslands (Callaham and Blair 1999; Callaham et al. 2001), and temperate forests (Scheu and Schaefer 1998; Lachnicht and Hendrix 2001; Bohlen et al. 2004a). Thus, many of the previously noted gaps in the understanding of the effects of earthworms on nutrient cycling processes in natural systems (Parmelee et al. 1998) have been filled. Many studies continue to focus on exotic earthworm and peregrine species, especially in regions where these species dominate the earthworm fauna. Some more recent research has begun to examine the influences of native species of earthworms on nutrient cycling properties (e.g., Callaham and Hendrix 1998; Callaham et al. 2001; Lachnicht and Hendrix 2001). Native communities of earthworms often inhabit ecosystems that have been less disturbed by humans. These communities offer unique opportunities for research that may elucidate differences in the effects of native and exotic species on nutrient cycling. Such research is also badly needed from the standpoint of conservation of the native earthworm fauna. Although much research still focuses on the effects of lumbricids on nutrient cycling, there has been a substantial increase in research on the effect of earthworms on soil fertility and nutrient cycling in the tropics and the importance of other earthworm families, such as the Megascolecidae and Glossoscolecidae (Lavelle et al. 1999). Lavelle et al. (see Chapter 8, this volume) highlight the importance of spatial and temporal scales in addressing the roles of earthworms in nutrient cycling processes. The majority of research has relied on laboratory microcosms or small, manipulated field enclosures that typically span short periods of days to months and, occasionally, to years. The problem of expanding from small spatial and short temporal scales to long-term effects at the field or landscape level is a major barrier to an improved understanding of how earthworms affect nutrient cycles (Blair et al. 1995b). One fruitful approach for investigating the longer-term effects of earthworms on nutrient cycling involves the direct introduction or invasion of earthworms into new habitats (see Chapter 5, this volume). Examining changes in nutrient cycling or storage through time following such earthworm introductions provides a rigorous basis for understanding the integrated effects of earthworms on nutrient cycling. Lavelle et al. (see Chapter 8, this volume) stress the importance of earthworm feeding behaviors, spatial distribution patterns, and the change in earthworm communities during ecological succession, all of which may contribute to the long-term effects of earthworms on nutrient cycling processes. They stress the need to determine the quality and quantity of soil organic matter ingested by earthworms because earthworms may alter the ratio of labile to recalcitrant organic matter and © 2004 by CRC Press LLC

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influence long-term storage of soil C and N. They also summarize research examining how the horizontal distribution of earthworms can influence nutrient cycles. In their study, earthworms existed in patches 20 to 40 m in diameter that were not synchronized with the population dynamics of other patches (i.e., juveniles and adults had different distribution patterns) and could be composed of different species. Such patch dynamics in the distribution of earthworms results in areas of high activity, followed by possible local extinctions and a cessation of activity or increased activity of species occupying different ecological niches. This dynamic patchiness of earthworm communities may have important consequences for soil structure (see Chapter 11, this volume) and spatial and temporal patterns in nutrient cycling processes. Linking the spatial distribution patterns of earthworm populations and communities with underlying soil or hydrological characteristics is an important area of research for examining possible feedbacks between soil physicochemical and biological properties. In a geostatistical analysis of earthworm populations in a temperate grassland, there was a strong correlation between the distribution of the populations of adult earthworms (Lumbricus terrestris and Aporrectodea caliginosa) and soil hydromorphic characteristics (Cannavacciuolo et al. 1998). In that study, the total biomass of adult earthworms was correlated positively with a well-drained area and negatively with a more poorly drained area. Similar to results from Lavelle et al. (Chapter 8, this volume), adult and immature individuals of L. terrestris were concentrated in distinct, non-overlapping patches. A study of the distribution of L. terrestris in relation to artificial subsurface drainage in a grass field in Finland showed that the total number of individuals was twice as high and the total biomass five times as high above the drains as between the drains (Nuutinen et al. 2001). Such a strong correlation between the distribution of adult L. terrestris and soil hydrology demonstrates a clear positive feedback between soil biological and hydromorphic characteristics; well-drained soils enhance populations of deep-burrowing earthworms (L. terrestris), which in turn enhance soil drainage. There are exciting opportunities for exploring feedbacks between the spatial distribution of earthworms and underlying geomorphic and hydromorphic characteristics and how such feedbacks influence nutrient cycling, solute movement, and other processes at the landscape scale. Examining changes in earthworm communities through succession provides another approach to studying the longer-term effects of earthworms on nutrient cycling and soil fertility. There is also the opportunity to examine changes in soil properties as earthworms colonize new habitats. The invasion of earthworms into previously uncolonized soil has altered the size and distribution of soil microbial biomass, plant biomass, and leaching of nutrients as well as soil C and N content and soil profile development (Alban and Berry 1994; Scheu and Parkinson 1994a,b; Bohlen et al. 2004a,b; Groffman et al. 2004; Suárez et al. 2004). Simulation models also can be useful for understanding the longer term effects of earthworms on soil organic matter dynamics. For example, using the CENTURY carbon model, Lavelle et al. (1998) showed that earthworms could promote the stabilization of soil organic matter and maintain organic pools and soil structure at longer time scales (10 to 50 years). Gilot (1997) reported on the effects of earthworms on yam production and soil characteristics in Africa. Earthworms altered soil structure by decreasing the percentage of small aggregates and increasing the percentage of larger ones. These larger aggregates, which are primarily aging earthworm casts, may have contributed to the protection of soil C, as indicated by a 5% decrease in C mineralization after 3 years. In this way, the presence of earthworms generally enhanced crop productivity. For assessing the short-term influence of earthworms on nutrient and organic matter turnover, Devliegher and Verstraete (1997) distinguished between a nutrient enrichment process (NEP) and a gut-associated process (GAP). Earthworms, such as L. terrestris, incorporate and mix surface organic matter with soil and increase biological activity and nutrient availability by such mixing (NEP). However, they also assimilate nutrients from soil and organic matter as these materials pass through their guts (GAP). The authors designed an experiment with Lumbricus terrestris in which the treatments included earthworms plus lettuce on the soil surface, lettuce manually mixed into © 2004 by CRC Press LLC

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the soil with no earthworms, or lettuce left on the surface without earthworms to distinguish between the importance of GAP and NEP. They concluded that GAP reduced microbial biomass and activity, soil nitrate, and the total N content of the crop, but NEP increased in biological activity and nutrient availability. It is likely that the relative importance of GAP and NEP will vary among earthworm species and in soils with different levels of fertility. Indeed, Blair et al. (1995b) noted that stimulatory effects of earthworms on soil biological and chemical activity appeared to be greater in soils lower in C and N content. Blair et al. (1997) suggested a possible mechanism by which earthworm-microbial interactions can influence soil N availability. In field plots of maize in which earthworm populations were reduced by 70%, microbial biomass N was significantly higher than in plots with ambient earthworm populations. After 6 years of manipulating earthworm populations, microbial biomass C in plots with increased earthworm population was 10% lower than in control plots and 17% lower than in plots with decreased earthworm populations (Domínguez et al. 2004). Microcosm experiments yielded similar results; microcosms with earthworms had less microbial biomass N compared with control microcosms without earthworms (Bohlen and Edwards 1995). Soil nitrate levels were higher in plots with increased earthworm populations but only in systems fertilized with inorganic fertilizer, not in systems fertilized with organic fertilizers. Similarly, results from a long-term tillage experiment in Georgia in the United States, in which 15N-labeled crop residues were added to plots with or without added earthworm populations, demonstrated that earthworm activity decreased the standing stocks of microbial biomass and increased turnover of microbial N (Hendrix et al. 1998). Together, these results indicate that earthworms can increase N availability by reducing microbial immobilization and enhancing mineralization. However, earthworms do not always reduce microbial biomass, and their influence can vary with soil type or with differences in the quantity or quality of available organic matter (Wolters and Joergensen 1992; Saetre 1998). The effects of earthworms on nitrogen cycling were important topics at the Fifth International Symposium on Earthworm Ecology (Edwards 1998) and continue to be an active area of study. Bouché et al. (1997) used 15N to develop models to quantify the direct role of earthworms in the N cycle. Steinberg et al. (1997) examined the role of earthworms in affecting the availability of N in forests. They reported that earthworms increased potential net N mineralization rates in both urban and rural oak forests. In similar forest sites from the same region, Burtelow et al. (1998) showed that areas with earthworm populations had greater soil denitrification activity than did areas without earthworms. Other, more recent, studies of earthworm invasion of forests in the northeastern United States did not report large effects of earthworms on N mineralization (Groffman et al. 2004). From the studies discussed above, it is apparent that in the majority of cases of earthworm activity to accelerated transformations of N, often increasing N availability. At the ecosystem level, there are several possible fates for N that is made available by earthworms. Some of this N is available for plant uptake (Devliegher and Verstraete 1997; see Chapter 2, this volume). However, there is also a possibility of increased losses of N because of increased leaching and denitrification. Results from our field research in northern Ohio indicated that earthworms increased the leaching of N below 45 cm, primarily by increasing the volume of water leaching through the profile (Domínguez et al. 2004). In a similar study in southern Ohio, there were larger volumes of leachate at a depth of 45 cm in plots with earthworms added and much higher concentrations of dissolved organic nitrogen (DON) in the leachate compared with control plots with ambient earthworm populations (Subler et al. 1997). The effects of earthworms on the quantity, quality, and movement of soil DON provides another previously unexplored avenue for research. Increased macropore formation due to earthworm burrows increased leaching (e.g., Lachnicht et al. 1997). The importance of earthworms in soil respiration and organic matter dynamics continues to be a central theme in investigating the biogeochemical effects of earthworms. Hendriksen (1997) used pathway analysis to show that earthworms enhanced microbial respiration beyond simply incorporating dung and increasing soil C and water content. Although earthworms are known to increase © 2004 by CRC Press LLC

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the decomposition rates of surface-applied litter (e.g., Tian et al. 1997), little is known about the fate of this material once it is incorporated into the soil. Much of the incorporated surface litter resides in belowground earthworm casts, and a significant portion lines earthworm burrows, indicating the importance of these microsites for subsequent C transformations (Jégou et al. 1998). McCartney et al. (1997) observed strong seasonal effects of earthworms on the coarse and intermediate-size classes of soil organic matter, with amounts of these size classes greater in plots with decreased earthworm plots relative to control plots. The overall question of whether earthworms increase or decrease net C storage is still unresolved and may depend on the temporal and spatial scale at which the question is asked.

CONCEPTUAL MODELS We present a series of conceptual models as an initial step toward developing more sophisticated simulation and system-level models for predicting the influences of earthworms on nutrient cycling processes. Our conceptual models emphasize the kinds of information that need to be incorporated into such models, provide a basis for exploring fundamental questions about the influences of earthworms on biogeochemical cycles, and may help direct the development of mechanistic simulation and ecosystem-level nutrient budget models. Such simulation models may be used to assess how earthworms, by their respiration and excretion of C and N and through their effects on microbial turnover, processing of organic matter, and aggregate formation, affect the production of CO2 and the availability of N in the soil. Nutrient budget models can help determine whether earthworms contribute to sustainability or degradation of ecosystems by examining those processes that affect storage of C and N within the system or loss of C and N from the system. Both types of models need to incorporate spatial and temporal patterns and abiotic constraints (Lavelle et al. 1998).

MECHANISTIC MODELS Our first conceptual model explores mechanisms by which earthworms affect system-level C flux (Figure 9.1). There are five major components by which earthworm activity can affect soil respiration: (1) earthworm respiration, (2) mucus production, (3) microbial turnover, (4) processing of litter and soil organic matter, and (5) changed soil structure (aggregate formation, burrows). These processes are not mutually exclusive, and most involve interrelated processes that influence organic matter mineralization and heterotrophic microbial activity. Although the direct contribution of earthworm respiration to the system flux of C is generally a small proportion of the total heterotrophic respiration (e.g., 5 to 6%; Lee 1985), when earthworm populations are large, their direct contribution to system respiration can be as high as 30% (Hendrix et al. 1987). There are insufficient data available to generalize about the contribution of earthworm respiration to total soil respiration, and more field-based studies are needed under a variety of environmental conditions. Advances in this area could also be made by modeling C flux based on known earthworm respiration rates in relation to temperature combined with detailed analysis of earthworm population dynamics in the field. There is very little information on rates of C excretion in earthworm mucus, which can be a particularly labile form of C. Using 14C-labeled earthworms, Scheu (1991) demonstrated that the production of mucus was a significant pathway of C loss from earthworms and exceeded the C lost by their respiration. Much of the available data demonstrating the importance of earthworm-microbial interactions in soil CO2 production come from microbiological studies of fresh casts and laboratory studies. Earthworm casts typically have higher microbial biomass and respiration rates than surrounding soil (Scheu 1987; Lavelle et al. 1992). This increase of microbial activity in fresh casts is greater than would be predicted by enrichment of soil with added organic matter and is probably stimulated by the readily available C sources in the casts (Tiunov and Scheu 2000). However, the situation is more complex because earthworms also seem to increase microbial turnover. © 2004 by CRC Press LLC

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Respiration and Mucus

+

Microbial Turnover

+

Litter Processing

+

Soil Organic Matter

+

CO2

Loss

Aggregate Formation

FIGURE 9.1 Conceptual model to examine the mechanisms through which earthworms can affect systemlevel flux of C. Note: + indicates an increase in CO2 flux because of earthworms; − indicates a decrease in output; and ± indicates that the effect of earthworms in CO2 output can result in either an increase or a decrease.

Scheu (1987) observed a simultaneous decrease in microbial biomass and increase in microbial respiration in earthworm casts after 2 weeks. Similarly, Wolters and Joergensen (1992) observed a lower microbial biomass but a higher microbial activity per unit biomass in soils after 21 days of earthworm activity. The simultaneous increase in respiration and decrease in microbial biomass suggests that earthworm-worked soil may contain a smaller, but more metabolically active, microbial community than soil without earthworms. Although the short-term dynamics of microbial activity in casts are becoming better documented, there is very little information on the respiration of older casts (Martin 1991). Investigations need to be initiated on how earthworms influence microbial respiration and turnover under field conditions over longer periods of time. It is well known that earthworms increase the rates of disappearance of surface litter, but less is known about how this process affects soil respiration rates. Burial of litter increases its decomposition rate because the litter becomes fragmented and is placed in a more favorable environment for microbial activity (e.g., Beare et al. 1992). Mixing of litter with soil by earthworms may create an even more favorable environment because of the high moisture content and availability of nutrients of fresh casts. It is possible that earthworms may enhance microbial activity beyond that due strictly to the input of litter C (Hendriksen 1997). Bohlen et al. (1997) reported that litter associated with L. terrestris middens had a higher microbial biomass and microbial activity than litter not associated with middens. The movement of litter into soil organic matter pools and the effects of earthworms on different fractions of soil organic matter are also of great interest. Whether earthworms feed on labile or recalcitrant soil organic matter pools has important implications for the long-term storage of C and N. The available evidence suggests that earthworms ingest different pools of soil C selectively (Parmelee et al. 1990; Martin et al. 1992a,b; McCartney et al. 1997), but the effects on ecosystem-level fluxes of C of this selective feeding are unknown. A final component of the mechanistic model is the effect of earthworms on soil aggregate formation. Although the overall effect of earthworms on other components of the model is to increase system fluxes of C, the formation of soil aggregates by earthworms may decrease the availability of C and increase the stabilization of soil organic matter (e.g., Gilot 1997; Scullion and Malik 2000). However, few studies have examined this phenomenon. The results of Martin (1991) were consistent with other research that reported increased short-term mineralization of organic

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matter in earthworm casts. In contrast, long-term results from the same study revealed that respiration rates were lower in earthworm casts than in the surrounding soil, indicating increased protection of organic matter in the stable soil aggregates created by earthworm casting. The effects of casts on C mineralization or stabilization are complicated further by the factors that influence stability of the casts, such as the timing of wetting and drying cycles, soil texture, and cast age. An example of the complexity of these dynamics was shown in a laboratory study, which indicated that earthworm casts could lose more nutrients than uningested soil when there was an increasing intensity of wetting-drying cycles (McInerney and Bolger 2000). The extent and timing of increased protection or loss of C from earthworm casts needs further study. Even less well studied is the influence of earthworm burrows, or the drilosphere, on C mineralization. The drilosphere is enriched in soluble organic C, which stimulates microbial activity (Parkin and Berry 1999). Görres et al. (2001) showed that C mineralization rates were significantly greater in earthworm burrows (41 µg CO2-C g−1 d−1) than in earthworm casts (31 µg CO2-C g−1 d−1). Specific C mineralization rates (µg CO2-C mg C−1 d−1) were also greater in earthworm burrows (2.56) than in casts (1.86), indicating that the greater respiration rates in burrows were not simply because of enrichment of burrow walls with organic matter, but because of a greater turnover rate of the C present. Thus, much of the stimulatory effect of earthworms on soil microbial respiration seems to be because of enhancement of microbial activity at specific microsites within the soil, and the net effects of earthworms on total soil C mineralization is the integrated sum of these microsite effects. In the longer term, C flux is influenced by the substantial changes that take place in these microenvironments over time as casts and burrows age and overall soil structure is modified by earthworm activity (Brown et al. 2000). One factor not considered in the conceptual model of C flux is the potential effect of earthworms on the loss of dissolved and particulate forms of soil C. Earthworm casts are susceptible to erosion, especially in cultivated soils and other erosion-prone systems. In a temperate maize agroecosystem, the erosion of earthworm casts contributed to annual sediment losses that could amount to up to 32.6 g C m−2 year−1 of organic C, which compares with the amounts of C lost from earthworm respiration in some systems (Binet and Le Bayon 1999; Le Bayon et al. 2002). Few researchers have examined the influence of earthworms on losses of dissolved C, but earthworms can increase the leaching losses of DON, which presumably would be linked to losses of dissolved organic C (Subler et al. 1997; Domínguez et al. 2004). Losses of dissolved C can represent a substantial proportion of total C loss in some ecosystems, and the effects of earthworms on such losses are not well known. A mechanistic model that is similar to the one presented for C can be developed for the effects of earthworms on the availability or loss of N (Figure 9.2). Generally, the same components incorporated into the carbon model are also included in the nitrogen model. An important distinction is an earthworm excretion compartment. Although the contribution of earthworms to total soil respiration is often considered small, estimates of earthworm excretion of N can be quite large (18 to 50 kg N ha−l; Lee 1983). The state of knowledge and need for further research to improve the understanding of the influence of earthworms on N cycling, microbial turnover, burial of litter, and aggregate formation are the same as for carbon. However, the potential fate of N in the N model differs from that of C in the C model. Although C is eventually lost from the system, mainly as CO2, increased concentrations of available N can lead to a loss of N from the system through leaching, overland flow, and gaseous flux or retention in the system by microbial or plant uptake. In addition, because N availability can affect both plant productivity and decomposition, there are important feedbacks between changes in N availability and net C exchange that need to be explored in future research. Each component of the models, with the exception of earthworm respiration, interacts strongly with the other components. It is the interactions among microbial activity, excretion of mucus, processing of organic matter, and protection of C and N from microbial attack in stable aggregates

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Excretion and Mucus

+

Microbial Turnover

+

Litter Processing

+

Soil Organic Matter

+

Leaching Available Nitrogen

Uptake Runoff Denitrification

Aggregate Formation

FIGURE 9.2 Conceptual model to examine the mechanisms through which earthworms can affect systemlevel availability and fate of N. Note: + indicates that earthworms increase the availability of N; − indicates a decrease in availability; and ± indicates that the effect of earthworms can result in either an increase or a decrease [in N availability].

that determine the overall role of earthworms in ecosystem fluxes of C and N. Experiments that address more than one component will provide greater understanding of the mechanisms by which earthworms affect the cycling of C and N than will experiments or studies addressing a single component. The greatest insights into the effects of earthworms on storage or loss of C and N at the ecosystem scale will probably come from studies that operate on large temporal or spatial scales. Furthermore, investigating earthworm invasions and their effects on nutrient cycling and soil organic matter dynamics can provide insights on the integrated effects of earthworms at these larger spatial scales (Alban and Berry 1994; Bohlen et al. 2004a,b).

ECOSYSTEM BUDGET MODELS We have incorporated concepts from the conceptual mechanistic models into conceptual ecosystem models that compare nutrient fluxes in the presence (Figure 9.3) and the absence (Figure 9.4) of earthworms in an agroecosystem. These models emphasize the major pathways by which earthworms change the retention and loss of C and N, incorporating the effects of earthworms on soil biological, chemical, and physical processes. To determine the roles of earthworms in agroecosystem sustainability, it may be necessary to focus on processes by which earthworms increase or decrease the storage or loss of nutrients and how they influence productivity and nutrient uptake by crops. Quantification of these nutrient budget models could provide data essential for determining whether earthworms are always beneficial organisms in the context of agroecosystem sustainability or whether they may sometimes have deleterious effects. The presence of earthworms can change the sizes of various nutrient pools and fluxes of C and N significantly (Figure 9.3). Earthworms reduce pool sizes of surface litter (Bohlen et al. 1997), coarse and particulate soil organic matter (Parmelee et al. 1990; McCartney et al. 1997), and microbial biomass (Blair et al. 1997). Through interactions of earthworms with the microbial community and by processing organic matter, earthworms can increase the system flux of CO2 (gaseous C loss), as discussed in the section on mechanistic models (Figure 9.1). These same interactions, coupled with earthworm excretion, can also lead to increased availability of N. In a long-term earthworm manipulation study in Ohio, the increases in available N corresponded to increased concentrations of nitrate in some years (Blair et al. 1997). Available N can be retained within an ecosystem by microbial or plant uptake, although we did not observe increased plant N uptake associated with increased earthworm populations in our field experiments

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Gaseous Loss

Crop

Fertilizer Litter

Runoff

Root Soil Organic C and N

Microbial Biomass

Available C and N Burrow

Earthworm Plant Uptake Stable Aggregates

Matrix and Bypass

FIGURE 9.3 Ecosystem budget model to examine pools and ßuxes of C and N in the presence of earthworms. Note: Bold boxes indicate pools and ßuxes in which earthworms are predicted to have a particularly signiÞcant impact.

(Stinner et al. 1997). (Mechanisms by which earthworms inßuence plant uptake of nutrients are addressed further by Brown and Edwards in Chapter 2 of this volume.) There is also a potential for N to be lost from the system through leaching. We observed increased nitrate levels deeper in the soil (Blair et al. 1997) and increased leachate volume and total N leaching below 45 cm in plots with increased earthworm populations relative to plots with decreased populations (Domínguez et al. 2004). In a similar study, DON concentrations were greater in soil leachates in plots with added earthworms than in control plots with ambient populations (Subler et al. 1997). The increased movement of N through the soil proÞle may be because of greater numbers of earthworm burrows (Lachnicht et al. 1997) and greater bypass or preferential ßow. A 15-year tillage study in Finland showed that the saturated hydraulic conductivity of soil was correlated positively with earthworm burrows, and that deeper burrows not continuous to the surface could act as preferential ßow paths at the topsoil-subsoil interface (Pitkänen and Nuutinen 1998). It is clear that earthworm burrows can have persistent effects on soil drainage, but a better understanding is needed of the long-term effect of burrows on the potential loss of nutrients through leaching. Earthworm cast production may also play an important role in system-level ßuxes of C and N. Earthworm casts are rich in available C and N and may include anaerobic microsites; these are conditions that favor denitriÞcation (Svensson et al. 1986). DenitriÞcation rates are typically greater in earthworm casts than in surrounding soil (Elliott et al. 1990, 1991). Earthworm casts deposited on the surface of English pastures were estimated to contribute from 12 to 26% of total denitriÞcation in those pastures (Knight et al. 1992). We found that denitriÞcation rates in middens of L. terrestris, which include a mixture of surface litter and earthworm casts, were four to ten times greater than in surrounding bare soil and could contribute up to 20% of total denitriÞcation in a corn agroecosystem (P. Bohlen, unpublished data). Although denitriÞcation

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Gaseous Loss

Crop

Fertilizer Litter

Runoff

Root

Soil Organic C and N

Plant Uptake

Microbial Biomass

Available C and N

Stable Aggregates

Matrix and Bypass

FIGURE 9.4 Ecosystem budget model to examine pools and ßuxes of C and N in the absence of earthworms. Note: Bold boxes indicate pools and ßuxes that are predicted to be affected most by the absence of earthworms.

from casts can contribute to overall losses of N from the system, the formation of stable soil aggregates from casts could lead to longer term protection of C and N. In the absence of earthworms (Figure 9.4), our model predicts that surface litter and soil organic C and N pools are larger than when earthworms are present. The microbial biomass C and N pool is also larger but with a predicted slower turnover rate. The increase in size of these pools, coupled with longer residence times, could lead to a decrease in available C and N. As a consequence, there would be less loss of C and N in gaseous and leaching ßuxes. Leaching, in general, would be expected to be less because of the absence of earthworm burrows, which contribute to bypass ßow. However, the absence of earthworm burrows, which open to the soil surface, could also lead to increased overland ßow and associated loss of nutrients in runoff. The role of earthworms in effecting overland ßow of nutrients has rarely been investigated, but Sharpley et al. (1979) reported increased losses of N and P in surface runoff in pastures without earthworms (treated with carbaryl) compared with that in pastures with earthworms present. Before it can be determined whether earthworms contribute to a net gain or loss of nutrients from a system, their inßuence on these various nutrient pools and ßuxes needs further quantiÞcation.

FUTURE EXPERIMENTS To provide data to validate the simulation and nutrient budget models, we suggest experimental approaches that are Þeld based and that use manipulated earthworm communities, stable isotopes (13C and 15N), and, when possible, extend for several years or use a chronosequence to explore effects on nutrient cycling over longer time scales. To investigate the effects of earthworms on microbial activity and turnover, microbial biomass could be labeled with 13C-labeled glucose and 15N-labeled ammonium © 2004 by CRC Press LLC

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sulfate or ammonium nitrate in Þeld plots with different populations of earthworms. To examine how earthworms inßuence the redistribution and lability of surface and soil organic matter, double labeling of plants with 13C and 15N holds much promise. Over time, the soil could be analyzed for organic fractions with different mineralization kinetics. Estimates of surface casting by earthworms are not difÞcult to make, but creative methodologies need to be developed to quantify amounts of belowground cast production in the Þeld. The use of ßuorescent dyes may be suitable. Equally important is assessment of the length of time over which casts maintain their structural integrity and how C and N cycling processes change as casts age. Finally, more information is required on how earthworms contribute to gaseous, leaching, and overland ßow ßuxes of nutrients.

A HIERARCHICAL APPROACH Determining the role of earthworms in nutrient cycling processes can unify many different aspects of the biology and ecology of earthworms. Studies at all levels of the ecological hierarchy can provide information that will contribute to a better understanding of how earthworms inßuence biogeochemical cycles at larger spatial and temporal scales (Figure 9.5).

INDIVIDUAL EARTHWORMS Information on the physiological ecology of earthworms can help in understanding the type and amounts of organic matter that earthworms consume and assimilate. Fundamental to an understanding of the biology of any soil-inhabiting organism is knowledge of its C and N assimilation efÞciency. QuantiÞcation of C and N assimilation efÞciencies is important because these could

Landscape

Spatial Patterns Source-Sink Relationships

Ecosystem

Storage vs. Loss of C and N

Community

C and N Flux Ecological Groups

Population

Biomass and Growth Reproduction and Mortality

Individual

Assimilation and Excretion Gut Microbiology

FIGURE 9.5 Diagram showing a hierarchical approach for investigating the effects of earthworms on nutrient cycling processes.

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allow for estimates of the amounts and types of C that need to be consumed by earthworms to support their populations. Very little information is available on the C and N assimilation efficiencies of earthworms, and there are no data for a wide range of common earthworm species. Carbon assimilation efficiency is assumed to be quite low (e.g., 2 to 6%) (Bolton and Phillipson 1976), but this remains to be confirmed for organic matter of different quality and for different earthworm species. There is also a need to determine the N assimilation efficiencies of earthworms, particularly relating to their role in the N cycle (Binet and Trehen 1992). Although there is certainly enough N in the soil to satisfy the demand by earthworms for N, much of it is considered unavailable. Nevertheless, earthworms excrete copious amounts of N in their urine and mucus, and N turnover appears to be quite rapid (Ferriere and Bouché 1985; Barois et al. 1987; Binet and Trehen 1992). Therefore, earthworms do not appear to conserve N efficiently, which suggests that they either ingest materials with high N content selectively or have very high N assimilation efficiencies. Using a budget approach with 15N, Binet and Trehen (1992) calculated an N assimilation efficiency of 27%, much higher than estimates of C assimilation efficiencies. A more detailed quantification of earthworm assimilation efficiencies was done using two species of field-collected earthworms in laboratory experiments with 15N-labeled soybean and rye litter (Whalen and Parmelee 1999). The N assimilation efficiency determined by this method ranged from 10 to 26% for Aporrectodea tuberculata and 25 to 30% for L. terrestris, which is very close to the assimilation efficiencies determined by Binet and Trehen using a budgetary approach. As with C, assimilation rates of N need to be determined for a wide range of earthworm species that occupy different ecological niches and for food resources of different quality. It is likely that the absence of research on these critical areas has been hampered by a lack of suitable methodologies, but the use of stable isotopes has considerable promise for quantifying C and N assimilation efficiencies (Whalen and Janzen 2002, 2003; Schmidt et al. 2003). The nitrogenous compounds that earthworms excrete in their urine and mucus may provide a particularly labile source of N for soil microorganisms. Earthworm urine is composed primarily of ammonium and urea, and the mucus is composed of mucoproteins with a low C:N ratio of 3.8 (Scheu 1991). There are few realistic estimates of rates of urine or mucus excretion by earthworms. Early estimates of urine and mucus excretion rates were based on the method of Needham (1957), who placed earthworms in a small volume of water and analyzed the water for N content after 24 hours. Whether such estimates are realistic remains to be confirmed. Nevertheless, it appears that the amounts of N excreted in earthworm urine and mucus could be substantial (Dash and Patra 1979; Lee 1983). In our chapter in the first edition of this book (Parmelee et al. 1998), we suggested that 15N would be a useful tool in tracer studies for quantifying earthworm N excretion rates; in fact, this tool has been applied to N excretion studies since that time. Nitrogen excretion rates of 274 to 744 mg N g−1 fresh earthworm tissues were estimated from experiments in which 15N-labeled earthworms were incubated for 48 hours in unlabeled soil (Whalen et al. 2000). Combining these excretion rates with data for earthworm population and biomass data from the field resulted in estimates of total annual N excretion of 41.5 kg N ha−1 in a corn agroecosystem in Ohio. That estimate was equivalent to 22% of crop uptake and compared well with an estimate of 30 to 40 kg N ha−1 year−1 excreted by a population of L. terrestris in temperate woodland (Satchell 1963) and 29 to 36 kg N ha−1 year−1 estimated from 15N turnover in temperate grassland (Curry et al. 1995). Although results from labeling experiments must carefully consider assumptions used to calculate N excretion and turnover (Schmidt et al. 2003; Whalen and Janzen 2003), this approach still holds considerable promise. Another particularly intriguing area of research is the microbiology of the earthworm gut. Microbial processes in the earthworm gut may be important in increasing the assimilation efficiencies of C and N. Barois and Lavelle (1986) first proposed the hypothesis that earthworms and soil microorganisms have developed a mutualistic relationship to exploit soil organic matter reserves © 2004 by CRC Press LLC

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better. As soil enters the earthworm gut, moisture levels, pH, and water-soluble C content increase. The soil is then mixed in the gizzard, followed by a significant increase in microbial activity. Water and solutes are reabsorbed in the hindgut, and soil is egested with the C content reduced by as much as 19%. The “priming” of the microbial community with mucus leads to more complete and efficient extraction of nutrients. The commonly observed burst of microbiological activity that occurs in freshly deposited earthworm casts may merely be an after effect of the earthworm gut processes. Although there is increasing evidence that this hypothesis is correct for both temperate and tropical earthworms, research into earthworm gut microbiology will continue to provide valuable information on how earthworms affect nutrient cycling processes. Studies of the microbiology, and nutrient dynamics of casts from different species of earthworms will be another fruitful research area.

EARTHWORM POPULATIONS Earthworm population size, growth, reproduction, and mortality all have significant consequences for C and N cycling, and estimates of population density and biomass in various habitats remain an important area of research. Although there have been many surveys on earthworm populations from a wide variety of ecosystems, many of these studies suffer from shortfalls such as inadequate sampling frequency and reliance on qualitative sampling techniques. A problem that makes it difficult to compare earthworm biomass estimates from different sites and studies is the inconsistency with which earthworm population biomass is reported. The majority of population studies report biomass as either fresh or dry weight, often with no indication of whether gut contents are included in the reported values; this makes it difficult to compare data from different studies. It would be helpful if all studies expressed earthworm population biomass as ash-free dry weight (AFDW; earthworms ashed at 500°C for 4 hours) per unit area. It is also appropriate to express biomass as grams C per unit area. As part of a continued effort to quantify earthworm biomass, we also need to develop long-term data sets to evaluate temporal variability in earthworm population size. And, as Lavelle et al. suggest (Chapter 8 of this volume), greater attention needs to be paid to the spatial or “patch” dynamics of earthworm populations. Earthworm population growth and reproduction rates are key to determining the amounts of C and N that flow directly through earthworm tissues. Unfortunately, earthworm growth rates are unknown for many species and conditions, and those that are available are often based on laboratory experiments under different optimum environmental conditions. Growth rates need to be determined under field conditions for a range of earthworm species. Similarly, very little is known about rates of earthworm cocoon production in the field. Although cocoon biomass is only a small fraction of the total earthworm biomass, knowledge of the rates of cocoon production can provide information about overall earthworm reproductive rates and can indicate periods of high earthworm activity and the potential rapid turnover of earthworm tissue. Earthworms ingest organic matter with relatively wide C:N ratios and convert it to earthworm tissues of lower C:N ratios (Syers and Springett 1984). In effect, this accelerates the cycling of nutrients in soil, particularly N. There is evidence from field studies of selective feeding by earthworms on organic materials with low C:N ratios, thereby leaving behind a pool of organic material with a higher C:N ratio (Bohlen et al. 1997; Ketterings et al. 1997). Although some earthworm N is excreted in urine and mucus, significant amounts are also returned to the soil in the form of dead tissues. Because earthworm tissue is highly labile, dead earthworms can be an important input of N into the soil. Satchell (1967) reported that over 70% of the N in dead earthworm tissue was mineralized in less than 20 days, and that 60 to 70 kg of N ha−1 was returned to the soil annually. Christensen (1988) also reported high seasonal inputs of N from dead earthworms. Whalen et al. (1999) used 15N-labeled earthworms in a microcosm experiment to trace the movement of N from dead earthworms into the soil and its eventual uptake by plants and soil microbes. Two days after earthworms were incorporated into the soil, 40% of earthworm N was © 2004 by CRC Press LLC

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incorporated into the microbial biomass. Nearly 40 to 50% of the earthworm N moved into soil organic N pools during the first week, but this declined to less than 19% after 16 days. After 16 days, 70% of earthworm N was incorporated into plant biomass. Clearly, the N stored in earthworm tissues enters the labile soil N pool rapidly, where it is available for uptake by microorganisms and plants.

EARTHWORM COMMUNITIES Much of the information gathered at lower levels in the ecological hierarchy can be integrated at the level of the earthworm community. Estimates of earthworm community secondary production can quantify the direct role of earthworms in the C and N cycle (Parmelee and Crossley 1988). Secondary production estimates the C and N fluxes through earthworm tissues as the sum of the production of new earthworm tissue and the losses through earthworm mortality. Secondary production of earthworms during any given time interval is calculated as the growth rate times the change in biomass: P = IGR [(Bf + Bi)/2]t, where P equals production, IGR is the instantaneous growth rate, and Bf and Bi are the final and initial standing stock biomasses, respectively, in grams AFDW m−2 observed over a time interval t measured in days (Romanovsky and Polishchuk 1982). The values for all time intervals can then be summed to calculate annual secondary production. To obtain an estimate of the C or N flux through earthworm biomass, the annual earthworm community secondary production estimate is multiplied by the average percentage C or N in earthworm tissue. Using this method in a no-tillage agroecosystem, we estimated that there was a 40-kg ha−1 year−l flux of N through earthworms (Parmelee and Crossley 1988). When estimates of N excreted in urine and mucus were included, the total N flux through earthworms increased to 63 kg N ha−l year−l, indicating that N flux through the earthworm community was significant for the no-tillage system, and earthworms could process 50% of the N input to soils from plant residues, accounting for 38% of the N uptake by plants. Furthermore, the N flux through earthworm biomass exceeded the losses from the system by denitrification and leaching. Although this example illustrates the direct importance of earthworms in C and N cycles, it may still be inadequate for a general conclusion because it relies on laboratory earthworm growth rate data and Needham’s (1957) possibly inadequate estimates of urine and mucus production. A more recent study assessing two different methods for estimating earthworm secondary production and N flux in agroecosystems showed that N flux through the earthworm community was 18 to 55 kg N ha−1 year−1 (Whalen and Parmallee 2000). The lower estimates were for inorganically fertilized systems, and the higher estimates were for organically fertilized systems. The higher estimates are within the range of the earlier estimates of Parmelee and Crossley (1988) and represent up to 30% of crop N uptake. As discussed, reliable estimates of rates of aboveground earthworm cast production exist, but aboveground casting may constitute only a small proportion of the total cast production. Studies that quantify belowground casting rates for earthworm communities are needed.

EARTHWORMS

IN

ECOSYSTEMS

The effects of earthworms on nutrient cycling processes at the individual, population, and community levels of the hierarchy are integrated fully at the ecosystem level. Much of the research needed at the ecosystem level has been discussed in terms of ecosystem budget models (Figure 9.3 and Figure 9.4). The main question of further interest concerns how earthworms affect the balance between processes that lead to conservation or storage of nutrients vs. losses from the system. Processes that lead to storage of C and N include plant uptake and protection in stable soil aggregates formed by earthworm casts. Another possible mechanism (as discussed by Lavelle et al. in Chapter 8) may involve the types of organic material that earthworms ingest. From our field studies in Ohio,

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we observed that earthworms can selectively ingest organic material that has a low C:N ratio (Bohlen et al. 1997; Ketterings et al. 1997). Consequently, the remaining organic material has a higher C:N ratio, with the potential for slower rates of decay and greater immobilization of nutrients. Loss pathways include a more rapid turnover of microbial biomass and greater production of CO2, denitrification, leaching, and runoff. Studies of the effects of earthworm invasions or introductions into new habitats provide further insights into the effects of earthworms on the net storage or losses of nutrients at the ecosystem scale. In these situations, whole system changes in storage and distribution of nutrients can be compared before and after earthworm invasions or between invaded and uncolonized sites. Studies in temperate forests indicated that earthworm invasions can lead to a decrease in soil C storage (Alban and Berry 1994; Bohlen et al. 2004a,b). Similarly, introduction of earthworms to pastures in New Zealand led to losses of 300 to 1000 kg C ha year−1 (Stout and Goh 1980). Less information is available on the effects of earthworm invasions on the losses or retention of N. However, earthworm invasion of northern forests in the United States apparently did not alter total soil N significantly, indicating that any N that was mobilized by earthworm invasion was retained within these N-limited systems (Bohlen et al. 2004a). Furthermore, earthworms altered the distribution and abundance of fine roots and increased total soil microbial biomass, thereby changing the potential patterns of uptake and turnover of nutrients in the plant and microbial community (Fisk et al. 2004; Groffman et al. 2004). Invasion or intentional introductions of earthworms into new environments will continue to provide insights on the effects of earthworms on nutrient cycling at the ecosystem scale (see Chapter 5, this volume).

EARTHWORMS

IN

LANDSCAPES

The effects of earthworms on nutrient cycling processes at the landscape level have yet to be explored fully, but a much better understanding of the spatial dynamics of earthworm populations is emerging. Lavelle et al. (see Chapter 8, this volume) present an intriguing concept of earthworm patch dynamics in which patches of different species and different size classes of earthworms move through the soil, thereby altering soil structure and nutrient cycling processes and contributing to soil heterogeneity. Poier and Richter (1992) investigated the spatial distribution of earthworms and identified patch sizes of 20 to 50 m and found correlations between earthworm populations and soil carbon and aggregate densities. Hendrix et al. (1992) examined the populations and distribution of earthworms in relation to landscape factors in the southeastern United States. They concluded that earthworm populations were related to soil textural properties, quantity and quality of plant inputs, and standing stocks of soil organic matter. Callaham and Blair (1999) reported that the distribution and relative abundance of native and exotic earthworms in temperate grassland was influenced by land use practices, including burning, mowing, and nutrient additions. Similarly, Bohlen et al. (1995) found that the earthworm community structure in seven small agricultural watersheds was influenced by cropping patterns, geographic location, and tillage. Examination of the presence and absence of earthworms in a northern temperate forest revealed a spatial pattern in which invasive earthworms were present along the edges of the forest near previous agricultural land use but were encountered less commonly in the forest interior, which gave a picture of invasion dynamics at the landscape scale (Bohlen et al. 2004a). More such studies are needed, and the next step should be to link spatial distribution patterns to the effects of earthworms on nutrient cycling properties and feedbacks between soil physical and hydrological properties and earthworm populations. The ultimate goal of landscape-level research should be to identify the major source vs. sink relationships for changes in soil C and N pools. For example, the influence of earthworms on soil erosion and surface runoff may also influence the losses or retention of nutrients at the landscape level (Sharpley et al. 1979). As

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landscape-level studies progress, it may become possible to assess the roles of earthworms in biogeochemical cycles at regional or even global scales.

CONCLUSIONS There is sufficient field research evidence to suggest that earthworms play a major role in soil nutrient cycling processes. Research evaluating the effects of earthworms on biogeochemistry is increasingly interdisciplinary, combining the various aspects of the effects of earthworms on soil biological, chemical, and physical processes. Researchers active in all aspects of earthworm ecology can contribute to a better understanding of the role of earthworms in nutrient cycling processes in both natural and managed ecosystems. However, there continues to be a need for research that crosses traditional boundaries and includes collaborations with scientists from other disciplines. The understanding of the effects of earthworms at the ecosystem and landscape scales still lags behind the many studies at smaller scales. The effects of earthworms on nutrient cycling at the ecosystem level cannot be fully appreciated unless results are scaled up from short-term, smallscale studies and unless data are integrated from different levels of the ecological hierarchy or across temporal and spatial scales.

ACKNOWLEDGMENTS Mike Allen, Scott Subler, and Joann Whalen provided helpful comments on earlier versions of this chapter. Many of the ideas presented were developed through interactions with Mike Allen, Clive Edwards, Dave McCartney, Ben Stinner, and Scott Subler. Clive Edwards reviewed and revised the manuscript. Our research has been supported by grants from the National Science Foundation (DEB-9O20461) and the U.S. Department of Agriculture (NRI-9402520).

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Part V Effects of Earthworms on Soil Physical Properties and Function

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the Effects of 10 Quantifying Earthworms on Soil Aggregation and Porosity Martin J. Shipitalo North Appalachian Experimental Watershed, U.S. Department of Agriculture, Agricultural Research Service, Coshocton, Ohio, U.S.A.

Reneé-Claire Le Bayon Department of Plant Ecology, Neuchâtel University, Neuchâtel, Switzerland

CONTENTS Introduction ....................................................................................................................................183 Ecological Classification of Earthworms ......................................................................................184 Aggregation ....................................................................................................................................184 Ingestion Rates and Properties of Casts.................................................................................184 Remolding of Soil Aggregates by Earthworms .....................................................................185 Measurement of the Stabilization of Aggregates in Casts.....................................................186 Stabilization of Aggregates in Casts: Physical, Chemical, and Biological Processes ....................................................................................................186 Role of Organic Matter...........................................................................................................187 Surface Casting, Soil Erosion, and Nutrient Transport .........................................................188 Porosity and Infiltration .................................................................................................................190 Characterization of Burrow Morphology ...............................................................................190 Effects of Earthworm Burrows on Infiltration .......................................................................190 Effects of Earthworm Burrows on Water Quality..................................................................193 Conclusions ....................................................................................................................................194 References ......................................................................................................................................194

INTRODUCTION The potential for earthworms to improve soil aggregation and porosity and the subsequent effects of these changes in soil structure on plant growth and soil hydrology were perhaps first recognized by Gilbert White in 1777 when he wrote “worms seem to be great promoters of vegetation, which would proceed but lamely without them; by boring, perforating, and loosening the soil, and rendering it pervious to rains and the fibres of plants; by drawing straws and stalks of leaves and twigs into it; and, most of all, by throwing up such infinite numbers of lumps of earth called wormcasts, which, being their excrement, is a fine manure for grain and grass” (White 1789). Before these observations, earthworms were often regarded as pests by farmers and detrimental to crop growth.

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One hundred years later, the first scientific observations on the effects of earthworms on soil structure were conducted by Darwin (1881) and centered mostly on how earthworms contribute to the geologic evolution of soils and landscapes. Like White, he recognized that earthworms promote the growth of vegetation by creating an intimate mixture of organic and mineral matter that aids in water retention and nutrient release and provides a medium suitable for root proliferation. He also recognized that deep-burrowing earthworms affect water movement in the soil and “materially aid in its drainage.” Darwin also postulated that earthworm activity can have negative aspects by contributing to “denudation” (soil erosion) by both wind and water. This was based on observations that casting activity by earthworms can result in the deposition of weakly aggregated material at the soil surface that can flow or be washed or blown downslope. In the ensuing years since these pioneering naturalists published their findings, a number of scientific studies have confirmed their observations, and there are now detailed data on the effects of earthworms on soil aggregation and soil porosity. We are also beginning to understand the chemical and physical processes by which earthworms affect soil structure and the consequences of their activity, both positive and negative, and the interrelationships between soil management and earthworm activity.

ECOLOGICAL CLASSIFICATION OF EARTHWORMS Earthworms affect soil physical properties when they ingest and excrete soil to construct burrows and as part of their feeding activities. Because different earthworm species have different ecological strategies, their effects on soil aggregation and porosity can vary considerably. Most earthworms are placed in one of three ecological groups: epigeic, anecic, or endogeic (Bouché 1977). Epigeic species of earthworms generally forage within accumulations of organic matter and rarely burrow into or ingest much soil. Typical habitats include forest litter or manure piles; thus, they have little direct effect on the structure of mineral soils. For example, Hamilton and Dindal (1989) noted that the epigeic earthworm Eisenia fetida had no effect on aggregation in a sludgeamended soil. Anecic earthworm species normally live in permanent or semipermanent burrows that can extend deep into the soil. They feed primarily on decaying surface organic residues, which they frequently pull into their burrows or mix with excrement to form a midden. The midden promotes further decay of the incorporated organic residues and covers the burrow entrance. Endogeic earthworm species burrow extensively belowground and obtain their nutrition by ingesting a mixture of soil and organic matter. They form extensively branched, subhorizontal networks of burrows in search of food, but most of their activity is in the upper 10 to 15 cm, where organic matter levels are generally highest. Portions of their burrows are often occluded with their casts, and they occasionally cast on the soil surface. These classifications are not absolute because the behavior of many species is intermediate to these groupings and can vary with environmental conditions (Edwards and Bohlen 1996).

AGGREGATION INGESTION RATES

AND

PROPERTIES

OF

CASTS

Although earthworms feed on decaying organic matter and the microorganisms that colonize it, the material ingested by endogeic and anecic species during feeding and burrowing is predominantly mineral matter. This mineral and organic material is mixed thoroughly in their digestive tracts and excreted as casts on the soil surface or belowground, depending on the species of earthworm, location of the food source, and soil bulk density (Binet and Le Bayon 1999). The amount of soil ingested is highly dependent on the size, composition, and activity of the earthworm population and is hard to measure accurately because subsurface activity is difficult © 2004 by CRC Press LLC

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to monitor. Nevertheless, estimated soil ingestion rates for earthworms in temperate regions are usually less than 100 Mg ha−1 year−1 (Tomlin et al. 1995). In tropical areas, such as the Ivory Coast, where climatic conditions are less likely to seasonally inhibit activity, Lavelle et al. (1989) reported a cast production rate of 1200 Mg ha−1 year−1. High cast-production rates such as this are attributable to the fact that geophageous, endogeic earthworms can ingest 5 to 30 times their body weight per day (Lavelle 1988). According to Lee (1985), earthworms can process up to 25% of the Ah horizon (organo-mineral soil layer enriched in humified organic matter) in 1 year and thus can be important aggregate-forming agents through the production of casts in the soil and on the surface. In a laboratory study, Ziegler and Zech (1992) showed that E. fetida could bind up to two thirds of the beech litter and unstructured artificial soil into 200- to 2000- µm diameter aggregates in 446 days. Earthworm casts, deposited on the burrow walls, within the burrow, or on the soil surface (Brown et al. 2000), usually contain more clay and less sand than the surrounding soil because of selective ingestion, with this effect more prominent with endogeic species, which tend to be smaller than anecic species of earthworms. This concentration of fine particles in earthworm casts may need to be taken into account when using methods such as dispersible clay or turbidity to compare the stability of casts with uningested soil. Moreover, the relative differences in texture between casts and uningested soil are probably dependent on the coarseness of the parent soil. For example, Shipitalo and Protz (1988) noted that casts of Lumbricus rubellus, an epigeic/endogeic species, contained less sand than those of L. terrestris (anecic species), and both had less sand than the uningested soil (18% sand). Schrader and Zhang (1997) reported only small differences in the texture between the casts of L. terrestris and Aporrectodea caliginosa (endogeic) and the parent soils with initial sand contents less than 4%. Likewise, the amount of organic matter incorporated into casts is dependent on whether the earthworms are actively burrowing or feeding and the food source. Shipitalo et al. (1988) reported that food ingestion rates and organic carbon contents of casts were higher for more palatable food sources, as reflected in earthworm weight gains, and that casts of L. rubellus were generally higher in organic carbon than those of L. terrestris. In a study by Schrader and Zhang (1997), however, L. terrestris casts were enriched in organic carbon to a greater extent than A. caliginosa casts. Specific organic compounds such as reducing sugars, amino sugars, phenolic materials (Mora et al. 2003), and carbohydrates (Scullion and Malik 2000) can also be concentrated by earthworms in their casts. Earthworm casts also usually have higher bulk density than the uningested soil (Edwards and Bohlen 1996; Görres et al. 2001), unless the soil is already compacted (Joschko et al. 1989), and are higher in pH, contain more available nutrients, and have higher levels of microbial activity.

REMOLDING

OF

SOIL AGGREGATES

BY

EARTHWORMS

The muscular contractions of the earthworm crop and gizzard, the peristalsis of the gut wall, and contractions of the body wall create a great range of pressures that mechanically disrupt soil microaggregates during passage through the digestive tract. The mean pressure applied to soil by Aporrectodea rosea was estimated as 259 Pa (McKenzie and Dexter 1987). For L. terrestris, Newell (1950) reported that the average coelom pressure was 1.6 kPa in segment 28 and 0.8 kPa near the tail region. Such pressures, concomitant to the addition of large amounts of watery mucus (Barois et al. 1993), can lead to the mobilization of clay (Marinissen et al. 1996) and the disruption of existing interparticle water and cation bridges in the aggregates (Shipitalo and Protz 1988, 1989). Conversely, soil remolding also brings clay minerals into close association with newly formed or released bonding agents originating from the ingested organic matter (Shipitalo and Protz 1989). Consequently, the soil fabric is reorganized in the posterior intestine of earthworms (Barois et al. 1993), with resistant organic fragments becoming the foci for new microaggregates (Shipitalo and Protz 1989). Earthworm gut transit time probably also affects the degree of microaggregate disruption. Reportedly it takes 2 to 24 hours for soil to pass through the digestive tract of lumbricid earthworms (Barley 1959; Piearce 1972; Bolton and Phillipson 1976). © 2004 by CRC Press LLC

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To gain further insights into the physical processes occurring in the earthworm gut, a number of researchers have compared the stability of artificial casts with those made by earthworms. In some cases, the artificial casts were less stable than natural casts, which may be related to applying forces to the soil greater than those normally encountered within the earthworm gut (Zhang and Schrader 1993; Hindell et al. 1997a; Schrader and Zhang 1997). This may be a particular concern when artificial casts are made by forcing soil material through a syringe (Hindell et al. 1997a). Zhang and Schrader (1993) also suggested that earthworm casts were more readily stabilized than artificial casts because the organic and mineral fractions were mixed more intimately within the earthworms than when the soil was artificially remolded. On the other hand, Marinissen and Dexter (1990) reported that fresh casts produced by A. caliginosa were up to two times more dispersible than artificial casts made by extruding the same soil through a syringe. In their work, they suggested that the results were probably attributable to less intensive remolding in the artificial casts than in the earthworm casts. Despite the difficulties in replicating the physical forces encountered by soil material during passage through earthworms, Hindell et al. (1997a) pointed out that artificial casts can be a useful model against which changes in soil structure that result from earthworm activity can be tested.

MEASUREMENT

OF THE

STABILIZATION

OF

AGGREGATES

IN

CASTS

Prior to the mid-1980s, most studies suggested that freshly excreted earthworm casts were immediately more stable than uningested soil (Hopp and Hopkins 1946; Dutt 1948; Swaby 1950; Teotia et al. 1950; Parle 1963; Lal and DeVleeschauwer 1982; Lal and Akinremi 1983). These results, however, were mainly attributable to the fact that the samples were dried before analysis. Most recent studies indicated that fresh, moist casts are less water-stable than uningested soil because of the intense remolding that occurs during passage through earthworms (Shipitalo and Protz 1988; Marinissen and Dexter 1990; Barois et al. 1993; Schrader and Zhang 1997; Decaëns et al. 2001). As casts age, they are stabilized by a combination of physical, chemical, and biological processes, which explains why some casts can persist at the soil surface for more than a year when protected from raindrop impact and trampling by animals (Decaëns 2000). Although several studies have concluded that aging or drying fresh casts reduced their dispersibility (Shipitalo and Protz 1988; Marinissen and Dexter 1990), this conclusion has not been universal (Haynes and Fraser 1998). To understand these seemingly contradictory findings, the methodology used to measure the stability of earthworm casts (e.g., wet or dry sieving, clay dispersion, turbidimetric analysis), as well as the effects of other treatments to which the casts have been subjected (wetting and drying cycles, simulated rainfalls, sterilization and chemical treatments), must be taken into account. Water-stable aggregation is an index for aggregate stability under wet conditions, whereas tensile strength (determined by a crush test) is an index for aggregate stability under dry conditions. The tensile strength of aggregates is influenced by their water and clay contents (Gill 1959) and decreases with increased porosity (Dexter et al. 1984). In studies in which tensile strength was measured (Schrader and Zhang 1997; Garvin et al. 2001), earthworm casts were significantly stronger than natural aggregates. The tensile strength of casts, however, appears to be species dependent because Flegel et al. (1998) observed a lower tensile strength for the casts of Dendrobaena octaedra compared with those of L. terrestris. Similarly, Schrader and Zhang (1997) noted that water-stable aggregation was significantly higher in casts of L. terrestris than in casts of A. caliginosa.

STABILIZATION OF AGGREGATES BIOLOGICAL PROCESSES

IN

CASTS: PHYSICAL, CHEMICAL,

AND

One of the physical processes thought to contribute to the stabilization of casts with age is thixotropic hardening (Shipitalo and Protz 1989). Thixotropic or age hardening is described by Utomo and Dexter (1981) as a rearrangement of particles and water films and a restoration of

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edge-to-edge contacts between clay domains with time but without water loss. These processes are reversible, and organic matter reportedly slows or has no effect on thixotropic hardening (Blake and Gilman 1970; Molope et al. 1985). As Marinissen et al. (1996) pointed out, however, thixotropic process normally occur within hours, so they cannot be solely responsible for the stabilization that occurs in casts over much longer time frames. In addition to the physical processes, chemical processes can contribute to stabilization of earthworm casts. Earthworms are known to secrete amorphous calcium carbonate (C.A. Edwards and Bohlen 1996), a possible binding agent (Tisdall and Oades 1982). In addition, by using selective chemical pretreatments, Shipitalo and Protz (1989) gathered indirect evidence that calcium and, to a lesser extent, magnesium are involved in the clay-polyvalent cation-organic matter linkages that stabilize soil microaggregates within casts. In fact, the change in CaCO3 content in casts compared with uningested soil (Zhang and Schrader 1993) might be useful as an index of the capacity of various earthworm species to bond soil particles and reform new stable aggregates. When they compared casts from various soils, Schrader and Zhang (1997) found positive correlations between tensile strength and the clay and CaCO3 contents of the soil. These parameters, however, correlated negatively to water-stable aggregation. Hindell et al. (1997a) hypothesized that the greater dispersibility of artificial casts compared with natural casts was because of a greater loss of calcium ions from the artificial casts, which reduced coagulation of clay particles. The production of microbial polysaccharides in casts (Chapman and Lynch 1985; Emerson et al. 1986; Robertson et al. 1991) and polysaccharides that are added to casts in the mucus secreted by earthworms and by mucilages produced by microorganisms living in their digestive tract (Barois and Lavelle 1986; Kristufek et al. 1992) may also affect aggregate stability. However, the role these polysaccharides play in cast aggregation is still uncertain. Hindell et al. (1997a) suggested that the secretion of soluble carbohydrates in the earthworm gut initially facilitates the dispersion of clay. On the other hand, Swaby (1950) showed that, as populations of intestinal bacteria increased, the production of gums and glues increased, and cast stability increased. Altemüller and Joschko (1992) also showed that carbohydrates produced by bacteria can serve as cementing agents, and Flegel et al. (1998) reported a significant correlation between phosphomonoesterase activity and the water-stable aggregation of earthworm casts. Other research, however, has demonstrated that microbial activity is not necessary for casts to stabilize (Marinissen and Dexter 1990; Marinissen et al. 1996; Haynes and Fraser 1998). In fact, microbial activity may be reduced in some casts because of limited gaseous exchange caused by their high bulk density (Blanchart et al. 1993). In some instances, poor correlations of aggregate stability in casts with the size of the microbial populations and polysaccharide content are probably because the arrangement and location of these constituents within casts is more important than the absolute quantities (Shipitalo and Protz 1989; Haynes and Fraser 1998). Microbial activity can also physically stabilize earthworm casts. Fungal hyphae have been reported to stabilize soil aggregates and casts (Tisdall and Oades 1982; Molope et al. 1987; Marinissen and Dexter 1990; Lee and Foster 1991; Tisdall 1991; Tisdall et al. 1997; Kabir and Koide 2002). Using scanning electron microscopy, Haynes and Fraser (1998) observed that fungal hyphae emanating from within casts enmeshed aggregates. Tiunov and Scheu (2000) found that most fungi can survive passage through L. terrestris, but the dominance structure of the fungal community changes with time and remains cast specific for up to 100 days.

ROLE

OF

ORGANIC MATTER

In most studies, the amounts and source of organic matter incorporated into the soil by earthworms has been shown to have a significant affect on aggregate stability within casts; thus, a positive correlation between cast organic carbon content and cast stability is frequently reported (Shipitalo and Protz 1988; Zhang and Schrader 1993; Schrader and Zhang 1997; Flegel et al. 1998). Earthworms play a large role in litter comminution and its repartitioning into the smaller aggregate size © 2004 by CRC Press LLC

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fractions. For example, the organic matter in the fraction larger than 2000 µm decreased from 97 to 27% in the presence of the E. fetida, leading to a predominant 630- to 2000-µm fraction after 446 days of incubation (Ziegler and Zech 1992). This litter-derived organic matter can serve as a bonding agent or promote microbial activity that leads to the production of bonding agents (Guggenberger et al. 1996). Beare et al. (1994) suggested that the incorporation of organic matter promotes the formation of stable microaggregates within macroaggregates. Kladivko et al. (1986) found that, after drying, aggregate stability was determined mainly by the type of plant remains, although the effect of earthworms was still significant. Once incorporated into casts and if not subject to further disturbances, the organic matter can persist for many years (McInerney et al. 2001), with organic carbon persistence and dynamics in earthworm casts dependent on complex interactions among soil texture, temperature, and wetting cycles (McInerney and Bolger 2000). However, in some instances, cast stabilization has been observed in the absence of a source of organic residue (Marinissen and Dexter 1990; Marinissen et al. 1996). Similarly, although Haynes and Fraser (1998) observed fragments of decomposing organic material adhering to aggregate surfaces, they noted stabilization in the absence of a source of organic residue. It is likely that the type and extent of bonding will depend on properties of the soil materials and on the quality and the quantity of the ingested organic debris. Thus, several physical, chemical, and biological mechanisms probably contribute to the stabilization of aggregates within casts, and their relative importance can vary under different conditions and with different earthworm species. The continued stability of these aggregates can be influenced by wetting and drying cycles and whether other soil organisms disrupt them. Successive wetting and drying cycles contribute to the stability of natural aggregates by creating bonds of different nature between the contact points of soil particles over time (Dexter et al. 1988). In casts, Marinissen and Dexter (1990) assumed that the effects of drying-rewetting would be more persistent with time than the effects of fungal hyphae. In newly remolded aggregates, Utomo and Dexter (1982) showed that wetting and drying increased the percentage of water-stable aggregates two- to fourfold. Nevertheless, Hindell et al. (1997b) reported opposite results for initially air-dried casts and uningested soil. Air-dried samples slaked severely when immersed in water, and they speculated that surface casts are the most subject to slaking following sudden rain or irrigation. In a laboratory microcosm study, Shaw and Pawluk (1986) noted that soil structure development was maximized when anecic and endogeic earthworm species were allowed to interact. In a field study in a tropical region, however, Blanchart et al. (1997) noted that small eudrilid earthworms accelerated the destruction of aggregates created by larger earthworms. This prevented accumulation of large casts at the soil surface and, in some cases, led to the formation of a compact and impermeable layer and to negative effects on plant growth (Blanchart et al. 1999; Chauvel et al. 1999). In temperate region soils, Ge et al. (2001) noted that casts near the soil surface degraded rapidly unless protected by a mulch cover, and Shuster et al. (2000) noted that foraging and midden building by anecic earthworm species reduced residue cover and exposed more soil and casts to raindrop impact.

SURFACE CASTING, SOIL EROSION,

AND

NUTRIENT TRANSPORT

Although Darwin (1881) speculated that earthworms contribute to soil erosion, it is now known that the net effect of their activity on soil losses depends on a number of interacting factors. By burrowing into the soil and creating macropores, earthworms can increase infiltration rates 2- to 15-fold, which should lead to a reduction in runoff (Ehlers 1975; Joschko et al. 1989; W.M. Edwards et al. 1990; Kladivko and Timmenga 1990; Bouché and Al-Addan 1997; Willoughby et al. 1997). This, in turn, should contribute to a reduction in soil loss. In addition, earthworms can increase surface roughness by casting on the soil surface, and their burrowing activity can disrupt soil crusts, which should further increase infiltration and reduce runoff (Kladivko et al. 1986).

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However, because freshly deposited casts are of low stability, they are subject to dispersion if deposited on the soil surface and not protected from raindrop impacts (Van Hoof 1983). This detached material is then subject to transport, which can contribute to soil losses and the loss of sediment-associated nitrogen (Parle 1963; Binet and Tréhen 1992; Buck et al. 1999), phosphorus (Graff 1970; Sharpley and Syers 1976; Ganeshamurthy et al. 1998), and potassium (Tiwari et al. 1989; Ganeshamurthy et al. 1998). Hence, surface casts have been shown to be a source of sediment and particulate and dissolved P in surface runoff from a permanent pasture (Sharpley and Syers 1976; Sharpley et al. 1979). In temperate regions, surface-casting activity can increase sediment losses from fields used to grow row crops, particularly in maize fields, where compacted soils can contribute to increased surface casting activity and increased surface runoff (Binet and Le Bayon 1999; Le Bayon and Binet 1999). In a tropical forest, Nooren et al. (1995) estimated that 0.12 kg m−2 year−1 of organic suspended sediment originated from the disintegration of earthworm casts because of the combined effect of rain splash and surface runoff. Field plot research conducted by Le Bayon and Binet (2001) using simulated rainfall highlighted the complexity of the dynamic interrelationships among earthworm surface casting activity, runoff, and erosion in temperate region agroecosystems. Although these experiments confirmed some of the findings of static measurements, such as cast water stability and tensile strength, they also indicated that a number of other factors must be taken into account to understand the potential contributions of earthworm casts to soil erosion and nutrient transport. They found that recently deposited earthworm casts were more susceptible to dispersion by raindrops and transport in surface runoff than older casts, probably because of enhanced stability with time caused by the mechanisms discussed earlier. The resistance of casts to dispersion with time also appeared to be species dependent (anecic vs. endogeic). Casts were enriched in particulate phosphorus compared with uningested soil. Nevertheless, soil and particulate phosphorus losses in runoff were less from plots with earthworm casts than from control plots without surface casts. This was attributed to a reduction in surface runoff because of enhanced infiltration from earthworm burrowing activity and the casts acting as a physical barrier to runoff by increasing surface roughness, thereby ponding water and further delaying the onset of runoff (Figure 10.1). By following the fate of individual earthworm casts, they also determined that cast morphology (base/height ratio), weight, bulk density, and abundance could affect cast susceptibility to dispersion and transport, factors that would not be evident based on water stability or tensile strength measurements. Moreover, factors external to earthworm activity, such as rainfall intensity and slope, can also affect the fate of surface casts (Figure 10.1). rainfall impact physical processes raindrops intensity physical brake feedback breakdown dispersion deposition/suspension of particles

runoff

hypodermic infiltration

percolation 4.5% slope

FIGURE 10.1 Interrelationships between earthworm surface casting activity, rainfall, surface runoff, infiltration, and soil erosion. (From Le Bayon and Binet, 2001.)

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POROSITY AND INFILTRATION CHARACTERIZATION

OF

BURROW MORPHOLOGY

Because they burrow extensively into mineral soil, endogeic and anecic earthworms can substantially alter soil porosity. Estimates of the number of burrows in temperate region soils range as high as 100 to 800 m−2 (Lavelle 1988). Although these burrows usually account for a small fraction of the soil volume, because of their continuity, stability, and relatively large size compared with pores formed by most other mechanisms, these macropores have the potential to greatly affect the movement of air, water, and solutes. Moreover, unlike cracks, earthworm burrows tend to remain open and continue to function as preferential flow paths under wet soil conditions (Friend and Chan 1995) and are less likely to be closed by vehicle-induction compaction than other soil macropores (Alakukku et al. 2002). Quantifying burrow numbers and morphology is difficult, however, and a number of techniques have been used to accomplish this task. A commonly used technique is to count the number of burrows open at various depths in the soil and measure their diameters. These counts can be performed manually or by taking photographs and using image analysis techniques (Shipitalo and Protz 1987; Edwards et al. 1988a, 1988b). If accurate information on burrow continuity is to be obtained, the observations must be made at relatively narrowly spaced vertical intervals. This can be accomplished by serially sectioning the soil either in situ or with impregnated soil thin sections (Ehlers 1975; Ligthart et al. 1993; McKenzie and Dexter 1993; Schrader 1993; Hirth et al. 1996; Ligthart 1997; Pitkänen and Nuutinen 1997; Sveistrup et al. 1997; Springett and Gray 1998). These techniques, however, are laborious and are often only partially successful, particularly in the soil layers near the surface, because of loose soil aggregates and interference by plant roots (Ligthart et al. 1993). For example, McKenzie and Dexter (1993) were successful only 20% of the time when they used a grid coordinate system and manual excavation to measure earthworm burrow geometry. A modification of the excavation technique that reduces some of the difficulties encountered in trying to track the continuity of individual earthworm burrows and that can result in more accurate characterization of their morphology is through the use of replicas of burrows made in situ (Figure 10.2). These replicas can be made using materials such as molten lead (Teotia et al. 1950), plaster (Bouma et al. 1982; Wang et al. 1994), wax (Smettem 1986), or fiberglass resin (Shipitalo and Butt 1999; Shipitalo and Gibbs 2000). Like excavation, this technique is not always successful because of an inability to fill the burrows completely with the impregnating media. Additional drawbacks are that the technique works best only on burrows of relatively large diameter (i.e., >5 mm), and removal of the replicas is a tedious operation. More recently, x-ray computed tomography has been used to characterize earthworm burrow morphology (Golabi et al. 1995; Daniel et al. 1997; Perret et al. 1997; Capowiez et al. 1998; Jegou et al. 1998; Langmaack et al. 1999). A major advantage of this technique, compared with procedures involving excavation, is that the soil is not disturbed during the analysis. Consequently, the dynamics of burrow construction can be investigated. Besides the limited availability and high expense of this equipment, a major disadvantage is that the resolution of the current generation of equipment is such that only the morphology of large diameter burrows can be accurately assessed. Another concern is that the samples are usually obtained by incubating earthworms in columns of soil that will fit within the instruments rather than examining burrows formed by earthworms in the field under natural conditions. As Springett and Gray (1998) noted when they manually excavated burrows, there can be major differences between those formed in laboratory columns and those formed in the field because of restriction of the available space in columns.

EFFECTS

OF

EARTHWORM BURROWS

ON INFILTRATION

A variety of field and laboratory techniques have been used to determine the effects of earthworm burrows on infiltration. These techniques include dye and tracer studies, studies in which movement © 2004 by CRC Press LLC

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FIGURE 10.2 Example of a fiberglass replica of an L. terrestris burrow photographed against the backdrop of a monolith of the soil from which it was obtained. Approximate length 1 m.

of water coming from individual earthworm burrows is monitored, and studies in which the overall effect of earthworm presence on infiltration rate of the bulk soil is investigated. These studies have shown that burrows made by anecic and endogeic species of earthworms can effectively conduct water (Zachmann et al. 1987; Joschko et al. 1992; Trojan and Linden 1992; Shipitalo et al. 2000). © 2004 by CRC Press LLC

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Because most of their activity is confined to surface soil horizons, however, endogeic earthworms probably do not directly influence water movement deep into the profile (Ela et al. 1992). The fact that portions of their burrows are often occluded with casts probably further limits their effectiveness in water transport. Thus, most research has centered on the effects of anecic earthworm species on infiltration and on L. terrestris in particular. Burrows created by L. terrestris are normally single, nearly vertical channels, up to 12 mm in diameter and 2.4 m deep (Edwards and Bohlen 1996). These burrows can have several entrances directly underneath the midden, but these usually coalesce into a single channel within the upper few centimeters of soil. Nevertheless, Shipitalo and Butt (1999) and Shipitalo and Gibbs (2000) found that about 5% of the L. terrestris burrows they investigated were Y shaped, with the two channels intersecting as deep as 69 cm below the soil surface. One method that has been used to investigate water movement through natural L. terrestris burrows in the field involves placing surface-vented collection bottles beneath individual burrows 30 to 50 cm below the soil surface (Edwards et al. 1989; Shipitalo et al. 1994). Because the portions of the burrows above the samplers are not disturbed, this technique can be used to investigate infiltration into burrows with intact middens. Although middens would seem to inhibit entry of water, these studies indicated that L. terrestris burrows could transmit substantial amounts of water. In fact, Darwin (1881) did not consider middens to be a barrier to water movement. These studies also indicated that the fraction of rainfall collected increased with rainfall intensity. With an intense rainfall on a dry soil surface, Edwards et al. (1989) estimated that the monitored burrows collected 10% of the rainfall and an average of 13 times more water than expected based on the diameter of the burrows at the soil surface. Problems with the bottle sample technique include concern that interception of flow with the samplers may allow more water to move through the burrows than would naturally occur because infiltration characteristics of the soil surrounding the lower reaches of the burrow might limit infiltration (Lee and Foster 1991; Golabi et al. 1995). In addition, after initially high rates of infiltration, soil air pressure might restrict further water entry under field conditions (Linden and Dixon 1976; Edwards et al. 1979; Baird 1997), a consequence precluded by the sampler design. These concerns appear to be unfounded in most soils under most conditions because procedures in which infiltration has been measured by introducing water directly into the openings of individual L. terrestris burrows at the soil surface have demonstrated average infiltration rates in the range of several hundred milliliters per minute, well in excess of the amounts measured using the bottle sampler technique, for soils in Germany (Ehlers 1975), the Netherlands (Bouma et al. 1982), Wisconsin (Wang et al. 1994), the U.K., and Ohio (Shipitalo and Butt 1999). Moreover, the study by Shipitalo and Butt (1999) indicted that the presence of live L. terrestris in the burrows did not have detectable effects on infiltration. This addressed the concern of Lee and Foster (1991) that anecic earthworms might tightly seal their burrows with their bodies and limit infiltration. In fact, Shipitalo and Butt (1999) speculated that occupied burrows might be more effective in transmitting water than are abandoned burrows because they are more likely to maintain near-surface continuity. The effects of earthworms on infiltration have also been investigated in the laboratory using intact or repacked soil columns with resident or inoculated earthworms. Although these studies have provided insight into mechanisms affecting infiltration, one concern, particularly with repacked soil columns inoculated with earthworms, is that the burrows formed are not representative of those constructed under more natural conditions (Springett and Gray 1998). Similarly, studies in which artificially constructed macropores are used to investigate water movement through earthworm burrows can have significant limitations and must be interpreted with caution (Joschko et al. 1989; Roth and Joschko 1991; Ela et al. 1992; Li and Ghodrati 1995). In this case, an additional limitation is that the artificial burrows lack the organic matter-rich lining or drilosphere, composed of earthworm excrement, mucus secretions, and plant remains, that can affect water and chemical movement (Stehouwer et al. 1993, 1994). © 2004 by CRC Press LLC

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Models have also been used to determine the impact of earthworm burrows and other macropores on inÞltration (Ehlers 1975; W.M. Edwards et al. 1979; Smettem and Collis-George 1985; Smettem 1986; Wang et al. 1994, Li and Ghodrati 1995). Although this approach is useful in investigating the factors affecting inÞltration in earthworm burrows, collection of the burrow data needed to obtain input parameters (i.e., burrow depth, length, diameter, volume) for the models is difÞcult. Moreover, although most of these models indicate that the aforementioned parameters should affect inÞltration capacity, Shipitalo and Butt (1999) were unable to detect any signiÞcant correlations between these geometrical properties and inÞltration rates through L. terrestris burrows. In addition, not all earthworm burrows conduct water (Ela et al. 1992; Trojan and Linden 1992; Shipitalo et al. 2000). In fact, Bouma et al. (1982) stated that theoretical models are unlikely to predict inÞltration in earthworm burrows successfully given the complexity and variability of the morphological factors affecting hydraulic performance.

EFFECTS

OF

EARTHWORM BURROWS

ON

WATER QUALITY

Increased inÞltration attributable to earthworm activity in soils is generally regarded as beneÞcial because it can reduce surface runoff, thereby increasing plant-available water and reducing the potential for overland transport of sediment, nutrients, and agrochemicals (Shipitalo et al. 2000). Earthworm burrows can also increase the efÞciency of subsurface drainage systems (Urbánek and ù 1992) and may help restore the inÞltration capacity of clogged septic system leach beds Dolezal (Jones et al. 1993). However, this increased inÞltration can increase the quantity and rate of solute movement through the soil proÞle. This is of particular concern with L. terrestris burrows because they are often deep enough to penetrate the entire soil proÞle (Figure 10.2). Thus, solutes transported through these burrows can rapidly bypass the upper reaches of the proÞle, where uptake is most likely to occur and biological activity and the potential for degradation are greatest. In addition, because the velocity at which water moves through these macropores is much greater than when the entire soil matrix is involved in the ßow process, the amount of soil a solute encounters and its contact time with the soil are reduced. It is difÞcult, however, to quantify the effects of earthworm burrows on chemical transport because, as just discussed, it is difÞcult to measure their effects on inÞltration. An additional complication is that the burrow linings can serve as both a source and a sink for various solutes. For example, Edwards et al. (1992b) found that when nitrate-free water was poured into L. terrestris burrows and immediately collected 45 cm below the soil surface, it contained as much as 40 mg of nitrate-nitrogen per liter. They speculated that the nitrate originated from the decomposition of the organic matter lining the burrows. This contention is supported by the work of Parkin and Berry (1999), in which higher microbial populations as well as higher nitriÞcation and denitriÞcation rates were noted in L. terrestris burrow linings than in bulk soil. Edwards et al. (1992b) also noted a Þvefold reduction in the concentration of alachlor and a ninefold reduction in the concentration of atrazine when solutions of these two herbicides were poured into burrows and collected at the bottom. When these solutions were poured through man-made artiÞcial burrows, the concentrations were reduced by only about half. In this instance, the decreased herbicide concentrations were attributed to sorption of the herbicides by the organic matter-rich linings of the burrows, a contention supported by the work of Stehouwer et al. (1993, 1994). For this reason, chemical tracers are often used to investigate solute movement in earthworm burrows. The results of a number of Þeld and laboratory chemical transport and tracer studies suggested that earthworm burrows can increase overall water movement through the soil and contribute to a slight increase in the leaching of surface-applied agrochemicals, particularly when intense storms occur shortly after application on residue-covered no-till soils (Germann et al. 1984; Bicki and Guo 1991; Edwards et al. 1992a; Trojan and Linden 1992). The potential for this to occur is greatly reduced with time (Edwards et al. 1993, 1997; Logsdon 1995) and low intensity intervening rainfalls

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(Shipitalo et al. 1990; Propes et al. 1993; Golabi et al. 1995). Ingestion of herbicide-coated residues by earthworms can also reduce leaching losses (Farenhorst et al., 2000a,b). Other potential adverse effects of earthworm burrows on water quality and water utilization include fostering nonuniform distribution of water during furrow irrigation and loss of water through unlined irrigation ditches. Possible remedies for these concerns include compacting the ditches and removing vegetation to reduce earthworm burrowing (Kemper et al. 1988) and adding ammonia at low rates to the irrigation water to repel earthworms (Trout and Johnson 1989). Burrowing by earthworms can also contribute to leakage of earthen-lined manure storage lagoons by increasing the hydraulic conductivity of the berms (McCurdy and McSweeney 1993). Presumably, the procedures used to reduce water movement through unlined irrigation ditches would help alleviate this concern. Earthworm burrows can also affect the movement of the constituents in animal wastes applied to soils. Joergensen et al. (1998) noted greater movement of fecal indicator organisms in cattle slurry applied to grassland than to plowed soil, which they attributed to greater numbers of L. terrestris burrows in the grassland. Similarly, the results of a study by Shipitalo and Gibbs (2000) suggested that L. terrestris burrows, close to subsurface drains, can contribute to rapid movement of injected animal wastes off-site. In this instance, rapid movement of the tracer to the buried drains was limited to burrows 0.5 m to either side of the drain. This suggests that disrupting the burrows in this region prior to slurry application or avoiding application in this region might reduce this concern.

CONCLUSIONS Despite the large number of studies that have been conducted on the effects of earthworm activity on soil structure, a number of important gaps in knowledge remain. Factors contributing to this problem include a lack of appropriate techniques to assess aggregation and porosity and ofteninappropriate extrapolation of laboratory findings to the field. In general, earthworm activity improves soil aggregation, but their casting activity initially destabilizes the soil. Although laboratory studies can elucidate some of the factors affecting the improvement of aggregation with time, only when the fate of earthworm casts is investigated in the field or in microcosms that reflect the complexity of natural systems and managed agroecosytems will a more complete understanding be obtained. One approach that shows promise is to manipulate earthworm populations in long-term field plots to assess the effects of different population levels on soil structural dynamics (Bohlen et al. 1995). Similarly, quantification of earthworm burrow morphology and the effects of earthworm burrows on water movement and water quality are hampered by limitations in methodology. Earthworm burrows, particularly those formed by anecic species of earthworms, can function as preferential flow pathways. Although enhanced infiltration is normally desirable, in rare instances it can result in increased chemical movement through the soil or inappropriate distribution of irrigation water and liquid animal wastes. Although there are some management options available to reduce this concern, the dynamics of water movement through earthworm burrows at the field scale are still poorly understood. Once problems with limited resolution are overcome, x-ray computed tomography holds considerable promise for increasing knowledge of the mechanisms affecting water and solute transport in earthworm burrows.

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Roth, C.H. and M. Joschko. 1991. A note on the reduction of runoff from crusted soils by earthworm burrows and artificial channels, Z. Pflanzenernahr. Bodenk., 154:101–105. Schrader, S. 1993. Semi-automatic image analysis of earthworm activity in 2D soil sections, Geoderma, 56:257–264. Schrader, S. and H. Zhang. 1997. Earthworm casting: stabilization or destabilization of soil structure? Soil Biol. Biochem., 29:469–475. Scullion, J. and A. Malik. 2000. Earthworm activity affecting organic matter, aggregation and microbial activity in soils restored after opencast mining for coal, Soil Biol. Biochem., 32:119–126. Sharpley, A.N. and J.K. Syers. 1976. Potential role of earthworm casts for the phosphorus enrichment of runoff waters, Soil Biol. Biochem., 8:341–346. Sharpley, A.N., J.K. Syers, and J.A. Springett. 1979. Effect of surface-casting earthworms on the transport of phosphorus and nitrogen in surface runoff from pasture, Soil Biol. Biochem., 11:459–462. Shaw, C. and S. Pawluk. 1986. The development of soil structure by Octolasion tyrtaeum, Aporrectodea turgida and Lumbricus terrestris in parent materials belonging to different textural classes, Pedobiologia, 29:327–339. Shipitalo, M.J. and K.R. Butt. 1999. Occupancy and geometrical properties of Lumbricus terrestris L. burrows affecting infiltration, Pedobiologia, 43:782–794. Shipitalo, M.J., W.A. Dick, and W.M. Edwards. 2000. Conservation tillage and macropore factors that affect water movement and the fate of chemicals, Soil Tillage Res., 53:167–183. Shipitalo, M.J., W.M. Edwards, W.A. Dick, and L.B. Owens. 1990. Initial storm effects on macropore transport of surface-applied chemicals in no-till soil, Soil Sci. Soc. Am. J., 54:1530–1536. Shipitalo, M.J., W.M. Edwards, and C.E. Redmond. 1994. Comparison of water movement and quality in earthworm burrows and pan lysimeters, J. Environ. Qual., 23:1345–1351. Shipitalo, M.J. and Gibbs. 2000. Potential of earthworm burrows to transmit injected animal wastes to tile drains, Soil Sci. Soc. Am. J., 64:2103–2109. Shipitalo, M.J. and R. Protz. 1987. Comparison of morphology and porosity of a soil under conventional and zero tillage, Can. J. Soil Sci., 67:445–456. Shipitalo, M.J. and R. Protz. 1988. Factors influencing the dispersibility of clay in worm casts, Soil Sci. Soc. Am. J., 52:764–769. Shipitalo, M.J. and R. Protz. 1989. Chemistry and micromorphology of aggregation in earthworm casts, Geoderma, 45:357–374. Shipitalo, M.J., R. Protz, and A.D. Tomlin. 1988. Effect of diet on the feeding and casting activity of Lumbricus terrestris and L. rubellus in laboratory culture, Soil Biol. Biochem., 20:233–237. Shuster, W.D., S. Subler, and E.L. McCoy. 2000. Foraging by deep-burrow earthworms degrades surface soil structure of a fluventic Hapludoll in Ohio, Soil Tillage Res., 54:179–189. Smettem, K.R.J. 1986. Analysis of water flow from cylindrical macropores, Soil Sci. Soc. Am. J., 1139–1142. Smettem, K.R.J. and N. Collis-George. 1985. The influence of cylindrical macropores on steady-state infiltration in a soil under pasture, J. Hydrol., 52:107–114. Springett, J.A. and R.A.J. Gray. 1998. Burrowing behaviour of the New Zealand indigenous earthworm Octochaetus multiporus (Megascolecidae: Oligochaeta), N.Z. J. Ecol., 22:95–97. Stehouwer, R.C., W.A. Dick, and S.J. Traina. 1993. Characteristics of earthworm burrow lining affecting atrazine sorption, J. Environ. Qual., 22:181–185. Stehouwer, R.C., W.A. Dick, and S.J. Traina. 1994. Sorption and retention of herbicides in vertically oriented earthworm and artificial burrows, J. Environ. Qual., 23:286–292. Sveistrup, T.E., T.K. Haraldsen, and F. Engelstad. 1997. Earthworm channels in cultivated clayey and loamy Norwegian soils, Soil Tillage Res., 43:251–262. Swaby, R.J. 1950. The influence of earthworms on soil aggregation, J. Soil Sci., 1:195–197. Teotia, S.P., F.L. Duley, and T.M. McCalla. 1950. Effect of Stubble Mulching on Number and Activity of Earthworms, Nebraska Agric. Exp. Station Bull. No. 165, University of Nebraska, Lincoln, NE. Tisdall, J.M. 1991. Fungal hyphae and structural stability of soil, Aust. J. Soil Res., 29:729–743. Tisdall, J.M. and J.M. Oades. 1982. Organic matter and water stable aggregates in soils, J. Soil Sci., 33:141–163. Tisdall, J.M., S.E. Smith, and P. Rengasamy. 1997. Aggregation of soil by fungal hyphae, Aust. J. Soil Res., 35:55–60.

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Tiunov, A.V. and S. Scheu. 2000. Microfungal communities in soil, litter and casts of Lumbricus terrestris L. (Lumbricidae): a laboratory experiment, Appl. Soil Ecol., 14:17–26. Tiwari, S.C., B.K. Tiwari, and R.R. Mishra. 1989. Microbial populations, enzyme activities and nitrogenphosphorus-potassium enrichment in earthworm casts and in the surrounding soil of a pineapple plantation, Biol. Fertil. Soils, 8:178–182. Tomlin, A.D., M.J. Shipitalo, W.M. Edwards, and R. Protz. 1995. Earthworms and their inßuence on soil structure and inÞltration, in P.F. Hendrix, Ed., Earthworm Ecology and Biogeography in North America, Lewis Publishers, Boca Raton, FL, pp. 159–183. Trojan, M.D. and D.R. Linden. 1992. Microrelief and rainfall effects on water and solute movement in earthworm burrows, Soil Sci. Soc. Am. J., 56:727–733. Trout, T.J. and G.S. Johnson. 1989. Earthworms and furrow irrigation, Trans. ASAE, 32:1594–1598. ù 1992. Review of some case studies on the abundance and on the hydraulic efÞciency Urbánek, J. and F. Dolezal. of earthworm channels in Czechoslovak soils, with reference to the subsurface pipe drainage, Soil Biol. Biochem., 24:1563–1571. Utomo, W.H. and A.R. Dexter. 1981. Age hardening of agricultural top soils, J. Soil Sci., 32:335–350. Utomo, W.H. and A.R. Dexter. 1982. Changes in soil aggregate water stability induced by wetting drying cycles in non-saturated soil, J. Soil Sci., 33:623–637. Van Hoof, P. 1983. Earthworm activity as a cause of splash erosion in a Luxembourg forest, Geoderma, 31:195–204. Wang, D., J.M. Norman, B. Lowery, and K. McSweeney. 1994. Nondestructive determination of hydrogeometrical characteristics of soil macropores, Soil Sci. Soc. Am. J., 58:294–303. White, G. 1789. The Natural History of Selborne, Benjamin White, London. Willoughby, G.L., E.J. Kladivko, and M.R. Savabi. 1997. Seasonal variations in inÞltration rate under no-till and conventional (disk) tillage systems as affected by Lumbricus terrestris activity, Soil Biol. Biochem., 29:481–484. Zachmann, J.E., D.R. Linden, and C.E. Clapp. 1987. Macroporous inÞltration and redistribution as affected by earthworms, tillage and residue, Soil Sci. Soc. Am. J., 51:1580–1586. Zhang, H. and S. Schrader. 1993. Earthworm effects on selected physical and chemical properties of soil aggregates, Biol. Fertil. Soils, 15:229–234. Ziegler, F. and W. Zech. 1992. Formation of water-stable aggregates through the action of earthworms. Implications from laboratory experiments, Pedobiologia, 36:91–96.

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of Earthworms on Soil 11 Effects Organization André Kretzschmar INRA-Biometrié, Avignon, France

CONTENTS The Effects of Earthworms on Soil Organization: Interactions between Earthworms and Soil Structure .................................................................................................201 Soil Porosity Caused by Earthworms.....................................................................................201 The Roles of Earthworm-Originated Soil Structures.............................................................202 Do Earthworm Burrows Affect Specific Soil Functions or Structures?................................202 Soil-Structure–Dependent Effects of Earthworm Burrows....................................................203 Earthworm Behavior Is Not Stable ........................................................................................204 Interactions between Earthworm Burrows and Soil Properties ....................................................204 Interactions of Earthworm Burrows with Other Soil Pores...................................................204 Interactions between Earthworm Burrows and Mass Transfer Processes .............................205 Interactions between Earthworm Burrows and Other Soil Biological Components.............206 Observation of Earthworm Interactions with Soils: Methodological Considerations ..................207 From Two to Three Dimensions.............................................................................................207 Topology and Functional Problems in Comparing Earthworm Burrow Patterns .................207 Problems of Scale ...................................................................................................................208 Conclusions ....................................................................................................................................208 Interactions between Earthworms and Soil Formation ..........................................................208 Monitoring Introductions of Earthworms to New Sites ........................................................208 References ......................................................................................................................................209

Earthworms have major impacts on the structure and organization of soils. Interactions between earthworm activities and behavior interact strongly with physical and biological soil properties and soil structures. Examples of such interactions are given in this chapter to demonstrate how careful observations and analysis can avoid a “chicken-and-egg” dispute. In addition, there is a discussion of how technical measures are used to address the question of the spatial stability of the interactions described.

THE EFFECTS OF EARTHWORMS ON SOIL ORGANIZATION: INTERACTIONS BETWEEN EARTHWORMS AND SOIL STRUCTURE SOIL POROSITY CAUSED

BY

EARTHWORMS

Endogeic species of earthworms contribute to soil porosity by burrowing and ingesting soil. Their burrowing creates large pore systems with complex structures that reflect the earthworm’s behavior

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and adaptations in correspondence with the soil environmental conditions (Lamparsky et al. 1987; Kretzschmar 1987; Kretzschmar et al. 1992). Ingestion of soil by earthworms creates structures of smaller size, mainly related to casts, which are deposited either on the soil surface or within the soil profile, and to aggregation properties of casts. The microporosity of earthworm casts results from the evolution and content of bound or free organic compounds after casts are deposited (Blanchard 1992); cast pore properties spread into the soil matrix and contribute partially to the soil matrix porosity characteristics that interfere with macropores. The two types of structures described belong to the two main categories of soil pores: macropores, which are roughly defined as larger than 0.5 mm diameter, and matrix porosity. Macropores are created either by physical processes (clay swelling, freezing, etc.) or by biological processes (roots, meso invertebrates, earthworms, etc.). Matrix pores are considered part of the basic soil fabric that defines the soil as a porous medium.

THE ROLES

OF

EARTHWORM-ORIGINATED SOIL STRUCTURES

Macropores formed as earthworm burrows are expected to improve mass transfers in soil at the profile scale. The extent of this improvement is based on connectivity and tortuosity of the earthworm burrows on the one hand (Kretzschmar 1987; Joschko et al. 1989; Edwards et al. 1992) and connections between the burrow walls and the matrix porosity in the vicinity of these macropores on the other hand (Kretzschmar 1987; Babel and Kretzschmar 1994). Earthworm burrows are not the principal source of soil macropores; an earthworm burrow system could be responsible for only 1 to 2% of the total porosity. The importance of earthworm burrows in mass transfer relies on two main factors. First, they are constructed in a cylindrical form with lightly compacted walls with at least several coatings of mucopolysaccharides (Kretzschmar 1987). Earthworm burrows are consolidated structures that stay open even when soil moisture is at high levels (e.g., when the soil matrix porosity is saturated with water, as are the macropores created by plant roots) and when clay swelling has closed most of the larger cracks in soils. At this stage, air-filled porosity is at its lowest level (Monnier 1992); the burrows might represent about 20% of the total air-filled porosity. Second, in temperate climates, a noticeable proportion of the burrow systems created by earthworm populations shows strong anisotropic orientations marked by a vertical direction. In this context, earthworm burrows are regarded as preferential vertical pathways for gravitic mass transfer (water and partial solutes) (Ehlers 1975; Shipita1o et al. 1990; Edwards et al. 1992; Smettem 1992). Earthworm burrows are also expected to influence the spatial distributions of other living organisms, such as roots and microorganisms. The distribution of plant roots can be affected by the presence of earthworm burrows. The density of roots in the vicinity of pores (cracks and burrows) is not the same as farther away from these pores (Krebs et al. 1994); nevertheless, no experimental evidence has shown that roots grow preferentially toward earthworm burrows or entered holes filled with earthworm casts (Hirth et al. 1998). The distribution of microorganisms associated with earthworm burrows is due partially to the organic carbon sources transferred through the earthworm burrow system, that is, mechanical transport of organic debris from the surface litter, mucus deposition along the burrow walls, or infillings of old burrows with casts. The distribution of microorganisms is also because of transfer of microorganisms through transit and the dispersion of microorganisms with water flowing through the larger burrows. Evidence of dispersion of microorganisms through earthworm burrows has been demonstrated in experimental conditions (Reddell and Spain 1991; Stephens et al. 1994) (see Chapter 10, this volume).

DO EARTHWORM BURROWS AFFECT SPECIFIC SOIL FUNCTIONS

OR

STRUCTURES?

The effects of earthworm burrows on soil structures and soil functions cannot be regarded as a simple causal relationship. When transport properties are measured in soils where earthworm burrows exist, it is not certain whether the burrow systems in these soils are so extensive because

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they provide suitable conditions for development of earthworm populations or if the earthworms develop a burrow system that improves mass transfer properties in this soil and then respond to soil environmental limitations. Observations of the development of earthworm burrow systems in cultivated soils previously deprived of any earthworm populations (e.g., pastures at Manatuke Experimental Station, Gisborne, New Zealand; personal observations) show interesting strategies of soil colonization by earthworms, but it is not possible to find any earthworm burrows (mainly caused by Aporrectodea caliginosa) occurring below the depth reached by the plant roots. Conversely, the burrow systems observed in a pasture in northern France (Kretzschmar 1982) contained a high density of earthworm burrows at a depth where very few roots were present. In these two cases, it is difficult to be sure whether the mechanism having the primary effects on soil structure is the effect of existing root system on the development of earthworm burrows or the opposite. The answer is almost certainly that roots and earthworms develop together, and in forming burrows that influence root patterns, earthworms adapt their strategies to suit the pedological environment. When studying the possible effects of the presence of earthworm burrows opening at the soil surface on the pattern of surface cracks under experimental conditions (Chadoeuf et al. 1994), the best that could be proved was that there are spatial relationships between the distribution of earthworm burrows and the distribution of cracks; none was assessed as the prime factor, and the conclusion was that the density of earthworm burrows in the vicinity of cracks (distance less than 2.5 cm) was greater than the density that would be predicted if the burrows had been distributed following a stationary Poisson pattern. It was not possible to decide whether the burrows usually opened closer to the cracks or if the cracks developed closer to the earthworm burrows (experimental conditions could be changed without solving this dilemma). This chicken-and-egg type of question demonstrates the limits of searching for causal relationships between earthworm activities and soil physical and biological properties, for which the former could cause variations in the latter.

SOIL-STRUCTURE–DEPENDENT EFFECTS

OF

EARTHWORM BURROWS

In temperate zones, soils with a “natural good structure,” “high fertility,” and “no earthworms” are known in North America and New Zealand (as well as tropical soils). It is then questionable how to define an improvement in soil physical properties in those types of soil after earthworms are introduced because the introduction of earthworms is often made at the same time as the introduction of new soil management practices. The improvements in pasture productivity in New Zealand after the widespread and monitored introduction of northern European lumbricid species of earthworms brought remarkable and stable increases in dry matter production (Springett 1985) in pastures that were then managed (i.e., fertilized and grazed) in the same way as northern European pastures. Earthworm populations developing in the pastures of South Australia have the advantage of living in soils receiving lime applications for pH control (Baker et al. 1992). Experimental evidence is available on the conditions under which earthworm activities in soils create significant modifications in soil properties. When a soil’s air-filled porosity is in the median zone (i.e., 15% soil volume), the development of an earthworm burrow system does not bring significant improvements in gaseous diffusion compared with soils with 25% air-filled porosity; conversely, when air-filled porosity is down to about 10%, conditions at which gaseous diffusion is severely limited (Glinski and Stepniewski 1985), a single earthworm burrow can dramatically change the rates of gas diffusion and its effects (Kretzschmar and Monestiez 1992; Kretzschmar and Ladd 1993). If the presence of earthworm burrows could be simulated as a short circuit in the mass transfer pathways in the soil matrix, the efficiency of this short circuit would depend entirely on the degree of connectivity associated with the burrows; furthermore, this connectivity must be considered from different views. In the soil profile, such a short circuit could join two zones of equal permeability crossing a less-permeable zone; at the scale of the soil matrix in the earthworm burrow vicinity, © 2004 by CRC Press LLC

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the connectivity of burrow volume with the pore space of the matrix through the soil will govern the extent of influence of the burrows. These two types of interactions have been described as axial and radial effects in the case of gaseous diffusion (Arah and Ball 1994).

EARTHWORM BEHAVIOR IS NOT STABLE Earthworm species may change their behavior greatly when they are transferred from one soil environment to another, despite apparently similar environmental conditions. It seems that there are no specific behavior patterns at the level of species or even at the level of “ecological” groups, as defined by Lee (1959, 1985), Bouché (1977), and Lavelle (1983). Aporrectodea caliginosa, a strictly endogeic species that makes short, disconnected burrows in soils in the Northern Hemisphere, makes long vertical burrows in the temperate soils of the Southern Hemisphere, a burrow shape that was reported by other authors to characterize anecic earthworm species. Similar observations can be made about Octolasium cyaneum, which has burrow patterns that are very different in prealpine pastures in Switzerland and in forested hills north of Adelaide in South Australia (personal observations). Lumbricus terrestris, which is considered a species with a strong potential for population development in orchard soils (Daniel 1992), was observed to be unable to spread in similar soil types and cultivation conditions (brown soil, orchards in organic farming system) in New Zealand (Springett, personal communication and personal observations, 1991). Interestingly, preliminary results about the burrow patterns of earthworm species introduced into homogenized soil plots set in field conditions seem to demonstrate that Allolobophora longa and A. caliginosa have the expected burrow patterns, that is, patterns similar to the burrow patterns they develop in their native environments (Springett, personal communication, 1991). Although the apparent similarity of the soil conditions might be mainly caused by a lack of suitable parameters to describe functional differences, such large variations in behavior indicate the ecological plasticity of the species that can be used to respond to local soil conditions. Moreover, the best examples of such a behavioral plasticity are observed among these northern European lumbricid earthworm species introduced into temperate America or the temperate Southern Hemisphere. Introduced species have to face soil conditions that result from interactions between natural local soil evolution and introduced soil management practices. Such interactions could develop a soil organization that might be regarded as deeply different from the soil conditions of places where the introduced species originated (see Chapters 5 and 13, this volume). It should be noted that, almost without exception, northern European earthworm species perform well when introduced at the same time as new agricultural management practices are adopted, despite variations in soil types and climates. The question is whether northern European earthworm species, which resist or could adapt to northern European soil management, are at the same time those species with the best aptitude to colonize new areas. Simple experiments could confirm the ecological significance of the behavioral plasticity of earthworms, for example, the intentional introduction of Southern Hemisphere earthworm species into northern European areas. For instance, what would be the population success of Microscolex dubius, which is well established in the temperate area of Australia, if it were introduced into reclaimed polders in the Netherlands?

INTERACTIONS BETWEEN EARTHWORM BURROWS AND SOIL PROPERTIES INTERACTIONS

OF

EARTHWORM BURROWS

WITH

OTHER SOIL PORES

A large variety of structures develop at the interface of the soil matrix and earthworm burrows. Essentially, walls that are lined, cracked walls in empty burrows, and burrows filled partially or completely with casts can be observed. Graff (1970) described the lining of earthworm burrow © 2004 by CRC Press LLC

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walls with organic material by L. terrestris. Lumbricus terrestris pulls leaves and large fragments of organic debris into its burrows, creating the black burrow lining observed by Graff, which resulted from the decomposition of organic material. In the burrows of many other earthworm species, a fine black layer is visible on the surface of the burrow walls. Reused burrows are lined by a layer of mucus produced from the accumulation of a mucus deposit about 10- to 15-µ thick each time an earthworm passes. Older burrows, which can be considered abandoned, present cracks that connect the burrows to porosity in the soil matrix (Kretzschmar 1987). Typical cracking patterns of the compacted layers around the burrows have been observed for the burrows of Megascolides australis (Babel and Kretzschmar 1994). These different patterns govern the contribution of burrow systems to overall soil transfer whether they are connected to the soil surface or not. The presence of macropores and their connections with the surrounding soil porosity create heterogeneous patterns of connectivity. It has been shown in simulated earthworm burrow systems that the mean length of the burrows plays a role in transfer properties at the scale of the soil profile; the average distance between burrows (governed by the burrow number) controls the mass transfer at the local scale of a given horizon (Kretzschmar 1988). The patterns of distribution of cracks around earthworm burrows have been described by Krebs et al. (1994). It is nevertheless still difficult to quantify interactions between these two components of the soil pore systems. Nevertheless, a similar interaction has been estimated for the distribution of earthworm burrows that open at the surface and the surface cracks during the drying phase under experimental conditions (Chadoeuf et al. 1994). The density of surface open burrows at a distance of less than 2.5 cm from cracks is greater than would be expected if the distribution of burrows and cracks were independent. The possibility of observing such “attraction-like” interactions within the soil profile between cracks and earthworm burrows depends on the soil structural context, mainly the presence of roots. More interestingly, the intensity of this interaction could summarize the history of the use of burrows both by earthworms and in water transfer.

INTERACTIONS

BETWEEN

EARTHWORM BURROWS

AND

MASS TRANSFER PROCESSES

The presence of large numbers of earthworms has been correlated with high soil hydraulic conductivity (Ehlers 1975); this observation has been confirmed many times with earthworms introduced into pastures or into differently cultivated plots (Springett et al. 1992), into experimental compacted columns (Joschko et al. 1992), and even when earthworms were expected to be present. Clothier and Vogeler (1994) reported improvements in soil conductivity that ranged from 0.3 µm−l to 1.2 µm−l under disk permeameters with salt solution because the earthworms are attracted to or repelled from the soil surface by salt solutions. The contribution of macropores to hydraulic conductivity can also be shown by theoretical considerations (Smettem 1992). Nevertheless, the effects of earthworm burrows depends on the type of connectivity that they attain within the burrow system itself or between the burrow system and pore spaces. Francis et al. (1994) reported differential effects of A. caliginosa and O. cyaneum in opening top soil and subsoil burrows, respectively. Although the total porosity was the same in both cases, the effects of the latter were much greater (Ksat = 573 mm h−l with O. cyaneum compared with 103 mm h−l with no earthworms) than the effects of the former (Ksat = 729 mm h−l with A. caligonosa compared with Ksat 439 mm h−l with no earthworms). The gaseous diffusivity was improved in the presence of earthworm burrows only when air-filled porosity was below a threshold (Kretzschmar and Monestiez 1992). Thus, it is inappropriate to attempt to measure the effect of earthworm burrows on mass transfer because such experiments tend to make this measurement only under conditions when they are visible or obvious; in essence, the absence of visible effects of burrows could be governed by opposite interactions because earthworm burrow volume is totally connected to the matrix pores, and the burrow volume is totally isolated from the matrix pores because of the impermeable walls.

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The global contribution of earthworm burrows to mass transfer does not describe the way that they interact with the soil structure. The effectiveness of short circuits caused by earthworm burrows is shown poorly by increases in the gas diffusion coefficient. This effectiveness is demonstrated clearly when the evolution of tracer gas concentrations with time is compared with the theory of gas diffusion following the non-steady-state equation (= second Fick’s law). In homogeneous soil cores, the data fit perfectly with the theoretical diffusion calculations. When macropores were present in the cores, the “distance” between the observed data and theory showed the effects of macropores. Gaseous diffusion in the presence of macropores no longer followed the Fickian diffusion process (Kretzschmar 1997). Modeling and simulations of specific gaseous diffusion because of macropores associated with basic diffusion caused by the matrix are not yet available.

INTERACTIONS COMPONENTS

BETWEEN

EARTHWORM BURROWS

AND

OTHER SOIL BIOLOGICAL

The interactions between roots and macropores, especially earthworm burrows, are easy to observe in a soil profile; nevertheless, it is not so easy when an attempt is made to follow these interactions in experimental conditions (Hirth et al. 1998). Observations of the soil profile have led to overestimates of the presence of roots in earthworm burrows because, although roots in burrows are easy to see, almost the whole root system is concealed in the soil matrix. Statistical estimations of spatial interactions between roots and fine cracks have been attempted on polished blocks of soil (i.e., in two dimensions) (Krebs et al. 1994). There is a large variability in these estimations within the same soil horizon and for the same plant species; moreover, it has been shown neatly that the interaction between roots and fine cracks is specific for each plant species, even when they are mixed, as in a pasture. Unpublished data from work by Brown and Kretzschmar (2001) showed at least that the global effects of earthworms seem to randomize root densities and then to favor the efficiency of soil exploration by roots. A three-dimensional estimation of the probability of crossing between earthworm burrow systems and root systems is not available. Earthworm burrows have been said to pave the way for plant roots; however, it is more realistic to expect that roots have a scouting function in exploring soil depth, and that earthworms follow the way and stabilize and develop the root-originated structures. The actual soil organization observed in an aged soil structure depicts an equilibrium reached through the joint development of roots and earthworms, each interacting with the other in response to soil conditions. Interactions between earthworm burrows and microbial activity have also been described either by measuring the microbial biomass in the vicinity of burrow walls (Loquet et al. 1977) or by showing the CO2 release associated with the presence of earthworm burrows. The chemical and organic properties of burrow walls are responsible for the development of microbial activities, together with the interactions of burrows with mass transfer, between the soil surface and soil matrix, through earthworm burrow systems. Water, solutes, and gas contents or concentration gradients are higher in the macropores than in the soil matrix, and the dynamics of microbial activities are closely related to the dynamics of these gradients. Observed effects of earthworms on CO2 release are dependent on the interactions of macropores with the surrounding porosity (Kretzschmar and Ladd 1993). CO2 release results from the combined effects of macropores on gaseous diffusion because of their connections to the matrix pore and to the soil surface and the enrichment of organic material along the earthworm burrow walls. The microbial population distribution interacts with that of the burrow system by way of both active and passive transport. Microorganisms are washed down and along the burrow walls when water flows intensively in vertical burrows; they are also transported into the guts of earthworms and deposited at the place where the earthworms cast.

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OBSERVATION OF EARTHWORM INTERACTIONS WITH SOILS: METHODOLOGICAL CONSIDERATIONS FROM TWO

TO

THREE DIMENSIONS

Most of the distribution data on earthworm burrows are available in only two dimensions, from the pioneer work of Ehlers (1975) to the latest computerized axial tomographic (CAT) scan observations. As connectivity and tortuosity have been reported to be the primary characteristics of the way earthworm burrows will interact with mass transfer properties, a three-dimensional reconstruction is a necessary step for studying burrow systems more fully. Several attempts to do this have been made, and new developments use x-ray CAT scans (Daniel et al. 1997). Although such three-dimensional reconstruction views the entire earthworm burrow system structure for the first time, it also raises two essential issues: •

•

What can be said about the similarity of two burrow system patterns? What are the pattern characteristics that describe the burrow system at a given site, and which characteristics can be associated with soil physical properties, such as mass transfer, or with biological activities such as root distributions? The three-dimensional reconstruction of earthworm burrow systems does not give evidence of their distribution in a three-dimensional connectivity of this system with the soil matrix. The spatial interactions between biological components (root, earthworms, microorganisms) requires that the distribution of any of these components should be described in three dimensions; because they develop at different scales, description methods should be compatible.

TOPOLOGY AND FUNCTIONAL PROBLEMS BURROW PATTERNS

IN

COMPARING EARTHWORM

Topological and stereological analyses (Kretzschmar 1988) seem to be insufficient to describe the specificity of earthworm burrow system patterns. The simplest case has been described for Lumbricus species (Lamparsky et al. 1987) for which individual earthworms lived in a single burrow and developed a few branches around it. The necessity to describe the whole burrow system is the major difficulty faced in this case. It has been shown that complex and continuous burrow patterns could be developed by single individual Megascolides australis (Kretzschmar and Aries 1992). When the earthworm burrow system results from the activity of a monospecific population (without the possibility of identifying individual earthworm territory) or from multispecific earthworm populations, where even the territory and pattern specificity of the burrows do not make it possible to distinguish burrows from each species, comparison of burrow system patterns requires that the topological and functional characteristics should not only rely on their geometrical distribution, but also take into account the earthworm behavior that governs such patterns. New approaches using Co-labeled earthworms together with x-ray tomography have opened new ways to understand, under artificial conditions, the differences between behavioral trajectories and the resulting burrow system (Jegou et al. 1999; Capowiez et al. 2001). Behavioral studies of earthworm movements and related functions under natural conditions are extremely rare and should be one of the research priorities for a better understanding of the development of earthworm populations in soils.

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Because biological organisms interact not only with each other, but also with soil physical properties (mainly the soil pore systems), it is necessary to describe the spatial distribution of each and to establish if their specific distributions can be considered independently. Functional links between living organisms and soil pore systems can occur at any scale, to such an extent that it is not possible to consider, at any place in the soil profile, the whole range of scales at which these interactions take place (i.e., ranging from microns to centimeters). Observation techniques are unable to deal with such a wide range; it is then necessary to develop a technical series in which, in the same sample, soil organization can be observed in such a way that the different levels of scale can be related to each other. It is possible to take large undisturbed soil cores (150- to 200mm diameter) and describe them with a CAT scanner at a resolution of 1 mm (Joschko et al. 1991; Daniel and Kretzschmar 1997). The same core could be impregnated and cut into 2-mm slices from which x-ray images can be taken at a resolution of 500 to 300 µm. Finally, the faces of these slices, once polished, could be observed under a microscope at a resolution of 50 to 30 µm (Vogel and Kretzschmar 1996). At any scale, specific models can be designed for an analysis of the pore system properties and their interactions with distributions of living organisms and activities (Monestiez et al. 1993; Krebs et al. 1994; Vogel and Kretzschmar 1996).

CONCLUSIONS INTERACTIONS

BETWEEN

EARTHWORMS

AND

SOIL FORMATION

As biological functions and physical properties interact, with the distribution and the seasonal variations in earthworm burrow distributions, earthworm burrows are seen as part of the development of soil organization, which is a characteristic of soils under given climatic, relief, and parentrock conditions (Dokuchaev 1883). Dokuchaev’s definition of soil genesis fixed the context of soil typology. The scientific literature has been generous in books and articles on soil-forming factors without any definite improvements of Dokuchaev’s statement. From the interactions between factors, soil organization is derived as a concept for the soil geometrical, functional, and topological properties. Babel et al. (1995) proposed an interesting definition of soil fabrics that is based on soil morphological features. They described soils based on three characteristics: place, pathways, and boundaries. These characteristics are attached to each object or function and are valid at any scale. It is probable that these characteristics would perfectly fit the description of the interactions addressed here. Soil-forming factors would be replaced by soil-forming interactions, for which places, pathways, and boundaries would be attached to geometrical, functional, and topological properties of these interactions, respectively.

MONITORING INTRODUCTIONS

OF

EARTHWORMS

TO

NEW SITES

The rationale of the intentional introduction of earthworms is based on an assumption of a beneficial effect of earthworms on soil fertility. However, transplanted species cannot always be expected to behave as they did in their original habitat. Earthworm activities should be regarded as dependent on soil organization in the same way that the density dependence of predator–prey relationships is described. In other words, these activities could not be understood without the ability to identify the specific interactions that take place in a given location (native habitat or introduction area). The question of reactions of soil organization to earthworm introductions should be regarded in terms of soil dynamics equilibrium: Will the introduced earthworms be able to change the characteristics of the actual organization of soils where they are introduced to such an extent that they bring about a new equilibrium, that is, a new state of organization involving and relying on a new set of

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interactions? The success of earthworm introductions (when both expected soil properties are obtained and introduced earthworm populations reach the expected and stable level of density) results from the conßict, or the synergy, of both soil and earthworm behavior plasticity.

REFERENCES Arah, J.R.M. and B.C. Ball. 1994. A functional model of soil porosity used to interpret measurements of gas diffusion, Eur. J. Soil Sci., 45, 135–144. Babel, U. and A. Kretzschmar. 1994. Micromorphological observations of casts and burrow walls of the Gippsland giant earthworm (Megascolides australis McCoy 1878), Ninth International Working Meeting on Soil Micromorphology, Townsville, Australia, July 1992, in Soil Micromorphology: Studies in Management and Genesis, A.J. Ringrose-Voase and G.S. Humphreys, Eds., Elsevier, Amsterdam, pp. 451–457. Babel, U., H. Vogel, M. Krebs, G. Leithold, and C. Hemmann. 1995. Morphological investigations on genesis and functions of soil fabric-places, pathways, boundaries, in Soil Structure. Its Development and Function, K.H. Hartge and B.A. Stewart, Eds., Adv. Soil Sci., Lewis Publishers, Boca Raton, FL, pp. 11–30. Baker, G., J. BuckerÞeld., R. Grey-Gardner, R. Merry, and B. Doube. 1992. The abundance and diversity of earthworm in pasture soils in the Fleurie Peninsula, South Australia, Soil Biol. Biochem., 24, 1539–1544. Blanchard, E. 1992. Restoration by earthworms (Megascolecidae) of the macroaggregate structure of a destructured savanna soil under Þeld conditions, Soil Biol. Biochem., 24, 1587–1594. Bouché, M. 1977. Statregies lombriciennes, Ecol. Bull., 25, 122–132. Capowiez, Y., P. Renault, and L. Belzuncez. 2001. Three-dimensional trajectories of 60Co-labelled earthworms in artiÞcial cores of soil, Eur. J. Soil Sci., 52, 365–375. Chadoeuf, J., A. Kretzschmar, M. Goulard, and K.R.J. Smettem. 1994. Description of the spatial interaction between earthworms burrows and cracks at the soil surface, Ninth International Working Meeting on Soil Micromorphology, Townsville, Australia, July 1992, in Soil Micromorphology: Studies in Management and Genesis, A.J. Ringrose-Voase and G.S. Humphreys, Eds., Elsevier, Amsterdam, the Netherlands, pp. 521–530. Clothier, B. and I. Vogeler. 1994. Soil physics under pressure (still a can of worms!), WISPAS, Hortresearch, 58, 4. Daniel, O. 1992. Population dynamics of Lumbricus terrestris L. (Oligochaeta: Lumbricidae) in a meadow, Soil Biol. Biochem., 24, 1425–1431. Daniel, O., A. Kretzschmar, Y. Capowiez, L. Kohli, and J. Zeyer. 1997. Computer-assisted tomography of macro porosity and its application to study the activity of earthworms Aporectgodea nocturna, Eur. J. Soil Sci., 48(4), 727–737. Dokuchaev, V.V. 1883. Russian Chernozem, St. Petersburg. Edwards, W.M., M.J. Shipiltalo, S.J. Traina, C.A. Edwards, and L.B. Owens. 1992. Role of Lumbricus terrestris (L.) burrows on quality of inÞltrating water, Soil Biol. Biochem., 24, 1555–1562. Ehlers, W. 1975. Observations on earthworm channels and inÞltration on tilled and untilled loess soils, Soil Sci., 119, 242–249. Francis, G., T. Fraser, and W. Jian. 1994. The worms that turned, WISPAS, Hortresearch, 58, 3–4. Glinski, J. and W. Stepniewski. 1985. Soil Aeration and Its Role for Plants, CRC Press, Boca Raton, FL, p. 229. Graff, O. 1970. Effect of different mulching materials on the nutrient content of earthworm tunnels in the subsoil, Pedobiologia, 10, 305–319. Hirth, J.R., B.M. McKenzie, and J.M. Tisdall. 1998. Roots of perennial ryegrass (Lolium perenne) inßuence the burrowing of the endogeic earthworm, Aporrectodea rosea, Soil Biol. Biochem., 30, 2181–2183. Jegou, D., V. Hallaire, D. Cluzeau, and P. Trehen. 1999. Characterization of burrow system of the earthworm Lumbricus terrestris and Aporrectodea giardi using x-ray computed tomography and image analysis, Biol. Fertil. Soils, 29, 314–318. Joschko, M., H. Diestel, and O. Larink. 1989. Assessment of earthworm burrowing efÞciency in compacted soil with a combination of morphological and soil physical measurements, Biol. Fertil. Soils, 8, 191–196. © 2004 by CRC Press LLC

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Joschko, M., O. Graff, P.C. Muller, K. Kotzke, P. Linder, D.P. Pretschner, and O. Larink. 1991. A nondestructive method for the morphological assessment of earthworm burrow systems in three dimensions by x-ray computed tomography, Biol. Fertil. Soils, 2, 88–92. Joschko, M., W. Sochtig, and O. Laring. 1992. Functional relationships between earthworm burrows and soil water movement in column experiments, Soil Biol. Biochem., 24, 1545–1548. Krebs, M., A. Kretzschmar, U. Babel, J. Chadoeuf, and M. Goulard. 1994. Investigations on distribution patterns in soil: basic and relative distributions of roots, channels and cracks, Ninth International Working Meeting on Soil Micromorphology, Townsville, Australia, July 1992, in Soil Micromorphology: Studies in Management and Genesis, A.J. Ringrose-Voase and G.S. Humphreys, Eds., Elsevier, Amsterdam, the Netherlands, pp. 437–449. Kretzschmar, A. 1982. Description des galeries de vers de terre et variations saisonnieres des reseaux (observations en conditions naturelles), Rev. Ecol. Biol. Sol, 19, 579–591. Kretzschmar, A. 1987. Caracteristiques micromorphologiques de l’activite des lombriciens, in Seventh International Workshop on Soil Micromorphology, N. Fedorov, Ed., AFES, Paris, pp. 325–330. Kretzschmar, A. 1988. Structural parameters and functional patterns of simulated earthworm burrow systems, Biol. Fertil. Soils, 6, 252–261. Kretzschmar, A. and F. Aries. 1992. Analysis of the structure of the burrow system of the giant Gippsland earthworm Megascolides australis, McCoy (1878) using 3D-images, Soil Biol. Biochem., 24, 1583–1586. Kretzschmar, A. and J.N. Ladd. 1993. Decomposition of 14C-labelled plant material in soil: the influence of substrate location, soil compaction and earthworm number, Soil Biol. Biochem., 25, 803–809. Kretzschmar, A. and P. Monestiez. 1992. Physical control of soil biological properties due to endogeic earthworm behaviours, Soil Biol. Biochem., 24, 1609–1614. Lamparsky F., A. Kobel-Lamparsky, and R. Kaffenberger. 1987. The burrow of Lumbricus badensis and Lumbricus polyphemus, in On Earthworms, A.M. Bonvicini and P. Amodeo, Eds., Mucchi, Modena, Italy, pp. 131–140. Lavelle, P. 1983. The structure of earthworm communities, in Earthworm Ecology, J.E. Satchell, Ed., Chapman & Hall, London, pp. 449–466. Lee, K.E. 1959. The Earthworm Fauna of New Zealand, New Zealand Department Sci. Industr. Res. Bull., New Zealand, pp. 130. Lee, K.E. 1985. Earthworms. Their Ecology and Relationships with Soils and Land Uses, Academic Press, Sydney, Australia. Loquet, M., T. Bhatnagar, M. Bouche, and J. Rouelle. 1977. Essai d’estimation de I’influence ecologique des lombriciens sur les micro-organismes, Pedologia, 17, 400–417. Monestiez, P., A. Kretzschmar, and J. Chadoeuf. 1993. Modeling natural burrow systems in soil by fibre process: Monte-Carlo test on independence of fibre characteristics, Acta Stereol. (YUG), 12, 237–242. Monnier, G. 1992. L’activite des vers de terre du point de vue de la physique du sol, Soil Biol. Biochem., 24, 1197–1200. Reddell, P.R. and A.V. Spain. 1991. Earthworms as vectors of viable propagules of mycorrhizal fungi, Soil Biol. Biochem., 23, 767–774. Shipitalo M.J., W.M. Edwards, W.A. Dick, and L.B. Owens. 1990. Initial storm effects on macropore transport of surface-applied chemicals in no-till soil, Soil Sci. Soc. Am. J., 54, 530–536. Smettem, K.R.J. 1992. The relation of earthworms to soil hydraulic properties, Soil Biol. Biochem., 24, 1539–1544. Springett, J.A. 1985. Effect of introducing Allobophora longa (Ude) on root distribution and some soil properties in New Zealand pastures, in Ecological Interactions in Soils: Plants, Microbes and Animals, A.H. Fitter et al., Eds., Blackwell, Oxford, U.K., pp. 399–405. Springett, J.A., R.A.J. Gray, and J.B. Reid. 1992. Effect of introducing earthworms into horticultural land previously denuded of earthworms, Soil Biol. Biochem., 24, 1615–1622. Stephens, P.M., C.W. Davoren, B.M. Doube, and M.H. Ryder. 1994. Ability of the earthworm Aporrectodea rosea and Aporrectodea trapezoides to reduce take-all under greenhouse and field conditions, Soil Biol. Biochem., 26, 1291–1297. Vogel, H. and A. Kretzschmar. 1996. Topological characterisation of pore space in soilsample preparation and digital image processing, Geoderma, 73, 23–38.

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Part VI Interactions of Earthworms with Microorganisms, Invertebrates, and Plants

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Interactions 12 Functional between Earthworms, Microorganisms, Organic Matter, and Plants George G. Brown Embrapa Soya, Londrina, Brazil

Bernard M. Doube Wood Duck Cellars, Bridgewater, South Australia, Australia

CONTENTS Introduction ....................................................................................................................................213 Organic Matter and Microbial Communities ................................................................................215 Annual Organic Inputs and Decomposition Processes ..........................................................215 Vertical Carbon Gradients, Patchiness, and Quality of Residues in Soil..............................217 Successions of Microorganisms during the Decomposition Processes .................................217 Feeding Behavior of Earthworms...........................................................................................218 Interactions between Earthworms and Microbial Communities...................................................218 The Earthworm Drilosphere: Microscale Interactions...........................................................218 Food Preferences and Dietary Requirements ...............................................................218 The Fate of Microorganisms in the Intestines of Earthworms ....................................221 Microbiological Composition and Activity in Earthworm Burrows and Casts ...........222 Successional Processes within Casts ............................................................................223 Mesoscale Interactions between Earthworms and Microorganisms ......................................224 Macroscale Interactions between Earthworms and Microorganisms ....................................229 Summary and Conclusions ............................................................................................................230 References ......................................................................................................................................231

INTRODUCTION From the moment that soil is consumed by or enters into contact with an earthworm, either superficially or internally, physicochemical and microbiological changes take place. Furthermore, when seeds germinate, they immediately come into contact with soil microorganisms, and as the plant roots grow, microorganisms promote changes in the soil physicochemical and microbiological environment. The three-way plant-microbe-invertebrate interactions that follow have profound effects on the growth and development of plants, soil microorganisms, and invertebrate communities. The intensity of the spatial and temporal interactions in subterranean food webs (Anderson 1988) can be influenced strongly by plant diversity, as well as by the abiotic environmental 213 © 2004 by CRC Press LLC

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MIDDEN

Intimate mixing of mineral and organic particles Litter comminution and decomposition Microbial and faunal proliferation External rumen

DRILOSPHERE COMPONENTS CO2 N2

BURROW SYSTEM Burrows open to soil surface

EARTHWORM SURFACE

Nutrient absorption, leaching

EARTHWORM INTERNAL PROCESSES

Open burrows/macroporosity Walls (Changes in C, N, microbial populations and activity)

FOREGUT Exonephridial (external) N excretion

Food selection/ ingestion

Digestion of fungal hyphae, bacteria, trophozoites, algae Intestinal mucus (assimilable C and N)

Mostly below ground Changes in microbial biomass, activity, diversity and successional processes Changes in the concentration of nutrients (C, N, some macro and micro) Microporosity Aggregation (C protection)

MIDGUT Antifungal and External mucus antibacterial secretions Assoc. N2 fixation secretion (along (antibiotics?) (e.g., Chlostridia?) whole body surface)

Crop

Calciferous glands (CaCO3 secretion, pH increase)

Gizzard (grinding)

CASTS Soil surface

Roots growing in burrow and into cast Closed burrows (cast-filled)

Mucus and other secretions Respiration

Aestivation/diapause chamber

N2O

Gases, water

HINDGUT

Reassimilation of C C

Tissue production (nutrient immobilization)

Enteronephridial N excretion (into the gut)

Bacterial stimulation, enzyme production, digestion of organic compounds

Changes in P & K solubility, ammonification (Org. N NH3)

Metabolic wastes

Feces egestion (casts)

Figure 12.1 Schematic representation of the drilosphere components and their relationships with the external and internal earthworm environment, microorganisms, and organic matter. (Modified from Brown et al. 2000.)

characteristics (parent material, soil type, climatic regime). These interactions occur in a number of key biological spheres that are the foci of intense microbial activity associated with the decomposition of organic residues, including those from the rhizosphere (roots), detritusphere (surface detritus), drilosphere (earthworms), myrmecosphere (ants), and termitosphere (termites) (Lavelle 2002). Some of the functional interactions between microorganisms, earthworms, and the structures in the drilosphere they generate (Figure 12.1) can affect soil organic matter (OM) dynamics (production, decomposition, stabilization) at various spatiotemporal scales, from the earthworm gut to its burrows and casts (Lavelle 1997; Brown et al. 2000; see Chapter 8, this volume). The spatial scales at which soil organisms act are determined mainly by their size and mode of operation. Three spatial scales can be recognized based on animal size (micro, meso, and macro; Wallwork 1970), and the processes that occur at each scale are crucial to primary production, decomposition, nutrient cycling, soil aggregation, and root health. At the microscale, there are algae and bacteria (which are unable to move long distances except if carried by water or larger soil organisms) and fungi (in which hyphal growth provides the capacity to colonize new soil). Still at a microscale, but increasing in size and spatial influence, the micro-food web includes the microfauna, such as nematodes, protozoa, rotifers, and other organisms that feed mainly on the microflora and are important in regulating nutrient cycling, particularly in the rhizosphere (e.g., Clarholm 1985; Ingham et al. 1985). At the mesoscale, larger organisms such as enchytraeids and micro- and mesoarthropods feed on litter, microorganisms, and invertebrates and are important in accelerating nutrient cycling and in the small-scale dispersal of microorganisms (Hassal et al. 1987). Finally at the macroscale, there are larger invertebrates, such as earthworms, termites, macroarthropods and ants; these are termed ecosystem engineers (Lavelle et al. 1997). They disperse microorganisms, produce important physical structures (mounds, burrows, casts, pellets) that may occupy or modify a great portion of the soil volume (and the microbial communities found therein), and therefore regulate many soil processes. © 2004 by CRC Press LLC

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These interactions must be observed and considered not only at different scales, but also in a number of directions. Typically, studies of interactions have focused on one-way processes, commonly disregarding the possibility of multifaceted interactions. For instance, the one-way effects of earthworms on plant growth have been investigated intensively (see Chapter 2, this volume), but there may be other interactions, such as between earthworms and root pathogens or beneficial soil microorganisms (Doube et al. 1994d,e), that also affect plant growth. Further, the corresponding effects of plants and microbial processes on earthworms are rarely considered in the same studies. This chapter primarily concerns interactions among the soil biota and their effects on plants in temperate and subtropical agricultural systems, with an emphasis on cropped soils. We consider interactions between earthworms, soil microorganisms (bacteria, fungi, algae, and protozoa), and plants (roots, seeds, and aboveground parts), and examine them at three different spatial scales: micro (gut, burrows, and casts), meso (whole soil), and macro (field), although there are few data on some topics, particularly on interactions at larger scales, and comparative data concerning the different life strategies of earthworms are lacking. At each scale, we address the two-way interactions between earthworms and other organisms or processes and attempt to reach some conclusions about the functional significance of these interactions. The theme is that the composition and activity of the microbial communities responsible for the decomposition of organic substrates in soil can be affected significantly by the activity of soil macroinvertebrates, especially earthworms, which therefore can play a vital role in regulating the microbial processes that maintain the biological health of soils.

ORGANIC MATTER AND MICROBIAL COMMUNITIES In this section, we consider the amounts, types, and distribution of organic materials entering soil systems and the succession of microorganisms that promote OM decomposition. These inputs and microbial processes are the basis of the soil food webs that support and are modified by earthworms (Killham 1994; Brown 1995; Lavelle and Spain 2001).

ANNUAL ORGANIC INPUTS

AND

DECOMPOSITION PROCESSES

Dead organic residues comprise the major part of soil OM (Figure 12.2). These are derived mainly from dead animals and three plant sources: surface plant residues or litter (which may be incorporated into soils by tillage or biotic processes), dead roots and sloughed cells, and exudates leaked into the rhizosphere from living roots (Lynch and Whipps 1990). Living OM makes up only a relatively small proportion of the total OM in soils (Figure 12.2), and soil microorganisms comprise the major part of the living biomass. Nevertheless, interactions between the three living soil components (roots, macrofauna, and microorganisms) have major influences on soil processes. Annual inputs of surface organic residues into cropping systems can be as high as 40 t ha−1 (e.g., in sugar cane), although the mass of surface residues is commonly much lower (e.g., ca. 2 to 6 t ha−1 in temperate cereal production). In agricultural systems, aboveground and subsurface plant residue inputs are commonly of a similar order of magnitude, with roughly half (range 35 to 80%) of the net photosynthate transferred to the roots of plants and entering the soil as roots or root exudates (Zwartz et al. 1994). During the first year in the soil, between 60 and 75% of the introduced carbon is respired and therefore lost to the pool of soil OM (Jenkinson and Ladd 1981), with the remainder persisting in organic-soil associations and living tissues (Figure 12.2). The living tissues, which are part of the soil food web, include autotrophic algae and a vast array of heterotrophic primary and secondary decomposers. For instance, in temperate cereal fields, the primary decomposers consist mainly of bacteria (30 to 90% by weight) and fungi (5 to 70% by weight), which are grazed by a variety of secondary decomposers, such as protozoans (0.6 to 6.0% of the total biomass) and earthworms (0 to 14% of the total biomass) (Brussard et al. 1990). The microbial biomass (bacteria, fungi, algae, rotifers, © 2004 by CRC Press LLC

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Soil Soil organic matter 0.1–10%

Mineral content >90%

Living 15%

Nonliving 85%

Root biomass 5–15%

Faunal biomass 5–10%

Microbial biomass 75–90%

Humus 70–85%

Inert carbon 5–20%

Macro-organic matter 10–30%

FIGURE 12.2 Composition of soil organic matter in a typical fertile soil. (Adapted from Pankhurst et al. 1997.)

protozoa) in soil in temperate and subtropical regions occupies less than 1% of the surface soil volume and is commonly on the order of 0.5 to 3 t ha−1 of biomass; that of the macrofauna is commonly much lower (Brussard et al. 1990; Coleman et al. 1994), although higher values have been recorded. For example, Doube et al. (1994a) measured 1.15 t ha−1 biomass of the earthworm Aporrectodea trapezoides in soil under a canola crop in Australia, and Barois et al. (1988) reported up to 4 t ha−1 biomass of earthworms (mostly Polypheretima elongata) in a tropical pasture in Martinique. The amounts of carbon in the soil biota, nonetheless, are only a small percentage (<0.5 to 3.0%) of the total organic carbon in soil (Haines and Uren 1990; Gupta 1994; Sparling et al. 1994; Lavelle and Spain 2001). The remaining organic carbon is bound in nonliving associations in soil. The succession of microorganisms associated with the decomposition of each different type of organic residue is similar in many respects but also has unique features that are dependent on their origin (e.g., leaf tissue is buried with phyloplane species) and composition (e.g., all root exudates leaked into the rhizosphere are readily assimilable) (Lavelle and Spain 2001). The microbial agents responsible for decomposition of organic residues also vary with region, location, and management practices. For example, at Lovinkhoeve in the Netherlands, Brussard et al. (1990) reported that bacteria constituted by far the greatest portion of the microbial biomass (93 to 95%), with fungi representing only about 5%. Similarly, in America at Horseshoe Bend in Georgia, Hendrix et al. (1986, 1987) reported that bacteria predominated (60 to 76% biomass), with fungi making up only 17 to 22% of the microbial biomass. In contrast, Andren et al. (1988, 1990) stated that at Kjetteslinge, Sweden, fungi dominated in a barley soil (64 to 69% of microbial biomass), and bacteria made up only about 30% of the total biomass. In a forest soil in Canada, bacterial:fungal ratios were 0.1 to 0.2 in the soil surface layers (Scheu and Parkinson 1994). In the surface soils of agroecosystems (where most earthworms feed and grow), both bacteria and fungi appear to be more abundant under reduced tillage than under conventional cultivation (Hendrix et al. 1986; Brussard et al. 1990; Coleman et al. 1994). Absence of tillage appears to increase the importance of fungi relative to bacteria as primary decomposers and hence as a source for the food web. Consequently, fungivorous nematodes and earthworms are correspondingly more abundant (Coleman et al. 1994). Such later increases in biomass appear to be in response to increases in the biomass of a specific resource base (e.g., fungi), but the degree to which the composition of the microbial community is influenced by the corresponding increase in predatory or bioturbating organisms (e.g., earthworms) is unknown. © 2004 by CRC Press LLC

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VERTICAL CARBON GRADIENTS, PATCHINESS,

AND

QUALITY

OF

RESIDUES

IN

217

SOIL

There are strong vertical gradients in the concentrations of OM and microbial biomass in the upper soil layers. These gradients are strongest where there is no tillage, since tillage redistributes organic residues through the soil profile. Some species of earthworms bury surface plant residues and are more common in direct drilled (DD) or no-till (NT) fields than in conventionally cultivated (CC) fields (Chan 2001). Despite this, soil carbon gradients are commonly, but not always (e.g., Brussard et al. 1990), steeper in DD/NT than in CC cereal fields (Hendrix et al. 1986; Haines and Uren 1990), suggesting that tillage is more effective than biological processes in redistributing surface OM residues through the soil profile. The patchiness of the distribution of OM residues through the topsoil varies with the spatial scale examined (Robertson 1994). Patchiness is extremely high at a small spatial scale (e.g., that of a root, a fragment of buried OM, or an earthworm cast) but decreases as the spatial scale increases and the microsite differences become integrated into a relatively homogeneous pattern of OM distribution (e.g., at the paddock or field scale). The degree to which earthworm activity alters this pattern is understood poorly, but fresh earthworm casts are clearly the foci of new and high microbial activity. The quality of organic inputs into soil, which vary with type of vegetation and management practices, has an important bearing on the composition and functioning of microbial communities. For example, the leaves and stalks of cereal crops contain large amounts of cellulose, hemicellulose, and lignin and have high C:N ratios, whereas in the rhizosphere, the inputs from roots are soluble carbohydrates, amino acids, and the like, with few recalcitrant residues (Bowen and Harper 1989; Moody et al. 1995; Lynch and Whipps 1990).

SUCCESSIONS

OF

MICROORGANISMS

DURING THE

DECOMPOSITION PROCESSES

There is a clear succession of microorganisms associated with the decomposition of fragments of organic residues in litter and soil (Ponge 1991; Robinson et al. 1993; Moody et al. 1995) and in the rhizosphere (Zwartz et al. 1994). The species composition of the microbial communities is determined by the nature of their substrates and the colonists (which include phyloplane species, i.e., those introduced to soil with plant residues) and soil residents found near the organic residues. The initial colonists of decomposing organic residues (e.g., cereal straw) are microbial species that can exploit soluble organic compounds (carbohydrates, organic acids, amino acids) and are dominated by fungi such as Mucor, Pythium, and Penicillium species (Harper and Lynch 1985; Bowen and Harper 1989; Moody et al. 1995) and bacteria such as Pseudomonas species. The second successional phase in the decomposition of organic residues is dominated by fungi such as Trichoderma, Fusarium, and Chaetomium species and bacteria such as Bacillus species, which have the capacity to digest cellulose and hemicellulose (Garnett 1981; Moody et al. 1995), although much still remains undecomposed at the end of this phase. The third phase of decomposition is associated with the slow metabolism of recalcitrant organic residues with high lignin or polyphenol contents. Fungi such as white rot fungi (Lavelle et al. 1993) and basidiomycetes commonly dominate this phase of the successional process, although cellulolytic fungi that can decompose delignified polysaccharides (e.g., Fusarium and Trichoderma spp.) may also occur at this stage (Bowen 1990; Bowen and Harper 1990; Moody et al. 1995). This flush of primary decomposers induces corresponding increases in the abundance of secondary decomposers (protozoa, bacterivorous and fungivorous nematodes) (Anderson 1994; Zwarz et al. 1994). In the rhizosphere, similar biological processes occur in response to root exudation and sloughed dead root cells (Zwarz et al. 1994). Rhizosphere populations of bacteria and protozoans may be many times higher than in surrounding soil, but fungi commonly show a proportionately smaller increase or no increase in relation to bulk soil (Newman 1985; Gilbert et al. 1993; Gupta 1994; Zwarz et al. 1994).

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EARTHWORMS

Three ecological strategies of earthworms have been recognized (Bouché 1977; Lavelle 1988). Epigeic species are litter dwelling and active primarily in the detritusphere; they feed on fresh organic materials and are important in comminution of litter and its decomposition. Endogeic species, which live in the soil (from the top few centimeters down to depths greater than 2 m), consume large quantities of soil and organic residues and some surface litter and are responsible for pronounced changes in the soil physical structure (Lavelle et al. 1989). Anecic species live in all strata of soil, normally in permanent vertical burrows that open to the soil surface. They feed on and bury surface litter and hence are important in modifying soil gaseous and water regimes (Edwards et al. 1990). Obviously, the consequences of the interactions between earthworms and microbial communities vary substantially with the ecological category to which the earthworms belong.

INTERACTIONS BETWEEN EARTHWORMS AND MICROBIAL COMMUNITIES Here, we examine the interactions between earthworms and microbial communities at three spatial scales. At a microscale (that of the earthworm gut or intestine, burrow lining, or casts), we examine the food preferences of earthworms, the fate of microorganisms in the intestines of earthworms, and the chemical and biological composition of casts and successional processes within them. At a mesoscale, which integrates the drilosphere with the surrounding soil, we consider the ways in which earthworm activity influences whole soil characteristics and functions, such as the distribution of microorganisms, soil respiration, microbial biomass, bacterial:fungal ratios, and the way these processes alter soil fertility and the incidence and severity of root diseases. These can be analyzed using microcosms or in small-scale field trials. At a macroscale (e.g., that of a field), few data are available, but earthworm-induced changes in microbial function have the potential to influence broader-scale processes such as soil structure (affecting water infiltration and erosion), microbial diversity, patterns of plant abundance, soil productivity, and crop yields. Conversely, at the last two spatial scales, the composition of the microbial community has the capacity to affect the distribution and abundance of earthworm communities.

THE EARTHWORM DRILOSPHERE: MICROSCALE INTERACTIONS The earthworm populations and the soil volume and microbial and invertebrate populations influenced directly or indirectly by earthworm activities were termed the drilosphere by Lavelle (1988). The drilosphere includes five main components (Brown et al. 2000): the internal microenvironment of the earthworm gut; the earthworm surface in contact with the soil; surface and belowground earthworm casts; middens; and burrows, galleries, or diapause chambers (open and closed) (Figure 12.1). Food Preferences and Dietary Requirements Many species of earthworms consume a mixture of soil and OM. For example, Doube et al. (1997) showed that species from all of the functional groups (the epigeic Lumbricus rubellus, the endogeic Aporrectodea caliginosa, and the anecic Lumbricus terrestris and Aporrectodea longa) preferred a mixture of soil and OM over pure OM. Furthermore, increased levels of clay (Barois et al. 1999), humus (Makulec and Kusinska 1995), and organic carbon (Doube et al. 1994b; Figure 12.3) in earthworm casts compared with the average of the surrounding soil in which earthworms have been living indicate that they feed selectively on patches of soils that are relatively rich in OM. Judas

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4.5 4 3.5 % Carbon

3 2.5 2 1.5 1 0.5 0 0–2.5

2.5–5.0

10–20

Soil Depth (cm) Parent soil

Casts (no dung)

Casts (plus dung)

FIGURE 12.3 Organic carbon levels in parent soil from three sampling depths and in the casts of A. trapezoides maintained in those soils in the presence and absence of added sheep dung. The sandy soil was taken from three depths (0 to 5 cm, 5 to 10 cm, and 20 to 30 cm) in a field near Monato, South Australia. (Adapted from Doube et al. 1994b.)

(1992) confirmed this by direct observation of earthworm gut contents. Using the same technique, Wolter and Scheu (1999) reported that the anecic species L. terrestris fed on soil microsites that were enriched in bacteria and fungi. Earthworms of different species and ecological categories differ greatly in their ability to digest various organic residues and assimilate nutrients from ingested OM (Lattaud et al. 1998, 1999), so selective feeding on OM- and microbe-rich microsites such as the rhizosphere (Spain et al. 1990; Brown et al. 2000) or decomposing plant litter (Wolter and Scheu 1999) may provide earthworms with additional soluble C sources, such as carbohydrates, root exudates, and those derived from microbial metabolism. In studies on the feeding behavior of several earthworm species (A. trapezoides, Aporrectodea rosea, Microscolex dubius, Eisenia fetida), use of an organic dye mixed with sheep dung (Doube and Davoren, unpublished) showed that, when offered buried or surface-deposited sheep dung, these species fed in an alternate manner on both dung and soil. Their castings thus consisted of a series of bands of dyed and undyed material, deposited sequentially. Organic carbon levels in the dyed portions (from the sheep dung) were 5 to 10 times higher than in the nondyed soil sections of the cast. This indicates that organic residue heterogeneity within casts is much higher than previously thought, and that differences in microbial activity may be correspondingly variable. The role of microorganisms as a source of earthworm nutrition is still a matter of debate. Various enzymes isolated from earthworm guts allow them to digest some bacteria and fungi, microinvertebrates (e.g., protozoa, nematodes), and partly decomposed plant debris (Brown et al. 2000). However, the amounts of microorganisms consumed and the ability of the earthworms to digest and assimilate microbial biomass varies greatly with the earthworm species, its ecological category, character of the food, and the environmental conditions in which the earthworms are living. For instance, Wolter and Scheu (1999) showed that levels of digestion of fungi and bacteria in the gut of L. terrestris were generally low and varied with the type of food substrate. Nevertheless, other authors (Dash et al. 1979a, 1986; Edwards and Fletcher 1988) have shown that fungi may be an important source of food for many earthworm species, but the level of dietary © 2004 by CRC Press LLC

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reliance on fungi, particularly among litter-feeding epigeic and anecic earthworm species (which feed on materials extensively colonized by fungi), remains uncertain. During passage through the earthworm gut, fungal hyphae (except those protected inside root fragments) (Reddell and Spain 1991a) are digested preferentially, whereas many (but not all) types of fungal spores pass through unharmed. This can have important consequences for the microbial processes and succession in earthworm casts (Tiunov and Scheu 1999; see Successional Processes within Casts, this chapter). Morgan (1988) studied the food value (for the epigeic species E. fetida) of pure cultures of a range of fungi. Earthworms gained or maintained weight when fed six of the eight fungi tested and died when fed the other two. Several other species of fungi (primarily toxin- or antibiotic-producing fungi such as Aspergillus spp., Fusarium spp. and Penicillium spp.) appeared to be detrimental to earthworms (Edwards and Fletcher 1988; Morgan 1988). Studies using six earthworm species and more than 10 soil and litter fungal species (Moody et al. 1995; Bonkowski et al. 2000) have shown that earthworms prefer, and partly digest, the rapid-growing fungi species typically associated with the early successional stages of decomposition (cellulolytic fungal species and those that consume soluble carbohydrates, e.g., Fusarium and Trichoderma spp.). On the other hand, fungal species appearing later in the succession (characteristically degraders of recalcitrant polymers, e.g., Basidiomycetes) were avoided and proved to be a poor food source for the earthworms. An examination of the survival of fungal spores during passage through the earthworm intestine showed that survival varies greatly depending on the characteristics of the spore and the earthworm gut environment (Brown 1995). For instance, Moody et al. (1996) observed that no viable spores of Fusarium lateritium survived passage through L. terrestris and A. longa; spores of Mucor hiemalis were more severely affected by passage through the guts of L. terrestris than through the guts of A. longa (10 and 28% survival, respectively). The survival of spores of yet another fungal species (Chaetomium globosum) was unaffected by passage through guts of either earthworm species. Reddell and Spain (1991a,b) demonstrated that spores of an actinomycete, Frankia sp., and spores of more than 20 mycorrhizal fungal species or groups remained intact after passage through the gut of Pontoscolex corethrurus, and most remained viable. On the other hand, spores of other species, such as Ulocladium botrytis, appear to have low germination rates in earthworm casts (fungistasis) (Striganova et al. 1989b); still others (Pithomyces chartarum) did not germinate after passage through earthworm guts (Keogh and Christensen 1976). Fungal digestion appears to show a gradient along the earthworm gut, in which fungi are digested mainly in the anterior and middle gut regions, with little digestion occurring in the hindgut (Gonzalez 1990; Tiwari et al. 1990). A similar pattern of digestion of protozoa along the earthworm gut has also been observed (Piearce and Phillips 1980). Free-living soil protozoa may also be important dietary elements for some earthworm species. For example, E. fetida was able to grow to maturity in sterilized soil recolonized by soil bacteria and fungi only after the addition of protozoa (Miles 1963). Flack and Hartenstein (1984) recorded large weight gains in earthworms after adding protozoa to their food. Bonkowski and Schaefer (1997) showed further that populations of the endogeic earthworm A. caliginosa increased in soils containing high densities of amoebae, and that earthworm growth rates increased with the addition of amoebae to soil. Similarly, Doube and Gupta (unpublished data) showed that A. trapezoides and E. fetida ate soil containing flagellate and cilliate protozoans preferentially, and both earthworm species responded positively to the fluid in which the protozoa had been living, indicating that the earthworms can respond to the products of protozoan activity, possibly low molecular weight compounds, as well as to the protozoa themselves. Protozoan protoplasm is highly assimilable by earthworms, and nonencysted forms (trophozoites) are digested preferentially while encysted forms seem to survive passage through the gut (Piearce and Phillips 1980; Rouelle 1983; Barois 1987; Gupta and Doube unpublished). However, not all protozoa are beneficial: Morgan (1988) reported that two species of protozoa (Tetrahymena pyriformis and Proterioochromonas minuta) killed E. fetida within 3 days in a restricted culture environment.

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The role of bacteria in the diet of earthworm species and the extent of species-specific feeding patterns and digestion are largely unknown. Some bacterial species may be digested; others may be little affected and survive passage through the gut; yet others may grow or become more active in the earthworm’s intestine. This phenomenon is discussed further in the following section. In the case of the compost earthworm E. fetida, 19 species of bacteria were digested in the gut, and this dietary addition increased the earthworm’s growth rate significantly (Flack and Hartenstein 1984; Hand et al. 1988). However, Morgan (1988) found that, of 12 bacterial species tested, only two allowed E. fetida to maintain weight; the earthworms either lost weight or died when they fed on the other species. Thorpe et al. (1993) showed that the bacteria Pseudomonas f l uorescens could proliferate in the gut of L. terrestris in the absence of competition from other gut-inhabiting microorganisms. Doube et al. (1994d) showed that rhizobia and Pseudomonas corrugata survived passage through the intestines of A. trapezoides and A. rosea. Different earthworm species may affect the same bacteria in different ways. For example, Schmidt et al. (1997) fed the same concentration of an inoculum of P. corrugata to four species of lumbricid earthworms, but the density of P. corrugata found in the fresh casts of A. longa was tenfold higher than that found in fresh casts of L. rubellus, A. caliginosa, and L. terrestris. On the other hand, bacteria such as Enterobacter aerogenes have been shown to infect and kill the tropical earthworm Hoplochaetella suctoria (Rao et al. 1983), and other species, such as Pseudomonas spp., Streptomyces spp., and Flavobacterium spp., that produce antimicrobial substances have killed the earthworm E. fetida in cultures (Hand et al. 1988). The role of algae and cyanobacteria (blue-green algae) in earthworm diets is still not clear because earthworms can be common in soils that contain a low biomass of algae (e.g., in Mediterranean climate regions), but laboratory trials suggest that these algae can play an important role in the nutrition of some species (e.g., E. fetida; Atlavinyté and Pociené 1973; Piearce 1978; Stamatiadis et al. 1994). The Fate of Microorganisms in the Intestines of Earthworms In the earthworm gut, various enzymes of microbial and earthworm origin are secreted, as well as intestinal mucus (a readily assimilable C source), CaCO3 (if calciferous glands are present), and bacteriostatic and microbicidal substances (including antibiotics of microbial origin and bacteriolysins, peroxidases, and phagocytoses of earthworm origin). All influence the ability of a particular ingested organism to survive passage through the earthworm gut (Brown 1995; Figure 12.1). Hence, different species of bacteria, fungi, protozoa, and algae may be affected in different ways, depending on the species of earthworm and the particular conditions created in their gut and the ability of these organisms to take advantage of, or resist, the gut conditions. The survivors (largely fungal and protozoan spores and resistant bacteria) provide inocula for microbial colonization of the earthworm casts (Dash et al. 1986; Spiers et al. 1986; Brown 1995). The microbial composition of the earthworm intestine contents has been considered to reflect that of the soil or ingested plant remains (Morgan 1988; Brown 1995), but there is evidence of the possible existence of an indigenous, autochthonous gut flora in some earthworm species (Jolly et al. 1993; Vinceslas-Akpa and Loquet 1995; Toyota and Kimura 2000). Furthermore, the numbers, biomass, and activity of the microbial communities in the earthworm gut have also been shown to be different from that in uningested soil (Fischer et al. 1995; Kristufek et al. 1995; Schönholzer et al. 1999). Hence, Lavelle et al. (1995) demonstrated the presence of a mutualistic digestive system in several tropical and temperate earthworms species in which soluble organic C, in the form of a low molecular weight mucus, was added in large quantities (5 to 80% of the dry weight of soil, depending on the species) into the foregut (Trigo et al. 1999; Brown et al. 2000) to “prime” the soil microflora. The mixture of mucus with ingested OM and the high water content and nearneutral pH of the foregut promoted the development of a microflora that could digest cellulose and

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other recalcitrant substances (which earthworms cannot digest by themselves) (Barois and Lavelle 1986; Lavelle and Gilot 1994). Most metabolites were reabsorbed in the earthworm hindgut. The highest relative mucus production (ratio of the soluble C in the foregut to that in the ingested food substrate) was observed in temperate earthworm species, probably because these have greater need to stimulate the microflora (because of lower annual mean temperatures) than those in tropical climates (Trigo et al. 1999). Higher rates of mucus production were also observed when earthworms fed on poorer-quality substrates. Thus, litter-feeding epigeic earthworm species feeding on higher-quality, less-recalcitrant materials (litter, fresh organic materials) may not need to stimulate their intestinal microflora with mucus secretions to the same extent as endogeic species that feed on poorer-quality (soil) material (Trigo et al. 1999). In addition, epigeic species have a more complete enzymatic system (including cellulase) than do endogeic species (Urbásek and Pizl 1991). However, a comprehensive description of the digestive system and the origin of different gut enzymes has so far been made for only seven species (five endogeic, one anecic, and one epigeic), and further research is needed, particularly for epigeic and anecic species (Brown et al. 2000). Ingested actinomycetes may inhibit the growth of other organisms (particularly fungi and Grampositive bacteria) in the earthworm gut by releasing antibiotics, leading to a predominance of antibiotic-resistant Gram-negative bacteria and other actinomycetes in earthworm guts (Kristufek et al. 1993). However, the activity of antibiotics has been tested only in vitro with specific test microorganisms (Ravasz et al. 1986; Kristufek et al. 1993) and has not yet been tested in vivo. Hence, the extent to which such an inhibition actually occurs in the guts of living earthworm is still unknown. Finally, in the microaerophilic guts of some earthworms (e.g., P. corethrurus and A. caliginosa), free-living Clostridia spp. may fix N2 (Barois et al. 1987; Striganova et al. 1989a), although this contribution to the overall N budget in casts is relatively small. Transit time of microorganisms through the earthworm gut may also be an important determining factor of the fate of ingested organisms in earthworm intestines (Scheu 1992; Brown 1995). This varies greatly among different species and furthermore is dependent on the quality of ingested materials (Hendriksen 1991) and on temperature (Barois and Lavelle 1986). Gut transit times range between 1 and 3 hours for Millsonia anomala (Martin et al. 1987), A. rosea (Bolton and Phillipson 1976), E. fetida (F. Hartenstein et al. 1981), A. caliginosa, and O. lacteum (Scheu 1992); 3 to 5 hours for Hormogaster elisae (Diaz Cosín et al. 2002); and maybe 8 hours or longer for L. terrestris (Parle 1963b; R. Hartenstein and Amico 1983), Lumbricus festivus (Hendriksen 1991), and L. rubellus (Daniel and Anderson 1992). In such a short transit, there is little potential for microbial multiplication (although there may be a large increase in activity and “awakening” of dormant bacteria), but with longer gut transit times, there may be sufficient time for microbial multiplication, particularly because many Lumbricus spp. tend to feed on litter or organic-rich materials, which already contain substantial microbial populations. Microbiological Composition and Activity in Earthworm Burrows and Casts Earthworms have high consumption rates, ranging from less than 1 time up to as much as 30 times their body (fresh) weight of soil per day in endogeic species (Lavelle 1988). Litter consumption rates may also be very high, representing from a small percentage up to more than 85% of the annual litter fall in sites containing large populations of anecic and epigeic earthworms (e.g., Knollenberg et al. 1985). However, assimilation efficiencies are generally low in endogeic species, ranging from only about 1% of the ingested C for A. rosea (Bolton and Phillipson 1976) and 3 to 19% for M. anomala and P. corethrurus (Martin and Lavelle 1992; Lavelle and Spain 2001). On the other hand, assimilation efficiencies may be higher (30 to 75%) in litter-feeding species such as L. rubellus and L. terrestris (Dikschen and Topp 1987; Daniel 1991). Therefore, earthworm casts may have large amounts of OM that is not assimilated but may have been modified greatly by gut

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passage, both physically (by comminution, restructuring, movement from one location to another in soil) and chemically (reduction in C:N ratio, change in quality). As mentioned, when an earthworm ingests soil containing “dormant” microorganisms, these microorganisms may become activated by a “priming effect” (Lavelle et al. 1995) in the gut, which often continues for a short time in the casts because of the abundance of soluble C and other nutrient resources. Thus, in fresh casts of A. caliginosa, Scheu (1987) reported an increase of approximately 90% in microbial respiration rates over a 4-week incubation period. Given the higher microbial activity and biomass often observed in earthworm casts (Edwards and Bohlen 1996), higher OM decomposition rates in casts would be expected. However, the OM may also be “protected” from microbial attack in compact, water-stable castings of endogeic earthworm species, leading to lower OM decomposition rates (Martin 1991; Ketterings et al. 1997; Marinissen and Hillenaar 1997). Preferential assimilation by earthworms of the more labile soil OM fractions may also help explain the lower mineralization rates in casts because the remaining (more recalcitrant) OM in casts will have slower mineralization kinetics (McCartney et al. 1997). Thus, it is the balance of the short-term priming effects and the longer-term protection effects induced by an earthworm community in a given location that may ultimately determine whether earthworm activities are enhancing the conservation or destruction of OM in the drilosphere (Brown et al. 2000). Nevertheless, in the drilosphere, the short-term priming of microorganisms increases nutrient mineralization rates, releasing more plant-available N and P (e.g., Edwards and Bohlen 1996; Barois et al. 1999). Therefore, in terms of nutrients, earthworm casts and burrows (lined with carbon and protein-rich mucus with a C:N ratio below 6.0; Cortez and Bouché 1987; Scheu 1991; Schmidt et al. 1999) constitute very favorable microenvironments for microbial and invertebrate activity and for plant root growth (Graff 1971; Tiunov and Scheu 1999; Jégou et al. 2001). Consequently, populations and activity of microflora, microfauna, and other organisms may be higher in earthworm burrows and casts than in surrounding soil (Brown 1995; Tiunov and Scheu 1999). For instance, in a temperate grassland with a complex earthworm community, burrow walls supported up to 42% of the total soil aerobic N2-fixing bacteria, 13% of anaerobic N2 fixers, and 16% of denitrifying bacteria and had more ammonifiers, denitrifiers, free-living aerobic and anaerobic N2 fixers, and proteolytic bacteria (Bhatnagar 1975). Earthworm casts had more cellulolytic aerobes and hemicellulolytic, amylolytic, nitrifying, and denitrifying bacteria than the soil in which they lived (Bhatnagar 1975; Loquet et al. 1977). Therefore, earthworm casts and burrows may be important microsites for denitrification because they possess larger populations of denitrifiers, higher levels of soluble C and NO3 , and higher water contents than the surrounding soil (e.g., Elliott et al. 1990; Karsten and Drake 1997; Parkin and Berry 1999). Several species of fungi can also grow rapidly in earthworm casts (Parle 1963a) from the inocula remaining after passage through the gut (surviving spores and hyphal fragments) or from the surrounding soil. Algae also appear to be able to take advantage of the high nutritional value of earthworm casts. Of 19 species of algae found in casts of an unidentified earthworm species (probably a lumbricid), 6 had higher growth rates in casts than in uningested soil (Shtina et al. 1981). Protozoa (live and encysted) that survive passage through the earthworm gut may also feed on the increased numbers of bacteria and fungi found in earthworm casts and multiply rapidly, so that their numbers become higher than in uningested soil (Shaw and Pawluk 1986). Successional Processes within Casts Passage through the earthworm intestine results in the removal of some of the active stages of protozoans and some (less resistant) fungal spores, hyphae, and bacteria. The surviving, resistant microorganisms, together with those found in the earthworm burrow walls, provide inocula for colonization of the newly formed casts. The microbial successional processes in casts have been little studied, and although a number of authors have detailed the type and abundance of microorganisms found in casts (Brown 1995), © 2004 by CRC Press LLC

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there has been little reference to changing patterns of relative abundance in casts over time following their deposition. Tiunov and Scheu (2000) and Orazova et al. (2003) reported that the microfungal community of L. terrestris casts was significantly different from that of the surrounding soil and the litter and consisted of a mixture of both litter- and soil-inhabiting species. The abundance of some soil and litter fungal species decreased (particularly early colonizers of litter, e.g., Alternaria, Cladosporium); the number of others increased (especially Trichoderma spp.). During aging of the casts, the microfungal community changed, with some species decreasing and others increasing their relative dominance; other, new species began to colonize the casts, thereby increasing the fungal species diversity (Tiunov and Scheu 2000). Unpublished data (Gupta and Doube) on the composition of casts of A. trapezoides showed that fresh casts contain numerous bacteria but no active protozoa or fungal hyphae. During the 4 weeks after cast deposition, a succession of fungal and protozoan species appeared in the casts. From the above discussion, it is clear that there is still much to learn about earthworm nutrition and the fate of microorganisms in earthworm intestines and their subsequent development in the casts. Microbial species composition, metabolic activity, and microbial successional processes in casts are substantially different from the surrounding soil. As a consequence, when earthworms process substantial amounts of soil or OM, their activity can have a major influence on the form and scale of microbial processes in soil.

MESOSCALE INTERACTIONS

BETWEEN

EARTHWORMS

AND

MICROORGANISMS

In this section, we consider the ways in which the activity of earthworms influences the average composition and functioning of microbial communities in small patches of soil and the possibility that the composition of microbial communities at this scale influences earthworm populations. These effects are usually examined using microcosms or small-scale field trials. In studies aiming to understand the complex soil changes associated with the invasion of the Canadian Rocky Mountains by earthworms, Scheu and Parkinson (1994) showed, using uniform soil in microcosms, that the earthworm Dendrobaena octaedra decreased the microbial biomass (measured by substrate-induced respiration, SIR), whereas in layered soil, the activity of both D. octaedra and O. lacteum redistributed organic residues into the deeper soil layers, resulting in increased microbial biomass at these depths. In contrast, in the presence of both earthworm species, there was a decrease (ca. 40% after 12 weeks) in the overall microbial biomass of the organic-rich surface material, and in most situations, there was a marked increase in the bacterial:fungal ratios (bacteria accounted for 20 to 27% of the biomass compared with 10 to 17% in controls). In the same study (Scheu and Parkinson 1994), earthworm activity induced only minor immediate changes in the composition (species dominance) of the fungal community after 8 weeks, but such changes, if compounded over several years, could result in major changes in fungal community composition. In the deepest soil layer, after 8 and 12 weeks, bacterial:fungal ratios increased in the presence of earthworms, with bacteria making up 60 to 70% of the biomass compared with 50 to 55% in the earthworm-free soil. Similar earthworm-induced increases in bacterial:fungal ratios in soil have been demonstrated for several species by Parle (1963a) and Ausmus (1977) and for A. caliginosa by Wolters and Joergensen (1992) and Kollmannsperger (1952) (Scheu and Parkinson 1994). Scheu and Parkinson’s experiments were conducted in microcosms and were limited to 12 weeks. Nevertheless, over longer periods of time, the colonization of forest soils by litter- and soilfeeding earthworms had profound effects on the composition and functioning of the soil microbial community (see McLean and Parkinson 1997a, 1998, 2000). These effects were achieved through the redistribution of organic surface residues, enhanced disappearance of the H and F layers, reductions in microbial biomass, increasing relative dominance of particular fungal species (particularly rapid-growth fungi), favoring of bacteria, and stabilization of C in earthworm castings. This process accelerates the formation of mull soil with an AH layer rich in humus and covered by a thin layer of fragmented material (McLean and Parkinson 1997a,b). © 2004 by CRC Press LLC

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Other studies have shown that the microbial species favored by earthworm activity differ with environmental conditions and the earthworm community present. For example, feeding by O. lacteum on partly decomposed OM in peat-humus forests in Russia appeared to benefit sporeforming bacteria (e.g., Bacillus spp. that decompose recalcitrant substances), to the detriment of fluorescing bacteria, which generally use fresh and easily decomposed OM (Kozlovskaya and Zhdannikova 1961). In contrast, when L. rubellus fed on fresh substrates, fluorescing bacteria were favored over Bacillus spp. in podzolized soils, O. lacteum had the same effect on microbial development as L. rubellus because in these soils O. lacteum came to the surface and fed on newer materials. In the peat-humus soils later in the season (summer), the absence of fresh litter residues caused L. rubellus to feed on older residues and have an effect similar to that of O. lacteum in the same soil (Kozlovskaya and Zhdannikova 1961). Microbial and invertebrate successions on decomposing plant debris and other substrates are modified by earthworm-induced changes in the quality (physicochemical nature) of these resources. This occurs through ingestion and comminution of organic residues by the earthworms changing its C:N ratio and physicochemical characteristics, by creating heterogeneous microsites within the soil, and by grazing selectively on and dispersing particular organisms. For instance, in apple orchards (Raw 1962) and temperate deciduous forests (Knollenberg et al. 1985), it was reported that L. terrestris may bury more than 90% of the annual litterfall. Thus, in apple orchards, up to 22 species of fungi and 5 species of insects inhabiting the surface litter were buried and their populations reduced (Mills 1976; Niklas and Kennel 1981; Laing et al. 1986; Kennel 1990). Under these conditions, distinct changes in the microfloral succession are also likely, not only because some species are buried and their populations reduced, but also because the comminution of litter by anecic earthworm species (and also epigeic and some surface-feeding endogeic species) favors the development of r-selected (fast growing) fungi such as Phycomycetes (e.g., Mortierella and Mucor spp.), Ascomycetes, and Deuteromycetes (e.g., Phoma and Trichoderma spp.) capable of rapid exploitation of the easily assimilable materials found in earthworm casts (Visser 1985; Tiunov and Scheu 2000; Orazova et al. 2003). In particular, middens created by anecic species of earthworms are important sites of enhanced (hot spots) microbial and faunal activity and populations, and the presence of a large anecic population may enhance soil surface and subsurface heterogeneity (the spatiotemporal distribution of resources), thereby altering the spatial and temporal organization of soil communities (Brown 1995; Maraun et al. 1999; Tiunov and Kuznetsova 2000; Shuster et al. 2001). Nevertheless, selective grazing (or burial) on fast-growing fungi by earthworms (see previous sections) may reduce their competitive ability and allow slower-growing (k-selected) fungi such as Basidiomycetes to gain a competitive advantage. For instance, Scheu (1992) reported that lignindecomposing fungi (which occur later in the succession) decomposed lignin in earthworm feces only after a lag phase of 3 months. In a later experiment, Scheu (1993) observed an overall increase in lignin mineralization by a factor of 1.1 in soil columns with O. lacteum and 1.2 for those with L. castaneus. Selective grazing may also lead to an increase in populations of the selected microbial species in earthworm structures and, occasionally, in total fungal diversity as well (Tiwari and Mishra 1993). Similar processes may also apply for algae and protozoa. For instance, Gupta and Doube (unpublished data) recovered over 30 species of protozoa from the casts of field-collected A. trapezoides, a substantially increased level of diversity compared with surrounding soil on a gram-for-gram basis. In a mesocosm study, using a simulated forest floor and a combination of various soil invertebrates, Huhta et al. (1991) reported greater N mineralization in the presence of L. rubellus than by complex fauna in the absence of earthworms. Similarly, in lysimeters with different combinations of animals, Anderson et al. (1983) observed increased losses of Na, K, Ca, and mineral N from oak leaf litter and up to 60 times greater NH4 losses in the presence of L. rubellus than in its absence. These responses were presumed to be caused by earthworm-induced changes in microbial activity, although the organisms responsible were not characterized. © 2004 by CRC Press LLC

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Earthworm activity has also been shown to influence the distribution and activity of bacteria, fungi, and protozoa in soils. In laboratory experiments, the earthworm A. trapezoides was fed sheep manure containing Rhizobium trifolii in pots in which subterranean clover was growing (Doube et al. 1994c). Earthworm activity dispersed the surface-applied rhizobia through the soil, resulting in a fivefold increase in the total number of root nodules on the clover plants and a four- to sixfold increase in the number of nodules on the primary roots 2 to 8 cm below the surface. Stephens et al. (1994d) further showed that both A. trapezoides and M. dubius could disperse Rhizobium melliloti through soil, and root colonization of alfalfa by this species increased more than 100-fold in the presence of A. trapezoides. However, the survival of R. melliloti in soil in pots was related inversely to the earthworm population density and, at the highest density tested, was reduced by 99% after 40 days in the presence of A. trapezoides. The reasons for this are unclear, and these results need to be confirmed in the field before there is confidence that such processes occur in natural environments. The majority of root diseases of agricultural crops are caused by soil-borne fungi, and the possibility that earthworm-dispersed biological control agents (bacteria and fungi) can enhance the biological control of these fungal diseases prompted several studies in Australia (e.g., Doube et al. 1994d,e). One such study concerned the bacterium P. corrugata, which has been shown to control “take-all,” a serious fungal root disease of wheat plants caused by Gauemannomyces graminis var. tritici. Using columns filled with soil, Doube et al. (1995) showed that both A. trapezoides and A. rosea dispersed P. corrugata up to 20 cm in 8 days, and that large numbers (ca. log 6 g−1) were recovered from both the earthworm casts and the burrow walls. This work was supported by Stephens et al. (1993), who placed straw pellets (inoculated with P. corrugata) on the soil surface in the presence and absence of A. trapezoides. In the presence of the earthworm, the bacteria were dispersed through the soil (>log 4 g−1 soil at 9 cm after 9 days), and there was a substantial increase in the level of colonization of the roots of seedling wheat plants by the bacteria. However, attempts to evaluate experimentally the capacity of this earthworm-dispersed inoculum to control take-all have been frustrated because the earthworm presence alone also controlled the disease (Doube et al. 1994d). Stephens et al. (1994a,b) thoroughly examined such interactions in greenhouse and field studies using the earthworms A. trapezoides and A. rosea, the two most common species in the cropping soils of southern Australia. Both species caused significant reductions in the severity of the disease symptoms (root lesions) in both laboratory and field trials. These effects were also observed in two soils of contrasting texture (a red-brown earth and a calcareous sandy loam). Reduced severity of disease was associated with a corresponding increase in plant growth, but in some cases, the earthworm activity had no effect on plant growth. In pot trials, the level of disease control increased with higher earthworm populations, but in the field trials, the level of disease control (30 to 40% reduction in lesions) with the equivalent of 100 earthworms m−2 did not increase by tripling the earthworm population (300 m−2). Similar results were obtained for the effects of the activity of the earthworm A. trapezoides on the bare patch root disease of wheat caused by Rhizoctonia solani (Stephens et al. 1994c,e; Stephens and Davoren 1995, 1997). Laboratory and field trials demonstrated reduced numbers of root lesions (20 to 40% reduction) and, by inference, reduced activity of Rhizoctonia species in soil. Again, earthworm populations of 100 m−2 and 300 m−2 resulted in similar degrees of disease suppression. Why additional earthworm activity failed to increase the level of control of root disease is unknown. Although the earthworm populations used in these experiments were greater than those commonly found in cropped soils in southern Australia, it seems likely that earthworms have potential to help control root diseases of crops under some conditions. For example, Doube et al. (1994a) reported A. trapezoides at populations of 410 m−2 and 140 m−2 in soils under canola and wheat, respectively, in New South Wales, Australia. Moreover, from their results, it seems probable that earthworm activity can modify the microenvironment in soils in such a manner that its suitability for root-associated fungi (including pathogens) is affected. This has important © 2004 by CRC Press LLC

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POTENTIAL DECREASE IN PLANT YIELDS

Increased potential for plant damage

Dispersal of plant pathogenic fungi, bacteria, and parasitic nematodes

Increase in activity and populations of plant pathogens and parasites

EARTHWORMS

Litter and organic matter Decrease in activity breakdown, consumption, and populations of and digestion of plant pathogens microorganisms and parasites

Nutrient release and availability

Increased activity and dispersal of litter decomposing fungi and bacteria, microbivores, mycorrhizal fungi, N2 fixers, and biocontrol agents

Reduced potential for plant damage

POTENTIAL INCREASE IN PLANT YIELDS FIGURE 12.4 Functional interactions between earthworms and beneficial and adverse microorganisms affecting plant growth. (Modified from Edwards and Fletcher, 1988.)

consequences for the overall composition of the microbial community and for plant productivity (Figure 12.4). Clearly, earthworm activity influences microbial communities, potentially modifying populations and activity of both plant pathogens and beneficial microorganisms. If earthworm activities lead to increases in the populations and activity of litter-decomposing and mineralizing microbes, mycorrhizal fungi, N2 fixers, biocontrol agents, and mesofauna, then the consequences for plant productivity are likely positive. On the other hand, if earthworms disperse plant pathogenic microbes and increase their populations or activity, the potential for damage to plants by these microorganisms may increase. Although it is known that earthworm activity can affect microbial communities, the opposite (i.e., the influence of microorganisms and their activities on earthworm behavior and abundance) is not well known, and available evidence is somewhat circumstantial. Many species of fungi, bacteria, protozoa, and nematodes can parasitize earthworms, both internally and externally, often killing them and probably aiding in the rapid decomposition of earthworm tissue (Poinar 1978; Segun 1978; Dash et al. 1979b; Rao et al. 1983). On the other hand, high microbial activity in soil microsites, particularly the rhizosphere, attracts small invertebrates such as protozoa and nematodes and may also attract earthworms in a process that is still not well understood (Bonkowski and Schaefer 1997; Lavelle and Spain 2001). For example, endogeic earthworms were reported to be more abundant in the root zones of wheat plants (Rovira et al. 1987), corn © 2004 by CRC Press LLC

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FIGURE 12.5 Grazing by A. trapezoides on rhizosphere soil surrounding the roots of wheat seedlings: (left) controls showing ungrazed rhizosphere soil attached to the root; (right) bare roots exposed following removal of rhizosphere soil by earthworm feeding. (Photograph by J. Coppi.)

plants (Binet et al. 1997), and sugar cane (Spain et al. 1990) than in adjacent soils. Doube (unpublished data) found a population of about 100 earthworms m−2 (a mixture of A. trapezoides, A. rosea, and M. dubius) among the roots of Salvia verbenaca, whereas none were recovered from Salvia-free soil 1 m away. Root herbivory by the anecic earthworm L. terrestris has been inferred from pulse-labeling experiments (Cortez and Bouché 1992), from direct observation (Carpenter 1985; Shumway and Koide 1994), and indirectly by the assessment of root fragments in earthworm guts or casts. For example, Reddell and Spain (1991a) observed root fragments in the fresh casts of P. corethrurus. Similarly, Baylis et al. (1986) suggested that earthworms graze on living clover roots. In contrast, studies on endogeic earthworm species in a rhizotron (Gunn and Cherret 1993) and on A. trapezoides (Figure 12.5) suggest that endogeic species can consume the rhizosphere soil (possibly containing root hairs) but do not eat the roots. In both cases, the rhizosphere microorganisms will form at least part of the earthworm’s diet, and so earthworm activity has the potential to influence rhizosphere function. Endogeic earthworm species also tend to accumulate in soil patches with higher levels of OM (Hughes et al. 1994). Because plant roots and organic residues are both localities of relatively high levels of microbial activity, and earthworms recognize and consume some microorganisms selectively, it seems probable that endogeic earthworm species aggregate in such areas in response to microbial activity. Whether such behavior enhances the earthworm reproductive performance (and corresponding abundance) has not been demonstrated, but the positive links between organic residues, microbial activity, and earthworm food make such an association very likely.

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MICROORGANISMS

When CC cropping soils are converted to NT or DD fields, a change in the composition of the soil microbial communities occurs, with increasing fungal:bacterial ratios (Brussard et al. 1990; Coleman et al. 1994). Earthworms are also more abundant under NT (DD) than under CC (Chan 2001), and this has been attributed in part to changes in the composition of the decomposer community. Although such causal relationships can be inferred readily from observed changes in earthworm community composition, experimental field evidence supporting such conclusions is scarce. Several experiments that manipulated earthworm populations under field conditions have cast some light on this relationship. For instance, in small field plots in Ohio, Blair et al. (1997) observed a greater microbial biomass N in treatments with an experimentally reduced (by electroshocking) earthworm population, and Hendrix et al. (1998) reported lower microbial biomass N but higher turnover rates in earthworm addition treatments in Georgia. These results seem to imply that earthworm activity exerts a regulatory influence on microbial communities, resulting in a lower microbial biomass that is more metabolically active and has higher turnover rates (Edwards and Bohlen 1996). There is much evidence indicating the involvement of earthworm activity in the recycling of C, N, and P in soils (Syers et al. 1979a,b; Hendrix et al. 1987; Bostrom 1988; Parmelee et al. 1998; Chapter 8, this volume), which, by implication, suggests that earthworms also modify soil microbial functions. For instance, as much as 60% of the C losses from earthworms during their life span can be in the form of mucus secretions (soluble organic C) that act as important microbial stimulants (Scheu 1991). Carbon losses caused by earthworm respiration are generally not large but can be as much as 29% of the total heterotrophic respiration in DD systems (Hendrix et al. 1987). In an extensive study of the role of earthworms in C recycling in a Swedish alfalfa field, Bostrom and Lofs-Holmin (1988) reported that C flow through a population of A. caliginosa (with biomass of 33 kg C ha−1) in 1 year totaled 3.8 t C ha−1 in consumption, 3.8 t C ha−1 in fecal egestion (casts), 47 kg C ha−1 lost in respiration, 41 kg C ha−1 in dead tissues, 10 kg C ha−1 in cocoon production, and 40 kg C ha−1 in tissue production. Mucus production was not assessed in this study. Earthworm invasion (or newly inoculated) sites, in particular, provide the opportunity to study the effects of earthworms on elemental cycles without confounding effects of long-term previous earthworm populations. The results of several studies (Stout 1983; Alban and Berry 1994; Burtelow et al. 1998) showed that earthworms can induce rapid and large C and N losses, but that these are unlikely to continue indefinitely. More likely, a new equilibrium will be reached once the earthworm invasion process has ended and stabilized, in which case the total soil C stocks are lower, but with certain fractions and particle sizes of OM turning over more rapidly than in the previously uninvaded soils (Brown et al. 2000). Nitrogen flows through earthworm populations have been estimated by several authors for various agroecosystems (Marinissen and de Ruitter 1993); in the alfalfa field studied by Bostrom and Lofs-Holmin (1988), N flows reached as much as 516 kg N ha−1 year−1. Of this total, 10 kg N ha−1 were for tissue production, 2 kg N ha−1 were in excretions (mostly in plant-assimilable N forms), and 504 kg N ha−1 were in casts. Priming effects and indirectly increased N cycling (because of consumption and activation of microorganisms and invertebrates) are important but difficult to assess. In the experimental earthworm population reduction/increase experiment in Ohio, the addition of four earthworm species (mainly L. terrestris and A. tuberculata, but also L. rubellus and A. trapezoides; Bohlen et al. 1995) increased amounts of extractable mineral N (NH4 and NO3), significantly depending on the treatment (Blair et al. 1997). However, the addition of deep-burrowing earthworms (L. terrestris) to a nearby field site had no effect on mineral N concentrations but increased the dissolved organic N, potentially mineralizable N, and microbial biomass N, depending on the earthworm treatment (Subler et al. 1997). The increased soil leachate volumes (4- to 12fold greater) observed at this site, probably because of increased macropore flow (Lachnicht et al. 1997), significantly increased the leaching of dissolved organic N (Subler et al. 1997). Such different results obtained regarding influences of earthworms on soil microbial biomass and mineral N © 2004 by CRC Press LLC

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contents imply site- (spatially) or temporally specific effects arising from different soil types, management practices (OM or fertilizer additions, cropping systems), sampling dates, and composition of the earthworm communities. Effects of earthworms on denitrification are also important. For instance, Knight et al. (1992) reported three to five times greater denitrification rates in earthworm casts than in surrounding soil, contributing to as much as 29% of the total denitrification losses in a heavily N-fertilized (200 kg N ha−1) pasture in Devon, U.K. Although earthworms have very little direct effect on P cycling in soils (only trace quantities of P are excreted in liquid wastes) (Bahl 1947), they may have substantial indirect effects because of enhanced phosphatase (acid and alkaline) activity in earthworm casts, burrows, and guts; increased availability of P (organic and inorganic) in casts; reduction in Al-binding of P; and increased incorporation of surface-applied fertilizers and plant litter in the soil (Satchell and Martin 1984; Syers and Springett 1984; Alter and Mitchell 1992). Thus, as much as 9 to 13 kg ha−1 of organic plus inorganic P were accumulated in earthworm casts in 1 year in New Zealand pastures (Syers et al. 1979b). These casts, containing more P than does the surrounding soil, form microsites rich in P that can enhance microbial activity and root growth (Mouat and Keogh 1987), particularly in soils poor in phosphorus. The size and composition of earthworm populations, even within one locality, vary widely with soil type and management practices (e.g., cultivation or liming) (Lee 1985; Robinson et al. 1992), and it is possible that such population differences may reflect differences in the composition and abundance of the microbial communities that form the basis of the decomposer food web in soil. For example, it is widely recognized that microbial populations are influenced strongly by soil pH (Gupta 1994), and some species of earthworms are sensitive to soil pH and are more abundant in neutral to slightly acid soil (Lofs-Holmin 1986). However, the relationships between such changes in microbial populations and the reproductive success of earthworms are still not known. The type of plant community and the corresponding litter-soil-rihzosphere conditions can also have large-scale effects on earthworm populations. For instance, Boettcher and Kalisz (1991) reported changes in the earthworm community structure (abundance and species composition) in a vegetation sequence involving hemlock, rhododendron, and yellow poplar, with one earthworm species (Bimastus parvus) largely replacing another (Komarekiona eatoni) along the sequence. Similarly, in agroecosystems, earthworms appear to be more abundant in soils under certain crops (e.g., clover) (Weternacher and Graff 1987) or under certain grasses in meadows (Babel et al. 1992) or cereals (Edwards and Bohlen 1996). The possibility that earthworms may be used to introduce and disperse beneficial microorganisms through soil was reviewed by Doube et al. (1994e). They examined the possibility of using pellets of a mixture of earthworm food and beneficial microorganisms (rhizobia for root nodulation, pseudomonads for biological control of take-all, and Metarhizium sp. for the biological control of root-feeding scarab larvae). These microorganisms could be applied to the field and dispersed through soil as a consequence of the feeding activities of earthworms. The success of this process relies on developing a food that is attractive to earthworms, on survival of the microorganism in the food pellet and during passage through the earthworm gut, and on effective earthworm dispersal through the soil. Although only in the early stages of development, a number of these constraints have been examined in laboratory experiments, and Doube et al. (1994e) considered that this novel mechanism showed considerable promise. Success would require that earthworm activity alter the composition and functioning of microbial communities in soils on a broad scale.

SUMMARY AND CONCLUSIONS The microbial decomposition of organic residues in soil provides the energy and nutrients that promote and sustain the biological fertility of soils. Surface crop residue inputs into agricultural systems in temperate and subtropical regions are commonly on the order of 2 to 10 tons ha−1 year−1, and similar © 2004 by CRC Press LLC

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or even greater amounts enter the soil as dead roots or root exudates. Between 60 and 75% of the introduced C is lost through respiration during the first year in the soil. The microbial biomass in these systems is commonly on the order of 0.5 to 5.0 tons ha−1, and this comprises only 1 to 2% of the total soil carbon pool. The microalgae and the primary decomposers (bacteria, actinomycetes, fungi) provide the basis of the food web in soils and provide food for a wide diversity of secondary consumers (predators), including protozoa and earthworms. Earthworms also feed selectively on soil rich in organic materials (organic fragments and the rhizosphere) and gain some nutrients by digesting microorganisms or their by-products associated with the decomposition of organic materials. Furthermore, earthworms selectively consume some microbial species and digest fungal hyphae, trophozoite protozoa, algae, and some bacteria and disperse the survivors (spores and resistant structures) throughout the soil. Such dispersed survivors include bacteria (pseudomonads and rhizobia), fungi, and protozoa. As an additional adaptive strategy to live in poor soil environments, earthworms have developed a mutualistic digestive system with gut bacteria that allows them to obtain nutrients from the soil organic matter, their main source of food. Earthworms create (through selective feeding, casting, and burying organic residues) zones in soil of high microbial activity (e.g., plant roots and residue patches) and aggregate in these zones. In addition, their activity has been shown to decrease the severity of plant root diseases as well as disperse and increase populations or activity of some beneficial soil microorganisms such as Rhizobia and plant growth-promoting rhizo-bacteria (see Chapter 2). Because of these activities, earthworms, if present in moderate populations, are likely to have a substantial influence on the distribution, composition, and activity of the microbial communities responsible for the decomposition of organic residues and for the regulation of plant growth, and thus play a key role in regulating the processes that maintain soil health.

REFERENCES Alban, D.H. and E. Berry. 1994. Effects of earthworm invasion on morphology, carbon and nitrogen of a forest soil, Appl. Soil. Ecol., 1:243–249. Alter, D. and A. Mitchell. 1992. Use of vermicompost extract as an aluminum inhibitor in aqueous solutions, Commun. Soil Sci. Plant Anal., 23:231–239. Anderson, J.M. 1988. Spatiotemporal effects of invertebrates on soil processes, Biol. Fertil. Soils, 6:216–227. Anderson, J.M. 1994. Functional attributes of biodiversity in land use systems, in D.J. Greenland and I. Szabolcs, Eds., Soil Resilience and Sustainable Land Use, CAB International, Wallingford, U.K., pp. 267–290. Anderson, J.M., P. Ineson, and S.A. Huish. 1983. Nitrogen and cation release by macrofauna feeding on leaf litter and soil organic matter from deciduous woodlands, Soil Biol. Biochem., 15:463–467. Andren, O., T. Lindberg, K. Paustian, and T. Rosswall. 1990. Ecology of Arable Land: Organisms, Carbon and Nitrogen Cycling, Vol. 40, Munksgaard, Copenhagen. Andren, O., K. Paustian, and T. Rosswall. 1988. Soil biotic interactions in the functioning of agroecosystems, Agric. Ecosyst. Environ., 24:57–65. Atlavinyté, O. and C. Pociené. 1973. The effect of earthworms and their activity on the amount of algae in the soil, Pedobiologia, 13:445–455. Ausmus, B.S. 1977. Regulation of wood decomposition rates by arthropod and annelid populations, in U. Lohm and T. Persson, Eds., Soil Organisms as Components of Ecosystems, Vol. 25, Ecological Bulletins, Stockholm, Sweden, pp. 180–192. Babel, U., O. Ehrmann, and M. Krebs. 1992. Relationships between earthworms and some plant species in a meadow, Soil Biol. Biochem., 24:1477–1481. Bahl, K.N. 1947. Excretion in the Oligochaeta, Biol. Rev. Cambridge Philos. Soc., 22:109–147. Barois, I. 1987. Interactions Entre les Vers de Terre (Oligochaeta) Tropicaux Géophages et la Microflore pour l’Exploitation de la Matière Organique du Sol. Travaux des Chercheurs de la Station de Lamto, Vol. 7. Publication of the Laboratoires de Zoologie de l’ENS, Paris.

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Barois, I., P. Cadet, A. Albrecht, and P. Lavelle. 1988. Systèmes de culture et faune des sols. Quelques données, in C. Feller, Ed., Fertilité des Sols Dans les Agricultures Paysannes Caribéenes. Effets des Restitutions Organiques, ORSTOM-Martinique, Paris, pp. 85–95. Barois, I. and P. Lavelle. 1986. Changes in respiration rate and some physicochemical properties of a tropical soil during transit through Pontoscolex corethrurus (Glossoscolecidæ, Oligochæta), Soil Biol. Biochem., 18:539–514. Barois, I., P. Lavelle, M. Brossard, J. Tondoh, M.A. Martínez, J.P. Rossi, B.K. Senapati, A. Angeles, C. Fragoso, J.J. Jiménez, T. Decaëns, C. Lattaud, J. Kanyonyo, E. Blanchart, L. Chapuis-Lardy, G.G. Brown, and A.G. Moreno. 1999. Ecology of species with large environmental tolerance and/or extended distributions, in P. Lavelle, L. Brussaard, and P.F. Hendrix, Eds., Earthworm Management in Tropical Agroecosystems, CAB International, Wallingford, U.K., pp. 57–85. Barois, I., B. Verdier, P. Kaiser, A. Mariotti, P. Rangel, and P. Lavelle. 1987. Influence of the tropical earthworm Pontoscolex corethrurus (Glossoscolecidae) on the fixation and mineralization of nitrogen, in A.M.B. Pagliai and P. Omodeo, Eds., On Earthworms, Mucchi Editore, Modena, Italy, pp 151–158. Baylis, J.P., J.M. Cherrett, and J.B. Ford. 1986. A survey of the invertebrates feeding on living clover roots (Trifolium repens L.) using 32P as a radiotracer, Pedobiologia, 29:201–208. Bhatnagar, T. 1975. Lombriciens et humification: un aspect nouveau de l’incorporation microbienne d’azote induite par les vers de terre, in G. Kilbertus, O. Reisinger, A. Mourey, and J.A.C. da Fonseca, Eds., Humification et Biodégradation, Pierron, Sarreguemines, France, pp 169–182. Binet, F., V. Hallaire, and P. Curmi. 1997. Agricultural practices and the spatial distribution of earthworms in maize fields. Relationships between earthworm abundance, maize plants and soil compaction, Soil Biol. Biochem., 29:577–583. Blair, J.M., R.W. Parmelee, M.F. Allen, D.A. McCartney, and B.R. Stinner. 1997. Changes in soil N pools in response to earthworm population manipulations in agroecosystems with different N sources, Soil Biol. Biochem., 29:361–367. Boettcher, S.E. and P.J. Kalisz. 1991. Single-tree influence on earthworms in forest soils in eastern Kentucky, Soil Sci. Soc. Am. J., 55:882–865. Bohlen, P.J., R.W. Parmelee, J.M. Blair, C.A. Edwards, and B.R. Stinner. 1995. Efficacy of methods for manipulating earthworm populations in large-scale field experiments in agroecosystems, Soil Biol. Biochem., 27:993–999. Bolton, P.J. and J. Phillipson. 1976. Burrowing, feeding, egestion, and energy budget of Allolobophora rosea (Savigny) (Lumbricidae), Oecologia (Berlin), 23:225–245. Bonkowski, M., B.S. Griffiths, and K. Ritz. 2000. Food preferences of earthworms for soil fungi, Pedobiologia, 44:666–676. Bonkowski, M. and M. Schaefer. 1997. Interactions between earthworms and soil protozoa — a new component in the soil food web, Soil Biol. Biochem., 29:499–502. Boström, U. 1988. Ecology of Earthworms in Arable Land: Population Dynamics and Activity in Four Cropping Systems, Swedish University of Agricultural Sciences, Uppsala, Sweden. Boström, U. and A. Lofs-Holmin. 1988. Earthworm population dynamics and flows of carbon and nitrogen through Aporrectodea caliginosa (Lumbricidae) in four cropping systems, in Ecology of Earthworms in Arable Land: Population Dynamics and Activity in Four Cropping Systems, Institutionen for ekologi och miljovard, Report no. 34, Swedish University of Agricultural Sciences, Uppsala, Sweden. Bouche, M.B. 1977. Strategies lombriciennes, in U. Lohm and T. Persson, Eds., Soil Organisms as Components of Ecosystems. Ecological Bulletins (Stockholm), 25:122–132. Bowen, R.M. 1990. Decomposition of wheat straw by mixed cultures of fungi isolated from arable soils, Soil Biol. Biochem., 22:401–406. Bowen, R.M. and S.H.T. Harper. 1989. Fungal populations on wheat straw decomposing in arable soils, Mycol. Res., 93:47–54. Bowen, R.M. and S.H.T. Harper. 1990. Decomposition of wheat straw and related compounds by fungi isolated from straw in arable soils, Soil Biol. Biochem., 22:393–399. Brown, G.G. 1995. How do earthworms affect microfloral and faunal community diversity? Plant Soil, 170:209–231. Brown, G.G., I. Barois, and P. Lavelle. 1998. Drilosphere effects on soil organic matter dynamics and microbial activity: From priming to regulation, in Proceedings of the XVIth World Congress of Soil Science, IUSS, Montpellier, France, CD-Rom.

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Scheu, S. 1992. Automated measurement of the respiratory response of soil microcompartments: active microbial biomass in earthworm faeces, Soil Biol. Biochem., 24:1113–1118. Scheu, S. 1993. Litter microflora-soil macrofauna interactions in lignin decomposition: a laboratory experiment with 14C-labelled lignin, Soil Biol. Biochem., 25:1703–1711. Scheu, S. and D. Parkinson. 1994. Effects of earthworms on nutrient dynamics, carbon turnover and microorganisms in soil from cool temperate forests on the Canadian Rocky Mountains-laboratory studies, Appl. Soil Ecol., 1:113–125. Schmidt, O., B.M. Doube, M.H. Ryder, and K. Killham 1997. Population dynamics of Pseudomonas corrugata 2140R lux8 in earthworm food and in earthworm casts, Soil Biol. Biochem., 29:523–528. Schmidt, O., C.M. Scrimgeour, and J.P. Curry. 1999. Carbon and nitrogen stable isotope ratios in body tissue and mucus of feeding and fasting earthworms (Lumbricus festivus), Oecologia, 118:9–15. Schönholzer, F., D. Hahn, and J. Zeyer. 1999. Origins and fate of fungi and bacteria in the gut of Lumbricus terrestris L. studied by image analysis, FEMS Micbrobiol. Ecol., 28:235–248. Segun, A.O. 1978. Monocystid gregarine parasites of Nigerian earthworms, J. Protozool., 25:157–162. Shaw, C. and S. Pawluk. 1986. Faecal microbiology of Octolasion tyrtaeum, Aporrectodea turgida and Lumbricus terrestris and its relation to the carbon budgets of three artificial soils, Pedobiologia, 29:377–389. Shtina, É.A., L.S. Kozlovskaya, and K.A. Nekrasova. 1981. Relations of soil oligochaetes and algae, Sov. J. Ecol., 12:44–48. Shumway, D.L. and R.T. Koide. 1994. Seed preferences of Lumbricus terrestris L., Appl. Soil Ecol., 1:11–15. Shuster, W.D., S. Subler, and E.L. McCoy. 2001. Deep-burrowing earthworm additions changed the distribution of soil organic carbon in a chisel-tilled soil, Soil Biol. Biochem., 33:983–996. Spain, A.V., P.G. Saffigna, and A.D. Wood. 1990. Tissue carbon sources for Pontoscolex corethrurus (Oligochaeta: Glossoscolecidae) in a sugarcane ecosystem, Soil Biol. Biochem., 22:703–706. Sparling, G.P., P.B.S. Hart, J.A. August, and D.M. Leslie. 1994. A comparison of soil and microbial carbon, mitrogen, and phosphorus contents, and macroaggregate stability of a soil under native forest and after clearance for pasture and plantation forest, Biol. Fertil. Soils, 17:91–100. Spiers, G.A., D. Gagnon, G.E. Nason, E.C. Packee, and J.D. Lousier. 1986. Effects and importance of indigenous earthworms on decomposition and nutrient cycling in coastal forest systems, Can. J. Forest Res., 16:983–989. Stamatiadis, S., E.T. Nerantzis, E. Giannakopoulou, and L.M. Maniatis. 1994. The nutritive value of two species of microorganisms to the earthworm Eisenia fetida, Eur. J. Soil Biol., 30:177–185. Stephens, P.M. and C.W. Davoren. 1995. Effect of the lumbricid earthworm Aporrectodea trapezoides on wheat grain yield in the field in the presence or absence of Rhizoctonia solani and Gaeumannomyces graminis var. tritici, Soil Biol. Biochem., 28:561–567. Stephens, P.M. and C.W. Davoren. 1997. Influence of the earthworms Aporrectodea trapezoides and A. rosea on the disease severity of Rhizoctonia solani on subterranean clover and ryegrass, Soil Biol. Biochem., 29:511–516. Stephens, P.M., C.W. Davoren, B.M. Doube, and M.H. Ryder. 1994a. Ability of the lumbricid earthworms Aporrectodea rosea and Aporrectodea trapezoides to reduce the severity of take-all under greenhouse and field conditions, Soil Biol. Biochem., 26:1291–1297. Stephens, P.M., C.W. Davoren, M.H. Ryder, and B.M. Doube. 1993. Influence of the lumbricid earthworm Aporrectodea trapezoides on the colonization of wheat roots by Pseudomonas corrugata strain 2140R in soil, Soil Biol. Biochem., 25:1719–1724. Stephens, P.M., C.W. Davoren, M.H. Ryder, and B.M. Doube. 1994b. Greenhouse and field experiments demonstrating the ability of earthworms to reduce the severity of take-all on wheat, in C.E. Pankhurst, Ed., Soil Biota: Management in Sustainable Farming Systems (Poster Papers), CSIRO, East Melbourne, Australia, pp. 21–23. Stephens, P.M., C.W. Davoren, M.H. Ryder, and B.M. Doube. 1994c. Influence of the earthworms Aporrectodea rosea and Aporrectodea trapezoides on Rhizoctonia solani disease of wheat seedlings and the interaction with a surface mulch of cereal-pea straw, Soil Biol. Biochem., 26:1285–1287. Stephens, P.M., C.W. Davoren, M.H. Ryder, and B.M. Doube. 1994d. Influence of the earthworm Aporrectodea trapezoides (Lumbricidae) on the colonization of alfalfa (Medicago sativa L.) roots by Rhizobium meliloti L5-30R and the survival of R. meliloti in soil, Biol. Fertil. Soils, 18:63–70.

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Stephens, P.M., C.W. Davoren, M.H. Ryder, B.M. Doube, and R.L. Correll. 1994e. Field evidence for reduced severity of Rhizoctonia bare-patch disease of wheat, due to the presence of the earthworms Aporrectodea rosea and Aporrectodea trapezoides, Soil Biol. Biochem., 26:1495–1500. Stout, J.D. 1983. Organic matter turnover by earthworms, in J.E. Satchell, Ed., Earthworm Ecology: From Darwin to Vermiculture, Chapman & Hall, London, pp. 35–48. Striganova, B.R., T.D.P. Derimova, G.P. Mazantseva, and A.V. Tiunov. 1989a. Effects of earthworms on biological nitrogen fixation in the soil, Biol. Bull. Acad. Sci. USSR, 15:560–565. Striganova, B.R., O.E. Marfenina, and V.A. Ponomarenko. 1989b. Some aspects of the effect of earthworms on soil fungi, Biol. Bull. Acad. Sci. USSR, 15:460–463. Subler, S., C.M. Baranski, and C.A. Edwards. 1997. Earthworm additions increased short-term nitrogen availability and leaching in two grain-crop agroecosystems, Soil Biol. Biochem., 29:413–421. Syers, J.K. and J.A. Springett. 1984. Earthworms and soil fertility, Plant Soil, 76:93–104. Syers, J.K., A.N. Sharpley, and D.R. Keeney. 1979a. Cycling of nitrogen by surface-casting earthworms in a pasture ecosystem, Soil Biol. Biochem., 11:181–185. Syers, J.K., J.A. Springett, and A.N. Sharpley. 1979b. The role of earthworms in the cycling of phosphorus in pasture ecosystems, in T.K. Crosby and R.P. Pottinger, Eds., Proceedings of the Second Australasian Conference on Grassland Invertebrate Ecology, Government Printer, Wellington, New Zealand, pp. 47–49. Thorpe, I.S., K. Killham, J.I. Prosser, and L.A. Glover. 1993. Novel method for the study of the population dynamics of a genetically modified microorganism in the gut of the earthworm Lumbricus terrestris, Biol. Fertil. Soils, 15:55–59. Tiunov, A.V., and N.A. Kuznetsova. 2000. Environmental activity of anecic earthworms (Lumbricus terrestris L.) and spatial organization of soil communities, Biol. Bull. USSR, 27:510–518. Tiunov, A.V. and S. Scheu. 1999. Microbial respiration, biomass, biovolume and nutrient status in burrow walls of Lumbricus terrestris L. (Lumbricidae), Soil Biol. Biochem., 31:2039–2048. Tiunov, A.V. and S. Scheu. 2000. Microfungal communities in soil, litter and casts of Lumbricus terrestris L. (Lumbricidae): a laboratory experiment, Appl. Soil Ecol., 14:17–26. Tiwari, S.C. and R.R. Mishra. 1993. Fungal abundance and diversity in earthworm casts and in uningested soil, Biol. Fertil. Soils, 16:131–134. Tiwari, S.C., B.K. Tiwari, and R.R. Mishra. 1990. Microfungal species associated with the gut content and casts of Drawida assamensis Gates, Proc. Indian Acad. Sci. (Plant Sci.), 100:379–382. Toyota, K. and M. Kimura. 2000. Microbial community indigenous to the earthworm Eisenia foetida, Biol. Fertil. Soils, 31:187–190. Trigo, D., I. Barois, M.H. Garvín, E. Huerta, S. Irisson, and P. Lavelle. 1999. Mutualism between earthworms and soil microflora, Pedobiologia, 43:866–873. Urbásek, F. and V. Pizl. 1991. Activity of digestive enzymes in the gut of five earthworm species (Oligochaeta: Lumbricidae), Rev. Écol. Biol. Sol, 28:461–468. Vinceslas-Akpa, M. and M. Loquet. 1995. Observation in situ de la microflore liée au tube digestif de Eisenia fetida andrei (Lumbricidae), Eur. J. Soil Biol., 31:101–110. Visser, S. 1985. Role of the soil invertebrates in determining the composition of soil microbial communities, in A. Fitter, D. Atkinson, D.J. Read, and M.B. Usher, Eds., Ecological Interactions in Soil: Plants, Microbes and Animals, Blackwell Scientific Publications, Oxford, U.K., pp. 297–317. Wallwork, J.A. 1970. Ecology of Soil Animals, McGraw-Hill, London. Westernacher, E. and O. Graff. 1987. Orientation behavior of earthworms (Lumbricidae) towards different crops, Biol. Fertil. Soils, 3:131–133. Wolter, C. and S. Scheu. 1999. Changes in bacterial numbers and hyphal lengths during the gut passage through Lumbricus terrestris (Lumbricidae, Oligochaeta), Pedobiologia, 43:891–900. Wolters, V. and R.G. Joergensen. 1992. Microbial carbon turnover in beech forest soils worked by Aporrectodea caliginosa (Savigny) (Oligochaeta: Lumbricidae), Soil Biol. Biochem., 24:171–177. Zwartz, K.B., P.J. Kiukman, and J.A. van Veen. 1994. Rhizosphere protozoa: their significance in nutrient dynamics, in J.F. Darbyshire, Ed., Soil Protozoa, CAB International, Wallingford, U.K., pp. 93–122.

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of Earthworms on 13 Impacts Other Biota in Forest Soils, with Some Emphasis on Cool Temperate Montane Forests Dennis Parkinson Department of Biological Sciences, University of Calgary, Calgary, Alberta, Canada

Mary Ann McLean Department of Biology, Indiana State University, Terre Haute, Indiana, U.S.A.

Stefan Scheu Institute of Zoology, Darmstadt University of Technology, Darmstadt, Germany

CONTENTS Effects of Channeling, Comminution, and Mixing.......................................................................242 Effects of Earthworm Grazing, Gut Transit, and Casts on Soil Organisms.................................244 Roles of the Soil Biota in Development of Forest Humus Forms ...............................................248 Experimental Approaches to Interactions......................................................................................249 Laboratory Studies of Interactions .........................................................................................249 Field Studies of Interactions...................................................................................................250 Impacts of Earthworm Invasions of Two Montane Forests ..........................................................251 Mode of Invasion of North American Forest Soils by European Earthworms .....................251 Changes in Soil Structure by Colonizing Earthworms ..........................................................251 Changes in Soil Biota Caused by Colonizing Earthworms...................................................252 Soil Microorganisms .....................................................................................................252 Soil Fauna......................................................................................................................253 Changes in Ecosystem Processes and Plant Growth Caused by Earthworms ......................254 References ......................................................................................................................................254

Interactions of the soil fauna both between different groups of invertebrates and, particularly, with soil microorganisms are considered important in affecting soil processes (e.g., organic matter decomposition and nutrient cycling) and in influencing the community composition of various groups of the soil biota (Seastedt 1984; Visser 1985). The ways in which these functional and community structure effects are achieved are by (1) comminution of and channeling in organic debris and mixing of organic debris and mineral soil, (2) grazing on microbial tissues, and (3) dispersal of microbial propagules (Visser 1985).

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Earthworms form an important group of soil invertebrates and, although they show different ecological strategies, all facets of their life in organic and mineral soil layers can potentially have major impacts on other groups of soil organisms. (These are discussed, with particular reference to the effects of earthworms in temperate and subtropical agricultural systems, in Chapter 12 this volume.) The aim of this chapter is to provide information and comments on the influences of earthworms on the structures of microbial and faunal communities in forest soils, particularly in the organic layers of northern montane forests. At the outset, it must be admitted that, although substantial data are available on the effects of earthworms on various soil processes and on microbial biomass dynamics, accumulation of relevant detailed data on their effects on other soil biota has begun only relatively recently.

EFFECTS OF CHANNELING, COMMINUTION, AND MIXING Although different earthworm species differ in size and behavior, their activities have great consequences for the physical and chemical characteristics of organic and mineral soil (Lee and Foster 1991). These activities include the ingestion of soil and organic material (plus comminution of organic matter), the production of mucus and urine, and the intermixing of these materials; ejection of gut contents as casts; and the formation of earthworm burrow systems. Comminution of organic matter has important effects on decomposition rates (Swift et al. 1979) and on microbial growth (Gunnarsson et al. 1988). It is often considered an important result of earthworm activity despite the paucity of quantitative evidence. Schulmann and Tiunov (1999) found that Lumbricus terrestris selectively ingested sand from a mixture of sand and litter and hypothesized this increased organic matter fragmentation during passage through the gut. Earthworms are well known for their abilities to improve soil conditions, such as increased stability through cast production and the production of macropores (burrows), which enhance the movement of water and gases and root penetration through soil (Devliegher and Verstraete 1997; Francis and Fraser 1998). Shipitalo and LeBayon and Kretzschmar (see Chapter 10 and Chapter 11, respectively, this volume) provide detailed comments on earthworm-originated soil porosity and the effects of earthworm burrows on soil structures. This increase in soil macroporosity also enables smaller animals such as springtails (Marinissen and Bok 1988) and mites (Walter and Proctor 1999) to penetrate to deeper soil horizons. Earthworm burrowing activities range from those of species (epigeic earthworms) that inhabit surface organic layers and do not produce well-defined burrows, to species (anecic earthworms) that produce deep vertical burrows into the mineral soil, and finally to endogeic species that construct networks of burrows in the upper mineral soil. Although earthworm activities are generally considered beneficial for soil structure, it must be remembered that, in some soils, they can have adverse effects (e.g., Puttarudriah and Shivashankara Sastry 1961; Ester and van Rozen 2002). Earthworm burrows are lined by the drilosphere soil zone (Bouché 1975). This is described by Doube and Brown (see Chapter 12, this volume) and in detail by Brown et al. (2000). It is a zone generally richer in nitrogen, phosphorus, and humified organic material than the surrounding soil (Beare et al. 1994; Tiunov and Scheu 1999); moreover, burrow walls usually have a higher pH, presumably because of excretion of earthworm subcutaneous mucus (Schrader 1994). However, the detailed conditions in the drilosphere depend on the ecological group of the burrowing species and the type of organic material eaten, the amount of organic fragments or humified material in the drilosphere wall, and materials ejected in casts. The drilosphere appears to be a zone in which significant increases in populations of some important physiological groups of soil bacteria occur, such as nitrogen transformers (Bhatnagar 1975) and cellulolytic species (Tiunov and Kuznetsova 2000; Tiunov et al. 2001); also, increased enzyme activities and CO2 efflux were recorded by Loquet (1978).

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With minor variations, significantly increased basal respiration, total microbial biomass, and numbers, biomass, or biovolume of different soil microbial groups have usually been recorded from earthworm burrow walls (e.g., Tiunov and Scheu 1999; Tiunov and Kuznetsova 2000; and Brown et al. 2000). However, Görres et al. (1997) recorded a significant decrease in microbial biomass in this zone. Warcup (1965), using meticulous direct observations of arable soil blocks in a nonquantitative study, observed fungal sporulation in microbial biomass produced by Eisenia rosea and Allolobophora (= Aporrectodea) caliginosa, in substrates such as the remains of other soil invertebrates (e.g., animal parasites such as Beauveria, Conidiobolus, and Entomophthora spp.; and saprotrophs such as various Mucorales, Penicillium, Aspergillus, and Verticillium spp.); earthworm casts (fructifications of Dictyostelium mucoroides); and mineral particles (several cleistothecial Ascomycetes and some aphyllophoraceous Basidiomycetes). He also observed sporulation of some fungi as confined to organic fragments that abutted earthworm burrows (i.e., possibly the drilosophere), such as Dinemasporium, Chaetomium, Periconia, Trichoderma, Gonytrichum, Brachysporiella, and Endophragmia (= Phragmocephala) spp. Apart from fungi, actinomycetes were seen in the burrows growing on plant and animal debris, fungal hyphae, and humus particles. Tiunov and Dobrovolskaya (2002), in the first detailed study of fungal communities in earthworm burrow walls, examined these communities in L. terrestris burrows in lime and beech forests. The earthworm burrows in lime forest soil could be separated into lined (with dark cast material) and unlined types. They found Cylindrocarpon spp. were the most abundant fungi in the control soil (not affected by earthworm activity) in both forests, but populations declined sharply in burrow walls. Their data indicate that the three types of burrows they studied were colonized by different fungal species, and they concluded that this was the result of differences in the quantity and quality of the organic materials moved down the soil profile. In the lined burrows (lime forest), a considerable proportion of the fungi were typical litter-inhabiting species (e.g., Trichoderma koningii, Mucor hiemalis); interestingly, very few dematiaceous species (common early leaf litter colonizers) were isolated. As well as the role of earthworm burrows as “transit routes” for various smaller soil invertebrates, the drilosphere has other impacts on the numbers of soil invertebrates in its vicinity. The numbers of epigeic and endogeic earthworms (particularly Lumbricus rubellus and A. caliginosa) increased significantly in the zone of L. terrestris burrows, but this was not the case for a population of Aporrectodea rosea (Tiunov and Kuznetsova 2000). The same authors reported significantly larger populations of springtail species in these burrows, whereas other invertebrates appeared to avoid the burrows. Salmon (2001) showed that mucus plus urine from various earthworm species are attractive to the collembolan species Heteromurus nitidus, and earthworms that burrow walls lined with these materials could provide a favorable habitat for this species. Görres et al. (1997) reported that nematodes were more abundant in earthworm burrow walls than in surrounding soil. Tiunov et al. (2001) found decreased lengths of fungal mycelium in the upper burrow depths and hypothesized that this was the result of activities of the abundant fungivorous nematodes in this zone. Several authors have recorded increased populations of a range of soil invertebrates as a result of earthworm activities (i.e., in creating preferential “hot spots” of faunal activity). This was discussed in detail by Brown (1995) and referenced briefly by Tiunov and Kuznetsova (2000). Apart from their burrowing activities, earthworms can have other significant effects on the fabric of mineral soil and on the surface organic layers. Kubiena (1955) showed that, in mineral soils possessing sufficient water-stable binding substances, earthworms can create a “spongy fabric” (i.e., aggregates bound together producing a porous internal structure). This fabric, which has good aeration and water status, enhances the development of aerobic bacteria, actinomycetes, and fungi. The soil spaces also allow the development of a wide range of species of soil invertebrates. However, in some tundra soils, the upper mineral layers show a spongy fabric that has developed from

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processes other than faunal activity (Pawluk 1985). In compact soil fabrics, root development (with subsequent decomposition) and earthworm activities may allow the development of “channel fabrics.” Again, these effects on soil structure can enhance microbial and faunal development, and it could be hypothesized that this would include increases in their species diversity. Another major impact of earthworm activities is the incorporation of organic debris from the soil surface into the subsurface mineral soil layers. As Doube and Brown (see Chapter 12, this volume) have stated, the subsequent increases in the organic matter content in these mineral soil layers may allow changed soil structure, increased microbial biomass and activity, changed patterns of colonization by saprotrophic microorganisms on the decomposing organic matter, and increased diversity of microhabitats for soil invertebrates. The importance of earthworm casts in adding “processed” organic materials, surface organic materials, or mineral soil and thus affecting other soil biota is considered later. Anecic species of earthworms (e.g., L. terrestris) produce vertical burrows and remove litter from the forest floor and either take it down into the soil or effectively mix the litter material with large numbers of casts on the soil surface, thereby constructing middens, which are conspicuous structures on the forest floor. In these middens, there is increased microbial and faunal activity, which causes increased litter palatability for ingestion by earthworms and other soil invertebrates (Brown et al. 2000). Maraun et al. (1999) reviewed work on the effects of the middens of L. terrestris on other micro- and mesofauna in a beech forest. Also, they presented the results of a detailed ambitious study that assessed the importance of middens to other groups of soil invertebrates. At the outset of their study, they hypothesized that, as a result of higher microbial biomass in middens, there would be an increase in microbivorous micro- and mesoinvertebrates plus the occurrence of invertebrate species specific to middens. The values they reported for carbon and nitrogen, microbial biomass, and basal respiration were significantly higher in midden than in nonmidden material, and middens became, at least temporarily, a preferred habitat for several soil invertebrate species. However, other species either avoided middens or were affected detrimentally by their environmental conditions. Many species of organisms coexist in middens, and conditions there, including high microbial biomass, allow various potential prey species (e.g., Nematoda, Collembola) to achieve high population densities. This benefits predaceous groups (e.g., Gamasina, Uropodina). Finally, their initial speculation that midden microhabitats would allow increased niche diversity and hence high faunal diversity was not supported. Almost all midden species also occurred in nonmidden materials, but as mentioned, populations above the densities of individual species were higher in middens.

EFFECTS OF EARTHWORM GRAZING, GUT TRANSIT, AND CASTS ON SOIL ORGANISMS The effects of earthworm grazing on soil organisms are investigated by Brown (1995) and Doube and Brown (see Chapter 12 this volume); hence, the following comments are brief and attempt, if possible, to deal particularly with forest ecosystems. There is evidence from feeding choice studies that earthworms prefer softer rather than harder textured leaves (Heath and Arnold 1966; Wright 1972), small organic fragments rather than large ones (Judas 1992), and leaves with high nitrogen content and low amounts of secondary metabolites (Hendriksen 1990; Edwards and Bohlen 1996). It has been considered for some time that leaf decomposer microorganisms, particularly fungi, may be an important food source for many earthworm species (see reviews by Edwards and Fletcher 1988; Brown 1995; and Doube and Brown, Chapter 12 this volume). Therefore, it could be expected that fragmented leaf material, supporting a high microbial biomass, would be preferred food for earthworms. When studying the impacts of Dendrobaena octaedra during colonization of a lodgepole pine forest McLean et al. (1996) reported that, in comparison with material from other litter

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layers, fragmented pine needles that were heavily colonized by fungi (i.e., F2 layer needles) favored maximum growth of the earthworms. In coniferous forests, the organic layers contain large communities of saprotrophic fungi comprising diverse hyphal types, mycelial strands, rhizomorphs, and other fungal structures, together with ectomycorrhizal fungi on conifer “feeder” roots and arbuscular mycorrhizal fungi (AM) fungi associated with roots of shrubs (some of which also have Frankia nodules) and of herbaceous plants. These fungi represent a large potential food source for the epigeic earthworm species colonizing these environments. However, it has been considered that there is severe competition between microorganisms (particularly fungi) and earthworms for the available nutrients in the soil organic matter (Scheu and Schaefer 1998). Unlike the situation for forest litter, for which fungal-feeding–preference studies by collembolans and mites abound, and for attempts to evaluate the potential ecological impacts of these activities, there have been relatively few such studies for earthworms. Bonkowski et al. (2000) produced an effective review of such studies, and although few refer to forest ecosystems, some interesting generalizations could be made. In their study, the five earthworm species they used included epigeic, anecic, and endogeic species, and all had very similar feeding preferences for the nine fungal species tested. However, consumption of preferred fungi by detritivorous species was considerably greater than that by geophagous species. Basidiomycetes were generally refused as food, a result corresponding with the observation that cellulose and lignin decomposer basidiomycetes were also rejected as food (Moody et al. 1995). Data from comparable studies with Collembola are not so consistent; for instance, Visser and Whittaker (1977) showed that Onychiurus subtenuis preferred Cladosporium spp. and sterile dark fungal forms to two basidiomycetes, which were generally avoided and appeared to be toxic. However, P.J.A. Shaw (1988), working with eight ectomycorrhizal and four saprotrophic basidiomycete species from forest litter or soil, showed that three ectomycorrhizal fungi were preferred and were nutritive; two common saprotrophic species were eaten but were not nutritive; and two other fungal species were dangerously toxic to earthworms. Given this situation and that, as yet, only a restricted number of microbial species have been tested in earthworm feeding studies, it seems that a wider study of the palatability of basidiomycetes for earthworms is desirable before any generalizations can be made. This is particularly important for studies of coniferous forest ecosystems in which these fungi are of great ecological importance and earthworm invasions and spread are occurring rapidly. It has been suggested (Bonkowski et al. 2000) that fungi of early successional stages on decomposing plant debris are preferred and partially utilized by earthworms. These fungi serve primarily as indicators of food quality to earthworms and other soil invertebrates. The choice of species of fungi for earthworm feeding preference experiments, in the main, seem to focus on species from agricultural or grassland soils or crop residues. The sequence of fungi on such residues (see Chapter 12, this volume) is very different from that occurring during coniferous litter decomposition. In the L2 and F layers of the floor of a montane lodgepole pine forest, dematiaceous hyphomycetes, coelomycetes, and sterile forms often represent about 85% of the fungal hyphae. Basidiomycete mycelia increase in frequency of occurrence with increasing stages of litter decomposition, and in the H layer, the fungal community is often dominated by hyaline species of Mortierella, Penicillium, and Trichoderma. The F and H organic layers are permeated with ectomycorrhizal pine roots in which the fungi, mainly basidiomycete species, extend into organic layers. A similar pattern of succession of saprotrophic fungi has been reported during the decomposition of leaf litter in a montane aspen poplar forest (Visser and Parkinson 1975). Several studies have been made of feeding preferences by springtails and mites on saprotrophic fungi from these two forests (McLean et al. 1996) and have shown a common preference of these invertebrates for dark-pigmented fungal forms. However, McLean et al. (1996) could not find any effect of mesofaunal grazing on fungal species richness and diversity. Since the invasion of these

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forests by the epigeic earthworm species D. octaedra, there have been no fungal-feeding–preference studies for this species (however, the overall effect of this invasion is covered in the later section discussing changes in the soil biota caused by colonizing earthworms). During grazing on organic matter, earthworms ingest not only fungi, but also a range of other organisms (bacteria, algae, protozoa, nematodes) associated with that material (Edwards and Fletcher 1988; Brown 1995; see Chapter 12 this volume). Considerable attention has been given to the fate of the major components of the microbial community that arrive in earthworm guts as a result of the consumption of soil and plant litter. Following intake of these materials, they are mixed in the earthworm gizzard, which leads to a considerable increase in microbial activity. These microorganisms are exposed, in all parts of the earthworm gut, to conditions different from those in noningested soil (i.e., significantly higher water contents and pH conditions). Moreover, in the earthworm foregut, there is a high content of soluble organic matter derived from the input of intestinal mucus, which declines greatly in the mid- and hindgut (Trigo and Lavelle 1993, 1995). This mucus production, to various degrees, occurs in a wide range of earthworm species (Trigo et al. 1999). In all parts of the earthworm gut, there are much higher respiration rates than in uningested soil (Trigo and Lavelle 1993, 1995). However, there is no evidence of corresponding increases in microbial biomass (Scheu 1992; Daniel and Anderson 1992; Tiunov and Scheu 2000a.). This might indicate a decrease in the metabolic efficiency of the microbial community in the earthworm gut. These observations lead to a concept of a mutualistic digestion of organic matter in earthworm guts (i.e., the existing gut microbial community, which is frequently dormant because of lack of easily assimilable carbon resources) is activated by secreted intestinal mucus under the suitable gut temperature and moisture conditions, and the decomposition of organic matter in the gut contents is enhanced to the mutual benefit of both the microorganisms and the earthworm. This so-called paradox was documented and reviewed comprehensively by Brown et al. (2000). With respect to the fate of bacteria, particularly fungi, during transit through the earthworm gut, there has been considerable debate that usually stems from differences in the methods used for studying these organisms. Cultural studies, although very useful in some specific cases, have given way to other approaches (e.g., quantitative estimates of total bacterial and fungal communities). These approaches include direct observations of microorganisms by methods such as epifluorescence microscopy (e.g., Scheu and Parkinson 1994a; Kristufek et al. 1995), image analysis with epifluorescence microscopy (Schöholzer et al. 1999), and transmission electron microscopy (Kristufek et al. 1994) and the use of whole cell hybridization (Fischer et al. 1995). Bacterial populations in the earthworm crop or gizzard are usually higher than in the surrounding soil. During transit through the earthworm gut, these numbers have been reported to increase greatly (e.g., Parle 1963; Edwards and Bohlen 1996), and usually numbers in earthworm casts are substantially greater than those in the surrounding soil (Edwards and Fletcher 1988). These increases in bacterial populations may be caused not only by microbial growth, but also by bacterial spore germination (Fischer et al. 1995). Wolter and Scheu (1999) showed a somewhat different pattern of bacterial populations using L. terrestris cultured in field samples of beech forest soil. Here, there were increases in numbers of bacteria from the crop to the gizzard to the foregut, then no significant changes to the hindgut. As stated, there is evidence for change in total microbial biomass during transit through the earthworm gut. As well as increases in bacterial populations, there have been reports of increases in bacterial cell size in the earthworm gut (Fischer et al. 1997; Wolter and Scheu 1999). These increases parallel the increases in gut water content mentioned in this section. It appears that many groups of bacteria and actinomycetes can pass unharmed through the earthworm gut (e.g., Gram-negative organisms, spore formers, and actinomycetes during gut transit) (Edwards and Fletcher 1988). However, mortality of some species (e.g., Bacillus cereus var. mycoides, Serratia marcescens, Escherichia coli) has been recorded (Lee 1985). A high percentage of earthworm gut bacteria are ultimately dispersed in earthworm casts, and few are dispersed while adhering to the earthworm body surfaces. This dispersal is of © 2004 by CRC Press LLC

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considerable ecological, even economic, importance, and the importance of earthworm dispersal of beneficial microorganisms (e.g., N2-fixing species) and harmful microorganisms (e.g., soilborne pathogens) was reviewed in detail by Doube et al. (1994), Brown (1995), and Doube and Brown (see Chapter 12, this volume). Data on fungal biomass (i.e., hyphae and various types of species) in earthworm guts are quite variable, and this may be the result of the different study methods used. Thus, Parle (1963) recorded no increases in fungal population in the guts of three earthworm species. He used dilution plating, a method shown to isolate fungi present as spores selectively. H.K. Dash et al. (1986), using light microscopy measurements of fungal hyphal lengths with no distinction between live and dead hyphae, observed that the total hyphal lengths and lengths of hyphal fragments decreased during passage through the gut and was minimum in casts. Wolter and Scheu (1999), working in the field (beech forests), sampled earthworms and, using epifluorescence microscopy, found significantly larger hyphal lengths in the earthworm crop or gizzard than in the surrounding soil. Following this, there was no significant change in hyphal lengths during passage through the rest of the gut. However, in laboratory microcosms to which beech litter was added, there was a significant decline in hyphal lengths from the fore- to hindgut. Schöholzer et al. (1999), working with L. terrestris feeding on decomposing leaves of Taraxacum officinale and using image analysis and epifluorescence, found a large decline in fungal biomass in the foregut compared with that of the leaf material, and this biomass was completely digested during passage through the gut; this is not surprising if fungi are a major food component for earthworms. There are only a few data on changes, if any, that occur in the soil or litter fungal communities as they pass through earthworm guts. Domsch and Banse (1972) reported a large change in the fungal community structure in casts of L. terrestris in comparison with that of the ingested materials. However, Tiwari and Mishra (1993) reported no significant differences between the fungal communities of soil and those of earthworm casts. The impacts of earthworm activities on fungal populations in soils of montane forests are summarized later in this chapter, but these studies did not include the assessment of fungi in earthworm casts. Tiunov and Scheu (2000b) studied the fungal communities in a beechwood mineral soil, alone and with leaf litter, prior to ingestion by earthworms and in the fresh casts produced by L. terrestris (which allowed an estimate of the effect of gut passage on the fungal communities). They reported that only small changes had occurred in fungal communities during the passage of mineral soil alone. In soil plus litter systems, the fungal community composition in earthworm casts depended on the type of litter consumed and was different from that in surrounding soils. Certain fungi (e.g., Absidia cylindrospora, Cladosporium spp., Alternaria spp.) were affected detrimentally by gut passage through the gut, and the number of fungi isolated per organic matter particle decreased greatly in the casts. As for bacteria, the ability of fungal spores to maintain their viability during transit through the earthworm gut is of ecological interest. Considering the fungi as a group, there is a wide range of spore types (e.g., size, shape, septation, pigmentation, wall structure, etc.), some of which will affect their survival in the earthworm gut. As yet, there have been relatively few studies on spore survival in the earthworm gut, and in such studies, only a narrow range of fungal species have been addressed (none were fungi from coniferous forest litter layers). Dash et al. (1986) studied survival of fungal spores in the guts of tropical grassland earthworms. They found that 54 to 64% of the fungal species occurring in earthworm foreguts survived and were found in the casts. Potentially antibiotic-producing Aspergillus spp. and Penicillium spp. and the thick spore walled Thielavia terricola were permanent survivors of gut transit. Moody et al. (1996), working with the spores of several fungi associated with decomposing wheat straw, found that those of Fusarium lateritium and Agrocybe temulenta did not survive passage through the guts of L. terrestris; germination of spores of Trichoderma sp. and M. hiemalis was significantly reduced, and that of Chaetomium globosum was reduced but only slightly. However, passage through the gut of Aporrectodea longa caused a significant increase in fungal spore germination. © 2004 by CRC Press LLC

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Again, there is the possibility of earthworms as vectors not only of saprotrophic fungi, but also of plant root pathogens and of beneficial mycorrhizal fungi. Toyota and Kimura (1994) showed that, although the numbers of viable spores of the pathogen Fusarium oxysporum f. sp. Raphini decreased during passage through Pheretima sp. guts, there was effective spread of this fungus via the earthworm casts. Earthworms have been considered potential vectors of mycorrhizae, particularly vesicular arbuscular mycorrhizae (VAM) fungi. Pattinson et al. (1997) provided a brief review of this issue and considered that, although earthworms have the potential for this dispersal, they found little evidence for it. They not only found no evidence for VAM spore dispersal by Aporrectodea trapezoides, but also reported that this earthworm reduced the VAM colonization of Trifolium subterraneum roots. Lawrence et al. (2003) found a similar situation with VAM colonization rates of sugar-maple roots and suggested, as did Pattinson et al. (1997), that disruption of the extramatrical hyphae and hyphal networks of the VAM fungus was a strong possibility for the cause for this reduction in colonization. However, Rabatin and Stinner (1988) and Cavendar et al. (2003) reported that soil invertebrates, particularly springtails and, to a lesser extent, earthworms, had important positive impacts on the transmission of VAM fungi. Very little work seems to have been done on the effects of earthworms on ectomycorrhizal fungi. Reddell and Spain (1991a) reported live spores and hyphae of hypogeous ectomycorrhizal fungi were present at low frequencies in earthworm casts. Obviously, the interactions between earthworms and ectomycorrhizal fungi are a topic requiring more research.

ROLES OF THE SOIL BIOTA IN DEVELOPMENT OF FOREST HUMUS FORMS In forest ecosystems, the importance of soil invertebrates, usually with emphasis on the roles of earthworms, in the development and maintenance of soil structure has been recognized for more than 100 years. In temperate forests, organic matter (humus in the broad sense) accumulates on the soil surface, and there may be an organic-enriched upper mineral horizon. In classifying forest and heathland humus forms, Müller (1878) introduced the terms mull and mor, and Ramann (1911) added the term moder. Green et al. (1993) provided an account of the tortuous subsequent history of the development of terminology for humus forms and provided a proposal for their classification. In this chapter, a simplified view is taken, and only mor, mull, and moder forms are considered. The mor litter form is one in which organic matter accumulates on the soil surface and in which definable L, F, and H horizons can be distinguished. There is very little or no mixing of organic materials with the upper mineral soil. In the mull litter form, there is little accumulation of organic matter on the soil surface because there is incorporation of this material into the upper mineral soil. This humus form is generally found in forests on base-rich soils and in grasslands. In forests with the moder humus form, there is accumulation of organic materials on the soil surface, but the distinction between the F and H layers is frequently unclear, and in these layers, there may be incorporation of mineral particles. Also, there may be limited infiltration of humus material into the uppermost mineral soil layer. Moders have been considered intermediate in the complex gradient from mor to mull humus forms (Green et al. 1993). The type of humus form in a particular ecosystem is determined by a range of climatic, edaphic, and vegetation factors. However, it is well known that the soil biota play important, interactive roles in the development and maintenance of specific humus forms (e.g., Rusek 1985). In low pH, base-poor, high C:N conditions in mor humus forms, there is a large fungal biomass with frequent production of hyphal mats and hyphal strands, large arthropod and nematode populations, and a rarity or absence of earthworms; therefore, the mixing of the organic layers or mixing of organic matter into the mineral soil does not occur often. The L, F, and H layers formed in mor

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humus represent a chronosequence in litter decomposition. Consequently, many studies of fungal “succession” during this process have been made (Kjøller and Struwe 1982; Widden 1986; Hansen 1989). Similarly, there have been many studies of arthropod communities in this humus type (e.g., Hågvar and Abrahamsen 1980; Petersen 1982; Kaneko 1985; Seastedt et al. 1989), including studies of the feeding preferences of various invertebrate species (Dash and Cragg 1972; Moore et al. 1987; Shaw 1988; Klironomos et al. 1992) and of their interactions with the litter fungi. As stated, earthworms are generally considered relatively rare in mor humus, but some data on their occurrence in this humus form have been given by Piearce (1972), Huhta and Koskenniemi (1975), Muys and Lust (1992), and McLean and Parkinson (1997). In the mull humus forms with higher pH, base-rich and lower C:N conditions, fungal biomass is usually lower and bacterial biomass higher than in mor humus forms. Arthropod populations are low, and earthworm populations are high (Schaefer and Schauermann 1991). The activities of anecic and endogeic earthworms are major factors in removing organic materials from the soil surface and in maintaining the mixing of these materials into the mineral soil with subsequent, stimulatory effects on soil microorganisms. The nature of the moder humus form results, at least in part, from the activities of epigeic species of earthworms (Rusek 1985). Their activities, which cause mixing of organic materials and the incorporation of earthworm casts, create major changes in the surface organic layers that may have considerable effects on the communities of other litter-inhabiting invertebrates and microorganisms. From the foregoing comments and those by Doube and Brown (see Chapter 12, this volume), it can be seen that earthworms can affect significantly the physical and chemical conditions in soils and different humus forms; this may affect the structure of the communities of other soil organisms. However, detailed studies of these biological effects are rare because of experimental difficulties.

EXPERIMENTAL APPROACHES TO INTERACTIONS Just as in many grassland and agricultural ecosystems, in mull and moder forests, changes in populations and activities of earthworms in forests have been occurring over long time periods. As a result, a total complex litter/soil biota has evolved in coexistence with earthworms. Consequently, there appear to be relatively few laboratory and field approaches possible for assessing impacts of earthworms on other groups of soil organisms.

LABORATORY STUDIES

OF INTERACTIONS

The laboratory experimental approach involves the use of soil systems (micro- or mesocosms) of varying degrees of complexity; this allows experiments to be done under controlled temperature and moisture conditions. Thus, the variable field conditions of soil microclimate and the like are avoided. However, if laboratory experiments are oversimplified, it is difficult to interpret them in terms of actual field situations and conditions. In research into the effects of earthworms on soil processes and soil communities, serious attempts have been made to use laboratory microcosms or systems that reflect the realities of field situations. Nevertheless, it is often very difficult to use this approach for studies of litter/soil systems taken from locations in which earthworms have been active for long periods and have already affected soil structure, processes, and the rest of the soil biota. In a laboratory study of the litter/soil profile of a cool, temperate aspen forest, in the early stages of colonization by an epigeic earthworm species (D. octaedra), Scheu and Parkinson (1994b) showed that the presence of these earthworms produced microbial biomass changes.

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There are two major field approaches for studying the impacts of earthworms on the other soil biota. One approach involves the partial or, preferably, total elimination of earthworms from replicate field study plots for subsequent comparisons with plots in which no earthworm elimination has occurred or alternative earthworms have been additionally inoculated. There are a number of areas where this situation can be studied, such as areas where the partial or total elimination of earthworms is normally a chronic, unplanned side effect of human activities and areas where earthworm elimination is a direct, acute result of planned human activities or experiments. Examples of the former can be seen in areas subjected to various agricultural practices or to various forms of soil pollution. However, in many of these examples, there may be severe difficulties in experimental design and data interpretation (e.g., often it is difficult to establish appropriate, analogous unimpacted control areas); often, it is difficult to determine the real cause of observed changes in the soil biota (earthworm absence or physical/chemical impacts of disturbance itself). Another example of the former is areas where land-use changes have been or are being made. Yeates (1988) showed that, with a change in land use from grassland to Pinus radiata plantations of different densities, earthworm populations disappeared from soils under high tree densities. Yeates (1991) also reviewed the effects of historical land-use changes on the soil fauna in New Zealand. Muys et al. (1992) showed, as part of a larger study, that 18 years after changing a pasture to a Quercus palustris plantation, there was a decrease in earthworm biomass. However, neither the study by Yeates (1988) nor that of Muys et al. (1992) aimed to investigate the effects of changed earthworm populations on other groups of the soil biota. In all the foregoing examples, there was the problem of the timing of the experiments on the effects of changed earthworm population decline or elimination on other organisms. Lavelle (1988) posed the question “When earthworms disappear, how long does it take before significant changes, if any, occur in the soil structure and nutrient cycling?” This may well be extended to include changes in the communities of other soil organisms. In the situation of planned elimination of earthworms, two approaches have been taken: application of biocides and electroshocking. Bohlen et al. (1995) reviewed these briefly, and, to avoid problems in the use of biocides (e.g., their impacts on nontarget organisms, etc.), they used electroshocking. However, if this type of approach is to be effective in soil community studies, then it must be long term to allow the soil to “equilibrate” to a zero earthworm population state (i.e., again, the timing of experiments becomes important). The second major field approach is to study previously earthworm-free soils into which earthworms have been introduced purposely or into which unplanned invasions or colonization is occurring (see Chapter 5, this volume). Stockdill (1982) showed that lumbricid earthworms (A. caliginosa) introduced into pastures in New Zealand improved their soil qualities. Yeates (1981) showed that the introduction of these earthworms into pasture soils led to a 50% reduction in populations of soil nematodes plus a significant change from bacterial to fungal feeding species of nematodes. Unfortunately, the impacts of this change on the microbial community structure are unknown. Baker (see Chapter 14, this volume) provides a detailed account of the various impacts of earthworm introductions into Australian soils. For instance, Hoogerkamp et al. (1983) showed beneficial changes following the introduction of various mixtures of earthworm species into pastures, developing on reclaimed polder soils. However, they did not study the impacts of these earthworm introductions on the other soil biota, Marinissen and Bok (1988) showed that effects of earthworms on the structure of polder soils resulted in increased size, populations, and distribution of Collembola in those soils. Fortunately, there are still areas that have been earthworm free for centuries and where colonization by earthworms is beginning, presumably as an indirect result of human activities. Earthworm invasions of forests in the Kananaskis Valley (southern Alberta, Canada) have begun, and the impacts of this are discussed in detail in the following section. © 2004 by CRC Press LLC

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IMPACTS OF EARTHWORM INVASIONS OF TWO MONTANE FORESTS MODE OF INVASION EARTHWORMS

OF

NORTH AMERICAN FOREST SOILS

BY

EUROPEAN

With the advent of European settlers in North America, lumbricid earthworms began to invade the continent. During early settlements, earthworms were probably spread mainly in soil adhering to the roots of garden plants and trees. Later, the spread of earthworms increased, presumably because of the use of machinery in agriculture and forestry. Soil containing earthworm cocoons was probably transported on tires and human and animal feet, even over long distances. Without accidental longdistance transport, lumbricid earthworms certainly would not have been able to colonize the entire North American continent (see Chapter 5, this volume). Colonization of new habitats by earthworms is usually slow and in the horizontal range of 10 m per year (Hoogerkamp et al. 1983; Marinissen and Van den Bosch 1992). Colonization of eastern North America by earthworms went almost unnoticed by scientists and was well established when the monitoring of earthworm populations started (Reynolds 1977; see Chapter 4, this volume). By contrast, western Canada was little colonized by European lumbricids until recently. Forests of the mountain ranges of the Rocky Mountains in southern Alberta were free of earthworms until 1985 (Parkinson unpublished data). Since then, a range of forests has been colonized, and a number of lumbricid earthworm species have been successful invaders (Scheu and McLean 1993); the most vigorous is D. octaedra (Dymond et al. 1997; McLean and Parkinson 1997). Studying populations of earthworms along transects from forestry trunk roads into the forest, Dymond et al. (1997) concluded that earthworm colonization usually begins from the roads and moves into the forest in a wavelike manner. Earthworms may reach exceptionally high populations along the invasion line, and colonization may be very rapid. In deciduous (aspen) and conifer (pine) forests, D. octaedra may reach populations as high as 2000 to 3000 individuals m–2 (Dymond et al. 1997; McLean and Parkinson 1997). These invasions completely alter the soil structure and soil invertebrates and fungal community composition, decomposition, and mineralization processes and therefore the whole ecosystem. The climate of the areas where these invasions occurred is continental, with long cold winters (from −5 to −20°C), during which the organic layers of the forest floor are frozen solid (Parkinson 1988). It was expected, from the work of Holmstrup (1996) and Holmstrup and Zacharias (1996), that the survival of D. octaedra during such harsh conditions would be only as cocoons. However, sampling of the frozen organic layers, followed by careful thawing and then extraction, yielded not only a predominance of cocoons, but also small populations of small and large immatures, matures, and aclitellate earthworm adults, yielding numbers for all growth stages of about 200 m−2 (McLean unpublished data).

CHANGES

IN

SOIL STRUCTURE

BY

COLONIZING EARTHWORMS

Dendrobaena octaedra preferentially colonizes the organic layers of forests. Therefore, its effects on soil structure are usually restricted to the uppermost soil layers. However, in both deciduous and conifer forests, it strongly changes the structure of organic layers. Dendrobaena octaedra feeds preferentially on F-layer material and transforms it into casts or faecal pellets (Scheu and Parkinson 1994b; McLean et al. 1996; McLean and Parkinson 1997). In addition, some mineral soil material is incorporated into the organic layers, which eliminates the clear boundaries between organic layers and mineral soil that typically occur in forest soils of western Canada. Thorough mixing of the organic and mineral soil layers occurs only in the presence of mineral soil-dwelling species of earthworms such as A. caliginosa and L. terrestris. In the first study on the effects of lumbricid earthworms on the soil structure of Canadian forests, Langmaid (1964)

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documented changes from podzolic moder soils to mull-structured soils within 3 to 4 years. In the laboratory, C. Shaw and Pawluk (1986) observed a complete mixing of organic and mineral soil materials if both anecic and endogeic earthworm species were present. Scheu and Parkinson (1994b) showed that endogeic earthworm species may entirely mix the F and H material of aspen forest soil into the mineral soil, thereby transforming moder to mull humus. Alban and Berry (1994) documented that the organic layers of an aspen forest in Minnesota were completely incorporated into the mineral soil, thereby transforming the former E horizon into an A horizon. This process was caused mainly by an invasion of endogeic earthworm species (Aporrectodea tuberculata), but the epi-endogeic species L. rubellus, which also invaded this forest although in lower populations, presumably also contributed to the incorporation of the L-layer material into the mineral soil. However, detailed analyses of the time course of this mixing process and the interactions between earthworm species responsible for the soil transformations under field conditions are still lacking.

CHANGES

IN

SOIL BIOTA CAUSED

BY

COLONIZING EARTHWORMS

The profound changes in soil structure resulting from the invasion of forest soils by earthworms modify the habitat and distribution of resources for the diverse soil microorganisms and soil invertebrates and therefore the whole decomposer food web. Unfortunately, information about these changes is fragmentary at best. Soil Microorganisms Microbial biomass in the L and F layers appeared to be quite resistant to earthworm-mediated changes in soil structure and soil processes, although microbial respiration decreased with increasing earthworm biomass in pine forest soils (McLean and Parkinson 1997). However, significant differences in the fungal community structure and species composition because of earthworm activities were observed in both the laboratory and field. In a 6-month laboratory mesocosm experiment, the activities of the epigeic earthworm species D. octaedra caused changes in the numbers of fungal isolates per soil particle, which reflects the density of different fungal taxa on an individual soil particle and is an indication of the degree of competition between fungi. In this experiment, increasing earthworm activities resulted in an initial increase in the number of fungal isolates per particle followed by a decrease in the number of isolates per particle (McLean and Parkinson 1998b). This may be because of the following factors: (1) in the absence of earthworms, competition between fungi for nutrients may limit the density of fungi on each particle; (2) low levels of earthworm activities decrease fungal competition (and therefore increase the number of isolates per particle) through increases in spatial heterogeneity or nutrient release; and (3) high levels of earthworm activities decrease numbers of fungal isolates per particle by decreasing their spatial heterogeneity (i.e., complete homogenization of the organic layers) or by repeated disruption of fungal hyphae (McLean and Parkinson 1998a). In addition, the fungal communities in the soil organic layers became less similar over time at low earthworm population densities and more similar at high earthworm densities, probably reflecting increasing then decreasing spatial heterogeneity or the stimulatory effects of low levels of fungal disruption and the destructive effects of high levels of fungal disruption (McLean and Parkinson 1998b). Data from a 2-year field experiment supported the observations of the effects of earthworms on the fungal community in mesocosms (McLean and Parkinson 2000a). The fungal community in the FH layer, the layer of highest earthworm activity, became more similar to that in the Ah horizon over time and less similar to that in the L layer with high earthworm activity (McLean and Parkinson 2000a). In the FH layer, the high earthworm populations were correlated positively with fungal dominance and negatively with fungal species richness and diversity (McLean and Parkinson 2000a). Two pieces of evidence suggesting that the source of these effects on the fungal community

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was disruption by earthworm activities are (1) fast-growing fungal species such as Trichoderma polysporum were associated with the high earthworm population treatment, suggesting that these species were tolerant of the disruptive effects of earthworm activity; and (2) there was a decrease in the frequency of occurrence of several species of Zygomycetes, which have few septa and may be particularly susceptible to cell content leakage when their hyphae are damaged (McLean and Parkinson 2000a). Soil Fauna The transformation of soil organic layers by earthworms must have strong implications for the activities of other soil invertebrates. In moder soil, soil invertebrates colonize organic layers preferentially, and few penetrate into the mineral soil. In the few experimental studies to date, significant effects of D. octaedra activities on the abundance of microarthropods, oribatid community structures, and species composition were observed in both the laboratory and field. In a 6month laboratory mesocosm experiment, oribatid species richness and diversity were low and dominance by Oppiella nova was high in the absence of the epigeic earthworm species D. octaedra. Increasing earthworm populations and activities increased oribatid diversity and decreased their dominance (McLean and Parkinson 1998b). Similarity between the oribatid communities increased in response to high earthworm populations and over time at low earthworm populations (McLean and Parkinson 1998b). Other effects of earthworm activities were also consistent with the intermediate disturbance hypothesis, including the positive effect of earthworms on Collembola populations after 3 months, followed by a negative effect after 6 months, and higher populations of Collembola and some oribatid mite species in the treatment with earthworms, accompanied by either increased abundance in the control or decreased abundance in the treatment with increased earthworm biomass (McLean and Parkinson 1998b). These different effects of earthworm activities may be due initially to an increased microhabitat diversity as earthworms added casts, urine, and mucus to the materials in the soil organic layers, followed by decreased microhabitat diversity as the organic layers became more completely homogenized (McLean and Parkinson 1998b). Data from a 2-year field experiment supported the observations made in our laboratory experiment but showed more severe effects of D. octaedra activities on the oribatid community structure, species composition, and overall microarthropod populations (McLean and Parkinson 2000b). By the end of this experiment, in the plots with the highest earthworm populations, the L2 and the FH layers were replaced entirely by earthworm casts, thus overwhelmingly decreasing the microhabitat heterogeneity of this layer. Therefore, it was not surprising that, in the FH layer, the layer of maximum earthworm activities, earthworm biomass was correlated negatively with oribatid species richness, populations of 18 oribatid species (mostly small Brachchthoniidae and Oppioidea), the populations of adult and juvenile oribatids, astigmatids, Actinedida, mesostigmatids, and Arthropleona springtails (McLean and Parkinson 2000b). By contrast, in the L layer, which was physically much less diverse than the FH layer, earthworm biomass was correlated positively with oribatid species richness and diversity, reflecting the addition of new substrates and the increasing microhabitat diversity as a result of earthworm activities (McLean and Parkinson 2000b). Migge (2001), in a mesocosm study, reported even more dramatic changes in the soil microarthropod community because of the invasion of the aspen forest floor by endogeic and anecic earthworm species. Lumbricus terrestris reduced the density of oribatid mites and springtails to 5 and 25% of those in the control (without earthworms), respectively. Overall, our knowledge of earthworm-mediated changes in soil invertebrate community composition is still marginal. Detailed field studies are urgently needed and, because previous studies have been restricted to a small number of soil invertebrate groups, information on important groups such as enchytraeids and dipterans is completely lacking. Exploring the implications of the invasion of North American forest soils by European and Asian lumbricids on the structure and functioning of the soil animal food web is an imperative for future soil research. © 2004 by CRC Press LLC

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CHANGES IN ECOSYSTEM PROCESSES AND PLANT GROWTH CAUSED BY EARTHWORMS Earthworms are generally believed to increase nutrient mineralization and thereby affect plant growth beneficially. However, most of the information on earthworm-mediated changes in nutrient mobilization and plant growth comes from agricultural systems, or crop plants, and considerably less is known about these processes in forest systems and on the effect of earthworms on tree growth (Scheu 2004). Furthermore, it has been realized only recently that the effects of earthworms on the aboveground system is not restricted to plants but propagates into the aboveground food web. As documented by Scheu et al. (1999) and Bonkowski et al. (2001), earthworms can alter the palatability of plants to herbivores and therefore influence herbivore development. Wurst and Jones (2003) presented evidence that earthworms may even affect the performance of herbivore predators such as parasitoid wasps. Overall, there is an increasing awareness that the below- and aboveground communities are much more closely linked than previously assumed (Scheu 2001; Van der Putten et al. 2001; Scheu and Setälä 2002; Wardle 2002). However, most of the sparse information available on this subject does not come from forest systems. Invasion of forest soils in North America by European earthworm species probably not only affects soil structure and the indigenous soil biota, but also affects plants and therefore the whole aboveground system (see Chapter 5, this volume). Unfortunately, information on how the invasion of the belowground system by earthworms may affect the aboveground food web is limited. Scheu and Parkinson (1994c) reported the growth of an understory grass species of aspen forests increased in presence of the litter-dwelling earthworm species D. octaedra. However, overall, they suggested that the invasion of aspen forests by European earthworms is more likely to change the plant species composition rather than the overall productivity of these forests. Similarly, Alban and Berry (1994) concluded that the productivity of aspen forests responded very little to earthworm invasions. However, in laboratory experiments, Welke and Parkinson (2003) documented that Douglas fir seedling biomass increased in the presence of populations of A. trapezoides, but these results need to be confirmed in the field. Gundale (2002) confirmed that earthworm invasions of North American forests altered plant community composition in the understory. The results of this study are of particular concern because the invasion of forests resulted in extirpation of a rare fern species, Botrychium mormo. This extirpation was associated with a transformation of moder to mull humus, which presumably was mainly because of invasion by the epi-endogeic earthworm species L. rubellus. Considering the vast spatial range of the invasion of North American soils by European earthworm species, these interrelationships need considerable further attention. The implications of the invasion of North America by European species for ecosystem processes, plant growth, and plant community composition passed almost unnoticed until recently. Studies made so far have concentrated on the effects of invasions of deciduous forests; however, there is now increasing evidence that the invasion of coniferous forests is not restricted to epigeic earthworm species such as D. octaedra (cf. McLean and Parkinson 1997), which may have few effects on plant growth and plant community composition. Pine forests in the eastern front ranges of the Canadian Rocky Mountains recently have been invaded by endogeic earthworm species such as Aporrectodea sp. (Migge unpublished data). Moreover, a range of ecological groups of European earthworms is currently invading the temperate rain forests of western North America. The sites invaded include national parks with pristine flora and fauna, such as those on the Olympic Peninsula and Vancouver Island (Scheu personal observation).

REFERENCES Alban, D.H. and E.C. Berry. 1994. Effects of earthworm invasion on morphology, carbon and nitrogen of a forest soil, Appl. Soil Ecol., 1:243–249.

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Beare, M.H., D.C. Coleman, D.A. Crossley, Jr., P.F. Hendrix, and E.P. Odum. 1994. A hierarchical approach to evaluating the significance of soil biodiversity to biogeochemical cycling, Plant Soil, 170:5–22. Bhatnagar, T. 1975. Lombriciens et humification: un aspect nouveau de l’incorporation microbienne d’azote induite par les vers de terre, in K. Gilbertus, O. Reisinger, A. Mourey, and J.A. Cancela da Fonseca, Eds., Biodégradation et Humification, Pierron, Sarreguemines, France, pp. 169–182. Bohlen, P.J., R.W. Parmelee, J.M. Blair, C.A. Edwards, and B.R. Stinner. 1995. Efficacy of methods for manipulating earthworm populations in large-scale field experiments, Soil Biol. Biochem., 27:993–999. Bonkowski, M., I.E. Geoghegan, A.N.E. Birch, and B.S. Griffiths. 2001. Effects of soil decomposer invertebrates (protozoa and earthworms) on an above-ground phytophagous insect (cereal aphid) mediated through changes in the host plant, Oikos, 95:441–450. Bonkowski, M., B.S. Griffiths, and K. Ritz. 2000. Food preferences of earthworms for soil fungi, Pedobiologia, 44:666–676. Bouché, M. 1975. Action de la faune sur les états de la matière organique dans les écosystèmes, in K. Gilbertus, O. Reisinger, A. Mourey, and J.A. Cancela da Fonseca, Eds., Biodégradation et Humification, Pierron, Sarreguemines, France, pp. 157–168. Brown, G.G. 1995. How do earthworms affect microfloral and faunal community diversity? Plant Soil, 170: 209–231. Brown, G.G., I. Barois, and P. Lavelle. 2000. Regulation of soil organic matter and microbial activity in the drilosphere and the role of interactions with other edaphic functional domains, Eur. J. Soil. Biol., 36:177–198. Cavendar, N.D., R.M. Atiyeh, and M. Knee. 2003. Vermicompost stimulates mycorrhizal colonization of Sorgum bicolour at the expense of plant growth, Pedobiologia, 47:85–90. Daniel, O. and J.M. Anderson. 1992. Microbial biomass and activity in contrasting soil materials after passage through the gut of the earthworm Lumbricus rubellus Hoffmeister, Soil Biol. Biochem., 24:465–470. Dash, H.K., B.N. Beura, and M.C. Dash. 1986. Gut load, transit time, gut microflora and turnover of soil, plant and fungal material by some tropical earthworms, Pedobiologia, 29:13–20. Dash, M.C. and J.B. Cragg. 1972. Selection of microfungi by Enchytraeidae (Oligochaeta) and other members of the soil fauna, Pedobiologia, 12:282–286. Devliegher, W. and W. Verstraete. 1997. Microorganisms and soil physico-chemical conditions in the drilosphere of Lumbricus terrestris, Soil Biol. Biochem., 29:1721–1729. Domsch, K.H., and H.J. Banse. 1972. Mykologische Untersuchungen an Regenwurmexkrementen, Soil Biol. Biochem., 4:31–38. Doube, B.M., P.M. Stephens, C.W. Davoren, and M.H. Ryder. 1994. Interactions between earthworms, beneficial soil microorganisms and root pathogens, Appl. Soil Ecol., 1:3–10. Dymond, P., S. Scheu, and D. Parkinson. 1997. Density and distribution of Dendrobaena octaedra (Lumbricidae) in an aspen and pine forest in the Canadian Rocky Mountains (Alberta), Soil Biol. Biochem., 29:265–273. Edwards, C.A., and Bohlen, P.J. 1996. Biology and Ecology of Earthworms, 3rd ed., Chapman & Hall, London. Edwards, C.A. and K.E. Fletcher. 1988. Interactions between earthworms and microorganisms in organic matter breakdown, Agric. Ecosystems Environ., 24:235–247. Ester, A. and K. van Rozen. 2002. Earthworms (Aporrectodea spp.: Lumbricidae) cause soil structure problems in young Dutch polders, Eur. J. Soil Biol., 38:283–287. Fischer, K., D. Hahn, R.I. Amann, O. Daniel, and J. Zeyer. 1995. In situ analysis of the bacterial community in the gut of the earthworm Lumbricus terrestris by whole cell hybridization, Can. J. Microbiol., 41:666–673. Fischer, K., D. Hahn, W. Honerlage, and J. Zeyer. 1997. Effect of passage through the gut of the earthworm Lumbricus terrestris L. on Bacillus megaterium studied by the whole cell hybridization, Soil Biol. Biochem., 43:1149–1152. Francis, G.S. and P.M. Fraser. 1998. The effects of three earthworm species on soil macroporosity and hydraulic conductivity, Appl. Soil Ecol., 10:11–19. Görres, J.H., M.C. Savin, and J.A. Amador. 1997. Dynamics of carbon and nitrogen, microbial biomass and nematode abundance within and outside the burrow walls of anecic earthworms (Lumbricus terrestris), Soil Sci., 162:666–671. Green, R.N., R.L. Trowbridge, and K. Klimka. 1993. Towards a Taxonomic Classification of Humus Forms, Forest Sci. Monogr., 29, Soc. Am. Forestry, Bethesda, MD.

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Gundale, M.J. 2002. Influence of exotic earthworms on the soil organic horizon and the rare fern Botrychium mormo, Conserv. Biol., 16:1555–1561. Gunnarsson, T., P. Sundin, and A. Tunlid. 1988. Importance of leaf litter fragmentation for bacterial growth, Oikos, 52:303–308. Hågvar, S. and G. Abrahamsen. 1980. Colonisation by Enchytraeidae, Collembola and Acari in sterile soil samples with adjusted pH levels, Oikos, 34:245–258. Hansen, P.A. 1989. Species response curves along a mull/mor gradient in Swedish Beech forests, Vegetatio, 82:69–78. Heath, G.W. and M.K. Arnold. 1966. Studies in leaf-litter breakdown. II. Breakdown of “sun” and “shade” leaves, Pedobiologia, 6:238–243. Hendriksen, N.B. 1990. Leaf litter selection by detritivore and geophagous earthworms, Biol. Fertil. Soils, 9:17–21. Holmstrup, M. 1992. Cold hardiness strategy in cocoons of the lumbricid earthworm Dendrobaena octaedra (Savigny), Comp. Biochem. Physiol., 102A:49–54. Holmstrup, M. and K.E. Zachariassen. 1996. Physiology of cold hardiness in earthworms, Comp. Biochem. Physiol., 115A:91–101. Hoogerkamp, M., H. Rogaar, and H.J.P. Eijsackers. 1983. Effect of earthworms on grasslands on recently reclaimed polder soils in the Netherlands, in J.E. Satchell, Ed., Earthworm Ecology: From Darwin to Vermiculture, Chapman & Hall, London, pp. 85–105. Huhta, V. and A. Koskenniemi. 1975. Numbers, biomass and community respiration of soil invertebrates in spruce forests at two latitudes in Finland, Ann. Zool. Fenn., 12:164–182. Judas, M. 1992. Gut content analysis of earthworms in a beechwood, Soil Biol. Biochem., 24:1413–1417. Kaneko, N. 1985. A comparison of oribatid mite communities in two different soil types in a cool temperate forest in Japan, Pedobiologia, 28:255–264. Kjøller, A. and S. Struwe. 1982. Microfungi in ecosystems: fungal occurrence and activity in litter and soil, Oikos, 39:389–422. Klironomos, J., P. Widden, and I. Deslandes. 1992. Feeding preferences of the collembolan Folsomia candida in relation to microfungal successions on decaying litter, Soil Biol. Biochem., 24:685–692. Kristufek, V., V. Pizl, and K. Ravasz. 1995. Epifluorescent microscopy of earthworms intestinal bacteria, Acta Microbiol. Imumunol. Hung., 42:39–44. Kristufek, V., K. Tajovsky, and V. Pizl. 1994. Ultrastructural analysis of the intestinal content of earthworm Lumbricus rubellus Hoffm. (Annelida, Lumbricidae), Acta Microbiol. Imumunol. Hung., 41:283–290. Kubiena, W.L. 1955. Animal activity as a decisive factor in establishment of humus forms, in D.K.McE. Kevan, Ed., Soil Zoology, Butterworths Scientific, London, pp. 73–82. Langmaid, K.K. 1964. Some effects of earthworm invasion in virgin podsols, Can. J. Soil Sci., 44:34–37. Lavelle, P. 1988. Earthworm activities and the soil system, Biol. Fertil. Soils, 6:237–251. Lawrence, B., M.C. Fisk, T.J. Fahey, and E.R. Suárez. 2003. Influence of nonnative earthworms on mycorrhizal colonization of sugar maple (Acer saccharum), New Phytol., 157:145–153. Lee, K.E. 1985. Earthworms, Their Ecology and Relationships with Soils and Land Use, Academic Press, North Ryde, New South Wales, Australia. Lee, K.E. and R.C. Foster. 1991. Soil fauna and soil structure, Aust. J. Soil Res., 29:745–775. Loquet, M. 1978. The study of respiratory and enzymatic activities of earthworm-made pedological structures in a grassland soil at Citeaux, France, Sci. Proc. Roy. Dublin Soc., Ser. A6:207–214. Maraun, M., J. Alphei, M. Bonkowski, R. Buryn, S. Migge, M. Peter, M. Schaefer, and S. Scheu. 1999. Middens of the earthworm Lumbricus terrestris (Lumbricidae): microhabitats for micro- and mesofauna in forest soil, Pedobiologia, 43:276–287. Marinissen, J.C.Y. and J. Bok. 1988. Earthworm-amended soil structure: its influence on Collembola populations in grassland, Pedobiologia, 32:243–252. Marinissen, J.C.Y. and F. Van den Bosch. 1992. Colonization of new habitats by earthworms, Oecologia, 91:371–376. McLean, M.A., N. Kaneko, and D. Parkinson. 1996. Does selective grazing by mites and Collembola affect litter fungal community structure? Pedobiologia, 40:97–105. McLean, M.A., D.U. Kolodka, and D. Parkinson. 1996. Survival and growth of Dendrobaena octaedra in pine forest floor materials, Pedobiologia, 40:281–288.

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McLean, M.A. and D. Parkinson. 1997. Soil impacts of the epigeic earthworm Dendrobaena octaedra on organic matter and microbial activity in lodgepole pine forest, Can. J. For. Res., 27:1907–1913. McLean, M.A. and D. Parkinson. 1998a. Impacts of the epigeic earthworm Dendrobaena octaedra on microfungal community structure in pine forest floor — a mesocosm study, Appl. Soil Ecol., 8:61–75. McLean, M.A. and D. Parkinson. 1998b. Impacts of epigeic earthworm Dendrobaena octaedra on oribatid mite community diversity and microarthropod abundances in pine forest floor: a mesocosm study, Appl. Soil Ecol., 7:125–136. McLean, M.A. and D. Parkinson. 2000a. Field evidence of the effects of the epigeic earthworm Dendrobaena octaedra the microfungal community in pine forest floor, Soil Biol. Biochem., 32:351–360. McLean, M.A. and D. Parkinson. 2000b. Introduction of the epigeic earthworm Dendrobaena octaedra changes the oribatid community and microarthropod abundances in a pine forest, Soil Biol. Biochem., 32: 1671–1681. Migge, S. 2001. The Effect of Earthworm Invasion on Nutrient Turnover, Microorganisms and Microarthropods in Canadian Aspen Forest Soil, Technische Universität Darmstadt, Darmstadt, Germany. Moody, S.A., M.J.I. Briones, T.G. Piearce, and J. Dighton. 1995. Selective consumption of decomposing wheat straw by earthworms, Soil Biol. Biochem., 27:1209–1213. Moody, S.A., T.G. Piearce, and J. Dighton. 1996. Fate of some fungal spores associated with wheat straw decomposition on passage through the guts of Lumbricus terrestris and Aporrectodea longa, Soil Biol. Biochem., 28:533–537. Moore, J.C., E.R. Ingham, and D.C. Coleman. 1987. Inter- and intra-specific feeding selectivity of Folsomia candida (Willem) (Collembola, Isotomidae) on fungi, Biol. Fertil. Soils, 5:6–12. Müller, P.E. 1878. Studier over Skovjord, Tidskr. Skovbrug, 3:1–124. Muys, B. and N. Lust. 1992. Inventory of the earthworm communities and the state of litter decomposition in the forests of Flanders, Belgium, and its implications for forest management, Soil Biol. Biochem., 24:1677–1681. Muys, B., N. Lust, and Ph. Granval. 1992. Effects of grassland afforestation with different tree species on earthworm communities, litter decomposition and nutrient status, Soil Biol. Biochem., 24:1459–1466. Parkinson, D. 1988. Linkages between resource availability, microorganisms and soil invertebrates, Agric. Ecosyst. Environ., 24:21–32. Parle, J.N. 1963. Microorganisms in the intestines of earthworms, J. Gen. Microbiol., 31:1–11. Pattinson, G.S., S.E. Smith, and B.M. Doube. 1997. Earthworm Aporrectodea trapezoides had no effect on the dispersal of vesicular-arbuscular mycorrhizal fungus Glomus intraradices, Soil Biol. Biochem., 29:1079–1088. Pawluk, S. 1985. Soil micromorphology and soil fauna: problems and importance, Quaestiones Entomol., 21:473–496. Petersen, H. 1982. Structure and size of soil animal populations, Oikos, 39:306–329. Piearce, T.G. 1972. Acid intolerant and ubiquitous Lumbricidae in selected habitats in North Wales, J. Anim. Ecol., 41:397–410. Puttarudriah, M. and K.S. Shivashankara Sastry. 1961. A preliminary study of earthworm damage to crop growth, Mysore Agric. J., 36:2–11. Rabatin, S.C. and Stinner, B.R. 1988. Indirect effects of interactions between VAM fungi and soil-inhabiting invertebrates on plant processes, Agric. Ecosyst. Environ., 24:135–146. Ramann, E. 1911. Bodenkunde, 3rd ed., Julius Springer, Berlin. Reddell, P. and A.V. Spain. 1991a. Earthworms as vectors of viable propagules of mycorrhizal fungi, Soil Biol. Biochem., 23:767–774. Reddell, P. and A.V. Spain. 1991b. Transmission of infective Frankia (Actinomycetales) propagules in casts of the endogeic earthworm Pontoscolex corethrurus (Oligochaeta: Glossoscolecidae), Soil Biol. Biochem., 23:775–778. Reynolds, J.W. 1977. The Earthworms (Lumbricidae and Sparganophilidae) in Ontario, Life Sciences, Miscellancus Publications, Royal Ontario Museum (ROM), Ottawa, Canada. Rusek, J. 1985. Soil microstructures — contributions on specific soil organisms, Quaestiones Entomol., 21: 497–514. Salmon, S. 2001. Earthworm excreta (mucus and urine) affect the distribution of springtails in forest soils, Biol. Fertil. Soils, 34:304–310.

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Schaefer, M. and J. Schauermann. 1991. The soil fauna of beech forests: comparison between a mull and a moder soil, Pedobiologia, 34:299–314. Scheu, S. 1992. Automated measurement of the respiratory response of soil microorganisms: active microbial biomass in earthworm faeces, Soil Biol. Biochem., 24:113–118. Scheu, S. 2001. Plants and generalist predators as links between the below-ground and above-ground system, Basic Appl. Ecol., 2:3–13. Scheu, S. 2004. Effects of earthworms on plant growth: patterns and perspectives, Pedobiologia, in press. Scheu, S. and M.A. McLean. 1993. The earthworm (Lumbricidae) distribution in Alberta (Canada), Megadrilogica, 4:175–180. Scheu, S. and D. Parkinson. 1994a. Changes in bacterial and fungal biomass C, bacterial and fungal biovolume and ergosterol content after drying, remoistening and incubation of different layers of cool temperate forest soil, Soil Biol. Biochem., 26:1515–1525. Scheu, S. and D. Parkinson. 1994b. Effects of earthworms on nutrient dynamics, carbon turnover and microorganisms in soils from cool temperate forests of the Canadian Rocky Mountains — laboratory studies, Appl. Soil Ecol., 1:113–125. Scheu, S. and D. Parkinson. 1994c. Effects of an invasion of an aspen forest (Canada) by Dendrobaena octaedra (Lumbricidae) on plant growth, Ecology, 75:2348–2361. Scheu, S. and M. Schaefer. 1998. Bottom-up control of the soil macrofauna community in a beechwood on limestone: manipulation of food resources, Ecology, 79:1573–1585. Scheu, S. and H. Setälä. 2002. Multitrophic interactions in decomposer communities, in T. Tscharntke and B.A. Hawkins, Eds., Multitrophic Level Interactions, Cambridge University Press, Cambridge, U.K., pp. 223–264. Scheu, S., A. Theenhaus, and T.H. Jones. 1999. Links between the detritivore and the herbivore system: effects of earthworms and Collembola on plant growth and aphid development, Oecologia, 119:541–551. Schönholzer, F., D. Hahn, and J. Zeyer. 1999. Origins and fate of fungi and bacteria in the gut of Lumbricus terrestris L. studied by image analysis, FEMS Microbiol. Ecol., 28:235–248. Schrader, S. 1994. Influence of earthworms on the pH of their environment by cutaneous mucus secretion, Zool. Anz., 233:211–219. Schulmann, O.P. and A.V. Tiunov. 1999. Leaf litter fragmentation by the earthworm Lumbricus terrestris L., Pedobiologia, 43:453–458. Seastedt, T.R. 1984. The role of microarthropods in decomposition and mineralization processes, Ann. Rev. Entomol., 29:25–46. Seastedt, T.R., M.V. Reddy, and S.P. Cline. 1989. Microarthropods in decaying wood from temperate coniferous and deciduous forests, Pedobiologia, 33:69–77. Shaw, C. and S. Pawluk. 1986. The development of soil structure by Octolasion tyrtaeum, Aporrectodea turgida and Lumbricus terrestris in parent materials belonging to different textural classes, Pedobiologia, 29:327–339. Shaw, P.J.A. 1988. A consistent hierarchy of fungal feeding preferences of the Collembola Onychiurus armatus, Pedobiologia, 31:179–187. Stockdill, S.M.J. 1982. Effects of introduced earthworms on the productivity of New Zealand pastures, Pedobiologia, 24:29–35. Swift, M.J., O.W. Heal, and J.M. Anderson. 1979. Decomposition in Terrestrial Ecosystems, Blackwell Scientific, Oxford, U.K. Tiunov, A.V. and T.G. Dobrovolskaya. 2002. Fungal and bacterial communities in Lumbricus terrestris burrow walls: a laboratory experiment, Pedobiologia, 46:595–605. Tiunov, A.V., T.G. Dobrovolskaya, and L.M. Polyanskaya. 2001. Microbial complexes associated with inhabited and abandoned burrows of Lumbricus terrestris in a soddy-podzolic soil, Eurasian Soil Sci., 34:525–529. Tiunov, A.V. and N.A. Kuznetsova. 2000. Environmental activity of anecic earthworms (Lumbricus terrestris L.) and spatial organization of soil communities, Bio. Bull., 27:510–518. Tiunov, A.V. and S. Scheu. 1999. Microbial respiration, biomass, biovolume, and nutrient status in burrow walls of Lumbricus terrestris L., Soil Biol. Biochem., 31:2039–2048. Tiunov, A.V. and S. Scheu. 2000a. Microbial biomass, biovolume and respiration in Lumbricus terrestris L. cast material of different age, Soil Biol. Biochem., 32:265–275.

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Tiunov, A.V. and S. Scheu. 2000b. Microfungal communities in soil, litter and casts of Lumbricus terrestris L. (Lumbricidae): a laboratory experiment, Appl. Soil Ecol., 14:17–26. Tiwari, S.C. and R.R. Mishra. 1993. Fungal abundance and diversity in earthworm casts and in uningested soil, Biol. Fertil. Soils, 16:131–134. Toyota, K. and M. Kimura. 1994. Earthworms disseminate a soil-borne plant pathogen, Fusarium oxysporum f. sp. raphani, Biol. Fertil. Soils, 18:32–36. Trigo, D., M.H. Garvin, E. Huerta, S. Irisson, and P. Lavelle. 1999. Mutualism between earthworms and soil microflora, Pedobiologia, 43:866–873. Trigo, D. and P. Lavelle. 1993. Changes in respiration rate and some physicochemical properties of soil during gut transit through Allobophora moelleri (Lumbricidae, Oligochaeta), Biol. Fertil. Soils, 15:185–188. Trigo, D. and P. Lavelle. 1995. Soil changes during gut transit through Octolasion lacteum Orley (Lumbricidae, Oligochaeta), Acta Zool. Fenn., 196:129–131. Van der Putten, W.H., L.E.M. Vet, J.A. Harvey, and F.L. Wäckers. 2001. Linking above- and belowground multitrophic interactions of plants, herbivores, pathogens, and their antagonists, Trends Ecol. Evol., 16:547–554. Visser, S. 1985. Role of invertebrates in determining the composition of soil microbial communities, in A.H. Fitter, D. Atkinson, D.J. Read, and M.B. Usher, Ed., Ecological Interactions in Soil: Plants, Microbes and Animals, Blackwell Scientific, Oxford, pp. 297–317. Visser, S. and D. Parkinson. 1975. Fungal succession on aspen poplar leaf litter, Can. J. Bot., 53:1640–1651. Visser, S. and J.B. Whittaker. 1977. Feeding preferences for certain litter fungi by Onychiurus subtenuis (Collembola), Oikos, 29:320–325. Walter, D.E. and H.C. Proctor. 1999. Mites: Ecology, Evolution and Behaviour, CAB International, Wallingford, U.K. Warcup, J.H. 1965. Growth and reproduction of soil microorganisms in relation to substrate, in K.F. Baker and W.C. Snyder, Eds., Ecology of Soil-Borne Plant Pathogens, Prelude to Biological Control, University of California Press, Berkeley, pp. 52–68. Wardle, D.A. 2002. Linking the Aboveground and Belowground Components, Princeton University Press, Princeton, NJ. Welke, S.E. and D. Parkinson. 2003. Effect of Aporrectodea trapezoides activity on seedling growth Pseudotsuga menziesii, nutrient dynamics and microbial activity in different forest soils, For. Ecol. Manage., 173:169–186. Widden, P. 1986. Microfungal community structure from forest soils in southern Quebec, using discriminant function and factor analysis, Can. J. Bot., 64:1402–1412. Wolter, C. and S. Scheu. 1999. Changes in bacterial number and hyphal lengths during the gut passage through Lumbricus terrestris (Lumbricidae, Oligochaeta), Pedobiologia, 43:891–900. Wright, M.A. 1972. Factors governing ingestion by the earthworm Lumbricus terrestris L. with special reference to apple leaves, Ann. Appl. Biol., 70:175–188. Wurst, S. and T.H. Jones. 2003. Indirect effects of earthworms (Aporrectodea caliginosa) on an above-ground tritrophic interaction, Pedobiologia, 47:91–97. Yeates, G.W. 1981. Soil nematode populations depressed in the presence of earthworms, Pedobiologia, 22:191–195. Yeates, G.W. 1988. Earthworm and enchytraeid populations in a 13 year old agroforestry system, N.Z. J. For. Sci., 18:304–310. Yeates, G.W. 1991. Impact of historical changes in land use on the soil fauna, N.Z. J. Ecol., 15:99–106.

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Part VII Earthworms in Agroecosystems

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Earthworms as a 14 Managing Resource in Australian Pastures Geoff H. Baker CSIRO Entomology, Canberra, Australia

CONTENTS Introduction ....................................................................................................................................263 The Earthworm Fauna in Australia ...............................................................................................264 Effects of Earthworms on Soil Properties and Plant Productivity in Australia ...........................267 Effects of Agricultural Management Practices on Earthworms....................................................270 Introductions of Earthworm Taxa to New Areas...........................................................................275 The Potential for Introducing New Earthworm Species into Australia........................................277 Earthworms in Pastures in Northern Australia..............................................................................277 Earthworms as Indicators of the Sustainability of Agriculture: From the Farmer’s Perspective..................................................................................................278 Conclusions ....................................................................................................................................278 References ......................................................................................................................................279

INTRODUCTION Earthworms are well known for their abilities to improve soil structure, fertility, and agricultural production (Lee 1985; Edwards and Bohlen 1996). For example, research in New Zealand has shown that introduction of earthworms to pastures that are lacking them can enhance pasture production by up to 25% in the long term (Stockdill 1982). This improvement in grassland production in New Zealand resulted especially from the introduced earthworms feeding on a thick mat of dead organic matter that had accumulated at the soil surface. The breakdown of this organic mat returned nutrients to the soil and enhanced water infiltration. Comparable research in northern Tasmania (Temple-Smith 1991) and on-farm applications (D. Ford and B. Farquar personal communication 1992) have also demonstrated that similar increases in pasture production can be achieved in these regions using the same technologies as those used in New Zealand (inoculating pastures by spreading sods of soil containing earthworms). In addition, significant increases in plant production have resulted from introductions of earthworms into agricultural soils in several other countries (e.g., Ireland, Netherlands, United States) (Baker 1998a). Agricultural soils in southern mainland Australia are generally poor in structure and fertility. The work in New Zealand stimulated a flurry of research in the late 1980s and early 1990s aimed at improving the management of earthworms as a resource in agricultural soils in southeastern Australia (Temple-Smith and Pinkard 1996). Several factors further encouraged this expansion in earthworm research. Increased on-farm costs (fuel, labor, machinery), reduced values of agricultural

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products, and greater awareness of soil structural decline encouraged farmers to adopt reduced cultivation techniques. Under such practices, populations abundance of earthworms are enhanced (Rovira et al. 1987), and their presence in optimal numbers is needed to replace some of the benefits previously achieved by the plough. In addition, the increasing costs of inorganic fertilizers, as well as the pollution problems they cause (e.g., through leaching and erosion into waterways), has stimulated thought on more efficient and safer means of transfer of nutrients from natural sources to plants. Studies in other countries that have shown the potential for earthworms to help offset soil degradation, such as improved lime burial and reductions in soil acidity, improved water infiltration, and increased rates of breakdown of surface litter, leading to reduced surface runoff of phosphorus and nitrogen (Sharpley et al. 1979; Springett 1983), also heightened interest in introducing such benefits to Australia. During the late 1980s and through the 1990s, research on earthworms in southern Australia was aimed primarily at determining the distribution and abundance of earthworms in agricultural soils, measuring the effects of the most common species on soil properties and plant productivity, and identifying means by which the beneficial role of earthworms can be enhanced (e.g., optimal farm management practices, introduction of new taxa of earthworms) (Baker 1998a). This chapter considers briefly the progress that has been made on these topics and identifies gaps in knowledge that could, if filled, lead to benefits for the grazing industries, because the focus here is on pastures. There is also substantial literature available on earthworm ecology and activities in grain cropping systems in southern Australia (e.g., Buckerfield 1992; Baker et al. 1993c, 1997b, 2004a; Doube et al. 1994a, 1997; Stephens et al. 1994a, 1995). Such information is referred to here only if it is particularly relevant. The majority of earthworm research that has been done in Australia has been in the southern temperate and mediterranean climatic regions. Knowledge of the biology, role, and management of earthworms in tropical grazing systems in Australia is more or less uncharted territory. Parallels may be sought in the extensive studies of Lavelle and colleagues in Africa and Central and South America (Lavelle et al. 1999).

THE EARTHWORM FAUNA IN AUSTRALIA Extensive surveys have demonstrated that populations and species diversity of earthworms are generally low in soils used for pastures in southern Australia (Kingston and Temple-Smith 1989; Mele 1991; Baker et al. 1992b, 1994; Garnsey 1994b; Lobry de Bruyn and Kingston 1997; Baker 1998a; Mele and Carter 1999b). Populations of more than 1200 earthworms m–2 have been recorded occasionally, but most pastures have fewer than 200 earthworms m–2. Although up to six species have been recorded per field, more commonly only two to three species are found. The earthworm fauna is dominated by exotic species, most notably Lumbricidae (e.g., Aporrectodea caliginosa, Aporrectodea trapezoides, Aporrectodea rosea), which have been introduced accidentally from Europe. The dominant species vary regionally and among habitat types. For example, A. trapezoides is dominant in permanent pastures in South Australia (S.A.) and southern New South Wales (N.S.W.), but A. caliginosa is more abundant in similar pastures in western Victoria. In pasture-cereal rotations in the same regions, A. rosea is the dominant species. Indigenous, native species (Megascolecidae), which are common in undisturbed, native habitats, are generally much more rare in agricultural soils than are exotic species. Reasons for this relative rarity of native species in managed soils are poorly understood, but tillage and changes in shelter, food type, and soil fertility have been suggested as possibly important factors. Interestingly, native species are more common in pasture soils in Victoria and southern N.S.W. (e.g., 42% here are native) than they are in S.A. and southwestern West Australia (W.A.) (Figure 14.1). Reasons for this cline in species richness are not clear, but possibly it is correlated with differences in summer aridity. Summers in S.A. and southwestern W.A. are hot and dry, with most

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(a) Numbers of Earthworms (m–2)

700 600 500 400 300 200 100 0

(b) 900 Numbers of Earthworms (m–2)

800 700 600 500 400 300 200 100 0

(c)

450

Numbers of Earthworms (m–2)

400 350 300 250 200 150 100 50 0

FIGURE 14.1 Average abundance of exotic (open bars) and native (closed bars) earthworms in pastures in (a) the Mount Lofty Ranges, South Australia (113 sites); (b) near Mortlake, Victoria (163 sites); and (c) in the southern tablelands of New South Wales (104 sites). Sites are arranged in order from those with the least total earthworms to those with the most.

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rain falling during the winter months. Further east in Victoria and southern N.S.W., rain falls more evenly throughout the year, and perhaps the resulting moister soils for more months in the year are more suitable for building up earthworm communities. Intriguingly, native earthworms were rare when total earthworm populations were highest in Victoria and vice versa (Figure 14.1b). The sites with the highest total earthworm population corresponded with the regions with the highest rainfall and included a predominance of dairy pastures rather than sheep pastures, which were the norm elsewhere. Whether the geographical switch in bias of the earthworm communities reflects the abilities of exotic and native earthworm species to cope with climatic or agricultural management variables or whether the patterns (in particular for the native species) relate to preagricultural habitat types is not known. The possible role of competition among different earthworm species in determining the observed distribution patterns is also not understood. Although exotic species tend to dominate earthworm communities in pastures in southern Australia, many pastures totally lack exotic species. It may be that such absences of exotic species simply reflect a lack of opportunity to colonize them to date rather than that they are unsuitable habitats. Environmental factors that determine the geographic distribution and population of the earthworm fauna are poorly understood. Many autocorrelated variables (climatic, edaphic, land use) are related weakly to earthworm abundance. Rainfall and soil particle size, somewhat understandably, may explain most of the variance (Baker 1998a). Baker et al. (1997a) conducted a national survey, “Earthworms Downunder,” involving schoolchildren in 1992. This survey included sampling urban gardens as well as agricultural land. Lumbricid earthworms predominated in both habitat types in southern Australia, but north of the Tropic of Capricorn they were rare and were replaced by other introduced species such as Pontoscolex corethrurus (Glossoscolecidae). Thus, the most common earthworm species in disturbed land in Australia are the peregrine lumbricid species identified by Lee (1985) and are similar to those occurring in comparable habitats elsewhere in the world. The common earthworm species in agricultural soils in southern Australia are endogeic, feeding predominantly on decomposing organic matter that is already incorporated into the mineral soil layer. These earthworms are active in the top 10 cm of soil for about 4 to 5 months of the year (early winter to early spring), when the soils are moistest (Baker et al. 1992a, 1993c,d; Garnsey 1994b). During the drier summer months, most earthworms are inactive and deep in the soil. There are very few anecic species in the earthworm fauna (i.e., species that feed at the soil surface and burrow deeply during the active season). Such species have the potential to influence soil properties markedly at depth (e.g., porosity) (see Chapter 10, this volume), thus encouraging deeper penetration of water, nutrients, and rooting of plants. In contrast to this paucity of anecic earthworm species in pasture soils in southern Australia, earthworm communities in similar habitats in other parts of the world are commonly dominated by anecic species (e.g., 70% of the earthworm biomass) (Lavelle 1983). One anecic species, Aporrectodea longa, is common in pastures in northern Tasmania (Baker 1998a) (see further comments on the potential of extending the distribution of this species “Introductions of Earthworm Taxa to New Areas”). Epigeic earthworm species, those that live near the soil surface and feed on recently produced dead organic matter, are patchy in distribution and abundance in southern Australia, but Lumbricus rubellus can be very abundant under moist conditions. Microscolex dubius is more widespread but rarely abundant. It survives summer as resistant cocoons (i.e., eggs) in the dry surface soil (Doube and Auhl 1998). Several of the rarer native species of earthworms that occur in pastures are probably anecic, but little attention has been paid to the ecology and behavior of these species. The giant Gippsland earthworm Megascolides australis (Megascolecidae) is perhaps the best-known native earthworm in Australia because of its size and protected status, but even this species has been poorly studied (van Praagh 1992).

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EFFECTS OF EARTHWORMS ON SOIL PROPERTIES AND PLANT PRODUCTIVITY IN AUSTRALIA There is a great variety of ways in which earthworms can influence soil properties and plant productivity (Lee 1985; Lavelle 1988; Curry 1994; Edwards and Bohlen 1996). Several studies have been made of the influences of the most common earthworm species in agricultural soils in southern Australia: on soil structure (e.g., Barley 1959c; Doube et al. 1994b,e; Hindell et al. 1994a,b,c, 1997; Hirth et al. 1994, 1996; Friend and Chan 1995; Chan et al. 1997; Curry and Baker 1998), nutrient availability (Barley and Jennings 1959; Baker et al. 2004a), burial of surface organic matter and lime (Barley 1959b; Baker et al. 1993e, 1998b, 1999a), distribution of beneficial microorganisms (Stephens et al. 1993b, 1994a,b; Doube et al. 1994c,d; Stephens and Davoren 1994), reduction of incidence of root diseases (Stephens et al. 1993a, 1995; Stephens and Davoren 1997), and plant yield and quality (Abbott and Parker 1981; TempleSmith et al. 1993; Garnsey 1994a; Stephens et al. 1994a; Baker et al. 1997b, 1999b). Two of these topics are addressed in more detail in this chapter: the influence of earthworms on pasture production and the burial of surface-applied lime and organic matter. Laboratory and field experiments have shown that some exotic earthworm species (mostly Lumbricidae) can substantially improve the availability of soil nutrients and the quality and quantity of pasture and crop productivity in Australia. For example, Baker (1998a) demonstrated that the anecic earthworm A. longa could increase pasture yields by as much as 60% within field cages at one site in the Mt. Lofty Ranges, S.A., 5 months after introduction. Similar trials throughout S.A., Victoria, and southern N.S.W. also demonstrated that the endogeic species A. caliginosa and A. trapezoides, as well as A. longa, can increase pasture production substantially (Baker et al. 1996, 1999b, 2002b; Chan and Baker unpublished data) (Figure 14.2). The degree of increase varies

% Increase in Pasture Production

70 60

A. caliginosa A. trapezoides A. longa

50 40 30 20 10 0 8

15

30

45

Initial Numbers of Worms/Cage

FIGURE 14.2 Percentage increases in pasture production when varying numbers of earthworms were added to field cages in South Australia and Victoria. Comparisons were with controls (no earthworms added). Data are for approximately 5 months of exposure to the earthworms and are averages for seven sites. (Data redrawn from Baker et al. 1999b.)

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positively with earthworm population density and biomass. Aporrectodea caliginosa and A. trapezoides increased pasture production per unit of biomass more than A. longa, possibly because of greater exploitation of available resources by larger populations of the two smaller species. A few studies have shown that induced increases in pasture production are additive among earthworm species (Baker 1998b), that is, species (functional) diversity within earthworm communities matters. However, much more work is needed to substantiate this conclusion fully. The influence of native earthworms (Megascolecidae) on pasture production has been relatively poorly studied. In general, native earthworm species have failed to influence plant growth significantly (Baker et al. 1996; Baker 1998a and Baker unpublished data), but in one study (Baker et al. 2004a), Spenceriella macleayi did increase ryegrass growth slightly in a glasshouse experiment. However, a much larger native earthworm species, Spenceriella hamiltoni, had no effect on ryegrass growth in the same experiment. The majority of research on the effects of earthworms on pasture production in Australia either has not discriminated among pasture grass species or has focused on ryegrass. Research on grain crops in Australia (e.g., Baker et al. 2004a) and overseas (Brown et al. 1999) has demonstrated that the effects of earthworms vary among crop types. In the main, legumes are less responsive to earthworm activities than are cereals. The abilities of legumes to fix their own nitrogen may obviate the “need” for earthworms to some extent. Some studies have demonstrated how plant growth responses to earthworms can vary among sites and soil types, ranging from responses that are mostly positive, through neutral, to occasional negative ones (Doube et al. 1997; Baker et al. 1999b). Reasons for this variability in responses are not at all clear. Most studies of the abilities of earthworms to enhance the transfer of nutrients from fertilizers and organic residues to agricultural plants in Australia have been confined to grain crops (Stephens et al. 1994a; Baker, unpublished data; Baker and Amato 2001; Baker et al. 1997b, 2004a). The mechanisms are poorly understood, but some earthworm species (e.g., A. trapezoides) are known to release more N from decomposing organic matter into the soil than others (e.g., A. rosea). Whalen et al. (1999) showed that N from dead earthworms can be transferred rapidly to plants, and they argued for care in interpreting data on the effects of earthworms on plant growth in experiments in which earthworms die. Figure 14.3 illustrates responses in growth of wheat when different 12

12

Wheat Yield/Pot (g)

10

B

10

A

A

8

B

B AB

8

6

6

4

4

2

2

0

Earthworm Biomass/Pot (g)

Straw Grain Earthworms

0 0

2

4

6

8

10

Dead Earthworms/Pot

FIGURE 14.3 Average grain and straw yields of wheat when grown in a red-brown earth soil in flower pots (3 kg soil per pot) in the presence of varying numbers of dead worms (Aporrectodea trapezoides, killed in boiling water). Grain and straw yield are dry weights; worm biomass is fresh weight. Different letters at the tops of histograms indicate significant differences.

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numbers of dead earthworms (A. trapezoides) were inoculated into a red-brown earth soil. Straw yields (grams) responded to the dead earthworms (at six or more earthworms per pot), but the grain did not. The N content of both the straw and grain also increased significantly above that in the controls in response to six or more dead earthworms per pot. Of course, different responses may well occur for different plant species in different soil types based on the initial nutrient status. In a similar experiment in a yellow clay soil, ryegrass did not respond in terms of shoot growth when 2 to 10 dead A. trapezoides were added per pot. Some of the pasture yield increases illustrated in Figure 14.2 (particularly at the highest inoculation density) may be explained by a fertilizing effect of earthworms that did not survive the experiment, but at the lower earthworm inoculation densities, survival was sufficient to discount this explanation reasonably. Some Australian experiments have explored interactions between earthworms and pasture roots (Hirth et al. 1997, 1998), with those reported from similar studies in New Zealand (e.g., Springett and Gray 1997). Hirth et al. (1997) could find no evidence that ryegrass roots elongated in the direction of soil pores filled with earthworm casts. On the other hand, Hirth et al. (1998) reported that A. rosea was attracted to ryegrass roots. Springett and Gray (1997) argued that interactions between roots and earthworms were two way. As in many other countries, soil acidity is a major environmental problem in high-rainfall regions of Australia (Coventry 1985; Chartres et al. 1992). The use of ammonium-based nitrogenous fertilizers and nitrogen-fixing legumes has contributed significantly to soil acidification. Lime is applied commonly to the surface of the soil to offset acidity but is generally slow to be incorporated into the root zone where it is needed (Helyar 1991). It is often too costly or inappropriate to incorporate lime mechanically using tillage (e.g., in permanent pastures on steep slopes). Research in New Zealand (Stockdill and Cossens 1966; Springett 1983, 1985) has shown that some species of earthworms have the potential to bury lime and increase the soil pH. Similar experiments have recently been done in southeastern and southwestern Australia to determine the earthworm species most likely to be useful in reducing soil acidity in this way (Baker 1998a; Baker et al. 1993e, 1998b, 1999a; Chan and Baker unpublished data). The results suggested that the endogeic species A. trapezoides and A. caliginosa can be effective in burying surface-applied lime into the top few centimeters of soil, but the anecic species A. longa is much more effective in burying lime deeper into the profile (e.g., to a depth of 15 cm within a few months with high earthworm populations) (Figure 14.4). Some other earthworm species (e.g., A. rosea) are ineffective in lime transport. Baker et al. (1993e, 1999a) explained such differences between earthworm species in terms of their relative surface activities and depth of their burrows. Aporrectodea longa greatly disturbs the soil surface during its feeding and creates surface-venting pores down which lime particles can be washed by rainwater. Similar transport of other surface-applied materials (e.g., gypsum, fertilizers) is likely but has not been tested. Chan and Baker (unpublished data) have demonstrated that the activities of mixes of endogeic and anecic earthworm species can be complementary in terms of moving lime throughout the soil profile. Large amounts of cattle and sheep dung accumulate on the surface of Australian pastures (Waterhouse 1974). As well as fouling pasture growth and increasing fly populations, this dung represents an inefficient return of plant nutrients to the soil. Many species of exotic dung beetles have been introduced to Australia to encourage the burial of cattle dung (Tyndale-Biscoe 1990), but the role of earthworms in this process has been largely ignored (Ferrar 1975). Holter (1979) showed that A. longa, working in concert with dung beetles, was particularly effective in burying cattle dung in a Danish pasture, and Martin and Charles (1979) demonstrated that A. caliginosa and L. rubellus buried large amounts of both cattle and sheep dung in New Zealand pastures. The influence of earthworms on the burial of sheep dung has been measured in southern Australia on a few occasions in cage experiments (Baker 1998a). Aporrectodea longa is clearly very efficient at burying such dung, more so than several other earthworm species, especially some native species. However, the influence of earthworms on cattle dung burial and on any dung type at a field scale in Australian pastures remains unclear. © 2004 by CRC Press LLC

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6.5 Control A. caliginosa A. longa A. trapezoides

Soil pH

5.5

*

* *

4.5

*

*

3.5 0–2.5 cm

2.5–5 cm

5–10 cm

10–15 cm

15–20 cm

Soil Depth

FIGURE 14.4 Average soil pH (in CaCl2) at varying depths within field cages containing different earthworm species. Cages were maintained in a pasture on a sandy loam at Strathdownie, Victoria. There were 45 earthworms added per cage. All treatments had an equivalent of 4 t lime added to the soil surface. Soil pH was measured approximately 5 months after adding earthworms and lime. Asterisks indicate significant differences compared with data for the control treatment (no earthworms) at the same soil depth. Soil pH in the absence of lime is indicated by horizontal dashed lines. (Data redrawn from Baker et al. 1999a.)

Optimal use of key plant nutrients is central to sustainable agricultural production and catchment health, particularly the development of systems that maximize the uptake of applied fertilizer into plants and the recycling of nutrients from plant residues; minimizing off-site losses and nontarget impacts through leaching and surface runoff is viewed by farmers and the broader community as a high priority. Research in New Zealand (Sharpley et al. 1979) has shown that earthworms can bury surface organic matter in pastures rapidly and reduce N and P loss (leached out of the dead organic matter) from sloped agricultural fields (by factors of 4 to 8). Similar research to that in New Zealand has not been done in Australia, but it seems likely that similar benefits would accrue through the retention of nutrients on farms if earthworm communities were managed well. Certainly, opportunities exist for earthworm treatments to be imposed on surface-runoff trials that are currently under way in Australian pastures to test for impacts.

EFFECTS OF AGRICULTURAL MANAGEMENT PRACTICES ON EARTHWORMS It is well known worldwide that agricultural management practices such as drainage, irrigation, lime, fertilizer and slurry application, pesticide use, stocking rate, tillage, crop rotation, and stubble retention can influence earthworm populations and biomass (Lee 1985; Lavelle et al. 1989; Curry 1994; Fraser 1994; Edwards and Bohlen 1996). Rovira et al. (1987) demonstrated that populations of earthworms in a red-brown earth soil in S.A. were doubled by the direct drilling of cereals, in contrast with those under conventional cultivation. Fewer earthworms were found in soils under a lupin-wheat rotation than under a pasturewheat rotation. Haines and Uren (1990), Buckerfield (1993b, 1994), Buckerfield and Wiseman (1997), and Mele and Carter (1999a) provided further evidence that tillage reduces earthworm populations in Australia and that pastures in rotation with cereals and the retention of stubbles can increase earthworm abundance.

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Tillage can reduce earthworm populations in a variety of ways. Earthworms may be damaged directly by machinery or exposed to predation by birds or adverse weather, their burrow systems may be disrupted, or the availability of suitable food may be reduced (Edwards and Lofty 1982a; Springett 1983; Lee 1985). Baker et al. (1999c) inoculated earthworms into undisturbed and disturbed soil (tilled to a fine tilth) in a pasture in S.A. and recorded the survival of the earthworms. Although the disturbance greatly reduced the survival of A. longa and A. trapezoides, it had no effect on the survival of A. caliginosa and A. rosea. Baker et al. (1999c) speculated that the tillage could have induced a relative shortage of food (Andrewartha and Browning 1961) for earthworms that prefer to feed at microsites rich in organic matter. However, the biomass of the surviving earthworms did not differ significantly between treatments, and when Dalby (1996) distributed dung throughout the same soil evenly and in patches, he reported no differences in the survival and biomass of these same earthworm species. Waterlogging of soils is a significant problem in some high-rainfall zones of southeastern Australia (Reed and Cocks 1982). Underground drainage is expensive to install but can increase lucerne production significantly (Chin 1990). Drainage can increase earthworm populations significantly (Baker 1998a). Such increases in earthworm populations may contribute, at least in part, to observed increases in plant productivity under drainage. Without irrigation, pasture growth usually ceases during the hot, dry summer in southern Australia, and no earthworms are active in the root zone (Baker et al. 1992a, 1993c,d). Baker et al. (1993c) suggested that earthworm activity ceases in soils above approximately 150 kPa of water suction potential. With irrigation, several species can remain active during summer (e.g., A. trapezoides, A. caliginosa, A. rosea). Irrigated pastures in the Mt. Lofty Ranges of S.A. are dominated by A. caliginosa in winter, whereas dryland pastures are dominated by A. trapezoides. These differences reflect the greater dependence of A. caliginosa on moist soil (and its more northerly distribution in the European distributions of two earthworm species). Lumbricus rubellus occurs in small numbers in irrigated pastures in S.A. but has never been found in dryland fields. In eastern Australia, where soil moisture levels may be higher, L. rubellus is occasionally very abundant. Lobry de Bruyn and Kingston (1997) also demonstrated changes in earthworm community structure resulting from irrigation in northern Tasmania. Noble and Mills (1974) reported that populations of A. caliginosa increase with irrigation in pastures but decline over time under heavy irrigation. They attributed this population decline to increased earthworm surface activity and greater predation by birds. In Tasmania, Kingston (1989) and Lobry de Bruyn (1993) recorded decreases in populations of A. caliginosa after irrigation but increases in numbers of L. rubellus. These authors explained their results in terms of trampling-induced mortality for both species and increased parasitism by Diptera, which are overcome for L. rubellus by greatly enhanced summer survival and reproduction. Although earthworms avoid freshly limed soil (Doube et al. 1995), several authors (Edwards and Lofty 1977) have shown that liming an acid soil can increase earthworm abundance in the longer term. Springett and Syers (1984) argued that changes in pH per se influence earthworms rather than the availability of calcium. Edwards and Lofty (1977) concluded that population responses to lime are not likely to occur if the initial pH of the soil is above 4.5 to 5.0, a level above which most species are insensitive. Mixed results have been recorded in response to liming pastures in southeastern Australia. Baker (1992) found that liming a pasture on a clay loam soil in western Victoria had no overall impact on total earthworm populations 9 years later (rates of 0 to 10 t ha−1, pH range 4.5 to 5.6 at the time of earthworm sampling), but at the species level, there were increases in abundance of Octolasion cyaneum and M. dubius, decreases in Heteroporodrilus sp. and Spenceriella sp. with increased pH, and no significant changes in A. trapezoides and A. rosea. By contrast, Buckerfield (1994) reported that liming a pasture in S.A. and increasing the soil pH over a range similar to that of Baker (1992) increased populations of A. trapezoides.

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In field cages in S.A. and Victoria, the addition of 4 t lime ha −1 had no influence on the establishment of A. longa after 5 months in a range of soil types (initial pH 4.3 to 5.2) but reduced the survival of Spenceriella sp. at some sites (Baker et al. 1999c; Baker unpublished data). Garnsey (1994a) reported that the addition of lime (5 t ha−1) increased populations of A. trapezoides, L. rubellus, and A. longa in one Tasmanian pasture after 1 or 2 years (initial pH 5.9) but had no influence on earthworms in another pasture (initial pH 6.0). Garnsey attributed the earthworm responses he collected to an indirect effect that was mediated through increased clover production and hence improved residue food quality for the earthworms. In other studies, at several lime trial field sites in pastures in southern N.S.W., Baker and Chan (unpublished data) failed to detect a response to liming by the population and biomass of earthworms (Megascolecidae and Lumbricidae). The numerical responses of earthworms to liming seem likely to vary according to the soil type, earthworm species present, and the range of pH involved. Barley (1959a) showed that population and biomass of earthworms (probably mostly A. rosea and A. trapezoides) increased after the addition of superphosphate to a pasture in S.A. Barley argued this was because of an increase in plant productivity and hence available residue food (as decomposing plant material). Similarly, Fraser et al. (1994) reported that earthworm populations (mostly A. caliginosa and L. rubellus) increased with superphosphate use and plant productivity in a New Zealand pasture. However, such relationships are not always evident. Baker et al. (1993a,b, 1998a) were unable to demonstrate any changes in earthworm populations after superphosphate applications to pastures in Victoria and S.A. Food supply was possibly not limiting for earthworms in these last situations. Lee (1985) indicated that some fertilizers can acidify soils and hence reduce earthworm abundance. The additions of nitrogenous fertilizers and moderate amounts of manures and slurries usually increase earthworm populations (Gerard and Hay 1979; Edwards and Lofty 1982b; Curry 1994). However, excessive amounts of slurries may reduce earthworm populations. Disposal of human sewage sludge by environmentally acceptable means poses a major challenge worldwide. However, safe and profitable disposal of such sludge as biosolids has been achieved in Europe and North America through its addition to pastures (Smith 1996). The disposal of biosolids has also been considered in N.S.W. (Joshua et al. 1998). An experiment that began near Goulburn in 1992 to assess the benefits and risks associated with the application of dewatered biosolids (DWB) to pastures grazed by sheep was surveyed 7 years later to measure impacts of the sludge on the abundance and diversity of earthworms (Baker et al. 2002b). Application of DWB increased local earthworm populations (Figure 14.5). Earthworm species composition varied with the amount of DWB applied. Introductions of earthworms (A. longa and A. caliginosa), which were not present naturally at the site, were successful (in the short term) and unaffected by the DWB applications. The water repellency of sandy soils is a serious agricultural problem across southern Australia, leading to significant land degradation and losses in productivity (Bond 1969). One potential way of offsetting the effects of these nonwetting sands is to add a dispersible clay to assist in water infiltration and holding capacity (Ma’shum et al. 1989). A field trial in the southeast of S.A. (Baker et al. 1998a) in which different amounts of clay were added to a nonwetting sandy soil beneath a pasture demonstrated that populations and biomass of A. trapezoides increased after the addition of clay (Figure 14.6). Trampling by agricultural animal stock can squash earthworms that live near the soil surface, compact the soil, and return organic matter and nutrients to the soil in a different form (dung and urine) and spatial distribution than occurs with senescent plants. Thus, animal stock therefore can influence earthworm populations. However, surprisingly few data have been published on interactions between animal stocking rates and earthworm abundance. Lobry de Bruyn (1993) excluded dairy cattle from pastures in Tasmania and demonstrated that trampling reduced populations of both A. caliginosa (19%) and L. rubellus (25%). As well as the differences in earthworm populations because of the trampling, pasture growth was reduced in the untrampled © 2004 by CRC Press LLC

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Numbers of Earthworms (m–2)

70 60 50

Other spp. S. macleayi M. dubius A. trapezoides

b

b

b 40 30 20 10

a

0 0

30 60 Dry Biosolids (t/ha)

120

FIGURE 14.5 Average abundance of earthworms in pastures at Goulburn, New South Wales, with varying amounts of dewatered biosolids incorporated into the soil 7 years earlier. Different letters above the histograms indicate significant differences between treatments (for total earthworms). (Data redrawn from Baker 2002b.) 80

30 Earthworm Population Earthworm Biomass

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60 20

50 40

15

30 10 20 5

10 0

Biomass of Earthworms (m–2)

Numbers of Earthworms (m–2)

70

0 0

10

75

150

Clay (t/ha–1)

FIGURE 14.6 Average abundance and biomass of Aporrectodea trapezoides in a nonwetting sand below a pasture at Mingbool, South Australia, in treatments with varying amounts of added clay (0 to 150 t ha−1). Different letters above the histograms indicate significant differences between treatments. (Data redrawn from Baker et al. 1998a.)

plots, and the plant species composition changed (i.e., more weeds). The mechanism driving the change in earthworm populations was not clear. Nevertheless, both Kingston (1989) and Lobry de Bruyn (1993) suggested that mortality of A. caliginosa and L. rubellus in irrigated dairy pastures in Tasmania is caused, at least in part, by direct animal trampling effects, exacerbated by greater surface activity in moist soils, and by compaction of the soil, which renders it unsuitable for earthworm survival. With smaller domestic animals, Hutchinson and King (1980) observed that earthworm populations were highest at a stocking rate of 29 sheep ha−1 in pastures in northern N.S.W. This stocking © 2004 by CRC Press LLC

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rate corresponded with maximum grass primary productivity. On the other hand, Baker et al. (1993a,b) and Baker and Chan (unpublished data) could show no consistent patterns between earthworm populations and the stocking rates of sheep (range 5 to 23 ha−1) in several pastures in western Victoria, N.S.W., and S.A. Pizl (1992), Sochtig and Larink (1992), and Hansen and Engelstad (1999) demonstrated significant declines in earthworm populations after compaction from machinery traffic in orchards, cereal fields, and dairy pastures in the Czech Republic, Germany, and Norway, respectively. The only comparable Australian study is that of T. Ellis in S.A. (personal communication 1992), who demonstrated a reduction in earthworm populations beneath wheel tracks in a controlled-traffic cereal production trial. Lee (1985), Edwards and Bohlen (1992, 1996), and Curry (1994) provided detailed discussions of the effects of various pesticides on earthworm populations. It is generally accepted that most herbicides are not directly toxic to earthworms, but they may influence populations indirectly by changing plant productivity, food supply, and soil microclimate. Interestingly, Mele and Carter (1999a) found that heavy (c.f. recommended) application rates of post-emergent herbicides increased earthworm populations, but this may have been caused by increased weed residues. Some fungicides, such as benomyl, can be very toxic to earthworms and influence them indirectly by altering their food supply. Buckerfield (1993a) showed that the use of fungicides can alter the species composition of earthworm populations in S.A. pastures. Fumigants such as methyl bromide and many insecticides (e.g., organochlorines and carbamates) also kill earthworms. However, very few studies have been made along these lines in Australia. Choo and Baker (1998) assessed the influence of several commonly used pesticides, including endosulfan (insecticide) and fenamiphos (nematicide), on A. trapezoides in both the field and the laboratory and showed that its growth and reproduction were affected at recommended application rates (Figure 14.7). A worrying trend in southern S.A. has been the increased use of molluscicides to control introduced helicid and hygromiid snails, which are pests of grain crops and pastures (Baker 1989, 2002). These snails are particularly numerous where tillage is reduced and organic matter is retained, just the situation where earthworm populations are likely to increase. Some authors have argued 20

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FIGURE 14.7 Average numbers of cocoons/pot−1 and percentage of Aporrectodea trapezoides that were clitellate after 5 weeks of culture in the laboratory in a brown loam soil to which fenamiphos (nematicide), endosulfan (insecticide), methiocarb (molluscicide), and ridomil (fungicide) were applied at recommended rates. Numbers at the tops of the histograms indicate average weight change (g) earthworm−1 during a similar experiment in the same soil in the field. (Data redrawn from Choo and Baker 1998.)

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that metaldehyde baits for snail control are less toxic to earthworms than methiocarb baits (Bieri et al. 1989) and thus should be used preferentially. Other authors (Wellmann and Heimbach 1996; Choo and Baker 1998) have suggested that the use of methiocarb at recommended rates poses no threats to earthworms. Antiparasitic drugs, such as avermectins, are used widely on grazing ruminants in Australian pastures, and residues of these drugs are excreted in the feces. One of these, ivermectin, has been shown to have serious detrimental effects on dung-inhabiting arthropods, especially larvae of dung beetles and ßies (Strong and Wall 1994; Wardaugh et al. 1996). Gunn and Sadd (1994) found detrimental effects of ivermectin on the earthworm Eisenia fetida in the laboratory, but the few Þeld studies that have been made suggest that ivermectin is harmless to earthworms there (Sommer et al. 1992). However, Edwards et al. (2001) concluded that avermectins could have signiÞcant effects on earthworms and hence slowed the breakdown of dung in pastures. Another group of chemicals, milbemycins, appear to be much less harmful to dung-inhabiting ßies and beetles than the avermectins (Strong and Wall 1994; Wardhaugh et al. 1996). One of these, moxidectin, has been tested for its effect on A. longa in Australia, in both Þeld and laboratory studies, via sheep and cattle dung (Svendsen and Baker 2002). No lethal or sublethal effects were found.

INTRODUCTIONS OF EARTHWORM TAXA TO NEW AREAS The introduction of earthworms to soils that lack them has usually resulted in signiÞcant increases in plant productivity (see Chapter 2 this volume). However, occasional examples exist of negative impacts from earthworm introductions. For example, James (1991) reported that the introduction of A. caliginosa and O. cyaneum to tallgrass prairie in the United States had a negative inßuence on soil properties through a reduction in populations of more useful native earthworm species (Diplocardia spp.). McLean and Parkinson (1998, 2000) also demonstrated the signiÞcant impact that an invasion of Dendrobaena octaedra had on the microfauna and microßora in pine forests in Canada (see Chapter 13 this volume). A simple method of inoculating earthworms into unpopulated soils was developed by Stockdill (1982). This involved cutting shallow sods of soil from heavily earthworm-populated pastures and placing them on a grid in unpopulated sites. This method is especially suitable for inoculating epigeic and endogeic earthworm species and has been adopted successfully by farmers in Tasmania. Butt (1992, 1999) and Butt et al. (1992, 1995) developed an alternative method for mass producing and distributing deep-burrowing, anecic species such as L. terrestis by placing cultures in polyurethane bags with holes in them on a grid across a site. The rates of dispersal of earthworms after their introduction to new habitats have been measured by several authors in New Zealand and Europe (e.g., see Stockdill 1982; Marinissen and van den Bosch 1992; Stein et al. 1992) and spread varied between 2 and 15 m year−1 according to the relative fecundities and burrowing behaviors of the different species. Aporrectodea longa was released in replicated (n = 5) 2-m long trenches (approximately 0.3 m wide and 0.3 m deep, with soil replaced on top of the earthworms, n = 200 worms per trench) in a dairy pasture in Canberra, Australian Capital Territory. Three years later, populations of A. longa in the vicinity of the trenches were measured (Figure 14.8). The migration of A. longa was comparable with the recorded movements for other species of Lumbricidae. One A. longa was found (by chance) approximately 20 m from the nearest release site. The earthworm fauna in pasture soils in southern Australia is dominated by accidentally introduced species that are now distributed in patches. Although edaphic and climatic factors can explain much of this patchiness in distribution, it is reasonable to assume that many areas lack particular earthworm species do so because of a lack of opportunity for the earthworms to colonize these areas. Increasing the distributions of the most beneÞcial earthworm species through deliberate introductions to sites where they are thought likely to establish may prove very proÞtable. However,

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Frequency

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FIGURE 14.8 Frequency of Aporrectodea longa in 0.1-m2 soil samples dug to 0.1-m depth and hand sorted for earthworms at varying distances away from release trenches in a pasture in Canberra, Australian Capital Territory. Each data point represents A. longa occurrences (presence data only) in 10 soil samples (duplicates at each distance along transects running perpendicular to five trenches). Data were collected 3 years after earthworm release.

few such field introductions of earthworms have been reported to date, especially at field scale. Aporrectodea caliginosa and A. longa have been introduced into pastures in Tasmania with resultant increases in productivity (Temple-Smith et al. 1993; Garnsey 1994b). Introductions of A. caliginosa to irrigated pastures in N.S.W. led to a breakdown of a thick litter mat and a decline in soil bulk density (Noble et al. 1970), and introductions of Aporrectodea spp. and Eukerria saltensis (Ocnerodrilidae) into irrigated wheat in N.S.W. increased the air permeability of the soil (Blackwell and Blackwell 1989). The distribution of the deep-burrowing species A. longa is currently mainly restricted within Australia to Tasmania (Baker et al. 1997a). Baker (1998a) used climatic matching software to predict where A. longa might colonize within mainland Australia if given the chance; the study was based on regions of the world where it already exists. His predictions corresponded with large areas of southeastern and southwestern Australia (essentially areas with more than 600 mm annual rainfall). Baker and Whitby (2004) have since cautioned that this potential distribution is probably an overestimate. For example, their research on the environmental factors that control the development time for A. longa cocoons showed that the length of season during which the soil remains sufficiently moist in much of southeastern Australia is likely to be inadequate to support viable A. longa populations. In addition, many soils are currently too acidic for A. longa to survive because A. longa prefers soils with a pH above 4.5 (Baker and Whitby 2004), and many pastures in southeastern Australia have a soil pH well below this value. Nevertheless, Baker et al. (1999c), introduced A. longa successfully to several pasture sites in southeastern Australia and reported its relatively high rate of establishment compared with that of other, more widely spread species (e.g., A. caliginosa and A. trapezoides). They concluded that a lack of opportunity to colonize was a major reason for its absence. Recent research has concentrated on developing mass rearing methods for A. longa (e.g., using optimum soil types, temperature, soil moisture, pH, food type, population density, etc.) (Baker and Whitby 2004; G. Baker, unpublished data). For example, the optimum temperature for cocoon and earthworm development is approximately 15°C. The opportunity now exists for private industry (e.g., earthworm farmers, hitherto concentrating on vermicomposting) to follow up this research and breed A. longa to provide a greater availability to landholders.

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THE POTENTIAL FOR INTRODUCING NEW EARTHWORM SPECIES INTO AUSTRALIA The possibility also exists to introduce other earthworm species into southern Australia from climatically matched regions overseas. No deliberate attempts to do this have yet been attempted, although such a strategy was suggested by Barley (1959c) and Lee (1985). Broad-scale surveys and intensive, seasonal monitoring of field populations suggest that the current earthworm fauna in agricultural fields is represented poorly by deep-burrowing species. Virtually all earthworm activity is confined to the top 10 cm or less of soil during winter and spring. The further spread of A. longa within Australia will possibly help redress this limitation of the earthworm fauna, but it is unlikely that A. longa will successfully colonize the strictly mediterranean climatic regions. For instance, the natural distribution of A. longa does not extend into mediterranean regions of countries such as France. Instead, other anecic species such as Scherotheca spp. are commonly found there. These last species might also be considered for importation to Australia (Baker 1998a). The native European distributions of several of the lumbricid earthworm species, now found in Australia, are broad, ranging from the Mediterranean to Scandinavia (e.g., A. caliginosa and A. chlorotica). The majority of early European migrants to Australia, who presumably brought these lumbricid earthworms with them accidentally (e.g., in potted plants), were mainly from countries with cool, temperate climates, such as the U.K., Ireland, and Germany. It is sensible to question the likely suitability of strains of earthworms from such countries when faced with the warmer and drier habitats in much of southern Australia and whether Mediterranean strains of the same or other species might be more appropriate (Baker 1998a,b). Dyer et al. (1998) used techniques based on polymerase chain reaction to at least show that different geographic races of A. trapezoides can be recognized within Australia. Perhaps such techniques could also be used to trace the origins of Australian earthworm populations to Europe, check their ecological suitability, and select better ones. We actively select climatically sensible strains or varieties of agricultural plants and biocontrol agents, so why not also earthworms? There are risks attached to introducing taxa to new areas, whether these taxa come from overseas or from elsewhere within Australia. Issues that must be faced include the possibility that the new invaders might compete with the local earthworm fauna, disrupt ecosystem processes in nontarget areas (e.g., native forests and pastures, compared with improved pastures and croplands), and carry with them undesired diseases. These diseases can be controlled through rigorous quarantine procedures. Some preliminary studies have been completed that suggest A. longa is unlikely to invade native woodlands in southern Australia and to compete with native earthworms there (Dalby et al. 1998b). There is little doubt that A. longa will have some impact on the abundance of resident earthworms when introduced into pastures (Dalby et al. 1998a; Baker et al. 1999c; Baker et al. 2002a), but studies suggest that this potential impact is probably small, and that the overall abundance and, most importantly, the functional diversity are increased.

EARTHWORMS IN PASTURES IN NORTHERN AUSTRALIA The comments in preceding sections all refer to knowledge of the distribution, biology, and agricultural value of earthworms in southern temperate or mediterranean climatic zones in Australia. Very little is known of such topics for pastures in northern Australia. Baker et al. (1997a) documented the presence of some earthworm species in tropical grasslands, noting they differed from those in the south (e.g., the exotic P. corethrurus) and indicated that some of the more southern species (e.g., Aporrectodea spp.) can be found as far north as southern Queensland. Blakemore (1997) reported that the introductions of exotic and native earthworm species increased pasture production on brigalow soils by 64% within a year in southeastern Queensland. In other studies, Friend and Chan (1995) and Chan et al. (1997) showed that native earthworms (e.g., Heteroporodrilus mediterreus) can improve the structure (hydraulic properties) of native pasture soils in northwestern N.S.W.

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EARTHWORMS AS INDICATORS OF THE SUSTAINABILITY OF AGRICULTURE: FROM THE FARMER’S PERSPECTIVE Large populations of earthworms are popularly believed to reflect soil health. They often make up a large proportion of the biomass of the soil fauna, and they respond positively to many agricultural management practices. Earthworm populations and biomass are also correlated with a range of edaphic variables. It is therefore not surprising that earthworms have been suggested as potential indicators of the sustainability of agricultural practices that farmers might use (Oades and Walters 1994; Buckerfield et al. 1997). A useful indicator of soil sustainability must be attractive to farmers so that they will understand and adopt it, should be easy to measure reliably, and should be responsive to environmental change in a timely fashion. Farmers know that earthworms are generally beneficial. However, few are aware of the species they have on their land, and, like scientists, they are unclear just how many earthworms they need in their soil (i.e., what are the abundance thresholds they need to seek). It is important that farmers recognize the species of earthworms they have on their farms, realize the varying abilities of different species to influence soil properties and plant production, and note the strengths and weaknesses of their earthworm resource. Simple keys have been devised that farmers might use for the common earthworm species in Australia (e.g., Baker and Barrett 1994; Mele and Hollier 1995), but these are now out of print and should be revised and reprinted. Sampling for earthworms, whichever method is used, is notoriously labor intensive and fraught with inaccuracies (Baker and Lee 1992); for this reason (and others), some authors (Doube and Schmidt 1997; Lobry de Bruyn 1997; Baker 1999) have questioned the practicality of using earthworms as biological indicators. Soil moisture can vary within short periods of time and affect the numbers of earthworms collected (Baker et al. 1993c). Earthworms are usually distributed in patches within fields and vary in populations from year to year when no overt changes in management practices occur (Baker et al. 1992a, 1993b; Baker 1999). Spatial distribution patterns vary among species (Baker 1999). Farmers are busy people, but they must take care to collect sufficient earthworm samples to make their data meaningful and to enable detection of differences in populations through time and space. Some earthworms have a large reproductive potential (e.g., epigeic species), but the species of earthworms that predominate in Australian agricultural fields probably do not (Lee 1985). Although drastic physical or chemical disturbance might quickly reduce earthworm populations, recovery may well take several years, and farmers need to be appreciative of this phenomenon. The fact that the earthworm fauna of Australian agricultural habitats is dominated by introduced species raises an immediate question: How far have these species spread to occupy sites that are suitable for them? The answer is unknown, but it seems likely that there are many potential sites yet to be occupied by exotic earthworms. More than 40% of pastures in one region of western Victoria lack A. trapezoides, but there seems to be no good reason other than lack of opportunity to colonize that can be given for its absence from these pastures. Although the presence of large populations of a diverse community of earthworms is usually a healthy sign, there is a strong risk that low populations, or indeed absence, of earthworms might be misinterpreted as a “problem” at a particular site when the real problem is not with the soil per se, but the chance of earthworm dispersal to it. Some seemingly “healthy” soils in Australia lack earthworms (e.g., some kraznozems). That is not to say that these soils would not be more productive with the arrival of appropriate species of earthworms.

CONCLUSIONS Much progress has been made in earthworm ecology in Australian pastures in recent years. However, at least four major gaps are obvious in current knowledge: 1. Hardly anything is known of the earthworm fauna in subtropical and tropical systems and its ability to influence soil properties and pasture production there. © 2004 by CRC Press LLC

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2. In the southeastern states, native earthworms (e.g., Megascolecidae), although rarer than the introduced Lumbricidae, can still constitute 40% on average of the earthworm fauna (and in many instances, the majority). We know virtually nothing of the agricultural importance of this “resource.” 3. Although a substantial amount is known about the biology of the exotic earthworms in southern pasture systems, ecological linkages between these and other soil biota (with the exception of root diseases), aboveground pests, and beneficial invertebrates (e.g., via nutrient flows) remain unstudied (cf. Scheu et al. 1999 and Wurst and Jones 2003). If current aspirations toward establishing and harnessing improved functional biodiversity in these soils is to succeed, such linkages need to be understood far better. 4. Optimal use of water and key nutrients such as N and P, particularly in the development of systems that optimize uptake by agricultural plants while minimizing off-farm economic losses and environmental degradation through leaching and surface runoff, are high-priority topics across a wide range of agricultural industries, including those that are pasture based. Previous studies throughout the world have suggested that earthworms, if managed properly, can contribute to substantial improvements in efficient usage of nutrients on the farm (Lee 1985; Edwards and Bohlen 1996; Lavelle et al. 1999). Earthworms need to be utilized as “soil engineers” — as taxa that can substantially create the soil architecture that determines water and nutrient movements through profiles and the abilities of plant roots to access these. Like other macrofauna, such as dung beetles, it needs to be realized that earthworms can be, and have been elsewhere, manipulated in agricultural landscapes. Earthworms are the most obvious element of the macrofauna in pasture soils in southern Australia. The earthworm fauna in pastures in this region is similar to that of several other countries with temperate or mediterranean climates. The fauna is dominated by introduced Lumbricidae from Europe, particularly A. caliginosa, A. trapezoides, and A. rosea. Population numbers and species diversity are usually low. The geographic distributions of the most common species are patchy. Earthworm population abundance is correlated with a number of climatic and edaphic variables, most notably rainfall. The common species are most active in the top 10 cm of the soil profile during winter-spring. Deep-burrowing (anecic) species are rare, particularly in mainland Australia. A number of studies have shown that earthworms can improve soil properties, help offset soil degradation (e.g., burial of lime to reduce soil acidity), and increase pasture productivity in southern Australia. Different species differ in these abilities to improve soil fertility. The paucity of anecic species in mainland Australia may be partially compensated by introduction of the highly beneficial species A. longa from Tasmania. Recent research has developed means to mass rear this species and predict where it might best be established. Agricultural management practices can markedly influence earthworm populations and biomass. Examples given in this chapter include tillage, drainage, irrigation, lime and fertilizer application, stocking rates, and pesticide use. The use of earthworms as biological indicators of the sustainability of agricultural practices has been suggested by some authors. However, the patchy distributions of earthworms in space and time present very significant hurdles for the successful adoption of such an approach.

REFERENCES Abbott, I. and Parker, C.A. 1981. Interactions between earthworms and their soil environment, Soil Biol. Biochem., 13, 191–197. Andrewartha, H.G. and Browning, T.O. 1961. An analysis of the idea of “resource” in animal ecology, J. Theor. Biol., 1, 83–97.

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Baker, G.H. 1989. Damage, population dynamics, movement and control of pest helicid snails in southern Australia, in Slugs and Snails in World Agriculture, BCPC Monograph no. 41, I. Henderson, Ed., BCPC, Thornton Heath, U.K., pp. 175–185. Baker, G.H. 1992. Optimising earthworm activity in soils, in Proc. 33rd Annu. Conf. Grassl. Soc. Vic., pp. 59–66. Baker, G.H. 1997. Influence of the introduced anecic earthworm, Aporrectodea longa (Lumbricidae) on pasture production in south-western Victoria, in Soil Invertebrates in 1997, P.G. Allsopp, D.J. Rogers, and L.N. Robertson, Eds., BSES, Brisbane, Australia, pp. 21–25. Baker, G.H. 1998a. The ecology, management, and benefits of earthworms in agricultural soils, with particular reference to southern Australia, in Earthworm Ecology, C.A. Edwards, Ed., St. Lucie Press, Boca Raton, FL, pp. 229–257. Baker, G.H. 1998b. Recognising and responding to the influences of agriculture and other land-use practices on soil fauna in Australia, Appl. Soil Ecol. 9, 303–310. Baker, G.H. 1999. Spatial and temporal patterns in the abundance and biomass of earthworm populations in pastures in southern Australia, Pedobiologia, 43, 487–496. Baker, G.H. 2002. Helicidae and Hygromiidae as pests in cereal crops and pastures in southern Australia, in Molluscs as Crop Pests, G.M. Barker, Ed., CAB International, Wallingford, U.K., pp. 193–215. Baker, G.H. and Amato, M. 2001. Increasing earthworm activity benefits crop yields, Farming Ahead, 113, 52–53. Baker, G.H., Amato, M., and Ladd, J. 2004a. Influences of Aporrectodea trapezoides and A. rosea (Lumbricidae) on the uptake of nitrogen and yield of oats (Avena fatua) and lupins (Lupinus angustifolius), Pedobiologia, 47, in press. Baker, G.H. and Barrett, V.J. 1994. Earthworm Identifier, CSIRO, Melbourne, Australia. Baker, G.H., Barrett, V.J., Carter, P.J., Cayley, J.W.D., and Saul, G.R. 1993a. The influence of fertiliser on the abundance and diversity of earthworms in pastures in western Victoria, Proc. Seventh Aust. Conf., Grassl. Invertebrate Ecol., pp. 312–315. Baker, G.H., Barrett, V.J., Carter, P.J., Cayley, J.W.D., and Saul, G.R. 1993b. The influence of phosphate application and stocking rate on the abundance of earthworms in a Victorian pasture, in Proc. Sixth Australasian Conf. Grassl. Invert. Ecol., pp. 85–91. Baker, G.H., Barrett, V.J., Carter, P.J., Williams, P.M.L., and Buckerfield, J.C. 1993c. Seasonal changes in the abundance of earthworms (Annelida : Lumbricidae and Acanthodrilidae) in soils used for cereal and lucerne production in South Australia, Aust. J. Agric. Res., 44, 1291–1301. Baker, G.H., Barrett, V.J., Carter, P.J., and Woods, J.P. 1996. Method for caging earthworms for use in field experiments, Soil Biol. Biochem., 28, 331–339. Baker, G.H., Barrett, V.J., Grey-Gardner, R., and Buckerfield, J.C. 1992a. The life history and abundance of the introduced earthworms Aporrectodea trapezoides and A. caliginosa (Annelida: Lumbricidae) in pasture soils in the Mount Lofty Ranges, South Australia, Aust. J. Ecol., 17, 177–188. Baker, G.H., Barrett, V.J., Grey-Gardner, R., and Buckerfield, J.C. 1993d. Abundance and life history of native and introduced earthworms (Annelida: Megascolecidae and Lumbricidae) in pasture soils in the Mount Lofty Ranges, South Australia, Trans. Roy. Soc. S.A., 117, 47–53. Baker, G.H., Buckerfield, J.C., Grey-Gardner, R., Merry, R., and Doube, B.M. 1992b. The abundance and diversity of earthworms in pasture soils in the Fleurieu Peninsula, South Australia, Soil Biol. Biochem., 24, 1389–1395. Baker, G.H., Carter, P.J., and Barrett, V.J. 1999a. Influence of earthworms, Aporrectodea spp. (Lumbricidae), on lime burial in pasture soils in south-eastern Australia, Aust. J. Soil Res., 37, 831–845. Baker, G.H., Carter, P.J., and Barrett, V.J. 1999b. Influence of earthworms, Aporrectodea spp. (Lumbricidae), on pasture production in south-eastern Australia, Aust. J. Agric. Res., 50, 1247–1257. Baker, G.H., Carter, P.J., and Barrett, V.J. 1999c. Survival and biomass of exotic earthworms, Aporrectodea spp. (Lumbricidae), when introduced to pastures in south-eastern Australia, Aust. J. Agric. Res., 50, 1233–1245. Baker, G.H., Carter, P.J., Barrett, V.J., Hirth, J., Mele, P., and Gourley, C. 2002a. Does the deep-burrowing earthworm, Aporrectodea longa, compete with resident earthworm communities when introduced to pastures in south-eastern Australia? Eur. J. Soil Biol., 38, 39–42. Baker, G.H., Carter, P.J., Barrett, V.J., Kilpin, G.P., Buckerfield, J.C., and Dalby, P.R. 1994. The introduction and management of earthworms to improve soil structure and fertility in south-eastern Australia, in Soil Biota. Management in Sustainable Farming Systems, C.E. Pankhurst, B.M. Doube, V.V.S.R. Gupta, and P.R. Grace, Eds., Kluwer Academic Publishers, Dordrecht, the Netherlands, pp. 42–49.

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Stephens, P.M., Davoren, C.W., Doube, B.M., Ryder, M.H., Benger, A.M., and Neate, S.M. 1993a. Reduced severity of Rhizoctonia solani disease on wheat seedlings associated with the presence of the earthworm Aporrectodea trapezoides (Lumbricidae), Soil Biol. Biochem., 25, 1477–1484. Stephens, P.M., Davoren, C.W., Ryder, M.H., and Doube, B.M. 1993b. Influence of the lumbricid earthworm Aporrectodea trapezoides on the colonisation of wheat roots by Pseudomonas corrugata strain 2140R in soil, Soil Biol. Biochem., 25, 1719–1724. Stephens, P.M., Davoren, C.W., Ryder, M.H., and Doube, B.M. 1994b. Influence of the earthworm Aporrectodea trapezoides (Lumbricidae) on the colonisation of alfalfa (Medicago sativa L.) roots by Rhizobium meliloti L5-30R and the survival of R. meliloti L5-30R in soil, Biol. Fertil. Soils, 18, 63–70. Stephens, P.M., Davoren, C.W., Ryder, M.H., Doube, B.M., and Correll, R.J. 1995. Field evidence for reduced severity of Rhizoctonia bare patch disease of wheat, due to the presence of the earthworms Aporrectodea rosea and Aporrectodea trapezoides, Soil Biol. Biochem., 26, 1495–1500. Stockdill, S.M.J. 1982. Effects of introduced earthworms on the productivity of New Zealand pastures, Pedobiologia, 24, 29–35. Stockdill, S.M.J. and Cossens, G.G. 1966. The role of earthworms in pasture production and moisture conservation, in Proceedings of the New Zealand Grassland Assoc., pp. 168–183. Strong, L. and Wall, R. 1994. Effects of ivermectin and moxidectin on the insects of cattle dung, Bull. Ent. Res., 84, 403–409. Svendsen, T.S. and Baker, G.H. 2002. Survival and growth of Aporrectodea longa (Lumbricidae) fed on sheep and cow dung with and without moxidectin residues, Aust. J. Agric. Res., 53, 447–451. Temple-Smith, M.G. 1991. Earthworm populations and their effects on pasture production, Australian Institute of Agricultural Science Occasional Publ. No. 62, 11-5. Temple-Smith, M.G., Kingston, T.J., Furlonge, T.L., and Garnsey, R.B. 1993. The effect of introduction of the earthworms Aporrectodea caliginosa and Aporrectodea longa on pasture production in Tasmania, in Proc. Seventh Aust. Agron. Conf., p. 373. Temple-Smith, M.G. and Pinkard, T. 1996. The Role of Earthworms in Agriculture and Land Management, Department Primary Industry and Fisheries, Tasmania, Australia. Tyndale-Biscoe, M. 1990. Common Dung Beetles in Pastures in South-eastern Australia, CSIRO, Melbourne, Australia. van Praagh, B.D. 1992. The biology and conservation of the giant Gippsland earthworm Megascolides australis McCoy, 1878, Soil Biol. Biochem., 24, 1363–1368. Wardaugh, K.G., Holter, P., Whitby, W.A., and Shelly, K. 1996. Effects of drug residues in the faeces of cattle treated with injectable formulations of ivermectin and moxidectin on larvae of the bush fly, Musca vetustissima and the house fly, Musca domestica, Aust. Vet. J., 74, 370–374. Waterhouse, D.F. 1974. The biological control of dung, Sci. Am., 230, 101–108. Wellman, P. and Heimbach, F. 1996. The effects of methiocarb slug pellets on the earthworm Lumbricus terrestris in a laboratory test, in Slug and Snail Pests in Agriculture, BCPC Symposium Proceedings No. 66, I. Henderson, Ed., BCPC, Farnham, U.K., pp. 181–188. Whalen, J.K., Parmelee, R.W., McCartney, D.A., and Vanarsdale, J.L. 1999. Movement of N from decomposing earthworm tissue to soil, microbial and plant N pools, Soil Biol. Biochem., 31, 487–492. Wurst, S. and Jones, T.H. 2003. Indirect effects of earthworms (Aporrectodea caliginosa) on an above-ground tritrophic interaction, Pedobiologia, 47, 91–97.

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in 15 Earthworms Agroecosystems: Research Approaches Paul F. Hendrix Institute of Ecology, University of Georgia, Athens, GA, U.S.A.

Clive A. Edwards Soil Ecology Program, The Ohio State University, Columbus, OH, U.S.A.

CONTENTS Introduction ....................................................................................................................................287 Recent Trends.................................................................................................................................288 Future Research Needs ..................................................................................................................290 Conclusions ....................................................................................................................................292 Acknowledgments ..........................................................................................................................293 References ......................................................................................................................................293

INTRODUCTION The impacts of earthworms on soil processes and soil fertility and their potential importance in agriculture are topics that are currently of wide interest. There is vast literature, ranging from scientific treatises to popular articles in magazines and newspapers, on these subjects, and it would be futile to attempt to review all of these comprehensively. As an alternative to an extensive literature review of the role of earthworms in agroecosystems, this chapter presents a broad overview of current research, identifying trends and future directions for research. The first edition of this book was associated with the Fifth International Symposium on Earthworm Ecology (ISEE5), which was held at The Ohio State University in Columbus, OH, U.S.A. in 1994 (ISEE5) (Edwards 1997), and the chapter authors were either chairs or associate chairs of the eight chosen research areas and authorities on their subjects. The earlier symposia on earthworm ecology were at Grange-over-Sands, England, in 1981 (ISEE1) (Satchell 1983); Bologna, Italy, in 1985 (ISEE2) (Bonvicini et al. 1987); Hamburg, Germany, in 1987 (ISEE3); and Avignon, France, in 1990 (ISEE4) (Kretzschmar 1992). These symposia were very well attended by earthworm researchers and provided fertile ground for transmission of information and collaboration. Symposia subsequent to ISEE5 were held in Vigo, Spain, in 1998 (ISEE6) (Diaz Cosin et al. 1999) and in Cardiff, Wales, in 2002 (ISEE7). For information on earlier research publications, refer to Bibliography Volume 2, Workshop on the Role of Earthworms in the Stabilization of Organic Residues (Worden 1981) and A Bibliography of Earthworm Research (Satchell and Martin 1983). A number of comprehensive books on earthworms have been published. These include Earthworms: Their Ecology and Relationship with Soils and

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Land Uses (Lee 1985); the third edition of Biology and Ecology of Earthworms (Edwards and Bohlen 1996); Earthworms in Waste and Environmental Management (Edwards and Neuhauser 1988); Earthworm Ecology in Forest, Rangeland and Crop Ecosystems (Hendrix 1995); the first edition of Earthworm Ecology (Edwards 1998); and Earthworm Management in Tropical Agroecosystems (Lavelle et al. 1999). These extensive literature sources provide excellent pointers to the needs for future research, promising directions for investigations and problems needing attention. Based on these works, we can summarize the needs for earthworm research on agricultural topics.

RECENT TRENDS Implications for the roles of earthworms in agriculture can be found in nearly the entire range of topics covered in the various International Symposia on Earthworm Ecology. This is partly because many researchers are interested in the potential beneficial influences of earthworms on soil fertility and agricultural production, soil conservation, environmental quality, and the like. However, from a broader perspective, agricultural systems often tend to be much more easily studied than more complex natural ecosystems, and effects of earthworms on soil processes, fertility, and plant growth are more obvious in managed agroecosystems than in natural ecosystems. Thus, agricultural ecosystems provide a ready crucible in which to test ideas about earthworms, sometimes beyond their role in agriculture. Research topics relevant to agriculture in the International Symposia on Earthworm Ecology include earthworm systematics, biogeography, ecology, and behavior; effects on soil processes (nutrient and organic matter dynamics, soil structure, and hydrology); uses in waste management and land management; ecotoxicology; and interactions with other soil organisms. For example, a listing of such topics was compiled from the programs of symposia and is summarized in Table 15.1. This compilation also includes the number of presentations (both oral and posters) within each topic category. Although it is not intended that firm conclusions be drawn from such a relatively small sampling of the earthworm literature, a few new interesting trends can be seen. First, there has been a steady increase in the number of agriculturally related contributions (Table 15.1). These were 57% at ISEE4, 77% at ISEE5, 86% at ISEE6, and 86% at ISEE7. Whether these numbers represent a worldwide increase in research interest in earthworms in agroecosystems is not clear, but it seems to be a strong indicator of such a trend. There seems to be little change in research into biogeography and biodiversity in managed systems over the four symposia. There

TABLE 15.1 Number of Contributions within Various Topics Relevant to Agroecosystems Presented at the International Symposia on Earthworm Ecology (ISEE) Number of Contributions ISEE4 ISEE5 (Avignon) (Columbus) Biogeography and biodiversity in managed systems Management effects on abundance/distribution (tillage, cropping systems) Managed introductions and land use Nutrient cycling processes (decomposition, mineralization, microbiology) Soil physical properties (aggregation, porosity, infiltration, leaching) Effects on plant growth (rhizosphere effects, disease suppression) Waste processing and vermicomposting Ecotoxicology (effects of pesticides and heavy metals)

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11 20 7 13 12 3 14 11 91

12 15 5 19 16 9 21 14 111

ISEE6 (Vigo)

ISEE7 (Cardiff)

7 3 11 5 3 6 10 12 57

11 9 8 8 14 14 38 37 132

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appears to be a decreasing interest in the effects of agricultural management practices on earthworms and on earthworm introductions into land uses and nutrient cycling processes. By contrast, there seems to be increasing interest in the effects of earthworms on soil physical properties and on plant growth. The two most significant changes over the four symposia are the increasing interests in earthworm ecotoxicology and in organic waste management with earthworms. The increased number of presentations on earthworm ecotoxicology is particularly surprising because there have been three international conferences on earthworm ecotoxicology: in Sheffield, U.K., 1991 (GreigSmith et al. 1992); in Amsterdam, the Netherlands, 1997 (Shepard et al. 1998); and in Arhus, Denmark, 2001. Each was attended by more than 100 researchers. This suggests that there is a great increase in the use of earthworms as ecotoxicological tools; in laboratory tests; in artificial soils, microcosms, and terrestrial model ecosystems; and in field experiments (see Chapter 16 and Chapter 17, this volume). There also seems to be much interest in the use of earthworm physiology and other associated processes as biomarkers to assess the impact of pollutants on soil ecosystems (see Chapter 16, this volume). The increased interest in the use of earthworms in waste management and vermicomposting seems a real and significant trend, particularly in developing countries such as India. The presentations at the four symposia cover a wide range, including the biology and ecology of vermicomposting species of earthworms, methods of vermicomposting, nutrient transformations, and field and greenhouse studies of the effects of vermicomposts on plant growth (see Chapter 18 to Chapter 20, this volume). The greatest interest in this topic seems to be in the United States, U.K., Spain, Mexico, and Australia. The strong microbial connection with earthworm feeding and activity is emerging as a central theme in soil nutrient process studies. For example, it is clear that earthworm activity greatly accelerates crop residue decomposition and nutrient mineralization, and there is growing evidence that microorganisms serve as a primary source of nutrition for earthworms (Edwards and Fletcher 1988). In an ecological sense, this relationship might be considered a “keystone” association. With respect to organic residue quality, earthworms may selectively ingest high-quality residues (Bohlen et al. 1997), but their relative effects on decomposition may be greater on low-quality residues, which earthworms fragment, inoculate with microbes, and incorporate into soil. High-quality substrates tend to decompose rapidly even in the absence of earthworms. This idea has also been suggested for other soil invertebrates (e.g., microarthropods, millipedes, woodlice, etc.) involved in forest litter decomposition (Seastedt et al. 1983). Lavelle et al. (Chapter 8, this volume) make the interesting point that the cost of earthworm activity, in terms of the status of organic carbon in agroecosystems, must be considered because earthworms tend to stimulate microbial activity and accelerate litter turnover and possibly cause losses of soil carbon if adequate organic inputs are not maintained. The agricultural management implication of this possibility is that some “base” amounts of organic substrates must be added to agricultural soils periodically to feed microorganisms and earthworms, with additional amounts needed to transform organic matter into soils if there have been soil losses. As mentioned, substrate quality has a strong influence on interactions between earthworms and microorganisms (see Chapter 2 and Chapter 12, this volume). These observations also have implications for mixed residue management in agroecosystems. Although the mineralization of rapidly decomposing residues may contribute to short-term nutrient availability, more slowly decomposing materials (i.e., those with higher C:N ratios) may immobilize nutrients and add to nutrient pools with longer turnover times and more sustained nutrient availability (Tian et al. 1995). Further, very recalcitrant materials, especially those left on the soil surface, contribute to a “mulch effect,” which protects soil surfaces from erosive forces (e.g., rainfall impact, wind) and protects and ameliorates the soil microclimate (Lal 1991; Hendrix et al. 1992). With respect to effects of earthworms on plant growth, a number of interesting observations have been reported. Several papers have indicated that the effects of earthworms on plants differ with different earthworm species, different plant species, and different soil types (Chapter 8, this © 2004 by CRC Press LLC

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volume). For example, Gilot-Vilienave (1995), Baker et al. (1995a,b), and Blakemore (1995) all reported greater plant productivity or nutrient uptake in soils when earthworms were present compared with control soils without earthworms. However, Devliegher and Verstraete (1995) reported decreased plant growth in the presence of Lumbricus terrestris, and Baker et al. (1995) recorded no effect of Aporrectodea rosea on clover growth. Lavelle et al. (Chapter 8, this volume) reported that legumes generally did not seem to respond to earthworm inoculations or additions, whereas maize and rice and other crops did. Several possible mechanisms were proposed to account for increased plant growth in the presence of earthworms, including reduced incidence of plant diseases (i.e., earthworms may consume disease fungi, promote antagonistic microorganisms, or make soil inhibitory to them); increased availability of nutrients (e.g., through stimulation of mineralization); enhanced soil structure (e.g., reduced bulk density and greater porosity), thereby creating a more favorable environment for root growth; or combinations of these factors. Despite all the various beneficial effects (well documented and hypothesized) of earthworms on nutrient dynamics, soil structure, and plant growth, Parmelee et al. (1998) reviewed some aspects of earthworm activities that may be considered undesirable in agricultural systems. These include the removal or burial of surface organic residues that would otherwise protect soil surfaces from erosion; increasing soil erosion and surface sealing via freshly produced, unstable casts that are susceptible to impact of raindrops; perforating the walls of irrigation ditches, making them less able to carry water; increasing nitrogen losses through leaching and denitrification; and increasing carbon losses through enhanced microbial respiration. Bohlen et al. (Chapter 9, this volume) suggest that it is the net balance of the positive and negative effects of earthworms that is important in determining whether they may have any detrimental impacts in agroecosystems, although, overall, their contributions are certainly beneficial. As pointed out by Curry (1994) (Chapter 6, this volume), any management practices applied to soils are likely to have some effects on earthworm populations. The well-used diagram of Edwards and Lofty (1969) and Wallwork (1976) is modified in Figure 15.1 to illustrate this point. Thus, agricultural management practices can have positive or negative effects on the abundance and diversity of earthworms. Apart from the direct toxicity of some agricultural chemicals and fertilizers or substances contained in soil amendments (e.g., heavy metals in sewage sludge), these effects are primarily the result of changes in soil temperature, soil moisture, and organic matter quantity or quality, which are the principal driving variables of soil biological activity. Finally, an interesting trend comes from Australia, which has historically had a strong interest in earthworms: scientists studying earthworm ecology are developing a tradition of interacting closely with farmers and with the public (see Chapter 14, this volume). The “Earthworms Downunder” program (Baker et al. 1995a), organized involving young people in the collection, identification, and census of earthworm species nationwide, not only accumulated useful information for agricultural purposes, but also stimulated public interest in earthworm research and may serve to recruit new scientists into agricultural research.

FUTURE RESEARCH NEEDS Priorities for earthworm research relevant to agriculture include research into the influences of earthworms on soil processes and plant growth and the impacts of agricultural practices on earthworm populations (Hendrix 1995). With respect to the effects of earthworms on biogeochemical cycles, Blair et al. (1995) identified several key topics. They considered that a better understanding of the effects of earthworms on soil nutrient and organic matter dynamics should come from studies over a range of spatial scales, ranging from short-term physiological processes within the earthworm gut to long-term patterns of nutrient storage and loss in ecosystems (see Chapter 9, this volume). Equally important is the need for a clear linkage between results derived from short-term laboratory experiments and long-term field studies. In particular, means of translating microsite phenomena

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INCREASE IN EARTHWORM ABUNDANCE AND DIVERSITY

NO TILLAGE

EARTHWORM INOCULATION

ROTATIONS

LIMING

ORGANIC AMENDMENTS

LAND USE PRACTICE

TILLAGE

SINGLE CROP

TOXICANTS

ACIDIFICATION

RESIDUE REMOVAL

DECREASE IN EARTHWORM ABUNDANCE AND DIVERSITY

FIGURE 15.1 The effects of soil management on the abundance and diversity of earthworms in agricultural soils. (Adapted from Edwards and Lofty 1969; Wallwork 1976.)

(e.g., trace gas flux, nutrient mineralization) to ecosystem and landscape scales under field conditions need urgent attention (see Chapter 8, this volume). A useful way to examine effects of earthworms on nutrient cycling processes is to use isotopic tracers such as 13C, 14C, 32P, and 15N. Various studies using both natural and tracer-enriched materials have helped to quantify nutrient fluxes and turnover because of earthworm activities (Cortez et al. 1989; Binet and Trehen 1992; Martin et al. 1992; Spain et al. 1992; Zhang and Hendrix 1995). Further research utilizing these techniques can be expected to increase the understanding of the influences of earthworms on biogeochemical mechanisms and pathways in both agricultural and natural ecosystems. Interesting deviations from the widely held view of earthworm functional groupings (i.e., Bouché 1977), which have been expressed by various authors, suggest a possible need for revision of the current scheme. In particular, species-specific activities of earthworms may preclude their inclusion into such broad ecological categories. A functional reexamination of the commonly accepted ecological groupings may be appropriate for at least certain tropical and subtropical (i.e., non-European lumbricid) taxa. Furthermore, because much of our understanding of earthworm biology and ecology is based on European lumbricid species, hence the relative importance of indigenous vs. exotic species to ecosystem processes in given areas should be examined further (see Chapter 5, this volume). There may be cases for which indigenous species are better adapted and are more effective in enhancing nutrient cycling (James 1991) and possibly are of greater agricultural utility than the familiar lumbricids or other exotic species. Interactions of earthworms with both plant roots and other soil biota have many important implications in agroecosystems. Enhanced plant productivity because of earthworm activities has been discussed, but the mechanisms involved in these increases are not always clear. Research into earthworm-microbial interactions, which has been greatly neglected, deserves particular attention. However, identification of soil food webs, their roles in nutrient cycling, and the influence of

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earthworms on them also needs much further study. The importance of earthworms as key organisms controlling soil food webs also needs further research (Edwards 1999). The impacts of earthworm activity on soil structure in agroecosystems are receiving increasing attention. Tomlin et al. (1995) identified several potentially fruitful lines for further research. Although increased soil porosity and water infiltration are well-known consequences of earthworm activity, relatively little is known of the mechanisms by which solutes (e.g., agricultural chemicals) are transported through earthworm burrows. Because earthworm burrow linings may act as both sources and sinks for solutes, better understanding is needed of their physical, chemical, and microbial characteristics (e.g., organic chemical nature and exchange properties) (see Chapter 10 and Chapter 11, this volume). Soil aggregation may be one of the most important soil properties affected by earthworms in both agricultural and natural ecosystems. Under laboratory conditions, earthworms can enhance aggregate stability, depending on water content, organic matter quality, time since cast deposition, and so on (Shipitalo and Protz 1989). However, it is not certain if these results can be extrapolated into a generalized concept of earthworms improving soil structure under field conditions. The key to understanding this phenomenon may lie in better knowledge of the nature and dynamics of organic compounds and their interactions with soil minerals as influenced by earthworms. Both field and laboratory studies are needed to address this problem adequately, and this should be a profitable area for future research. The stability of earthworm casts also has implications for the storage and turnover of nutrients in soil and should receive more attention. Edwards et al. (1995) and Lee (1995) made a number of recommendations for various kinds of research specifically on earthworms in agroecosystems. Many ecological questions about earthworms remain unanswered, including details on basic life cycles, habitat requirements, and distribution and population patterns in many geographical areas. The basic biology and ecology of relatively few species are understood, and these are mainly lumbricids. However, Barrios et al. (1999) summarized the knowledge of the life histories of 59 species of earthworms very effectively. Projects such as the “Earthworms Downunder” study in Australia (Baker et al. 1995a) are needed in agricultural regions worldwide if the resource base available for future earthworm research is to be assessed. Such efforts will also provide base information for earthworm inoculation studies, which often are used in laboratory experiments but need to be done much more on a field scale to assess the potential for earthworms to enhance crop growth and improve soil properties in agroecosystems. A much better understanding of the impacts of various earthworm species and earthworm community assemblages on soil organic matter dynamics in agroecosystems, especially those under no tillage or receiving mainly organic inputs, is needed. Earthworms may be a primary mechanism for the incorporation and decomposition of organic residues in these systems. Also, it could prove fruitful to include earthworms in simulation models of soil organic matter turnover, which typically do not explicitly include the effects of the soil biota other than microorganisms and roots (see Chapter 9, this volume). Finally, there is a continuing need for studies into the effects of crop rotations and management systems (including use of cultivation, pesticides, fertilizers, irrigation, etc.) on earthworm abundance, diversity, and distribution. This is particularly true when new management practices and cropping systems are tested and management goals include the buildup and maintenance of earthworm populations. Much more information along these lines is needed to improve the ability to manage earthworms in the field as tools to improve soil fertility and productivity.

CONCLUSIONS The growing number of participants at each successive International Symposium on Earthworm Ecology attests to the increasing interest in many aspects of earthworm biology and ecology. This chapter has covered only some of the selected highlights of presentations at the international © 2004 by CRC Press LLC

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earthworm symposia; the reader is referred to other chapters in this volume and to the proceedings of these meetings for a wealth of details about current research on earthworms and agriculture. Despite the large literature body on earthworms, much remains to be discovered before the beneficial activities of earthworms in agriculture can be exploited fully. Clearly, excellent progress is being made in a number of areas. However, the ultimate test will be to demonstrate how successfully predictions can be made about effects of earthworms on ecosystem processes and about management of earthworm populations and communities to improve soil fertility. The simulation models at this stage are mostly empirical and almost perhaps adequate for some management purposes (e.g., site-, species-, or management-specific conditions). However, as the knowledge base grows and questions become much more focused and refined, better understanding of earthwormlinked processes and the development of mechanistic models that predict earthworm behaviors over wide ranges of environmental conditions should be expected. Steps in this direction have been made by Martin and Lavelle (1992), Monestiez and Kretzschmar (1992), Klok et al. (1995) and Bohlen et al. (Chapter 9, this volume).

ACKNOWLEDGMENTS The authors are grateful to G.G. Brown, M.A. Callaham, Jr., and D.A. Crossley, Jr. for helpful discussions during preparation of this chapter.

REFERENCES Baker, G.H., T. Thumlert, L. Meisel, P.I. Carter, and GP. Kilpin. 1995a. “Earthworms Downunder”: a survey of the earthworm fauna in urban and agricultural soils in Australia, Soil Biol. Biochem., 29, 589–598. Baker, G.H., P.M.L. Williams, P.I. Carter, and N.R. Long. 1995b. Influence of lumbricid earthworms on wheat and clover yield and quality, Soil Biol. Biochem., 29, 599–602. Barrios, I., P. Lavelle, M. Brossard, J. Tondol, M. Martinez, J.P. Rossi, B.K. Serapati, A. Angeles, C. Fragoso, J.J. Jimenez, T. Decaëns, C. Lattand, J. Kanyonyo, E. Blanchart, L. Chapius, G.G. Brown, and A. Moreno. 1999. Ecology of earthworm species with large environmental tolerance and/or extended distributions, in Earthworm Management in Tropical Agroecosystems, P. Lavelle, L. Brussaard, and P. Hendrix, Eds., CAB International, Wallingford, U.K., pp. 57–86. Binet, F. and P. Trehen. 1992. Experimental microcosm study of the role of Lumbricus terrestris (Oligochaeta: Lumbricidae) on nitrogen dynamics in cultivated soils, Soil Biol. Biochem. 24, 1501–1506. Blair, I.M., R.W. Parmelee, and P. Lavelle. 1995. Influences of earthworms on biogeochemistry, in Ecology and Biogeography of Earthworms in North America, P.F. Hendrix, Ed., Lewis Publishers, Boca Raton, FL, pp. 125–156. Blakemore, R.I. 1995. Earthworms and pasture production in south-east Queensland, Soil Biol. Biochem., 29, 603–608. Bohlen, P.J., R.W. Parmelee, and C.A. Edwards. 1997. Earthworms (Lumbricus terrestris) alter nutrient dynamics and microbial activity of surface litter in agroecosystems, Ecol. Appl., 7, 341–349. Bovicini, A.M. Paglian, and P. Omodeo, Eds. 1987. On Earthworms, Mucchi Editore, Modena, Italy. Bouché, M.B. 1977. Stratégies Lombriciennes, in Soil Organisms as Components of Ecosystems: Proceedings of the 6th International Colloquium on Soil Zoology, U. Lohm and T. Persson, Eds., Swedish Natural Science Research Council Ecological Bulletin No. 25, Stockholm, pp. 122–133. Cortez, J., R. Hameed, and M.B. Bouche. 1989. C and N transfer in soil with or without earthworms fed with 14C- and 15N-labelled wheat straw, Soil Biol. Biochem., 21, 491–497. Devliegher, W. and W. Verstraete. 1995. The effect of Lumbricus terrestris on soil in relation to plant growth, Soil Biol. Biochem., 24, 341–346. Diaz Cosin, D.J., J.B. Jesus, D. Trigo, and M.H. Garvin, Eds. 1999. Proceedings of Sixth International Symposium on Earthworm Ecology, 43, 481–908. Edwards, C.A., Ed. 1997. Proceedings of Fifth International Symposium on Earthworm Ecology, Soil Biol. Biochem. 29, 215–766.

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Edwards, C.A., Ed. 1998. Earthworm Ecology, St. Lucie Press, Boca Raton, FL. Edwards, C.A. 1999. Soil invertebrate controls and microbial interactions on organic matter dynamics in natural and agroecosystems, in Webmaster Functions in Agrosystems: Invertebrate, Climate and Substrate Quality, Coleman, D.C. and P. Hendrix, Eds., CABI Publ., Wallingford, U.K., pp. 141–159. Edwards, C.A. and P.J. Bohlen. 1996. Biology and Ecology of Earthworms, 3rd ed., Chapman & Hall, New York. Edwards, C.A., P.J. Bohlen, D.R. Linden, and S. Subler. 1995. Earthworms in agroecosystems, in Ecology and Biogeography of Earthworms in North America, P.F. Hendrix, Ed., Lewis Publishers, Boca Raton, FL, pp. 185–214. Edwards, C.A. and K.E. Fletcher. 1988 Interactions between earthworm and microorganisms in organic matter breakdown, Agric. Ecosyst. Environ., 24(1–3), 235–249. Edwards, C.A. and J.R. Lofty. 1969. The influence of agricultural practice on soil microarthropod populations, in The Soil Ecosystem, J.G. Sheals, Ed., Systematics Association, London, pp. 237–247. Edwards, C.A. and E.F. Neuhauser. Eds. 1988. Earthworms in Waste and Environmental Management, SPB Academic Publishing, The Hague, the Netherlands, pp. 21–31. Gilot-Vilienave, C. 1995. Effects of inoculation with the tropical endogeic earthworm M. anomala on soil characteristics and yam production in Cote d’Ivoire, Soil Biol. Biochem., 24, 353–360. Greig-Smith, P.W., H. Becker, P.J. Edwards, and F. Heimbach. Eds. 1992. Ecotoxicology of Earthworms, Intercept, Andover, U.K. Hendrix, P.F., Ed. 1995. Earthworm Ecology and Biogeography, Lewis Publishers, Boca Raton, FL. Hendrix, P.F., D.C. Coleman, and D.A. Crossley, Jr. 1992. Using knowledge of soil nutrient cycling processes to design sustainable agriculture, J. Sustainable Agric., 2, 63–82. James, S.W. 1991. Soil, nitrogen, phosphorus, and organic matter processing by earthworms in tallgrass prairie, Ecology, 72, 2101–2109. Klok, C., A.M. de Roos, H.M. Baveco, J.C.Y. Marinissen, and W. Ma. 1995. Modelling population dynamics of the earthworm Lumbricus rubellus. I. Identification of most sensitive life history parameters, Soil Biol. Biochem., 24, 287–294. Kretzschmar, A., Ed. 1992. Proceedings of 4th International Symposium on Earthworm Ecology, Soil Biol. Biochem., 24, 1193–1773. Lal, R. 1991. Soil conservation and biodiversity, in The Biodiversity of Microorganisms and Invertebrates: Its Role in Sustainable Agriculture, D.L. Hawksworth, Ed., CAB International, Wallingford, U.K., pp. 89–104. Lavelle, P., L. Brussaard, and P.F. Hendrix. Eds. 1999. Earthworm Management in Tropical Agroecosystems, CAB International, Wallingford, U.K. Lee, K.E. 1985. Earthworms: Their Ecology and Relationships with Soil and Land Use, Academic Press, Sydney, 411 pp. Lee, K.E. 1995. Earthworms and sustainable land use, in Ecology and Biogeography of Earthworms in North America, P.F. Hendrix, Ed., Lewis Publishers, Boca Raton, FL, pp. 215–234. Martin, S. and P. Lavelle. 1992. A simulation model of vertical movements of an earthworm population (Millsonia anomala amodeo, Megascolecidae) in an African savanna (Lamto, Ivory Coast), Soil Biol. Biochem., 24, 1419–1424. Martin, A., A. Mariotti, J. Balesdent, and P. Lavelle. 1992. Soil organic matter assimilation by a geophagous tropical earthworm based on delta 13C measurements, Ecology, 73, 118–128. Monestiez, P. and A. Kretzchmar. 1992. Estimation of the relationship between structural parameters of simulated burrow systems and their partitioning effect, Soil Biol. Biochem., 24, 1549–1554. Parmelee, R.W., P.J. Bohlen, and J.M. Blair. 1998. Earthworms and nutrient cycling processes integrating across the ecological hierarchy, in Earthworm Ecology, C.A. Edwards, Ed., St. Lucie Press, Boca Raton, FL, pp. 123–146. Satchell, J.E., Ed. 1983. Earthworm Ecology: From Darwin to Vermiculture, Chapman & Hall, London. Satchell, J.E. and K. Martin. 1983. A Bibliography of Earthworm Research, Pt. I, Institute of Terrestrial Ecology, Grange-over-Sands, U.K., 113 pp. Seastedt, T.R., D.A. Crossley, Jr., V. Meentemeyer, and J.B. Waide. 1983. A two year study of leaf litter decomposition as related to macroclimatic factors and microarthropod abundance in the southern Appalachians, Holarctic Ecol., 6, 11–16.

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Shephard, S., J. Berribridge, M. Holmstrup, and L. Posthuma. Eds. 1998. Advances in Earthworm Ecotoxicology, Society of Environmental Toxicology and Chemistry, Pensacola, FL. Shipitalo, M.J. and R. Protz. 1989. Chemistry and micromorphology of aggregation in earth-worm casts, Geoderma, 45, 357–374. Spain, A.V., P. Lavelle, and A. Mariotti. 1992. Stimulation of plant growth by tropical earthworms, Soil Biol. Biochem., 24, 1629–1634. Tian, G., B.T. Kong, and L. Brussaard. 1995. Earthworm activity and substrate chemical composition interaction during decomposition of plant residue, Soil Biol. Biochem., 24, 369–374. Tomlin, A.D., M.J. Shipitalo, W.M. Edwards, and R. Protz. 1995. Earthworms and their influence on soil structure and infiltration, in Ecology and Biogeography of Earthworms in North America, P.F. Hendrix, Ed., Lewis Publishers, Boca Raton, FL, pp. 157–181. Wallwork, J.A. 1976. The Distribution and Diversity of Soil Fauna, Academic Press, London. Worden, D.D. 1981. Bibliography, Vol. 2. Workshop on the Rate of Earthworms in the Stabilization of Organic Residues, Beech Leaf Press, Kalamazoo, MI. Zhang, Q.L. and P.F. Hendrix. 1995. Earthworm (Lumbricus rubellus and Aporrectodea caliginosa) effects on carbon flux in soil, Soil Sci. Soc. Am. J., 59, 816–823.

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Part VIII Earthworms and Environmental Pollution

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as Test Organisms 16 Earthworms in Ecotoxicological Assessment of Toxicant Impacts on Ecosystems Adriaan J. Reinecke and Sophié A. Reinecke Department of Zoology, University of Stellenbosch, Stellenbosch, South Africa

CONTENTS Introduction ....................................................................................................................................299 The Test Organisms .......................................................................................................................300 Acute Toxicity Testing with Earthworms ......................................................................................301 Chronic Toxicity Tests and Sublethal Effects ...............................................................................303 Earthworm Toxicity Testing for Regulatory Purposes ..................................................................304 The Use of Earthworm Biomarkers and Understanding Bioavailability......................................305 Estimating Environmental Exposure of Earthworms to Toxicants ...............................................308 Earthworms in Bioassays, Microcosms, and Model Ecosystems .................................................308 Field Toxicity Tests Using Earthworms ........................................................................................309 Residues in Earthworms and their Role as Biomonitors and Bioindicators ................................310 Risk Assessment Using Earthworms .............................................................................................311 Decisions on Managing Estimated Risk of Chemicals to Earthworms........................................314 Conclusions ....................................................................................................................................315 References ......................................................................................................................................316

INTRODUCTION Concerns about sustaining soil fertility in agricultural land, in which a variety of toxicants are used for crop protection; risks of chemicals leaching into drinking water; contamination of soil; and detrimental effects of contaminants on the nontarget living environment have grown. This is also evidenced by the realization that endocrine-disrupting effects of some toxicants can have severe implications for ecosystems as well as human health. This has also resulted in a strong and growing awareness of the importance of soil biodiversity and its role in providing ecosystem services (Bengtsson 1997; Daily 1997). Environmental scientists and legislators are today more aware of the need to protect the structure and functioning of soil ecosystems in which the decomposer community performs especially vital functions. Many toxic materials accumulate through food webs, and the detritivore-decomposer levels are often the first to be affected. Organic matter, which serves as a valuable resource for the soil biota, and the soil itself are the ultimate sinks for most contaminants. Legislation in many countries has focused the attention of scientists on the need for sensitive organisms from the soil environment for use in research, in environmental monitoring, as 299 © 2004 by CRC Press LLC

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indicators of contamination, and for regulatory toxicity testing, similar to developments in the aquatic field. Earthworms play an important role as “engineers” in many soils, in which they contribute to the complex processes of organic matter decomposition and affect aeration, water transport, and soil structure. Van Hook (1974) suggested that earthworms could serve as useful biological indicators of contamination because of the fairly consistent relationships among the concentrations of certain contaminants in earthworms. They can also play a valuable role as in situ sentinels for assessing biological risks from toxic components in terrestrial environments (Morgan 1992). As earthworms form an important component of the soil biota, especially because of their role in soil fertility, their protection is also important. Because earthworms are common in many soils, they are vulnerable to chemical and physical impacts on soils. Although earthworms are by no means the only important soil-dwelling invertebrates, they are selected for assessment of environmental risk for various reasons. They are widely distributed and beneficial in promoting soil fertility and serve as food for a variety of animals. Not only are they general representatives of the soil fauna, but also they are practical to breed and use in both laboratory and field toxicity tests. They are convenient to handle because of their relatively large size and can reach a relatively high biomass in some soils. Many species are suitable for captive breeding, which makes them readily available. Their behavior brings uncontaminated individuals into close contact with the soil, and they have relatively short life cycles, allowing longterm studies of successive generations. They are therefore very suitable for genetic studies of adaptive responses and could, for example, be used to evaluate the genetic erosion hypothesis by quantifying genetic variation in toxicant-stressed populations (Van Straalen and Timmermans 2002). New trends in environmental toxicity testing have not rendered earthworm testing obsolete. The use of aquatic toxicity tests for predicting the toxicity of soil samples (Gälli et al. 1994) was a useful development that was also followed by the pore-water concept (Van Gestel 1997). This method extracts an aqueous leachate from the soil, which is then bioassayed aquatically. However, it does not replace the need for actual tests with soil-dwelling organisms because compounds that are not sufficiently water soluble or adsorb very strongly on soil particles cannot be evaluated by aquatic bioassays. The use of earthworms for ecotoxicological evaluation has undergone considerable progress since the f i rst International Workshop on Earthworm Ecotoxicology was held in Sheffield in 1991. Two further international workshops, in the Netherlands in 1997 (Sheppard et al. 1998) and Denmark in 2001, have made comprehensive recommendations to researchers and regulators; these are discussed, but not repeated in detail here. Various new approaches have been adopted to refine and simplify procedures, which could lead to more ecologically relevant information because ecotoxicological test methods for soils are relatively underdeveloped (Van Straalen and Van Gestel 1993). However, various problems remain, and although some have been addressed by adopting compromises, there is an urgent need to develop acceptable approaches to field testing and the use of biomarkers. Some of these new approaches became evident at the Seventh International Symposium on Earthworm Ecology (ISEE), which took place in Cardiff, Wales, in 2002. This chapter attempts to discuss the latest approaches to the various problems associated with using earthworms in toxicity testing and to explain new routes of investigation, with some emphasis on the contributions to the International Workshops on Earthworm Ecotoxicology (see also Chapter 17 this volume).

THE TEST ORGANISMS Single-species tests have often been criticized severely for their lack of realism in terms of their modes of exposure and for difficulties in extrapolating such results to field conditions (Edwards 2002). There have been doubts expressed about using Eisenia fetida as a “typical” earthworm in single-species toxicity testing. However, in acute toxicity testing for regulatory purposes, the position of this species seems to be well established. The basic requirement of finding a species © 2004 by CRC Press LLC

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that is easy to rear and genetically homogenous (Bouché 1992) can be fulfilled by using individuals of this species. Heimbach (1984) was of the opinion that the sensitivity of E. fetida andrei is sufficiently comparable with that of Lumbricus terrestris in spite of the ecological differences between these species. Heimbach (1992) obtained good correlations between median lethal concentration (LC50) values from the artificial soil test (using this species) and a standardized field test. Callahan et al. (1994) suggested that E. fetida may be representative of four other earthworm species based on the concentration-response relationships for 62 chemicals and after applying the Weibull function. However, all these species are, to a large extent, representatives of the same ecological group, and extrapolation to earthworm species from other groups may be less successful. As reviewed by Edwards and Coulson (1992), in 42 pesticide comparisons by various researchers involving two to five earthworm species, there was no consistent toxicity relationship among species. However, it must be kept in mind that the toxicity tests were seldom conducted under identical experimental conditions. A factor of 10 was suggested to bring E. fetida in line with the most sensitive species (Heimbach 1992), an approach that allows the continued use of E. fetida in many toxicity tests. In the sense that E. fetida combines sensitivity, economic importance, and ecological relevance, it can be seen as the selected earthworm species for routine toxicity testing. Those who criticize E. fetida as not representative of most other soil-dwelling species, although generally valid, should keep in mind that a three-tier approach is customary in risk assessment. The first tier is merely an indicator in the procedure, and the existing guidelines are suitable for most risk assessments required for legislative purposes (C. Kula 1998). In chronic toxicity tests as well as bioaccumulation studies, the ideal would be to include a representative of each ecological type. The three main ecological types of earthworms (Bouché 1972) are the endogeic, epigeic, and anecic species. Sublethal effects on the growth and reproduction of earthworms have been measured in the laboratory (Lofs-Holmin 1980; N.A. Martin 1986; Van Gestel et al. 1989; Helling et al. 2000), but lack of sufficient comparative data on sublethal effects has made comparisons based on species susceptibility impossible. Reproduction toxicity tests on E. fetida (Van Gestel et al. 1989) have led the way in showing that this species could also be considered a standard laboratory earthworm species for studying sublethal effects. Edwards and Coulson (1992) advocated a program of cooperative research to compare the susceptibility of different species under identical conditions at different laboratories. This has not yet materialized fully, although the database is growing steadily. Edwards (2002) also advocates the use of soil microcosms and terrestrial model ecosystems (TMEs) to do more holistic integrated studies that are more field relevant than singlespecies tests. There is clearly a need to distinguish between the purposes of the various earthworm toxicity tests. Although E. fetida may be suitable for some tests, and there are also sufficient grounds to extrapolate to certain other species and soils, it can never serve as the sole representative species of the soil environment for assessing ecological risks. A much broader-based approach, although less practical for regulatory purposes, is required if an in-depth understanding of the effects of chemicals on soil ecosystems is required. In this context, the earthworm has only a limited contributory role to play. Earthworm researchers in the field of ecotoxicology should therefore not limit themselves to the requirements proposed for regulatory purposes but should follow a more holistic approach in recognizing that earthworms are but one component of a very complex soil environment, of which there is only a very limited understanding. Because E. fetida is a useful organism for toxicity testing does not necessarily make it a useful biological indicator species in the ecological sense or a more useful biomonitoring species.

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Acute toxicity studies are designed to express the potency of the toxicant in terms of a median lethal dose LD50 , causing the death of 50% of the universal population of the species exposed, under the defined conditions of the test. When indirect administration of a toxicant is done, the potency is expressed as the median lethal concentration LC50, which is more often the case in earthworm studies. The Workshops on Earthworm Ecotoxicology held in the Netherlands and Denmark concluded that acute earthworm toxicity testing has been well established with standardized end points, but there is still room for improvement (C. Kula 1998). Various researchers, including Van Gestel (1992) and Spurgeon and Weeks (1998), have pointed out the limitations of this test for purposes of extrapolation to the field. These tests per se do not provide sufficient information to predict the effect of a chemical applied in field situations. These acute artificial soil tests on earthworms are now conducted with well-established protocols (Reinecke 1992; Van Straalen and Van Gestel 1993), have preliminary screening value, and provide useful information on relative toxicity of chemicals. Edwards and Coulson (1992) recommended that initial screening, using the Organization for Economic Cooperation and Development (OECD) Guideline 207, based on the pioneering reports of Edwards (1983, 1984) on ring tests undertaken by 35 laboratories, remains the standard laboratory practice. Van Gestel (1992) supported this view based on the results of Van Gestel and Ma (1990) and Van Gestel et al. (1991) and after studying the influence of soil characteristics on the toxicity of chemicals. This approach was also supported by subsequent international workshops (Bembridge 1998). Edwards and Bohlen (1992) reviewed all published work on the toxicity of chemicals to earthworms involving the effects of more than 200 chemicals and ranked the chemicals as nontoxic, slightly toxic, moderately toxic, very toxic, and extremely toxic. To a large extent, the evaluation of the environmental risk of chemicals has until now relied on acute earthworm toxicity tests, although no clear foundation for the validity of such an approach existed. These laboratory tests have been conducted over a wide range of chemicals and earthworms (Goats and Edwards 1982; Neuhauser et al. 1983, 1986; Roberts and Dorough 1983, Vermeulen et al. 2001), but the application of these test results to environmental risk analysis is lacking. Callahan et al. (1994) followed a new approach toward comparing earthworm species toxicity by integrating an extensive database on the acute toxicity of chemicals to earthworms based on the Weibull function (Weibull 1951) used by Shirazi and Lowrie (1988) for fish. This technique generalized the relationship between the chemical and the organism in terms of two parameters: toxicity (scale factor k) and tolerance (form factor a). Their analysis provided an assessment of the relative tolerance of four earthworm species to each of 62 chemicals and in relationship to each other using two test protocols, the contact test and the artificial soil test. Their results also suggested that E. fetida may be representative of a group of species: Allolobophora tuberculata, Eudrilus eugeniae, and Perionyx excavatus. Although laboratory tests cannot fully simulate the structure of soil or the behavior of earthworms in the field, certain end points other than mortality can be obtained from these tests. The use of the same test protocol to study changes in body weight and reproduction can provide useful data when integrated with results of field tests, but does not preclude the need for the latter, when required. Acute toxicity tests in artificial soil can probably be extrapolated to natural soils, but data are still scarce (Spurgeon and Hopkin 1995). Acute toxicity tests are designed to identify very toxic chemicals that have immediate effects and not to determine acceptable environmental concentrations. They can, however, serve as a qualitative screen for detection of other toxicological effects and give a first estimate of the NOEC (no observed effect concentration) level for continuous exposure. These tests do not clear any chemical from further testing but do assist in practice in setting priorities for further testing of sublethal effects. The bottom line is that the OECD standard artificial soil test was not necessarily developed to enable the direct extrapolation of toxicity results to the field situation (Spurgeon and Weeks 1998).

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CHRONIC TOXICITY TESTS AND SUBLETHAL EFFECTS Living earthworms in the soil are exposed, more frequently over much longer periods of time, to chemical agents at levels lower than those that are fatal. To simulate such exposure, short-term (subchronic) and long-term (chronic) studies are required. The former are usually conducted over 14 days or longer, whereas the latter are conducted over 1 or 2 years. Sublethal earthworm toxicity testing has now been accepted as part of the routine regulatory testing scheme in Europe and is designed to detect subtle effects, such as disturbances in earthworm behavior, retarded development, lowered fertility, teratogenic effects, and the like that may cause population changes without necessarily leading to mortality. Data referring to the impairment of vital functions of organisms often have a greater ecological relevance than those obtained from measurements of acute toxicity (Ma 1984). Standardization of a chronic toxicity test for earthworms has not been achieved fully because procedures vary and standardization is limited. The ecological significance of the results is still uncertain, and reproducibility is limited. A variety of tests has been suggested. The pesticide test for sublethal effects (Kokta 1992), which was subjected to a ring test in Germany, and the artificial soil test proposed by Van Gestel et al. (1989) have been used with minor adjustments by various researchers. International Organization for Standardization (ISO) draft guideline DIS 11268-2 (Bembridge 1998) for an earthworm reproduction study provides an adequate method for a second-tier assessment, and previous workshops on earthworm ecotoxicology have considered it sufficient. Useful end points that have emerged from various studies and are mentioned in the recommendations of the 1997 earthworm ecotoxicology workshop (Bembridge 1998) include the following: • • •

• • • •

Growth of juveniles Changes in body weight of adults Reproduction Rate of cocoon production Number of hatchlings Hatching success Incubation time Sperm parameters Behavior Morphological effects Physiological effects Decomposition in the soil as an indirect measure of activity

Chronic toxicity tests are labor intensive and time consuming but seem to be the next logical step, within certain toxicity limits, after the acute toxicity level of a chemical has been established in screening tests (Kokta and Rothert 1992) if the estimated exposure concentration was high. However, it must be kept in mind that environmental behavior and other factors associated with a chemical may still require that chronic toxicity testing be done although a chemical has “passed” the acute toxicity test and has been rated as “probably a low risk.” The relevance of life history parameters in earthworm ecotoxicology was demonstrated clearly by Cluzeau et al. (1992) and Reinecke and Reinecke (1996) and also stressed by Forbes (personal communication). Cluzeau et al. showed that disturbances in nontarget earthworm population dynamics could affect the quality of cultivated soils. It is especially relevant in the context of population recovery after a disturbance (Edwards and Brown 1982). Helling et al. (2000) showed that juvenile earthworm growth and cocoon production can respond sensitively to very low levels of copper-containing fungicides, although the chemical is considered relatively harmless to earthworms.

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To date, earthworm behavior has not often been used as a criterion in ecotoxicological studies (Reinecke et al. 2002), and its value is not fully understood. Field studies showed that earthworms may migrate over large distances in response to environmental changes, and other studies demonstrated avoidance behavior when exposed to highly contaminated soils (Yeardley et al. 1996; HundRinke and Wiechering 2001). These aspects of earthworm behavior may be utilized as ecotoxicological end points. The Denmark earthworm ecotoxicology workshop (2001) recommended that more attention be given to understanding earthworm behavior, interpreting avoidance responses and surface migrations. This holds both for pesticide exposure testing and the risk assessment of contaminated land.

EARTHWORM TOXICITY TESTING FOR REGULATORY PURPOSES Over the years since the earlier workshops on earthworm ecotoxicology, laboratory toxicity testing has become further standardized in the guidelines ISO 11268-1 (acute earthworm toxicity test), ISO 11268-2 (earthworm reproduction toxicity test), and ISO 11268-3 (earthworm field toxicity test). Moreover, several other international bodies are currently developing their own testing guidelines (e.g., OECD and Environment Canada), basically applying principles similar to the ones laid down in the ISO guidelines. The earthworm ecotoxicology workshop held in Denmark in 2001 considered the possibilities of improving existing methods and the aspects needed for proper use of earthworm ecotoxicology data in risk assessment. In addition to the use of data to test the toxicity of new and existing chemicals and for the registration and admission of pesticides, there was a strong emphasis on the value of using earthworm toxicity methods for the assessment of the status of contaminated land. The Denmark workshop recommended that observations of performance (e.g., earthworm growth), burrowing activity, and other sublethal end points in the existing toxicity tests should be considered as potential end points in earthworm bioassays on contaminated soils. In the original test guideline of acute earthworm toxicity testing (OECD 1984), a filter paper contact test was also described. Exposure conditions using this method are vastly different from any exposure scenario in soil, bearing little relevance to soil toxicity (see, e.g., Heimbach 1984; Van Gestel and Van Dis 1988). Participants at the workshop in Denmark (2001), however, argued that the filter paper test has a role in the risk assessment procedure because it is useful to narrow down the range for toxicants prior to soil testing (see, e.g., Roberts and Dorough 1983). In addition, it still is a useful research tool when comparing different routes of exposure. In the artificial soil acute toxicity test on earthworms, weight changes of the earthworms used are usually recorded. Because adult earthworms are used in this test, weight loss or gain cannot and should not be interpreted as effects on growth. However, weight loss in acute toxicity tests on earthworms can serve as a check of the validity of the exposure test. Weight loss (when there is no mortality) can be used as an indication of a need for further testing. The Denmark workshop recommended a compilation of existing data on earthworm weight changes in acute toxicity tests to establish a “critical weight loss” that might trigger further testing. Attempts to demonstrate the development of resistance to chemicals such as heavy metals among earthworm populations in highly polluted areas experimentally have been only partly successful (Bengtsson et al. 1992; Spurgeon and Hopkin 1999, 2000; Reinecke et al. 1999; Morgan personal communication). Therefore, it remains unclear whether earthworms can develop resistance to long-term metal exposure. However, it remains an important issue if earthworms are, for regulatory purposes, to be used in standardized toxicity tests or even as bioindicators because the occurrence of resistant genotypes in a test population may bias results.

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THE USE OF EARTHWORM BIOMARKERS AND UNDERSTANDING BIOAVAILABILITY The use of sensitive biomarkers in toxicity tests using earthworms represents a fairly new approach (A.J. Reinecke and Reinecke 1998). Scott-Fordsmand and Weeks (1998) reviewed biomarkers comprehensively. A biomarker is defined as a xenobiotically induced variation in cellular or biochemical components or processes, structures, or functions that is measurable in a biological system or sample. As such, it constitutes a response to the presence of a stress factor. Relatively little research has been undertaken so far on biomarkers in earthworms. Biomarkers have been evaluated in L. terrestris and E. fetida (Goven et al. 1988; Rodriquez et al. 1989; Svendsen and Weeks 1997; S.A. Reinecke et al. 2002). For instance, a broad spectrum of xenobiotics can alter the immune function. The immunobiology of earthworms has been studied intensively in immunotoxicological studies (Fitzpatrick et al. 1990; Venables et al. 1992; Goven et al. 1994; Marino and Morgan 1998). Although a number of coelomocyte-based end points have been shown to be sensitive indicators of the sublethal toxicity of chemicals (Venables et al. 1992) and hold promise to provide cost-effective assessment of potential risks, evidence of ecological relevance is limited, although promising (Maboeta et al. 2002). The comet assay (Figure 16.1), measuring the effects of toxicants on deoxyribonucleic acid (DNA) integrity in earthworms, could also serve as a biomarker (S.A. Reinecke and Reinecke 2003). This represents a new approach, which may yield promising results if a battery of biomarkers can be employed, to predict toxic risks to organisms at higher levels of organization. An important new development is the application of modern molecular techniques to identify biomarkers in earthworms. Heavy metal-responsive genetic indexes (Stürzenbaum et al. 1998) may enhance future testing protocols. A number of molecular genetic techniques have already been utilized to identify genes responsive to certain types of exposure to toxicants. The value of biomarkers in risk assessment (Weeks 1995) has received increasing attention because of the problems experienced when trying to relate environmental concentrations of toxicants to their bioavailable fraction. The bioavailability and toxicity of a chemical may differ considerably in laboratory tests from those observed in the field. Biomarkers, on the other hand, respond to the dose of a toxicant entering the animal body.

FIGURE 16.1 The use of the comet assay as a biomarker. An example of changes in the DNA structure of nuclei of coelomocytes of the earthworm, Eisenia fetida, before and after the worms were exposed to nickel.

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The bioavailability problem may be addressed by measuring biomarker responses accurately. This would facilitate the setting of environmental quality criteria, for generic use by regulatory authorities, in risk assessment. The heterogeneity and complexity of soil and other systems may force us to think in terms of site-specific criteria for each unique situation (differentiated environmental quality criteria). This would require a completely different approach (Slijkerman et al. 2000) to the standardized one currently implied or accepted within the framework of existing laws and regulations in many developed countries. At the subcellular level, progress has been made in showing that certain biomarker responses, such as the neutral red retention time assay in earthworms (Weeks 1995; Svendsen and Weeks 1997; Spurgeon et al. 2000; Reinecke et al. 2002), can be used reliably to link changes in the permeability of the lysosomal membranes to ecologically relevant life cycle effects caused by certain substances. Although this does not constitute a definite link to structural and functional properties of the ecosystem, these authors demonstrated dose-related responses in earthworm weight change and reproduction that correlated with neutral red retention times. This neutral red retention technique is therefore well established for measuring the cellular toxicity of certain chemicals to earthworms and other invertebrates. It is sensitive and easy to use and has now been used for a few metals and organic chemicals on a range of earthworm species, such as Lumbricus terrestris, Lumbricus rubellus, Lumbricus custaneas, Aporrectodea rosea, Allolobophora caliginosa, Eisenia fetida, Eisenia andrei, and a Microchaetus sp. In all cases, reliable dose-response relationships were established by various researchers, emphasizing the reliability of the technique in spite of the fact that it relies heavily on subjective observations (S.A. Reinecke and Reinecke 1999). The use in ecotoxicology of stress proteins, metallothioneins, ultrastructural changes, and isozymes has also gained ground as potential tools for toxicant hazard identification and exposure assessment (Kammenga et al. 2000). The usefulness of these biomarkers should become clearer once a large enough database for different species of earthworms and soil and contaminant types becomes more readily available. Some biomarkers may not be practical for regular use in field monitoring, but others, such as the neutral red retention assay, may be used at low cost by a trained technician on organisms sampled from the field to provide a quick indication of toxic stress. The exposure of uncontaminated organisms with a known biomarker response under uncontaminated conditions could provide a more practical and realistic tool for measuring bioavailability in contaminated soil, or at least circumventing some of the problems in assessing toxicant bioavailability (Weeks 1997), than is currently appreciated. This can be done either by placing earthworms directly in the field in outdoor enclosures or microcosms or by exposing them to field soil in the laboratory and subsequently monitoring the changes in biomarker response. This may in the future provide an excellent biomonitoring tool for evaluating or at least estimating bioavailability and eventually for predicting earthworm population changes caused by continued toxic stresses. Biomarker responses provide an early indication that toxicant uptake and internal exposure has occurred, and that a toxic response may have been initiated. In comparison with chemical toxicity measurements, biomarkers have a striking advantage in that they are indicative of a reaction that has commenced and not just of a chemical concentration. This already gives the biomarker a higher relevancy in the assessment of toxicity than the measurement of a chemical concentration could. This is especially true not only for biochemical responses to toxicants (Stürzenbaum et al. 1998), but also for cellular responses such as changes in the lysosomal membrane stability that reflect various degrees of cytological damage in a dose-related manner. This response may vary among different earthworm species but is expected to reflect differences in the relative sensitivity of earthworm life cycle traits (Spurgeon et al. 2000). It is, however, a dose-related response and probably presents a more reliable reflection of the bioavailable fraction of the toxic substance taken up over time from the substrate, thereby causing the effect. Although nonspecificity is generally seen as a drawback in relating cause and effect, this response may nevertheless be useful precisely because it integrates the background toxicant concentrations, in both the organism and the substrate as well as other stress factors, with the effect © 2004 by CRC Press LLC

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of the bioavailable fraction of an environmental contaminant into the same combined response. Because it is an early response, it may, for certain substances at least, also reflect the effect of contamination age or time on bioavailability (Reinecke and Reinecke 2003). It may be able to differentiate whether equilibrium has been attained in the substrate matrix. It could also (under similar conditions of pH, soil texture, etc.) reveal differences in the bioavailability of a substance in a spiked substrate compared with that in a contaminated field sample containing exactly the same concentration of toxicant but with a long exposure history. Because the bioavailability of a substance can be dynamic and variable, simple biomarker responses could provide an ideal method for the regular assessment of changes, provided that response sensitivity is maintained and that practical feasibility of methodology and execution can be achieved. Contaminated soils usually contain mixtures of toxicants (Weltje 1998), causing multiple exposures, which in any case make the causal linking of a response to one toxicant in the mixture very difficult, if not impossible. The nonspecificity of some biomarkers that is generally perceived as a drawback in ecological risk identification may therefore still be utilized. Although the relative contribution of the different factors to the stress response cannot be singled out, it may still provide a practical and realistic tool for purposes of assessing the stresses caused by the bioavailability of a chemical, in conjunction with existing confounding factors, under field conditions. After all, if the protection of biodiversity is kept in mind, the mere establishment of stressful conditions, irrespective of their nature or origin, at an early stage is of primary importance. The idea of using biomarkers for the more accurate quantification of exposure to contaminants is gaining support (Kammenga et al. 2000; Spurgeon et al. 2000). Given the complexities involved in dealing with the various parameters that can affect the bioavailability of contaminants in soils and the difficulties in trying to incorporate them into a practical risk assessment procedure, alternative approaches are needed. Therefore, it is not surprising that a trimming of risk assessment procedures to a more manageable scale is sought by many regulatory authorities. This can be achieved by acknowledging the biological nature of the bioavailability concept and using a realistic biological tool such as earthworm biomarkers to assess toxicant bioavailability or to validate the estimated toxicant bioavailability values obtained by chemical fractionation methods such as the PBASE (Basta and Gradwohl 2000). The Earthworm Ecotoxicology Workshop held in Denmark in 2001 concluded that, for a biomarker or a battery of biomarkers to be useful and applicable under field conditions, they should comply with some basic requirements: 1. Chemicals: Information on specificity is needed about the range of hazards (including chemicals) that elicit a biomarker response; that is, biomarkers responding to a broad range of chemicals or other factors can be used in monitoring or screening studies, and biomarkers responding to a narrow range of compounds can be used in the case of known pollutants. Therefore, the mechanistic connections of the biomarker response to the desired end point must be known. 2. Linking to higher levels: Because the ultimate aim is to validate a marker for regular toxicity test use, there should be a correlation (linkage) between the biochemical or other marker response and deleterious changes to the population or community. The Denmark workshop recommended that biomarkers should be used that are connected to the life cycle parameters of earthworms apart from their possible linkage to pure compounds. It is therefore important to test biomarkers in field experiments with known earthworm population changes to see if the biomarkers possess predictive abilities. For example, is the biomarker responding even before a population response occurs? This was seen in the study of Maboeta et al. (2002), in which the neutral red retention response preceded declines in the earthworm populations. 3. Inherent variation and temporal aspects: For a biomarker response to be useful and reliable in the field, it should ideally have a low inherent variability with a limited, or © 2004 by CRC Press LLC

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at least a known, dependence on physiological and physicochemical conditions and a good baseline data set. The induction time and the persistence of a potential biomarker response should be known to estimate the likelihood and significance of detecting a true response (avoiding false positives) in field samples. The possibility of adaptation to stressful conditions could also influence the reliability (S.A. Reinecke et al. 1999; S.A. Reinecke and Reinecke 2003). The earthworm ecotoxicology workshop in 2001 recommended that confounding factors, such as drought, temperature, and a linkage between the biomarker and the physiological responses, should be clearly established. This includes species differences and the influence of both endogenous (reproductive cycle, life stage) and exogenous cycles (diurnal, annual). It is clear from the more recent literature that research on biomarker responses in earthworms has a strong bias toward the effects of metals on them, and biomarkers for assessing exposure or effects of organic and estrogenic compounds have received much less attention.

ESTIMATING ENVIRONMENTAL EXPOSURE OF EARTHWORMS TO TOXICANTS To obtain some idea of an exposure to the environmentally available chemical (in contrast to its bioavailability), the application rate (of the active ingredient) per unit area should be taken into account to determine the degree of exposure experienced by the earthworms in the soil. The actual exposure will also depend on the behavior of the earthworms, soil conditions, and characteristics of the particular chemical. Only the bioavailable fraction of the chemical is relevant, and only biological responses or chemical analysis of the earthworms will provide an indication whether the earthworms were effectively exposed to the chemical. By assuming that the soil is evenly packed and the chemical is evenly distributed, the critical depth limit can be estimated and will vary according to the soil type and soil cover. All the factors affecting exposure should be considered, but some are difficult to measure. Various authors have made different assumptions in determining the estimated environmental concentration (EEC). The following assumptions advanced by Kokta and Rothert (1992) and others are still valid for purposes of estimating exposure following liquid toxicant applications: 1. 2. 3. 4. 5.

There is an even distribution of the chemical in the upper 2.5 cm of soil. There is a bulk density of soil of 1.5g−1 cm3. The full amount of toxicant is used if applied to bare soil. 50% of the total toxicant is used if plant cover exists. For pesticides with fewer than three applications per season, the total amount applied is summed to allow for persistence.

EARTHWORMS IN BIOASSAYS, MICROCOSMS, AND MODEL ECOSYSTEMS There is currently considerable renewed interest in studies of the mesocosm, semifield, model ecosystem types (Römbke personal communication; Burrows and Edwards 2002; Edwards 2002); these were very often used in pesticide degradation studies in the 1970s. Model ecosystems (slightly more complex than microcosms), and which approximate real ecosystems, have been used by various authors. Design features and pros and cons were discussed by Van Straalen and Van Gestel (1993) and Edwards (2002). The effects of toxicants are assessed in systems of various complexities, simulating field conditions as closely as possible. Two or more interacting organisms are usually exposed simultaneously. The development of microcosms and model ecosystems as ecotoxicological tools has © 2004 by CRC Press LLC

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proceeded beyond the laboratory experimentation phase, and there seems to be much scope for incorporating them in routine regulatory testing when applicable. The soil microcosm test has various advantages over the standardized artificial soil test. It accounts for interactions with other organisms, can be used with different soil types, has high reproducibility, and can provide reproduction data. The more complex TME closely approximates field exposure conditions but is usually more expensive and labor intensive. The use of earthworms as tools in the detection of chemical contamination includes bioassays at contaminated sites in the field or in the laboratory using soil taken from the contaminated site in microcosms. Contaminated sites are often characterized by combining the effects of complex chemical mixtures rather than single chemicals. Adverse biological effects associated with a contaminated site can therefore seldom be ascribed to a single chemical. Determination of causal relationships between specific chemicals and observed effects is not possible because of the complexity and variability of biological responses. To determine the environmental impact associated with complex mixtures in a site, a toxicity-based approach rather than a chemically based approach should be adopted (Callahan and Linder 1992), thus incorporating biological information into the assessment process.

FIELD TOXICITY TESTS USING EARTHWORMS Field tests can use replicated experimental field plots or use continuous monitoring in the field, but they are expensive, labor intensive, and time consuming because the most accurate method for quantitative sampling of earthworms is still hand sorting. Field tests have obvious advantages, such as realism, because they are done under natural climatic conditions. Moreover, they can take several earthworm species into account because different species are exposed, and a broad database is obtained. The logical sequence of events would be to undertake these longer-term, labor-intensive tests only after the potential toxic effects of a chemical have been established reliably from both acute and chronic toxicity testing (first and second tier). The aim is to determine whether an ecologically significant earthworm population change will occur as a result of exposure to the chemical, taking full cognizance that earthworm populations can fluctuate enormously because of seasonal changes or cropping. The importance of a sensitive and carefully designed test is obvious. Test designs were proposed by Edwards (1992, 1998, 2002), H. Kula (1992), and Lofs (1992). The literature showed that the results of most field experiments are extremely variable and do not allow for meaningful comparisons because of poor design and inadequate replications and controls. The distribution of chemicals in soils is also extremely variable (Beyer 1992). Natural, seasonal changes in environmental factors could also cause large fluctuations in earthworm population densities, thereby masking possible effects of toxicants during certain periods. Field tests can also provide valuable data on the indirect effects of chemicals on earthworms through effects on food supply and soil cover. The limitations of field tests for assessing toxicity of chemicals to earthworms were discussed fully by Edwards (1992, 1998), who proposed a standardized field test design along the following lines: Site variables: At least six species of common earthworms with at least 100 earthworms per square meter Preferably a loam-based soil type No history of chemical use for at least 5 years Treatment variables: Clearly defined chemical with known physicochemical properties such as solubility, water/lipid partition coefficient, and volatility The highest recommended dose © 2004 by CRC Press LLC

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A minimum of four replicates, also for the control Plot size at least between 5 and 10 m2 Defined cropping Timing of the experiment when earthworms are most active, with a duration of at least 6 months, but preferably longer for persistent chemicals Use of toxic standards such as benomyl or carbaryl Determination of residues in earthworms and soil to assess bioconcentration factor The persistence of the chemical and the frequency of application are important considerations in designing field tests. Because the accurate assessment of earthworm populations is extremely difficult (Edwards 1991), a “worst-case approach” has been suggested (Lofs 1992) to determine a biologically significant effect. However, the necessity for valid controls is obvious and was discussed fully by Edwards (1998). For sampling, a combination of hand sorting and dilute formalin extraction could be used. Earthworms in the top 5 to 10 cm of soil are hand sorted, after which the quadrat is treated with dilute formalin. This method recovers both the species that live close to the soil surface (endogeic) as well as those that penetrate deep into the soil (anecic). Until more data can be obtained from well-designed and standardized field experiments, direct comparison of field data with data from standardized laboratory experiments must be done with care. To obtain a balanced judgment on the potential earthworm toxicity and environmental hazard of a chemical, a suitable databank of results from both laboratory and field experiments is needed. Heimbach (1992) has obtained promising results in such a comparative study. He obtained a good correlation (r = 0.86) between LC50 values from the artificial soil test and a standardized field test based on a test design by Edwards and Brown (1982). There is a growing body of evidence that LC50 tests in the laboratory could still be used, in spite of their known shortcomings, to predict field results by using a compensation factor. However, further evidence is needed before the validity of such an approach can be generally accepted. Because of their demands on labor and costs, complete detailed field tests are rare and therefore remain difficult to interpret. The time for recovery of earthworm populations after exposure is an important aspect of toxicity in field experiments (Sheppard et al. 1998). What constitutes a significant percentage negative change in populations that could affect their functioning? This question still remains unanswered and requires further study and discussion. It is generally accepted that TMEs (Morgan and Knacker 1994) and microcosms (Burrows and Edwards 2002) may be useful tools to bridge the gap between single-species laboratory tests and the field. Further research is needed to explore the potential usefulness of TMEs on the risk assessment of chemicals for earthworms.

RESIDUES IN EARTHWORMS AND THEIR ROLE AS BIOMONITORS AND BIOINDICATORS The use of earthworms as biomonitors and bioindicators may be accomplished in different ways. The concentration of a chemical in earthworms is sometimes a useful indicator of whether the chemical is near toxic levels in the environment. It is very important to analyze both the soil and the earthworms during field experiments to determine bioconcentration factors. The proximity of earthworms to the soil contaminants makes them useful monitoring organisms for the soil environment. A classic example of the use of earthworms as biological monitors was presented by Ma (1987); he showed that amounts of heavy metals in moles were correlated closely with concentrations in earthworms but not with the concentrations in soils. A study by Reinecke et al. (2000) obtained similar results for the transference of lead in the food chain from soil to earthworms to shrews. The presence of toxic amounts of a chemical in earthworms poses a serious risk of secondary © 2004 by CRC Press LLC

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poisoning of vertebrate predators. It is therefore important not only to know what the earthworm body burdens of any particular toxic chemicals are, but also to understand the various pathways of metabolism and detoxification. The use of organisms as indicators of toxicity implies that the organisms can indicate the presence or absence of a particular chemical or environmental factor or condition. The presence or absence of a particular species can tell us about the environment. Thus, the use of earthworms as indicators of environmental contamination is a topic of considerable interest but of limited practicality. Their uses as such indicators are limited because a variety of environmental influences could be responsible for the observed condition of the earthworm population, making the establishment of causal relationships between the earthworm population densities and the degree of pollution very difficult. The term indicator has also been used loosely to refer to biological monitors. Biological monitors (over and above their possible role as indicators) are usually available in abundance throughout areas of study (Martin and Coughtrey 1982). They respond to changes in the degree of pollution and may retain the pollutant progressively during the exposure period. Earthworms have been used extensively in environmental monitoring, especially as biological monitors of heavy metal and organochlorine insecticide pollution because they bioconcentrate these chemicals. A clear relationship has been demonstrated between the concentrations of some metals in earthworms and those in the surrounding soils. Concentrations of lead and cadmium in earthworms are related closely to those in soil, making earthworms good monitors for at least these two metals; data on others are mostly lacking or do not show such a clear relationship. The interpretation of results, however, should take interspecific variations into account as well as soil characteristics and other physical factors. The use of earthworms as biological monitors to determine the accumulated concentrations of pollutants in their tissue as an integrative estimate of the degrees of pollution is an attractive idea, but in practice it is fraught with difficulties (Morgan et al. 1992). Closely related earthworm species differ in the ways that they accumulate pollutants, and factors such as age, size, season, and diet influence rates of accumulation. Furthermore, the possibility of genetic adaptation to selection pressures complicates matters further. The use of slope-intercept plots (Morgan et al. 1992) can, under certain circumstances, provide insight into the bioavailability of pollutants. To use earthworms effectively as biomonitoring tools, it is important to assess the body burdens of any particular toxic chemical and to have data on the various pathways of concentration metabolism and detoxification. Chemical analyses over a wide range of doses and chemicals are an expensive undertaking, but when the vertebrate toxicity of a chemical is relatively high, such residue analyses should be undertaken, especially if the long-term persistence and the potential for bioaccumulation of a pollutant is high. Measurement of levels of pollutants in organisms indicates how much is present at a particular moment in time but does not indicate dynamic fluxes or rates of metabolic degradation of the pollutant. Considerable research has been undertaken on the uptake of trace contaminants into earthworms, but no standard protocols exist for measuring bioaccumulation in terrestrial ecosystems (Phillips 1993). More research along the lines of the modeling done by Connell and Markwell (1990) for bioaccumulation of lipophilic compounds into earthworms is needed.

RISK ASSESSMENT USING EARTHWORMS The objective of ecotoxicological risk assessment is to use all available toxicological and ecological information to estimate the possibility and probability that some undesired ecological event will occur (Wilson and Crouch 1987). These events or ecological end points (as opposed to toxicity end points) are not confined to specific taxa, are clouded with uncertainties, and require an ecosystem approach. The final estimated effect at the ecosystem level is usually expressed as a probability, and the identification of critical ecological end points is of prime importance (Bartell © 2004 by CRC Press LLC

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1 0.9 0.8

Species c

0.7

Species b

0.6 0.5 0.4 0.3 0.2

PAF

0.1 0 –3

–2

–1

Species A EQC

HC5

0

1 2 3 4 5 Log Concentration (mg/l–1) LC50/EC50/NOEC for dif. species SSD

FIGURE 16.2 The species sensitivity distribution (SSD). The dots show the data for the different species. Forward use (from the log concentration to the y-axis) yields the potentially affected fraction (PAF) as defined by Van Straalen and Van Leeuwen (2002). Inverse use (from the y-axis to the x-axis) provides the environmental quality criterion (EQC). In this case, it is shown as the hazardous concentration (HC) for 5% of the species. (Modified after Posthuma et al. 2002.)

et al. 1992). Risk assessment of chemicals is normally aimed at the protection of human health or at the protection of ecosystems (and organismal biodiversity) as such. The interrelationships of these two aims are obvious. Discharge limits and cleanup values for environmental protection are based on how much of a contaminant can be tolerated by the more highly sensitive species exposed to it. By looking at the distribution of sensitivities (Figure 16.2) of all tested species to particular chemicals and selecting a value that includes 95% of the variability, most species can be protected for most of the time (Van Straalen 2001). The use of distribution-based extrapolation methods has been criticized, but the outcome often seems to agree with data from field experiments. The use of earthworms in risk assessment programs is twofold. Apart from protecting earthworms as beneficial organisms, earthworms are used to obtain information on environmental quality and to ensure environmental safety in general. The overall aim is to derive “acceptable” concentrations of particular toxicants in soil. The earthworm is therefore used as a sensitive indicator or “responder.” The consensus seems to be that, although there is no single species of earthworm sensitive to all kinds of chemicals, the species E. fetida can still fulfill this role to some extent. However, extrapolating directly from earthworms to other species is seldom useful. Species from different taxa vary greatly in their response to toxicants. Van Straalen and Van Gestel (1993) approached this problem by applying a safety factor to data obtained with indicator species to ensure the protection of “all” species in an ecosystem. This safety factor is usually based on the variability of the toxicity data. Van Straalen and Denneman (1989) followed Kooijman (1987) in developing a new approach for estimating a benchmark concentration HCp (hazardous concentration for a selected percentage p). These statistical methods apply only to toxicity data for single species and provide a useful way of deriving environmental quality criteria (EQC; Figure 16.2). Application of such a model to earthworm toxicity data was attempted by Van Straalen (1990). A comparison between NOEC values of five chemicals in soil to E. fetida and other soil invertebrates suggested the use of a safety factor greater than 20, and preferably 100, to protect more sensitive species. The use of such types of extrapolation methods was discussed by Forbes and Forbes (1994) and Van Straalen (2001). Van Straalen (2001) admitted that extrapolation factors, safety factors, and uncertainty factors have limited scientific basis, lack flexibility, and often give unrealistically low © 2004 by CRC Press LLC

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protection values. He therefore preferred a distribution-based sensitivity threshold approach (Posthuma et al. 2002) as scientifically more rigorous. This approach assumes that the variability in sensitivities to toxicants among species is a function of inherent differences in species physiology and environmental exposure conditions. Therefore, when considering the role of earthworms in the context of risk assessment, it should be realized that they can serve only as a single potential species in the range of other useful species. However, once the position of a species in the distribution function (Figure 16.2) is known, its value as a test or indicator species may be derived. Species a in Figure 16.2 is clearly more sensitive, and it may be more useful rather than species b or c to set environmental quality criteria, provided that the same situations are under consideration. The same holds true for the opposite approach of risk assessment (to determine the potentially affected fraction, PAF). Risks of ecological damage also depend on what happens to a toxicant after it reaches the soil. Toxicity may be a misleading term, especially if the chemical is one that will degrade quickly or become detoxified. The extent of the initial environmental damage may be such that speedy recovery from the impact can minimize damage within a short period of time. The extent and duration of toxic effects are therefore important characteristics for categorizing chemicals (Kokta and Rothert 1992), but criteria by which recovery of organismal populations and communities can be measured are still needed. Regulatory schemes are operated in many countries to ensure the environmental safety of chemicals, especially pesticides. These schemes rely on experimental data obtained under both laboratory and field conditions. In some schemes, it is a requirement that the chemicals must also be examined for potential effects on earthworms (Greig-Smith 1992). The International Workshop on Ecotoxicology of Earthworms held in Sheffield in 1991 recommended that a general risk assessment should be carried out for each type of land on which a product is used and should be flexible. Consistent criteria for placing chemicals into categories of high, intermediate, low, and negligible risk to earthworms should be agreed on. Products should be labeled to indicate the degree of risk for earthworms. Most risk-assessment studies using earthworms have been on agricultural land or relatively small areas of contaminated land. In case of more diffuse pollution or large-scale use of chemicals, it may, however, be more appropriate to focus on larger spatial scales. This requires alternative approaches, for instance, including landscape, island, or metapopulation theories; definition of minimum viable population sizes; or assessment of rates of recovery and dispersal rates (Eijsackers, Chapter 17, this volume). A risk assessment scheme adopted by Kokta and Rothert (1992) still provides a useful procedure applicable to earthworms, but ecological risk assessment in a broader sense is concerned not only with effects of toxicants on earthworms, but also with endeavors to extrapolate their effects to complex ecosystems. Ecological risk assessment essentially requires an interdisciplinary approach, and earthworm researchers should remain aware of useful developments and approaches in other related ecotoxicological fields. The incorporation of biomarkers into risk assessment schemes (Figure 16.3) may contribute in the future toward a more scientific basis for decision making. Advances in expert systems and artificial intelligence capabilities will contribute in the future to the coordination of models and data. The use of mathematical models for forecasting ecological effects of chemicals (Bartell et al. 1992) is well established in the aquatic field but much less so for the soil environment. Bartell et al. (1992) considered that, although the original contention was that direct extrapolation of laboratory toxicity data to the field was ill advised because of the complexity of ecological systems, results from bioassays will continue to generate useful information because the effects measured in laboratory microcosms can be compared with results from mathematical models. Various simulation models have been proposed for the soil environment (Axelsen 1997; Van Wensem 1997). By applying these to earthworm communities, researchers in earthworm ecology and ecotoxicology can contribute to a better understanding of the basic functions of soil ecosystems and the effects of chemicals on them. This will contribute to the development of more useful risk assessment methodologies.

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Risk assessment

Hazard Identification

PNEC

PEC Risk characterization

Refined risk characterization

PEC/PNEC 1?

Current risk minimal (no further testing)

No Risk reduction measures

Can ratio PEC/PNEC be lowered by:

No

1. Refinements of information 2. Further testing? Yes

2. Bioaccumulation tests 3. Extended representation of trophic levels and/or Refined exposure assessment incorporating biomarkers and bioavailability levels and/or

Refined risk assessment

1. Chronic tests / field tests

Long-term monitoring (including use of biomarkers to assess changing bioavailability)

FIGURE 16.3 A risk assessment scheme indicating how biomarkers may in the future be incorporated in the assessment process. PEC = predicted environmental concentration; PNEL = predicted no effect concentration.

DECISIONS ON MANAGING ESTIMATED RISK OF CHEMICALS TO EARTHWORMS Once a regulatory scheme has been developed and it seems likely that earthworms are adversely affected by a chemical, the obvious next step is to protect earthworm populations. Most agricultural pesticides are used in soils in which earthworms may be affected. Because this is unavoidable, the second-best option would be to decide whether the effects will be “acceptably low.” Another possibility is to decide if any effects will be reversible within a reasonably short period of time and thus not affect the earthworm population too severely in the long term. Knowledge of earthworm life cycles and reproduction rates becomes relevant if population recovery is to be considered, but unfortunately knowledge of life cycles is available for relatively few species. Kokta and Rothert (1992) advocated that at least the pretreatment earthworm population level should be restored before the next application of the same chemical is undertaken, but this assumes that populations are at a steady level, which is usually not the case. Whatever the approach may be, making decisions will remain difficult even when a more scientific basis for decision making on pollutants is provided © 2004 by CRC Press LLC

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by researchers. This is because there are many conflicting interests requiring wise compromises. Good decisions on the protection of soil biodiversity and sustainable use of soils will be taken only if a thorough appreciation of the role of earthworms and their interactions with other organisms in soil exists.

CONCLUSIONS Earthworms play an important role as engineers in many soils, in which they contribute to the complex processes of organic matter decomposition and affect aeration, water transport, and soil structure. Their protection from polluting chemicals is therefore important. The use of earthworms for ecotoxicological evaluation has made considerable progress in the years since the first International Workshop on Earthworm Ecotoxicology was held in Sheffield in 1991. Two other international workshops, in the Netherlands in 1997 (Sheppard et al. 1998) and Denmark in 2001, made comprehensive recommendations to researchers and regulators. There is clearly a need to distinguish between the purposes and use of the various earthworm toxicity tests. Although E. fetida may be suitable for some toxicity tests and there are sufficient grounds to extrapolate such data to certain other species and soils, it can never serve as the sole representative species of the soil environment for the purpose of assessing ecological risks. Acute toxicity tests with earthworms are useful, but they are designed to identify very toxic chemicals that have immediate effects on earthworms and not to determine acceptable environmental concentrations. Such tests, however, can serve as a qualitative screen for detection of other toxicological effects and give a first estimate of the NOEC level for continuous exposure to the toxicant. Such tests do not necessarily absolve any chemical from further toxicity testing but do assist in practice in setting priorities for further testing of sublethal effects. The OECD standard artificial soil test was not developed to enable the direct extrapolation of toxicity results into field situations. Earthworm behavior is not often used in ecotoxicological studies because it is not fully understood. Some field studies reported that earthworms can migrate over large distances in response to environmental changes, although this is still open to debate; other studies demonstrated avoidance behavior of earthworms when exposed to highly contaminated soils. The Denmark workshop (2001) recommended that more attention be given to study earthworm behavior and interpret avoidance responses and surface migrations. The use of sensitive biomarkers in earthworm ecotoxicology represents a fairly new approach that may yield promising results if a suitable battery of biomarkers can be employed to predict toxic risks from pollutants at higher levels of organization. An important development is the application of modern molecular techniques to identify biomarkers in earthworms. Heavy metal responsive genetic indexes (Stürzenbaum et al. 1998) may in the future improve existing toxicity testing protocols. A number of molecular genetic techniques have already been utilized to identify those genes responsive to certain types of chemical exposure. The value of biomarkers in risk assessment has received increasing attention because of problems experienced when trying to relate environmental concentrations of toxicants with their bioavailable fraction. The bioavailability and toxicity of a chemical may differ considerably in laboratory tests compared with those observed in the field. Biomarkers, on the other hand, respond only to the dose entering the animal’s body, and the idea of using biomarkers for more accurate quantification of contaminant exposure is gaining support. To obtain a balanced judgment on the potential toxicity and environmental hazard of a chemical, a suitable databank of results from both laboratory and field experiments is needed. Because of their labor demands and cost, complete field tests are rare, and results remain difficult to interpret. The rates of recovery of earthworm populations, after exposure, are important in field tests. Microcosm studies can contribute to a more field-relevant understanding of toxic effects of chemicals on earthworms. The use of earthworms as biological monitors to determine accumulations of © 2004 by CRC Press LLC

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pollutants in their tissues as an integrative estimate of the degrees of pollution is an attractive idea fraught with practical difficulties. Closely related earthworm species differ in the way they accumulate pollutants, and factors such as age, size, season, and diet can influence rates of accumulation. To use earthworms effectively as biomonitoring tools, it is important to know their body burdens of any particular toxic chemical and to understand the various pathways of metabolism and detoxification. Ecological risk assessment essentially requires an interdisciplinary approach, and earthworm researchers should be aware of useful developments and approaches in other fields. The incorporation of earthworm biomarkers in risk assessment schemes for soils may in the future contribute to a more scientific basis in decision making.

REFERENCES Axelsen, J.A. 1997. A physiologically driven mathematical simulation model as a tool for extention of results from laboratory to ecosystem effects, in Ecological Risk Assessment of Contaminants in Soils, N.M. Van Straalen and H. Løkke, Eds., Chapman & Hall, London, pp. 233–250. Bartell, S.M., Gardener, R.H., and O’Neil, R.V. 1992. Biological Risk Estimation, Lewis Publishers, Boca Raton, FL. Basta, N. and Gradwohl, R. 2000. Estimation of Cd, Pb, and Zn bioavailability in smelter-contaminated soils by a sequential extraction procedure, J. Soil. Contam., 9, 149–164. Bembridge, J.D. 1998. Recommendations from the Second International Workshop on Earthworm Ecotoxicology, Amsterdam, Netherlands, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola, FL, pp. 389–398. Bengtsson, G. 1997. Dispersal, heterogeneity and resistance: challenging soil quality assessment, in Ecological Risk Assessment of Contaminants in Soil, N.M. Van Straalen and H. Løkke, Eds., Chapman & Hall, London, pp. 191–212. Bengtsson, G., Ek, H., and Rundgren, S. 1992. Evolutionary response of earthworms to long-term metal exposure, Oikos, 63, 289–297. Beyer, W.N. 1992. Relating results from earthworm toxicity tests to agricultural soil, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 109–115. Bouché, M.B. 1972. Lombriciens de France, Ecologie et Systimatique, Publ. Institut National de la Recherche Agronomique, Paris. Bouché, M.B. 1992. Earthworm species and ecotoxicological studies, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 20–35. Burrows, L. and Edwards, C.A. 2002 The use of integrated microcosms to predict the fate and effects of pesticides on soil ecosystems, Eur. J. Soil Biol., 38, 245–249. Callahan, C.A. and Linder, G. 1992 Assessment of contaminated soils using earthworm test procedures, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 187–196. Callahan, C.A., Shirazi, M.A., and Neuhauser, E.F. 1994. Comparative toxicity of chemicals to earthworms, Environ. Toxicol. Chem., 13, 291–298. Cluzeau, D., Lagarde, R., Texier, G., and Fayolle, L. 1992. Relevance of life-history parameters in earthworm ecotoxicology, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 225–229. Connell, D.W. and Markwell, R.D. 1990. Bio-accumulation in the soil to earthworm system, Chemosphere, 20, 91–100. Daily, G.C. 1997. Natures Services, Island Press, Washington, D.C. Edwards, C.A. 1983. Development of a Standardized Laboratory Method for Assessing the Toxicity of Chemical Substances to Earthworms, Report EUR 8714 EN, Environment and Quality of Life: Commission of the European Communities, Brussels, Belgium, 141 pp. Edwards, C.A. 1984. Report of the Second Stage of a Standardized Laboratory Method for Assessing the Toxicity of Chemical Substances to Earthworms, Report EUR 9360 EN, Environment and Quality of Life: Commission of the European Communities, Brussels, Belgium, 99 pp.

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Edwards, C.A. 1991. Methods for assessing populations of soil-inhabiting invertebrates, Agric. Ecosyst. Environ., 34, 145–176. Edwards, C.A. 1992. Testing the effects of chemicals on earthworms: the advantages and limitations of field tests, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 75–84. Edwards, C.A. 1998. Principles for the design of flexible earthworm field-toxicity experiments and interpretation of results, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola, FL, pp. 313–326. Edwards, C.A. 2002. Assessing the effects of environmental pollutants on soil organisms, communities, processes and ecosystems, Eur. J. Soil Biol., 38, 225–231. Edwards, C.A. and Bohlen, P.J. 1992. The effects of toxic chemicals on earthworms, Rev. Environ. Contam. Toxicol., 125, 23–99. Edwards, P.J. and Brown, S.M. 1982. Use of grassland plots to study the effects of pesticides on earthworms, Pedobiologia, 24, 145–150. Edwards, P.J. and Coulson, J.M. 1992. Choice of earthworm species for laboratory tests, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 36–43. Fitzpatrick, L.C., Goven, A.J., Venables, B.J., and Cooper, E.L. 1990. Earthworm immunoassays for evaluating biological effects of exposure to hazardous materials, in In situ Evaluation of Biological Hazards of Environmental Pollutants, S.S. Sandhu, Ed., Plenum Press, New York, pp. 119–129. Forbes, V.E. and Forbes, T.L. 1994. Ecotoxicology in Theory and Practice, Chapman & Hall, London. Gälli, R., Munz, C.D., and Scholtz, R. 1994. Evaluation and application of aquatic toxicity tests: use of the Microtox test for the prediction of toxicity based upon concentrations of contaminants in soil, Hydrobiology, 273, 179–189. Goats, G.C. and Edwards, C.A. 1982. The prediction of field toxicity of chemicals to earthworms by laboratory methods, in Earthworms in Waste and Environmental Management, C.A. Edwards and E.F. Nauhauser, Eds., SPB Academic Publishing, The Hague, the Netherlands, pp. 283–294. Goven, A.J., Venables, B.J., Fitzpatrick, L.C., and Cooper, E.L. 1988. An invertebrate model for analyzing effects of environmental xenobiotics on immunity, Clin. Ecol., 4, 150–154. Goven, A.J., S.C. Chen, L.C. Fitzpatrick, and J. Venables, 1994. Lysozyme activity in earthworms (Lumbricus terrestris coelomic fluid and coelomccytes. Enzyme assay for immunotoxicity of xenobiotics, Environ. Toxicol. Chem., 13, 607–613. Greig-Smith, P.W. 1992. Risk assessment approaches in the U.K. for the side effects of pesticides on earthworms, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 159–168. Heimbach, F. 1984. Correlations between three methods for determining the toxicity of chemicals to earthworms, Pest. Sci., 15, 605–611. Heimbach, F. 1992. Effects of pesticides on earthworms populations: comparison of results from laboratory and field tests, in Ecotoxicology of Earthworms, P.W. Grieg-Smith et al., Eds., Intercept, Hants, U.K., pp. 100–108. Helling, B., Reinecke, S.A., and Reinecke, A.J. 2000. Effects of the fungicide copper oxychloride on the growth and reproduction of Eisenia fetida (Oligochaeta), Ecotox. Environ. Saf., 46, 108–116. Hund-Rinke, K. and Wiechering, H. 2001. Earthworm avoidance test for soil assessments. An alternative for acute and reproduction tests, J. Soils Sediments, 1, 15–20. Kammenga, J.E., Dallinger, R., Donker, M.H., Kohler, H.R., Simonsen, V., Triebskorn, R., and Weeks, J.M. 2000. Biomarkers in terrestrial invertebrates for ecotoxicological soil risk assessment, Rev. Environ. Contam. Toxicol., 164, 93–147. Kokta, C. 1992. Measuring effects of chemicals in the laboratory: effect criteria and endpoints, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 55–62. Kokta, C. and Rothert, H. 1992. Hazard and risk assessment for effects of pesticides on earthworms — the approach in the Federal Republic of Germany, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 169–176. Kooijman, S.A.L.M. 1987. A safety factor for LC50 values allowing for differences in sensitivity among species, Water Res., 17, 527–538.

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Kula, C. 1998. Endpoints in laboratory testing with earthworms: experience with regard to regulatory decisions for plant protection products, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola, FL, pp. 3–14. Kula, H. 1992. Measuring effects of pesticides on earthworms in the field: test design and sampling methods, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 90–99. Lofs, A. 1992. Measuring effects of pesticides on earthworms in the field: effect criteria and endpoints, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 85–89. Lofs-Holmin, A. 1980. Measuring growth of earthworms as a method of testing sublethal toxicity of pesticides, Swed. J. Agric. Res., 10, 25–33. Ma, W.-C. 1984. Sublethal toxic effects of copper on growth, reproduction and litter breakdown activity in the earthworm Lumbricus rubellus, with observations on the influence of temperature and soil pH, Environ. Pollut. Ser. A, 33, 207–219. Ma, W.-C. 1987. Heavy metal accumulation in the mole, Talpa europa, and earthworms as an indicator of metal bioavailability in terrestrial environments, Bull. Environ. Contam. Toxicol., 39, 933–938. Maboeta, M.S., Reinecke, S.A., and Reinecke, A.J. 2002. The relation between lysosomal biomarker and population response in a field population of Microchaetus sp. (Oligochaeta) exposed to the fungicide copper oxychloride, Ecotox. Environ. Saf., 52, 280–287. Marino, F. and Morgan, A.J. 1998. Immunohistochemical detection of heat shock protein expression in stressed earthworms, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola, FL, pp. 199–224. Martin, M.H. and Coughtrey, P.J. 1982. Biological Monitoring of Heavy Metal Pollution. Land and Air, Applied Science Publishers, London. Martin, N.A. 1986. Toxicity of pesticides to Allolobophora caliginosa, N.Z. J. Agric. Res., 29, 699–706. Morgan, E. and Knacker, T. 1994. The role of laboratory terrestrial model ecosystems in the testing of potentially harmful substances, Ecotoxicology, 3, 213–233. Morgan, J.E., Morgan, A.J., and Corp, N. 1992. Assessing soil metal pollution with earthworms: indices derived from regression analyses, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 233–237. Neuhauser, E.F., Malecki, M.R., and Adnatra, M. 1986. Comparative toxicity of ten organic chemicals to four earthworm species, Comp. Biochem. Physiol., 83C, 197–200. Neuhauser, E.F., Malecki, M.R., and Loehr, R.C. 1983. Methods using earthworms for the evaluation of potentially toxic materials in soils, in Hazardous and Industrial Solid Waste Trading, Vol. 2, STP 805, R.A. Conway and W.P. Gulledge, Eds., American Society for Testing and Materials, Philadelphia, PA, pp. 313–320. Organization for Economic Cooperation and Development. 1984. Guideline for Testing of Chemicals 207. Earthworm, Acute Toxicity Tests, Organization for Economic Cooperation and Development, Paris. Phillips, D.J.H. 1993. Bioaccumulation, in Handbook of Ecotoxicology, Vol. 1, P. Calow, Ed., Blackwell Scientific, London, pp. 378–396. Posthuma, L., Traas, T., and Suter, G.W. 2002. General introduction to species sensitivity distributions, in Species Sensitivity Distributions in Ecotoxicology, L. Posthuma, G.W. Suter II, and T.P. Traas, Eds., Lewis Publishers, Boca Raton, FL, pp. 3–10. Reinecke, A.J. 1992. A review of ecotoxiological test methods using earthworms, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 7–19. Reinecke, A.J., Maboeta, M.S., Vermeulen, L.A., and Reinecke, S.A. 2002. Assessment of lead nitrate and Mancozeb toxicity in earthworms using the avoidance response, Bull. Environ. Toxicol. Contam., 68, 779–786. Reinecke, A.J. and Reinecke, S.A. 1998. The use of earthworms in ecotoxicological evaluation and risk assessment: new approaches, in Earthworm Ecology, C.A. Edwards, Ed., St. Lucie Press, Boca Raton, FL, pp. 237–293. Reinecke, A.J., Reinecke, S.A., Musibono, D.E., and Chapman, A. 2000. The transfer of lead (Pb) from earthworms to shrews (Myosorex varius), Arch. Environ. Contam. Toxicol., 39, 392–397.

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Reinecke, S.A., Helling, B., and Reinecke, A.J. 2002. Lysosomal response of earthworm (Eisenia fetida) coelomocytes to the fungicide copper oxychloride and relation to life-cycle parameters, Environ. Toxicol. Chem., 21, 1026–1031. Reinecke, S.A., Prinsloo, M.W., and Reinecke, A.J. 1999. Resistance of Eisenia fetida (Oligochaeta) to cadmium after long-term exposure, Ecotox. Environ. Saf., 42, 75–80. Reinecke, S.A. and Reinecke, A.J. 1996. The influence of heavy metals on the growth and reproduction of the compost worm Eisenia fetida (Oligochaeta), Pedobiologia, 40, 439–448. Reinecke, S.A. and Reinecke, A.J. 1999. Lysosomal response of earthworm coelomocytes induced by longterm experimental exposure to heavy metals, Pedobiologia, 43, 585–593. Reinecke, S.A. and Reinecke, A.J. 2003. The comet assay as biomarker of heavy metal genotoxicity in earthworms, Arch. Environ. Contam Toxicol., in press. Roberts, B.L. and Dorough, W.H. 1983. Relative toxicity of chemicals to the earthworm Eisenia foetida, Environ. Toxicol. Chem., 3, 66–78. Rodriquez, J., Venables, B.J., Fitzpatrick, L.C., Goven, A.J., and Cooper, E.L. 1989. Suppression of secretory rosette formation by PCBs in Lumbricus terrestris: an earthworm immunoassay for humoral immunotoxicity of xenobiotics, Environ. Toxicol. Chem., 8, 1201–1207. Scott-Fordsmand, J.J. and Weeks, J.M. 1998. Review of selected biomarkers in earthworms, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola, FL, pp. 173–189. Sheppard, S.C., Evenden, W.G., and Cornwell, T.C. 1998. Depuration and uptake kinetics of I, Cs, Mn, Zn and Cd by the earthworm (Lumbricus terrestris) in radiotracer-spiked litter, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola, FL, pp. 151–164. Shirazi, M.A. and Lowrie, L.N. 1988. Comparative toxicity based on similar asymptotic endpoints, Arch. Environ. Contam. Toxicol., 17, 273–280. Slijkerman, D.M.E., Van Gestel, C.A.M., and Van Straalen, N.M. 2000. Conceptueel kader voor de afleiding van ecotoxicologische risicogrenzen voor essentiele metalen, Rep. D00020, Institute of Ecological Science, Vrije Universiteit, Amsterdam, the Netherlands. Spurgeon, D.J. and Hopkin, S.P. 1995. Extrapolation of the laboratory-based OECD earthworm toxicity test to metal-contaminated field sites, Ecotoxicology, 4, 190–205. Spurgeon, D.J. and Hopkin, S.P. 1999. Tolerance of zinc in populations of the earthworm Lumbricus rubellus from uncontaminated and metal-contaminated ecosystems, Arch. Environ. Contam. Toxicol., 37, 332–337. Spurgeon, D.J. and Hopkin, S.P. 2000. The development of genetically inherited resistance to zinc in laboratoryselected generations of the earthworm Eisenia fetida, Environ. Pollut., 109, 193–201. Spurgeon, D.J., Svendsen, C., Rimmer, V.R., Hopkin, S.P., and Weeks, J.M. 2000. Relative sensitivity of lifecycle and biomarker responses in four earthworm species exposed to zinc, Environ. Toxicol. Chem., 19, 1800–1808. Spurgeon, D.J. and Weeks, J.M. 1998. Evaluation of factors influencing results from laboratory toxicity tests with earthworms, in Advances in Earthworm Ecotoxicology, S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., SETAC Press, Pensacola FL, pp. 15–25. Stürzenbaum, S.R., Kille, P., and Morgan, A.J. 1998. Identification of heavy metal induced changes in the expression patterns of the translationally controlled tumour protein (TCTP) in the earthworm Lumbricus rubellus, Biochem. Biophys. Acta, 1398, 294–304. Svendsen, C. and Weeks, J.M. 1997. Relevance and applicability of a simple earthworm biomarker of copper exposure: I. Links to ecological effects in a laboratory study with Eisenia andrei, Ecotox. Environ. Saf., 26, 72–79. Van Gestel, C.A.M. 1992. The influence of soil characteristics on the toxicity of chemicals for earthworms: a review, in Ecotoxicology of Earthworms, P.W. Greig-Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Intercept, Hants, U.K., pp. 44–54. Van Gestel, C.A.M. 1997. Scientific basis for extrapolating from soil ecotoxicity tests to field conditions and the use of bioassays, in Ecological Risk Assessment of Contaminants in Soils, N.M. Van Straalen and H. Løkke, Eds., Chapman & Hall, London, pp. 25–50. Van Gestel, C.A.M. and Ma, W.-C. 1990. An approach to quantitative structure-activity relationships (QSAR’s) in earthworm toxicity studies, Chemosphere, 21, 1023–1033.

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in Environmental 17 Earthworms Research Herman Eijsackers Alterra, Wageningen University and Research Centre and Institute of Ecological Sciences, Vrije Universiteit, Amsterdam, The Netherlands

CONTENTS Introduction ....................................................................................................................................321 Current Interest in Earthworm Ecotoxicology ..............................................................................321 Toxicokinetic Behavior by Earthworms (Availability, Uptake, Elimination, Bioaccumulation) .......................................................................................................................323 Testing Contaminant Toxicity with Earthworms...........................................................................327 Field Studies...................................................................................................................................330 Pesticide and Heavy Metal Toxicity Studies..........................................................................330 Land Improvement and Earthworms as Bioengineers ...........................................................332 Effects of Toxicants at Food Chain and Ecosystem Levels...................................................334 Risk Assessment Based on Earthworm Toxicology......................................................................336 Acknowledgment............................................................................................................................337 References ......................................................................................................................................337

INTRODUCTION Earthworms are increasingly widely used in environmental research. This chapter reviews the progress made in earthworm ecotoxicology over the past 20 years. It is related especially to the assessment of soil quality in relation to soil degradation because of physical and chemical stressors and to recovery processes, bioremediation, and restoration practices. After a brief introduction summarizing the attention given to earthworm ecotoxicological research during recent years, this overview starts with a discussion of uptake and elimination processes (bioaccumulation and toxicokinetics), followed by a review of different aspects of laboratory testing, field studies, and sampling programs (including land restoration). Finally, food chain transfer of pollutants is described as a process of species interaction in combination with integrated studies at a higher biological integration level (community, ecosystem).

CURRENT INTEREST IN EARTHWORM ECOTOXICOLOGY What makes earthworms a favorite “model animal” in ecotoxicological research? Given the general attitude of the public toward earthworms, those slimy, squirming creatures, it seems rather remarkable at first sight that earthworms are so popular. However, this same attitude indicates that earthworms are easily recognized. Their usefulness in nature is readily acknowledged as food for birds and bait for fish. Moreover, they have a well-defined and positive role in the formation of soil, which can be easily observed. Finally, from a technical-experimental point of view, they are easy to breed, culture, and handle.

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In the past few decades, some specifically soil ecotoxicity-oriented research programs have been developed or initiated. At the national level, in the Netherlands there was the Netherlands Integrated Soil Research Program (NISRP), and there is the current System-Oriented Ecotoxicological Research Program. In Sweden, there was the MATS program (Soil Biological Variables in Environmental Hazard Assessment), succeeded by ISA (Integrated Soil Analysis). In a general European context, there was the network SERAS (Soil Ecotoxicity Risk Assessment Systems) and the European Union (E.U.) SECOFA-SE (development, improvement, and standardization of test systems for assessing sublethal effects of chemicals on fauna in the soil ecosystem) (Løkke and Van Gestel 1998). All these programs included earthworm testing. The present interest in ecotoxicological research with earthworms, when reviewed by screening the programs of the annual Conferences of the Society for Environmental Toxicity and Chemistry (SETAC), reveals a somewhat different picture. In the SETAC conference in Sheffield, U.K. (1991), 7 papers on earthworm toxicity were presented of a total of 142 presentations (5%) and were combined with other soil ecotoxicological aspects (Donkers et al. 1994). Later SETAC conferences had lower percentages of earthworm papers: Lisbon, Portugal (1993), had 2%; and Houston, TX (1994), had only 5 of over 1500 contributions. During the most recent SETAC-Europe Symposium in Vienna (2002), of approximately 1400 presentations were 10 papers on earthworms. By contrast, the Earthworm Ecotoxicology Workshop that was organized in Århus, Denmark, during the summer of 2001 brought together many world specialists on earthworm toxicity from the research, consultancy, industry, and government fields. For earthworm research in general, there is a world audience; at the Seventh International Symposium on Earthworm Ecology in Cardiff, Wales, in 2002 was the latest example of bringing together some 200 earthworm specialists, with earthworm ecotoxicology well represented. Thus, earthworm research is a permanent element in environmental assessment studies. Much effort has been put into the development of earthworm tests; the European Union, the Organization for Economic Cooperation and Development (OECD), and the International Organization for Standardization (ISO) have all organized a number of international ring tests and standardization studies. In a study for OECD on the selection of a set of laboratory ecotoxicity tests, Van Gestel et al. (1997) observed that the earthworm tests scored at a relatively low level. It was rated 6th (of 44 tests in total) for the OECD tests, 13th for the tests of the U.S. Environmental Protection Agency, and 18th for the ISO test. Standardization was the main factor of criticism, but this was underrated in comparison with aquatic tests. Consequently, new tests developed by the International Organization for Biological Control and the National Institute for Human Health and Environmental Protection, which have a higher ecological relevance than the artificial soil/Eisenia fetida test, also scored at a very low level (38th, 39th, 40th). Hence, there is still a lot to do, as there is with respect to earthworm toxicity testing both in the laboratory and the field. When the recommendations of the three International Workshops on Earthworm Ecotoxicology — held in Sheffield in 1991 (Greig-Smith et al. 1992), Amsterdam in 1997 (Sheppard et al. 1998), and Århus in 2001 (Van Gestel and Weeks 2004) — are compared, a number of recommendations came up each time: • • • •

The need to develop proper end-point effect criteria The development of specialized earthworm tests, especially for effects on juvenile growth The necessity to define the effects of fluctuating environments related to exposure in a dynamic world The relevance of including earthworm behavioral ecology in testing, covering behavioral responses to contaminants, as well as mobility and dispersal capabilities in general

The extensive recommendations of the International Earthworm Ecotoxicology Workshop in Århus (Van Gestel and Weeks 2004) were discussed further, confirmed by a number of world specialists in earthworm toxicity research in Cardiff, and are summarized as follows:

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•

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Toxicant bioavailability is a leading aspect and should always be measured, preferably working with field soils (instead of artificial soils). Soils should be characterized fully by texture (percentages of sand, silt, and clay), pH, percentage organic carbon content, water-holding capacity, and cation exchange capacity. Chemical extraction techniques used to measure bioavailability, including the promising new biomimetic techniques (diffuse gradient in thin films (DGT), semi-permeable membrane devices (SPMD), solid phase micro extraction (SPME), should be verified properly and validated. Toxicant bioavailability models should include temporal and spatial variations, as well as internal distribution phenomena in earthworms. Life cycle parameters as suitable end points have to be combined on one hand with mechanistic biomarker responses and on the other hand with recorded population changes. Bioassays are important tools in both laboratory and field and should include the effects of factors like drought, temperature, and acidity. Also, Terrestrial Model Ecosystems (Knacker et al. 2004) could be useful in this perspective. For regulatory purposes, there are a number of additional promising approaches; these include behavioral aspects, critical body weight loss, and juvenile growth rates. A proper definition of recovery from toxicants would be most helpful in assessing the results of field tests, especially in the longer term. This also implicates an extended duration for field tests, depending, for instance, on negative responses caused by experimental treatments. Moreover, site-relevant species should always be included. A proper assessment of the impacts of pesticides applied at a broader scale or the implications of mixed contamination cannot be restricted to the field experiments but should be extended to an adjacent set of field plots, thereby including landscape ecological elements in the toxicity assessment.

TOXICOKINETIC BEHAVIOR BY EARTHWORMS (AVAILABILITY, UPTAKE, ELIMINATION, BIOACCUMULATION) Ecotoxicological work can be separated into a number of phases (Eijsackers, 1994). In the first phase, the main question is how the organism is exposed to a contaminant. This comprises three steps: how the contaminant is chemically available in the soil, by what routes and mechanisms earthworms take up the contaminant, and how it is internally processed in the earthworm (whether is it broken down, excreted, or stored). This can be summarized under headings of environmental chemistry and toxicokinetics, which are discussed in this section. The second phase deals with questions such as the following: What are the effects of a chemical on individual earthworms? What are the implications for earthworm populations? What impacts does this have on soil ecosystems and food chains? With respect to the bioavailability of toxicants, the three major steps, each with a different setting and characteristics, are (1) chemical availability; (2) biological uptake of the chemical into earthworms (this also includes exoenzymes of microorganisms, which are active outside the organisms and can “digest” the contaminant there, which can be interpreted as an external uptake-andprocessing mechanism of the compound); and (3) internal transport and processing in the earthworm body (Figure 17.1). Much progress has been made in the development and validation of the soil pore-water approach, which suggests that most contaminants are taken up from the pore water directly surrounding the earthworm through its skin. For the effects of heavy metals on earthworms, much work has already been done on factors influencing the uptake, internal distribution, and elimination of these toxicants (Beeby 1993). The characteristics of the metal, the soil, and the earthworm are of prime importance and play significant

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Soil Equilibria

Equilibration of Organisms with Their Environment

Redistribution within Organisms

Environmental Availability

Environmental Bioavailability

Toxicological Bioavailability

pH, %Clay, etc. Liquid Phase

C(eq)

K1

[Me] mmol/kg

Solid Phase

Circulation Target(s) Storage Organs(s)

C(0) time

KP

Toxicokinetic Parameters

Toxicodynamic Parameters

FIGURE 17.1 The three major processes in availability: (1) chemical availability of the compound bound to mineral or organic soil material or present in the soil pore water; (2) biological uptake processes into an organism (including exoenzymes of microorganisms); and (3) internal transport and processing (storage, degradation, or effective entry in a target organ). (Adapted from Posthuma personal communication.)

Class A

Borderline

Class B

Period 2

Li

Be

3

Na

Mg

4

K

Ca

5

Rb

Sr

6

Cs

Cr

Mn Fe Co Ni

Cu

Zn

Cd

Cu

Sn

Ag

Pb

Au

Hg

Pb

FIGURE 17.2 ClassiÞcation of metals according to biological working mechanisms (arranged according to periodic table). Class A: Macronutrients forming O bonds; class B: forming N and S bonds. (From Nieboer and Richardson 1980.)

roles in the overall effects. Nieboer and Richardson (1980) distinguished two groups of metals: one that tends to bind metals to oxygen sites in biological ligands and act as macronutrients and a second that forms strong bonds with N or S and exerts toxicity on earthworms by affecting protein structure and enzyme function (Figure 17.2). As a consequence, some heavy metals are regulated dynamically after uptake by earthworms (e.g., Zn) whereas others such as Cd are accumulated continuously and become tightly bound and immobilized. This accumulation is reinforced because of the low toxicity of Cd to earthworms. This type of continuous accumulation of heavy metals that remain in earthworms after a transfer to clean soil was observed for Cd by Van Gestel et al. (1993a). The metals in the group, which act as macronutrients, however, may become toxic under certain doses, conditions, and end points. In addition, Fischer and Molnar (1997) observed that Na, K, and Ba all affected earthworm cocoon production negatively, possibly by inßuencing soil/cocoon osmotic relations. Toxic metals can interact with trace metals as well as metal macronutrients. Substitution between these forms is an important process in this context, as demonstrated for Ca-Cd and CaPb interactions by Ma (1993). In general, invertebrates with a high Ca requirement actively accumulate heavy metals. Janssen et al. (1998) showed that potassium affects the uptake of 134Cs but not of Na, whereas Morgan and O’Reilly (2002) and Oste et al. (2002) reported a competitive inhibition of heavy metal accumulation by Mn. © 2004 by CRC Press LLC

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In most contaminated sites, a mixture of various contaminants will be present, so their combined toxicity to earthworms must also be assessed. Posthuma et al. (1993) tested a number of combinations of Cd, Zn, and Cu and observed both synergism and antagonism in earthworm toxicity. Antagonistic responses to chemicals mean that just adding up toxicities of separately tested metals can often overestimate the overall impact of the combined heavy metals under field conditions. Of the different soil factors affecting the toxicity of chemicals to earthworms (soil acidity, pH, cation exchange capacity, percentage organic matter), pH plays the most important role. It affects the health of different earthworm species directly and therefore their sensitivity to adverse impacts. An example is the feeding rate and metabolic turnover of Pb uptake by earthworms reported by Bengtsson and Rundgren (1992). It also influences the physicochemical binding and the biological uptake processes of toxicants by earthworms, as shown by Marinussen and Van der Zee (1997), and so indirectly influences the survival and functioning of various earthworm species. In an extensive study, Janssen et al. (1997a,b) measured the chemical availability kp and biological concentration factor (BCF) of toxicants to earthworms in different field soil types for four heavy metals (Cd, Zn, Pb, and Cu). They also measured a number of soil characteristics and correlated both the chemical availability and the BCF to earthworms with these soil factors by a multiple correlation analysis (Table 17.1). This impact of soil type on bioaccumulation in combination with different background levels of toxicants in various soil types could result in a metalloregion-specified derivation of environmental quality standards, as suggested by deGroot et al. (2002). Marinussen and Van der Zee (1997) discussed the importance of soil moisture to toxicity of chemicals to earthworms. With increased soil moisture, the Cu content of the pore water decreased, which illustrates that bioaccumulation of chemicals by earthworms has both physicochemical and

TABLE 17.1 Multiple Regression of Environmental Availability kp and Bioconcentration Factors (BCFs) in the Earthworm Eisenia fetida of Cadmium, Copper, Lead, and Zinca Log kp Cd Cu Pb Zn BCF Cd Cu Pb Zn

= = = =

0.48*pH 0.15*pH 0.24*pH 0.61*pH

+ + + −

0.28 0.45*log Feox − 0.71*log DOC + 1.33 0.40*log Feox + 1.98 0.65

= = = =

−0.43*pH + 1.36*log clay − 1.39*log OM + 3.19 −0.65*log Feox − 0.38*log clay + 1.38 −0.78*log clay − 0.45*log Feox + 0.46 −0.39*pH − 1.06*log Alox + 0.73*log clay + 3.04

a

Based on relation to various soil factors (arranged in decreasing order of importance), according to the following general formula: Log kp, respectively, log BCF = a*pH(CaCl2) + b*log OM + c*log clay +d*log Feox + e*log Alox +f* log DOC + g*log I + h. BCF = bioconcentration factor; OM = organic matter Source: Janssen et al. 1997a,b.

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biological aspects. Again, like pH, soil moisture (or desiccation) and other confounding factors like temperature (e.g., frost) can affect earthworm behavior and hence metabolism in general. Bauer and Römbke (1997) reported negative relationships for the effects of soil moisture on pesticide uptake by E. fetida for parathion. Holmstrup et al. (1998) studied the impact of frost combined with desiccation on the toxicity of copper to earthworm cocoons. The influence of pH is also important in the ecotoxicological assessment of different soil cleanup techniques. One of the cleanup methods is to wash the contaminated soil with acids, thereby extracting the heavy metals present in those soils. To test the ecological change in the cleaned soil, Van Gestel et al. (1993b, 2002) used some earthworm bioassays. They showed that these extractive treatments might even increase the uptake of metal residues by earthworms from remediated soils. In this context, there was an interesting observation by Cheng and Wong (2002) that earthworms (Pheretima species) had a decreasing effect on the pH of a red soil, thereby influencing the availability of Zn diethylene triamine pentaacetic acid (DTPA)- and NH2OH–HCl-extractable fractions) to earthworms in that soil. Biological factors also influence seasonal fluctuations in the uptake processes of chemicals by earthworms. Bengtsson and Rundgren (1992) reported, in a field experiment with Lumbricus terrestris, that the uptake of Pb was lower during wintertime than in summer. The steady state of the lead burden of the earthworms during the cold winter period indicated that uptake is an active process, probably related to feeding, in which soil temperature, pH, and moisture play important roles. Morgan and Morgan (1993) observed that the epigeic species Lumbricus rubellus accumulated a higher Zn concentration during winter and early spring when earthworm activity is high. The endogeic earthworm species Aporrectodea caliginosa accumulated lower Cd and Zn concentrations during diapause than when active, which may be explained by active elimination and a significantly higher Pb content in the earthworm. The higher Pb content was explained by greater retainment of Pb at the same time that biomass decreased. For the toxicity of organic compounds to earthworms, Van Gestel (1992) and Van Gestel and Ma (1993) developed a soil pore-water partitioning approach (Figure 17.3) and derived quantitative structure-activity relationships (QSARs) for a number of chlorinated hydrocarbons (Figure 17.4). This has been further modeled and validated by Jager et al. (1998, 2000, 2003a,b) for the toxicity of organic chemicals in general and extended to hydrophobic chemicals and dioxins/furans by Belfroid (1994) and Loonen (1994). In studies on the uptake of these chemicals by earthworms from water, moist soil, and soil plus food, they showed that the uptake proceeds in a monophasic way. By contrast, the elimination of chemicals by earthworms in soil is biphasic, with a slow second phase similar to the elimination rate in water (Belfroid et al. 1994). The first stage of fast elimination of chemicals could therefore be ascribed to emptying soil from the gut. Loonen (1994) observed that, in the presence of sediments, the aquatic earthworm species Lumbricus variegatus accumulated additional chemicals not accounted for by the soil-water partitioning model, suggesting also an active uptake from sediment particles. Such uptake was measured for soil in laboratory experiments by Belfroid (1994). Jager investigated further these two uptake and elimination pathways using ligatured earthworms (a tissue adhesive technique developed by Vijver et al. 2003) and observed that the gut route of elimination became more important compared with the skin route of loss with increasing hydrophobicity of the contaminant. It seems that, in organic rich soils, this gut route of loss of chemicals can be of greater importance. Relating these data to field conditions is still not possible (Belfroid et al. 1996). As a consequence, there is a strong argument for using the potential or critical body burden (or better, the critical body concentration) as an index of actual chemical exposure instead of using applied doses (as suggested by Lanno et al. 1997 and Fitzgerald et al. 1997). Loonen (1994) also observed, when repeating an accumulation study of toxicants in earthworms after a contact period of more than 2 years, that there was clearly decreased bioavailability (Figure 17.5). This indicates that contaminants may become bound to organic matter, mineral particles, and micropores and thereby become less available for uptake into earthworms or other biological process. This aging process is now broadly accepted and provides one of the main arguments why © 2004 by CRC Press LLC

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Soil Solid Phase

Soil Pore Water Kp = f(CEC, pH, %OM, %clay, etc.)

Adsorbed + Complexed Metal Species

Other Metal Species Organometal Complexes Free Metal Ions (Men+ )

BCF = f(pH, ion conc., etc.) n+ LC50 = f(pH, Me , etc.)

Soil Organism

FIGURE 17.3 The processes involved in the soil–water partitioning approach, worked out for heavy metals (OM = organic matter, BCF = bioconcentration factor). (After Van Gestel and Ma 1993.)

3

2

1

0

–1

+ +

● ●

▫ ▫ ▫ ▫

▫ chlorophenols Y = –1.00* ´ + 4.86 r 2 = 0.93

●

+ + + +

▫ ▫

●

+ +

+ chloroanilines Y = –1.04* ´ + 5.00 2 r = 0.95

▫ ▫ ▫

chlorobenzenes Y = –1.85* ´ + 7.62 r 2 = 0.98 2.5

▫ +▫ +

●

1.5

● ● ●

● ●

▫ ▫ ▫ ● ▫ ●

3.5

4.5

5.5

FIGURE 17.4 Relationships among the toxicity of chlorophenols, chloroanilines, and chlorobenzenes to Eisenia andrei and log Kow. For Lumbricus rubellus, the relation with chlorophenols is log LC50 (µmol L−1) = −0.72*log Kow + 4.46 r2 = 0.87. (After Van Gestal and Ma 1993.)

aged contaminants do not have as much ecotoxicological impacts as would be expected from the total content. For pesticides, this process is termed bound residues. It illustrates once more that the bioavailable fraction of the total toxicity content is of prime importance. Some studies have described the internal transport, more speciÞcally the storage of contaminants, in earthworms (Eijsackers 1989) but seldom also describe the ecological consequences of this distribution pattern. Bengtsson and Rundgren (1992) suggested that increased metal contents in the earthworm nerve cord, for instance, might primarily inßuence the transfer of stimuli and consequently earthworm mobility and behavior.

TESTING CONTAMINANT TOXICITY WITH EARTHWORMS When it is understood how the earthworm takes up a contaminant and distributes it through its body, the next question is what kind of effects on the earthworm will result. The concentration of contaminants in the earthworm reproductive organs may exert a different effect compared with their accumulation in the cerebral ganglion or in special storage organs. There has been a strong trend toward © 2004 by CRC Press LLC

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Cworm/Csediment

30 + + + 20

+ + ++ +

10

+ + +

+ +

+ +

+

+ + + + ++



+




+ 0 0 5

10
















15

20

25

30

time (days)

FIGURE 17.5 Organism/sediment ratios of L. variegatus 3 weeks (+) and 21 months (∆) after mixing sediment with tetrachlorodibenzo(para)dioxine (TCDD). (From Loonen, 1994.)

the development of tests of toxicity to earthworms at the physiological and biochemical level using biomarkers. Moreover, these various effects (test end points) can have different impacts on the ultimate survival of the earthworm or, when looking at the population level, on the intrinsic rate of population increase or maintenance of the earthworm populations. Next, there are different earthworm ecotypes (epigeic, endogeic, and anecic species), which exert different effects on the soil, so the impact of this on the earthworm community must be assessed. When the input of earthworms to ecosystem processes (turning over the soil, inßuencing nutrient release, drainage and aeration processes) is included, the assessment has to be at the ecosystem level as well. In a Þnal scaling-up, it must be realized that environmental impacts occur in a mosaic of different systems, which in practice means Þelds with different crops and pesticide treatments in agroecosystems. Much effort has been put into the standardization of earthworm tests. The European Community (Edwards 1983, 1984) OECD test using artiÞcial soil and E. fetida or Eisenia andrei has proven to be highly reproducible. By adding a small pellet of cow manure as food, the survival of the earthworms improves greatly. The nutritive value of the peat, clay, and sand mixture of the artiÞcial soil has proved to be rather low. An improved protocol setup has been developed (Van Gestel et al. 1989) in which a more sensitive chronic end point (reproduction instead of survival) is used, and a recovery period is introduced, which provides the opportunity to study the hatchability of the cocoons produced. Taken together with cocoon production, this gives a potential prediction of effects on population maintenance rates. Kula (1994) adapted the method for the assessment of pesticide toxicity by applying the pesticide to the soil surface. Comparing the effect of surface application of a compound with effects from mixing it through the total test soil mass produced reasonably similar results for the effects of benomyl (Table 17.2). The observation of earthworm behavioral responses to toxicants has received more attention recently, although the Þrst observation of active avoidance of copper-contaminated soil dates to the 1970s (Eijsackers 1981, 1989). Observations on such earthworm responses include pesticides (diazinon, mecoprop, carbaryl, and potassium-fatty acid salts) (Bauer and Römbke 1997; Slimak 1997; Reinecke and Reinecke 2002); brass powder-contaminated soil (Wentsel and Guelta 1988); ßy ash (Ma and Eijsackers 1989); and harbor sludge (Ma et al. 1997). Capowiez et al. (2002) studied earthworm behavior in situ using x-ray tomography and three-dimensional reconstruction. In a number of studies, earthworms responded to much lower amounts of toxicants than they might encounter in the Þeld (Mather and Christensen 1998), which demonstrates the suitability of these kinds of tests as early warning systems. Hence, toxicant avoidance should be incorporated as an assessment end point in test procedures, as suggested by Yeardly et al. (1996). © 2004 by CRC Press LLC

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TABLE 17.2 Effects of Benomyl after Surface Application or Total Mixture with the OECD Artificial Soil Test Control Number of juveniles Surface application Total contamination Body weight Surface application Total contamination

0.5 kg ha−1

2.5 kg ha−1

11.8 ± 3.6 11.8 ± 3.6

8.7 ± 1.7 6.9 ± 2.5

1.8 ± 0.4 0

124.4 ± 0.3 124.4 ± 0.3

130.1 ± 5.3 136.7 ± 3.6

65.0 ± 1.3 49.9 ± 2.0

Source: Adapted from Kula (1994).

The use of biomarkers is a new concept in earthworm toxicity testing. These have been developed based on enzyme functions (Goven et al. 1993, 1994), for coelomycetes with a cytometric assay (Brousseau et al. 1997), for hemoglobin content (Rozen and Mazur 1997), and for sperm production and fertility (Cikutovic et al. 1993; S.A. Reinecke and Reinecke 1997). For an overview of the potential uses of biomarkers in earthworm toxicity testing, see the work of Scott-Fordsman and Weeks (1998). A major challenge to the adoption of earthworm toxicity tests is still the question why tests using earthworms should be better than tests using other organisms. There is also the question whether earthworm biomarkers should be generally applicable to other soil organisms. For further improvement of earthworm toxicity testing, attention is needed to the statistical interpretation of dose-effect curves, intercompound testing, interspecies testing, testing real (low and repeated) field doses, and artificial vs. natural soil. In interpreting the results of earthworm ecotoxicity tests, it must be realized that, when using the LC50 (the concentration that results in toxicity to 50% of the test animals), the form of the doseeffect curves is not critical. However, in using more sensitive end points such as reproduction and use of end points like the no observed effect concentrations (NOECs), their form certainly matters. Spurgeon et al. (1994) reported a dose-response step function for Cd and a gradual decrease for Zn. Posthuma et al. (1993) reported a linear logarithmic model as the best description for the combined effects of Cd/Cu and Cd/Zn on the reproduction of earthworms and a hormesic model for the effect of Cu/Zn on earthworms. Especially when deriving NOEC values, the model that is used and the statistical variation affect the results considerably. Therefore, the design of the experiment and the variability in the resulting data set can influence the NOEC derivation greatly. Moreover, it is only an approximation, and as a consequence (e.g., for step functions) it is not always possible to derive NOECs. Hence, to understand and interpret fully the different response patterns of earthworms to chemicals, much more work is necessary. Studies of intercompound effects of organic compounds that influence the overall toxicity (e.g., Neuhauser et al. 1986; Van Gestel and Ma 1993; Kula 1994; Larink and Kula 1994; Callahan et al. 1994) and on the effects of heavy metals (Posthuma et al. 1993; Diaz-Lopez and Mancha 1994) showed that, for a number of compounds, generalized relationships can be observed. For generally acting (narcotic) organic compounds (Van Gestel and Ma 1993), it is possible to derive a QSAR, as shown in Figure 17.4. For assessing the toxicity of heavy metals to earthworms, Posthuma et al. (1993) developed an approach in which they “summed” heavy metals on the basis of their LC50 values expressed in toxicity units. Diaz-Lopez and Mancha (1994) investigated the effects of different anions and the additions of different fertilizers on the toxicity of copper. Various copper salts (sulfate, nitrate, and chlorides) have different toxicities to earthworms; moreover, the combination of NH4NO3 and CuSO4 produces a much greater toxic effect than these chemicals tested separately. Fischer and © 2004 by CRC Press LLC

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Molnar (1997) reported that a number of the macronutrient metals (NaCl, KCl, CaCl2, MgCl2, BaCl2, MnCl2, SnCl2, SrCl2, AlCl3, and Fe Cl3) reduced earthworm reproduction, and CoCl2, CuCl2, and NiCl2 proved to be acutely toxic to reproduction. Studies of the susceptibility of different earthworm species to chemicals produced results that are distinctly different. Neuhauser et al. (1986) claimed that the four species they tested showed good correlations in relative toxicity. Haque and Ebing (1983) reported a significant relationship between E. fetida and L. terrestris for susceptibility to a series of pesticides. Heimbach (1985) found, in general, a factor 2 to 3 difference in response when comparing E. fetida and L. terrestris. By contrast, Kula (1994) reported much greater differences between E. fetida and three other earthworm species (A. caliginosa, Allolobophora chlorotica, and Allolobophora longa), ranging from a factor of 4 to a factor of 80 in relative susceptibility. According to Kratz and Pöhhacker (1994), two closely related species such as Eisenia fetida fetida and Eisenia fetida andrei have distinctly different toxicity responses to carbendazim and phenmedipham. Ma and Bodt (1993) tested the toxicity of chemicals to six earthworm species and suggested a genus-related variation in toxicity to chlorpyrifos, with increasing susceptibility in the order Eisenia spp. < Aporrectodea spp. < Lumbricus spp. These differences were not associated with body weight, and it seems more likely that some physiological factor is involved. This conclusion was supported by the observation of Stenersen et al. (1992) that the different sensitivities of the closely related Eisenia species (E. fetida, E. andrei, and Eiseniella veneta) to carbaryl was related to the possession of a carbaryl-resistant enzyme by the first two species. For a more systematic comparison, the results of four studies of various pesticides and a number of different earthworm species are plotted in a comparable way in Figure 17.6. To this end, the LC50 values are plotted on the same scale irrespective of the absolute susceptibility (= LC50) for that pesticide. In general, E. fetida was always at the least sensitive end of the range, and the Allolobophora species was always at the most sensitive end. But, there was no fixed sequence in sensitivity relationship between the species. Hence, there is little basis for a taxonomically related sensitivity pattern, although it is clear that E. fetida is usually the least-sensitive species and hence the least-suitable test species. As was recommended at the Sheffield International Workshop on Earthworm Toxicity Testing (Greig-Smith et al. 1992) and was confirmed later by Kula (1994), field-related doses should be those used in tests and five times this dose. However, Springett and Gray (1992) showed that much lower doses (20, 10, and 5% of field rate) can be toxic when applied regularly for a longer period (100 days at 2-week intervals), which is in fact more relevant to the actual growing period of the earthworm and agronomic practices. The use of artificial soil as a substitute for natural soils has been scrutinized, and there are reports indicating good correlations as well as only a limited correlation between them. With respect to both heavy metals (Posthuma et al. 1993) and organic pesticides, the artificial soil test may overestimate as well as underestimate the toxicity to earthworms in real soils. Lanno et al. (1997) observed a marked influence of soil type, comparing three soil types, with respect to the impacts of benomyl, 2-chloroacetamide, diazinon, and pentachlorophenol on earthworms. Different toxicant availability may be a governing factor because both organic matter and clay adsorb chemicals. Van Gestel and Van Straalen (1994) stated that the organic matter content, in combination with soil sorption data, can provide a good basis for a translation formula from artificial soil to real soil conditions.

FIELD STUDIES PESTICIDE

AND

HEAVY METAL TOXICITY STUDIES

The ultimate relevance of all toxicity testing lies in proper maintenance of the earthworm population, as a key component of soil ecosystems, by turning over the soil continuously, thereby creating an optimal soil condition, degrading organic matter, and working it into the soil. In agronomic systems,

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Parathion (1) 600

331

Propoxur (1) 300

Aldicarb (2) 20

Carbaryl + Carbofuran (2) 80

291 E.f.

554 E.f.

>16 E.f.

16 E.f. 41 A.chl. 318 L.t. 232 A.c.

4-8 L.r.

119 A.I.

0

0

Dithiocarbamates (2) 40

54 A.chl. 36 L.t. 6.8 A.c. 3.8 A.I.

4 A.c. 4A.chl. 0

Paraquat (3)

16-32

Chloroacetamide (3) 80

4000

0

80 A.n. 75 E.f.

<4 L.r. <4 A.chl. <4 A.c.

Chlorpyrifos (4) 1200

1174 E.v. 1077 E.f.

3200 E.f. 778 A.I. 755 A.c.

54 L.t. 47 A.chl. 40 A.c. 16 L.r.

458 L.t. 1000 L.t.

8 A.chl.

580 A.c. 0

0

129 L.r. 0

0

FIGURE 17.6 Comparative plots of the sensitivities (LC50) of different earthworm species tested under comparable conditions in multicompound studies for various sets of pesticides. (Data from Kula et al. 1994; Stenerson et al. 1992; Ma and Eijsackers, 1989.)

a proper combination of management of these “underground processes” and chemical pest treatment should be achieved by proper, preventive toxicity test assessment procedures. Other than the standardized Þeld tests discussed and formulated at the ShefÞeld and Amsterdam Earthworm Toxicity Workshops (Edwards 1992, 1998; Heimbach 1992a,b; H. Kula 1992) limited progress has been made. Kula (1994) proposed a standardized approach to Þeld experimentation, with prescriptions for methods on representative grass and arable sites, plot sizes of 10 × 10 m with a 2-m protection zone with four replicates, sampling procedures, and intervals. He suggested L. terrestris and A. caliginosa as the most relevant species with sufÞcient population densities in these types of sites. Heimbach (1997) experimented with barriers between plots and included the indirect impact of a changed food supply (grass cut or mulched). Lateral movement of earthworms should be limited but cannot be neglected in Þeld sampling, whereas the impact of plowing and reseeding could be rather drastic. However, the observations of Mather and Christensen (1998) that there could be extensive lateral dispersal of earthworms call for caution in Þeld experiments. There is a need for proper validation of laboratory vs. Þeld toxicity tests on earthworms. Heimbach (1992a,b, 1994, 1997) made extensive studies comparing laboratory reproduction studies © 2004 by CRC Press LLC

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and field studies of earthworm toxicology for 10 pesticides. Assuming an even pesticide distribution in the upper 5 cm of the soil, the laboratory reproduction studies were reported to be at least 5 to 10 times more sensitive than the field responses of all earthworms. However, as stated, there is an impressive difference in sensitivity among various earthworm species. Jones and Hart (1998) reviewed laboratory-field comparisons for the toxicity of pesticides and underlined the impact of sensitivity differences among species. They considered that differences among soils are less important. Bembridge et al. (1997) showed, in a long-term experimental field treatment and by ordinations and principle response curve analysis, the different responses of various earthworm species to benomyl. Allolobophora spp. and Nicodrilus spp. proved to be the most sensitive to chemicals. Jones and Hart (1998) stressed the element of recovery by earthworms from exposure to chemicals. For pesticides with a half-life in soil of less than 50 days, earthworm population recovery occurred within a year, and it was possible to recover from reductions in numbers of other soil invertebrates of over 90%. With more persistent pesticides, the rate of recovery was reduced. This is effectively also the principal difference in assessing the impacts of persistent heavy and degradable organic contaminants. In correlating laboratory test results on heavy metal toxicity to earthworms with field sampling, Spurgeon et al. (1994, 1999), Posthuma et al. (1998), and Nahmani et al. (2002) reported on earthworm populations and diversity in the vicinity of large smelter works or deposits of smelter ashes. Spatial distribution in a dose-effect relationship was expressed by Spurgeon et al. (1994) based on the OECD 14-day LC50 values of the heavy metals present in terms of square kilometers from the source in the vicinity of the smelter works, which should be devoid of earthworms. For Zn, this was 75 km2; for Cu, it was 42 km2; and for Pb, it was 4.7 km2. Posthuma et al. (1998) plotted the numbers of earthworms found in a sigmoid distance-response curve, which enabled the calculation of an EC50 that could be compared with the results of standard OECD tests. Posthuma et al. (1998) and Spurgeon et al. (1994) reported an overestimation of the potential toxic effect of the heavy metals using the OECD test. Posthuma et al. (1998) explained this by noting a difference in the uptake rate of Cd accumulation by earthworms that they actually observed (Figure 17.7) and hence concluded a difference in bioavailability. This was confirmed by the observation of Spurgeon et al. (1994) that, despite their extrapolation that no earthworms should be present in the close vicinity of the smelter, they found some still present there. In addition, Posthuma et al. (1998) pointed at a possible antagonistic effect of binary mixtures of heavy metals compared with effects in the single-metal OECD test (see also the section on current interest in earthworm ecotoxicology). Clearly, these studies confirm for field conditions a difference in sensitivity among various earthworm species. The endogeic Allolobophora and Nicrodilus spp. appear to be the most sensitive, followed by epigeic species such as Lumbricus rubellus and L. castaneus, whereas organic matteroriented species like Dendrobaena veneta are sometimes found in soils with high heavy metal contents. Callahan and Linder (1992) sampled soil from a contaminated site and tested this in the laboratory or under outdoor conditions for toxicity to earthworms. This last approach was also used by Bengtsson et al. (1992) when they studied the possible adaptation of earthworms to long-term exposure to chemicals in the soil surrounding a copper brass mill dating from the 17th century. Bengtsson and Rundgren (1992) discussed the possibilities for recovery of the observed scattered and low-density Dendrobaena octaedra population in the vicinity of the old mill.

LAND IMPROVEMENT

AND

EARTHWORMS

AS

BIOENGINEERS

A rather new role of earthworms in environmental research can be related to the concept of their suggested impact as soil bioengineers (Jones et al. 1994). Earthworms contribute to a better soil environment by their burrowing behavior. This is especially useful not only in ameliorated soils (former mining soils); eroded soils; and remediated, formerly contaminated soils, but also in deposits of mine waste or contaminated sediments. In all these instances, earthworms were able to

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Cd in Enchytraeus crypticus OECD and Budel 4 (0.8 mg/kg)

22.5

Q0 A k1 Q eq BCF

20

mg–1/kg (dry wt)

17.5

OECD / Budel = 0.3175/0.069 = 1.092 / 1.046 = 0.050 / 0.137 = 22.158/7.724 = 27.698/9.655

15 one-compartment model OECD observations (s.e.) OECD one-compartment model Budel 4 soil observations (s.e.) Budel 4

12.5 10 7.5 5 2.5 0

0

5

10

15

20

25

time (days)

30

35

40

45

FIGURE 17.7 Uptake of zinc by Enchytraeus crypticus from OECD artiÞcial soil and Budel sandy Þeld soil showing reduced bioavailability in the Þeld soil. (After Posthuma et al. 1998.)

revive the soil by mixing it with microorganisms, stimulating drainage and aeration, and by improving the natural microbial breakdown processes of contaminants. A special and still-limited case is soil improvement in the framework of nature development as carried out in the Netherlands for contaminated riverbank topsoil or nutrient overloaded former arable soils. The contaminated or nutrient-loaded topsoil is taken away, and a new soil ecosystem can start to develop on the subsoil brought to the surface; earthworms have a prime role then in bringing back a natural soil structure. The key to success in all cases is the recolonization by the earthworms from outside the area or from local remnants of a degraded population (so-called refugia). This recolonization process theoretically includes earthworm dispersal, establishment, and population development. Dispersal is related mainly to mobility, adaptabilities, establishment, and population development with a potential for recovery or, more precisely, maintenance under suboptimal conditions. In their studies of contaminants close to a brass mill, Bengtsson and Rundgren (1992) attributed the maintenance of D. octaedra populations to either the immigration of juveniles from less-polluted areas or to spatial heterogeneity in metal distribution in the soil. Immigration is perhaps a more relevant explanation than has been suggested given the studies by Mather and Christensen (1988) about the surface mobility of earthworms, and this perspective of immigration or reimmigration to polluted sites needs further attention. It should be investigated particularly in relation to the restoration of acidiÞed and limed coniferous soils such as those studied by Rundgren (1995), harbor sludge deposits (Ma et al. 1997), or remediated sites as described by Bradshaw (1994) and Tamis et al. (1994). For rehabilitation of contaminated land, Ma and Eijsackers (1989) suggested a management program that included organic amendments and introduction or recolonization of earthworms in parallel with revegetation of the area. This was worked out in a more general way by Eijsackers (1995). When considering the impacts of earthworm recovery under these conditions, we have to be aware of indirect effects, confounding factors that inßuence the earthworms such as their sensitivity to toxic compounds. Very few data are published regarding these kinds of effects. Heimbach (1997) studied the implications of different kinds of management of Þeld plots (e.g., grass cover cut and removed, mulched, or plowed) on the toxicity to earthworms of applied chemicals. He incorporated factors such as changes in food availability, mechanical damage, and a change in soil structure as

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indirect effects in the test. He concluded that these last effects seemed to be more important than the impacts of the applied chemicals. In addition to improvements in physical soil conditions, earthworm activities can contribute to better chemical soil conditions. In a study of a peat polder soil contaminated with polycyclic aromatic hydrocarbons (PAHs), Eijsackers et al. (2001) showed that the introduction of earthworms could result in an accelerated breakdown of the PAHs present. Kersanté et al. (2002) observed another impact: earthworm activity promoted the adsorption of the pesticide atrazine to the ingested soil, thereby causing a slower microbial breakdown. Singer et al. (2001) observed an enhanced microbial breakdown of polychlorinated biphenyls by microorganisms introduced by the introduction of earthworms, which stimulated mixing of the microorganisms through the soil. Eijsackers and Doelman (2000) further specified this for removal of a number of contaminant groups (organochlorines (OCls), PAHs, and heavy metals) from a riverbank soil. Here, the soil biological processes are combined with physical forces caused by soil erosion from the river as well as the implications of inundation. For heavy metals, bioremediation is related mainly to differences in their availability under aerobic or anaerobic conditions. However, next to that, the consumption of soil by the earthworms may lead to changed availability of the heavy metals. Salisbury (1925, cited by Lee 1985) observed that earthworm casts had a higher pH than the surrounding soil. Research aimed specifically at assessing the role of earthworms in changing the bioavailability of contaminants resulted in mixed observations: Ablain et al. (2002) observed increased contaminant availability, whereas Zorn (personal communication) suggested decreased bioavailability. Another type of indirect relationship that deserves further study is between earthworms and soil structure. Tomlin et al. (1993) observed the impacts of three types of treated sewage sludge on soils and reported that earthworm burrows were lined with fecal material that had higher heavy metal concentrations than the surrounding soil. Moreover, soluble forms of heavy metals could be leached more easily through these burrows.

EFFECTS

OF

TOXICANTS

AT

FOOD CHAIN

AND

ECOSYSTEM LEVELS

The best-recognized use of earthworms for the general public is as fish bait. Many other animal species prefer earthworms as food, including songbirds, raptors (buzzards and tawny owl), various mammals (hedgehog, badgers, and boars), and amphibians and reptiles (Edwards and Bohlen 1996). Hence, earthworms act as an important route for transfer of contaminants from the soil to food chains in terrestrial ecosystems. They have a considerable ability to accumulate contaminants from the soil matrix into their body parts, greater than most other soil invertebrates, and therefore play an important role in food chain transfer of contaminants. With respect to contaminant food chain research, a number of models of the food chain transfer of toxicants have been developed. These start with rather simple models, such as those suggested by Romijn et al. (1991) and Luttik et al. (1993). In these models the maximal permissible concentration (MPC) of a toxicant is calculated by dividing the NOEC for these bird or mammal predators (NOECbird/mammal) by the bioconcentration factor (BCF) of the contaminant from the soil into the earthworm (BCFearthworm). MPCs were calculated for standard soil situations and compared with MPCs for terrestrial organisms. As a next step, the biological characteristics of the predator were introduced. Noppert et al. (1994) combined BCFs for earthworms with the uptake rate by their predators, including feeding rates, uptake efficiency, elimination rates, and life expectancy. Gorree and Tamis (1992) extended this concept by introducing the bioavailable proportion of a compound (the soil pore-water partitioning approach). Everts et al. (1993) introduced the basal existence and field metabolic rates of contaminants for predators, both in laboratory feeding experiments and (based on general assumptions) in the field, and developed a formula:

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EMR total bird FMR norm. condition E prey - * ------------------------------------------* -----------------------------------------NOEC prey = NOEC lab. food * ------------------E lab. food EMR species concern FMR peak activity where E is the energy content, FMR is the field metabolic rate, EMR is the existence metabolic rate, and BMR is the basal metabolic rate. These last three models all assume a stochastic response for exposure of earthworms to chemicals and thus introduce probability into the risk estimates derived. The third step in these model developments introduces earthworm population dynamic elements. A major problem is the need to validate the model parameters by field data. Ma and Van der Voet (1993) proposed a nonlinear, age-dependent relationship for the predators, in combination with the BCFearthworm and the bioavailable fraction of the chemical in the soil and theoretically derived chemical uptake rates for the various organs of the predators. It was possible in this way to predict target organ contaminant loads as functions of ingestion of the contaminant, its assimilation rate, and internal partitioning in the earthworm. A fourth step is to integrate these uptake mechanisms into ecosystem processes such as the cycling of biomass through an ecosystem. Effectively, it is the cycling of biomass that steers the accumulation rate of a contaminant through trophic levels in the food chain. This approach is used in the CATS model, which analyzes the cadmium cycle in a meadow ecosystem based on the C and N cycles and the resulting biomass transfers between the different pools in the system (Traas and Aldenberg 1992). These results and estimates can be expressed in different ways. They can be expressed as the relationship between potential environmental concentration (PEC) and the no effect concentration (NEC), the PEC/NEC ratio (Romijn et al. 1991; Luttik et al. 1993; Noppert et al. 1994). Using this, uptake of diazinon by earthworms can be described as a risk for passerine birds because of the sensitivity of these birds to this pesticide (Stephenson et al. 1997). But, it is also possible to define a potential risk area around a point pollution source (Gorree and Tamis 1992) or an exposure risk criteria for target organs (Ma and Van der Voet 1993). In all food chain contaminant uptake studies, the role of earthworms, or soil invertebrates in general, as the intermediary between soil and terrestrial predators proved crucial. BCFs for soil earthworms, in all cases, had the highest levels, as exemplified by Ma (1994) (Table 17.3). Because of their high bioaccumulation potential and relatively low sensitivity to certain toxic compounds, earthworms are major toxicity sources for many terrestrial predators (Table 17.4), as shown convincingly in the extensive model studies of Klok (2000).

TABLE 17.3 Concentration Factors (CFs) for Cadmium and Lead in Herbivore and Decomposer-Based Terrestrial Food Chains Species Group Coleoptera Carabidae Aranea Opiliones Lumbricidae Grass Soil (mg kg−1)

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CFcadmium

CFlead

1.3–6.5 1.5–9.5 18–86 5.5–23 32–136 0.5–2.5 0.2–2.9

0.2–0.5 0.09–0.3 0.3–0.5 0.1–0.8 1.3–3.0 0.07–0.3 20–130

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TABLE 17.4 Cadmium and Zinc Contents of Mole Kidneys, Earthworms (L. rubellus), and Soil in Heavy Metal-Contaminated Sandy Soils (Kempen Area and Clean Sandy [Two Areas]) and Clayey Reference Soils Cadmium

Mole kidney Earthworms Soil (0–10 cm)

Zinc

Contaminated Sand

Reference Sand

Reference Clay

Contaminated Sand

Reference Sand

Reference Clay

197 77 1.8

18/47 3/8 0.2/0.2

43 26 1.5

202 1599 103

131/118 286/426 35/46

132 725 272

Adapted from Klok 2000.

The function of earthworms as food for other organisms is important not only in the context of food chain transfer. When earthworm populations decrease because they are negatively affected by a specific contaminant, this may cause a food shortage for their predators. When this causes negative effects on the predators, these are considered “indirect effects.” There have been a number of field studies on the responses of earthworm populations to chemicals, but a detailed proper population dynamic analysis or modeling is still lacking. Klok et al. (1997) presented some results of a model that combined earthworm population development with energy budgets and distinguished among the various approaches in population biology to differentiate between the development stage of the earthworm or merely its size, to use defined discrete time steps or a continuous time frame, and to include density-dependent processes when necessary. Another approach is to model the structure of a soil system (Knoop and Traas 1993). Such a model has been built to stimulate the structure of an arable or grassland soil ecosystem. With this model, changes in C and N cycles under the influence of different management practices have been simulated (de Ruiter et al. 1993). Next, the removal of a specific microbivore or predatory group because of its relatively high sensitivity to or use of a specific toxic compound is simulated in terms of its consequences for C and N cycles and general biomass turnover (de Ruiter and Moore 1994) in the ecosystem. In a study by Marinissen and de Ruiter (1993), the impact of earthworms on such systems was analyzed. Therefore, it should be possible to assess the effects of removal of earthworms from the system in a similar way, but this has yet to be done. A weakness in the field validation of such ecosystem-structure studies is a lack of studies on interactions between earthworms and other groups of soil invertebrates (see Chapter 2 and Chapter 13, this volume). Hyvönen et al. (1994) published a study on interactions between earthworms and nematodes. Bengtsson et al. (1986) observed that nematode infections of earthworm cocoons decreased hatchability. Acidity and heavy metal contamination may also interfere with the impact of nematodes.

RISK ASSESSMENT BASED ON EARTHWORM TOXICOLOGY The ultimate goal of all the research described here is to assess the overall environmental impact of specific chemical contaminants (pesticides, chemical compounds, heavy metals); specific contaminated substrates (sewage sludge, building materials); or generally or specifically contaminated situations, such as abandoned waste dumps or contaminated and dredged harbor sludge. For such research aims, earthworm toxicology tests, bioassays, field studies, and monitoring programs could be useful components. It must be realized that there are some prerequisites for a proper ecological risk assessment. As Menzie et al. (1993) stated, “The ‘problem formulation’ phase of ecological risk assessment is © 2004 by CRC Press LLC

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probably the most important aspect. Sensitivity and field validation studies have often been insufficient to ensure the relevance of ecological exposure and effect models to real world situations.” Jepson (1993) claimed that the current emphasis on toxicological criteria at the expense of ecological factors can result in a detailed description of local toxic effects of a contaminant but neglect of the consequences at a population level. In this context, he recommended a long-term risk analysis at the scale of total earthworm populations, hence not just the field or the farm but the agricultural landscape. In this way, the total pattern and frequency of exposure of earthworms to toxic chemicals can be combined for pollutants with short or long persistence in soils, small or large application areas, and high or low application frequencies. Moreover, this can include the impacts of pollutants in temporary habitats or pollutants affecting dispersive invertebrates. The SCARAB (seeking confirmation about results of Boxworth) project reported by Tarrant et al. (1997) compared total agricultural management systems with normal vs. low inputs of pesticides over a long period. These kinds of studies may provide the data for the above-suggested long-term assessments of toxicant impacts at the landscape scale. Burger and Gochfeld (1992) called for incorporating appropriate temporal time scales and critical life stages into ecological risk assessment.

ACKNOWLEDGMENT The help of Mathilde Zorn in updating this chapter is gratefully acknowledged.

REFERENCES Ablain, F., C. Sahut, and D. Cluzeau. 2002. Earthworms activities and sewage sludge heavy metals availability, in Proceedings Sixth International Earthworm Ecology Symposium, Cardiff, in press. Bauer, C. and J. Römbke. 1997. Factors influencing the toxicity of two pesticides on three lumbricid species in laboratory tests, Soil Biol. Biochem., 29:705–708. Beeby, A. 1993. Toxic metal uptake and essential metal regulation in terrestrial invertebrates, in R. Dallinger and P.S. Rainbow, Eds., Ecotoxicology of Metals in Invertebrates, Lewis Publishers, Chelsea, MI. Belfroid, A.C. 1994. Toxicokinetics of Hydrophobic Chemicals in Earthworms, Ph.D. thesis, University of Utrecht, the Netherlands. Belfroid, A.C., J. Meiling, D. Sijm, J. Hermens, W. Seinen, and C.A.M. Van Gestel. 1994. Uptake of hydrophobic halogenated aromatic hydrocarbons from food by earthworms (Eisenia andrei), Arch. Environ. Contam. Toxicol., 27:260–265. Belfroid, A.C., D.T.H.M. Sijm, and C.A.M. Van Gestel. 1996. Bioavailability and toxicokinetics of hydrophobic aromatic compounds in benthic and terrestrial invertebrates, Environ. Rev., 4:276–299. Bembridge, J., T.J. Edwards, and P.J. Edwards. 1997. Variation in earthworm populations and methods for assessing responses to perturbations, in S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., Advances in Earthworm Ecotoxicology, SETAC Press, Pensacola, FL, pp. 341–352. Bengtsson, G., H. Ek, and S. Rundgren. 1992. Evolutionary response of earthworms to long-term metal exposure, Oikos, 63:289–297. Bengtsson, G., T. Gunnarsson, and S. Rundgren. 1986. Effects of metal pollution on the earthworm Dendrobaena rubida (Sav.) in acidified soils, Water Air Soil Pollut., 28:361–383. Bengtsson, G. and S. Rundgren. 1992. Seasonal variation of lead uptake in the earthworm Lumbricus terrestris and the influence of liming and acidification, Arch. Environ. Contam. Toxicol., 23:198–205. Bradshaw, A.D. 1994. Natural rehabilitation strategies, in H.J.P. Eijsackers and T. Hamers, Eds., Integrated Soil and Sediment Research: A Basis for Proper Protection, Kluwer, Dordrecht, the Netherlands, pp. 577–587. Brousseau, P., N. Fugère, J. Bernier, P. Coderre, D. Nadeau, G. Poirier, and M. Fournier. 1997. Evaluation of earthworm exposure to contaminated soil by cytometric assay of coelomycetes in Lumbricus terrestris (Oligochaeta), Soil Biol. Biochem., 29:681–684. Burger, J. and M. Gochfeld. 1992. Temporal scales in ecological risk assessment, Arch. Environ. Contam. Toxicol., 23:484–488.

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Callahan, C.A. and G. Linder. 1992. Assessment of contaminated soils using earthworm test procedures, in P.W. Greig Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Ecotoxicology of Earthworms. Intercept, Andover, U.K., pp. 187–196. Callahan, C.A., M.A. Shirazi, and E.F. Neuhauser. 1994. Comparative toxicity of chemicals to earthworms, Environ. Toxicol. Chem., 13:291–298. Capowiez, R.M.Y., C. Mazzia, and L. Belzunces. 2002. Modifications of the burrowing behavior of A. nocturna and A. chlorotica due to the presence of imidacloprid in the soil, in Proceedings Sixth International Earthworm Ecology Symposium, Cardiff, in press. Cheng, J.M. and M.H. Wong. 2002. Effects of earthworms on Zn fractionation in soils, Biol. Fertil. Soils, 36:72–78. Cikutovic, M.A., L.C. Fitzpatrick, B.J. Venables, and A.J. Goven. 1993. Sperm counts in earthworm (L. terrestris) as a biomarker for environmental toxicology — effects of cadmium and chlordane, Environ. Pollut., 81:123–125. de Groot, A., R. Fleuren, R. Baerselman, and W. Peijnenburg. 2002. Towards a new methodology of ecotoxicity testing of metal contaminated soils: factors affecting metal toxicity to earthworms, in Proceedings Sixth International Earthworm Ecology Symposium, Cardiff, in press. de Ruiter, P.C. and J.C. Moore. 1994. Simulation of the effects of contamination on the functioning of below ground food webs, in H.J.P. Eijsackers and T. Hamers, Eds., Integrated Soil and Sediment Research: A Basis for Proper Protection. Kluwer, Dordrecht, the Netherlands, pp. 309–312. de Ruiter, P.C., J.C. Moore, K.B. Zwart, L.A. Bouwman, J. Hassink, J. Bloem, J.A. de Vos, J.C.Y. Marinissen, W.A.M. Didden, G. Lebbink, and L. Brussaard. 1993. Simulation of nitrogen mineralization in the below-ground food webs of two winter wheat fields, J. Appl. Ecol., 30:95–106. Diaz-Lopez, G. and R. Mancha. 1994. Usefulness of testing with Eisenia foetida for the evaluation of agrochemicals in soil, in M. Donkers, H. Eijsackers, and F. Heimbach, Eds., Ecotoxicology of Soil Organisms, Lewis Publishers, Chelsea, MI, pp. 251–256. Donkers, M., H. Eijsackers, and F. Heimbach, Eds. 1994. Ecotoxicology of Soil Organisms, Lewis Publishers, Chelsea, MI. Edwards, C.A. 1983. Development of a Standard Laboratory Method for Assessing the Toxicity of Chemical Substances to Earthworms, Report EUR 8714 EN, Environment and Quality of Life Commission of the European Communities, Brussels, Belgium. Edwards, C.A. 1984. Report of a Second Stage in Development of a Standardized Laboratory Method for Assessing Toxicity of Chemicals to Earthworms, Report EUR 9360, EN, Environmental and Quality of Life Commission of European Communities, Brussels, Belgium. Edwards, C.A. 1992. Testing the effects of chemicals on earthworms: the advantages and limitations of field tests, in P.W. Greig Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Ecotoxicology of Earthworms, Intercept, Andover, U.K., pp. 75–84. Edwards, C.A. 1998. Principles for the design of flexible earthworm field toxicity experiments, in S. Sheppard, J. Beinbridge, M. Holinstrup, and L. Posthuma, Eds., Proceedings of the Second Workshop on Earthworm Ecotoxicology, SETAC Press, Pensacola, FL, pp. 313–326. Edwards, C.A. and P.J. Bohlen. 1996. Biology and Ecology of Earthworms, Chapman & Hall, London. Eijsackers, H. 1981. Effecten van koperhoudende varkensmest op regenwormen en op de kwaliteit van grasland, Landbouwkundig Tijdschr., 93:307–314. Eijsackers, H. 1989. The impact of heavy metals on terrestrial ecosystems: biological adaptation through behavioural and physiological avoidance, in O. Ravera, Ed., Ecological Assessment of Environmental Degradation, Pollution and Recovery, Elsevier, Amsterdam, the Netherlands, pp. 245–259. Eijsackers, H. 1994. Soil ecotoxicology: finding the way in a pitch dark labyrinth, in M. Donkers, H. Eijsackers, and F. Heimbach, Eds., Ecotoxicology of Soil Organisms, Lewis Publishers, Chelsea, MI, pp. 1–23. Eijsackers, H. 1995. How to manage accumulated contaminants, in W. Salomons and W.M. Stigliani, Eds., Biogeodynamics of Pollutants in Soils and Sediments, Springer-Verlag, Berlin, pp. 309–329. Eijsackers, H. and P. Doelman. 2000. Using natural cleaning processes in the river ecosystem: a new approach to environmental river management. Does natural attenuation out balance the risks of organic and inorganic contaminants in a river ecosystem? in A.J.M. Smits, P.H. Nienhuis, and R.S.E.W. Leuven, Eds., New Approaches to River Management, Backhuys Publishers, Leiden, the Netherlands, pp. 307–328.

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Eijsackers, H., C.A.M. Van Gestel, S. de Jonge, B. Muijs, and S. Slijkerman. 2001. Polycyclic aromatic hydrocarbon-polluted dredged peat sediments and earthworms: a mutual interference, Ecotoxicology, 10:35–50. Everts, J.W., Y. Eys, M. Ruys, J. Pijnenburg, H. Visser, and R. Luttik. 1993. Assessing the risk of biomagnification: a physiological approach, Sci. Total Environ., Suppl. 1993:1501–1506. Fischer, E. and L. Molnar. 1997. The effect of metal chlorides on growth and reproduction of Eisenia fetida (Oligochaeta Lumbricidae), Soil Biol. Biochem., 29:667–670. Fitzgerald, D., R.P. Lanno, A. Farwell, and D.G. Dixon. 1997. Critical body residues (CBRs) in the assessment of pentachlorophenol toxicity to three earthworm species in artificial soil, Soil Biol. Biochem., 29:685–688. Gorree, M. and W.L.M. Tamis. 1992. BIOMAG-2: Risk Assessment of Soil Pollution for Terrestrial Vertebrates; Adaptation and Evaluation [in Dutch], Centre Environmental Studies, Leiden Report 97, the Netherlands. Goven, A.J., S.C. Chen, L.C. Fitzpatrick, and B.J. Venables. 1994. Lysozyme activity in earthworms (L. terrestris) coelomic fluid and coelomocytes — enzyme array for immunotoxicity of xenobiotics, Environ. Toxicol. Chem., 13:607–613. Goven, A.J., G.S. Eyambe, L.C. Fitzpatrick, B.J. Venables, and E.L. Cooper. 1993. Cellular biomarkers for measuring toxicity of xenobiotics — effects of polychlorinated biphenyls on earthworm L. terrestris, Environ. Toxicol. Chem., 12:863–870. Greig-Smith, P.W., H. Becker, P.J. Edwards, and F. Heimbach. 1992. Ecotoxicology of Earthworms, Intercept, Andover, U.K. Haque, A. and W. Ebing. 1983. Toxicity determination of pesticides to earthworms in the soil substrate, Z. Pflanzenkrankheiten Pflanzenschutz, 90:395–408. Heimbach, F. 1985. Comparison of laboratory methods using E. fetida and L. terrestris for the assessment of the hazard of chemicals to earthworms, Z. Pflanzenkrankheiten Pflanzenschutz, 92:186–193. Heimbach, F. 1992a. Correlation between data from laboratory and field tests for investigating the toxicity of pesticides to earthworms, Soil Biol. Biochem., 24:1749–1753. Heimbach, F. 1992b. Effects of pesticides on earthworm populations: comparison of results from laboratory and field tests, in P.W. Greig Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Ecotoxicology of Earthworms, Intercept, Andover, U.K., pp. 100–108. Heimbach, F. 1994. Use of laboratory toxicity tests for the hazard assessment of chemicals to earthworms representing the soil fauna, in H.J.P. Eijsackers and T. Hamers, Eds., Integrated Soil and Sediment Research: A Basis for Proper Protection, Kluwer, Dordrecht, the Netherlands, pp. 299–302. Heimbach, H. 1997. Field tests on side effects of pesticides on earthworms: influence of plot size and cultivation practices, Soil Biol. Biochem., 29:671 – 676. Holmstrup, M., B.F. Petersen, and M.M. Larsen. 1998. Combined effects of Cu, desiccation, and frost on the viability of earthworm cocoons, Environ. Toxicol. Chem., 17:897–901. Hyvönen, R., S. Andersson, M. Clarholm, and T. Persson. 1994. Effects of lumbricids and enchytraeids on nematodes in limed and unlimed coniferous mor humus, Biol. Fertil. Soils, 17:201–205. Jager, T. 1998. Mechanistic approach for estimating bioconcentration of organic chemicals in earthworms (Oligochaeta), Environ. Toxicol. Chem., 17:2080–2090. Jager, T, R. Baerselman, E. Dijkman, A.C. de Groot, E.A. Hogendoorn, A. de Jong, J.A.W. Kruitbosch, and W.J.G.M. Peijnenburg. 2003a. Availability of polycyclic aromatic hydrocarbons to earthworms (Eisenia andrei, Oligochaeta) in Field-Polluted Soils and Soil-Sediment Mixtures, Environ. Toxicol. Chem., 22: 767–775. Jager, T., R.H.J.L. Fleuren, W. Roelofs, and A.C. de Groot. 2003b. Feeding activity of the earthworm Eisenia andrei in artificial soil, Soil Biol. Biochem., 35:313–322. Jager, T., P.A. Sanchez, B. Muijs, E. van der Velde, and L. Posthuma. 2000. Toxicokinetics of polycyclic aromatic hydrocarbons in Eisenia andrei (Oligochaeta) using spiked soil, Environ. Toxicol. Chem., 19:953–961. Janssen, R.P.T., W.J.G.M. Peijnenburg, L. Posthuma, and M.A.T.G. van der Hoop. 1997a. Equilibrium partitioning of heavy metals in Dutch field soils. 1. Relationships between metal partitioning coefficients and soil characteristics, Environ. Toxicol. Chem., 16:2470–2478. Janssen, R.P.T., L. Posthuma, R. Baerselman, H.A. den Hollander, R.P.M. van Veen, and W.J.G.M. Peijnenburg. 1997b. Equilibrium partitioning of heavy metals in Dutch field soils. 12. Prediction of metal accumulation in earthworms, Environ. Toxicol. Chem., 16:2479–2488.

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Jepson, P.C. 1993. Ecological insights into risk analysis: the side effects of pesticides as a case study, Sci. Total Environ., Suppl. 1993:1547–1566. Jones, A. and A.D.M. Hart. 1998. Comparison of laboratory toxicity tests for pesticides with field effects on earthworm populations: a review, in S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., Advances in Earthworm Ecotoxicology, SETAC Press, Pensacola, FL, pp. 247–267. Jones, C.G., J.H. Lawton, and M. Sachak. 1994. Organisms as ecosystem engineers, Oikos, 69:373–386. Klok, C. 2000. A Quest for the Role of Habitat Quality in Nature Conservation, Ph.D. thesis, University of Amsterdam, the Netherlands, p. 127. Klok, C., A.M. de Roos, H.M. Baveco, J.C.Y. Marinissen, and W. Ma. 1997. Modeling population dynamics of the earthworm Lumbricus rubellus. I. Identification of most sensitive life history parameters, Soil Biol. Biochem., 29:287–294. Knacker, T., C.A.M. Van Gestel, S.E. Jones, A.M.V.M. Soares, H.-J. Schallnass, B. Förstner, and C.A. Edwards. 2004. Ring-testing and field-validation of a terrestrial model ecosystem (TME) — an instrument for testing potentially harmful substances: conceptual approach and study design, Ecotoxicology, 13:5–23. Knoop, J. and Th. Traas. 1993. Acidification, changing land use, and cadmium mobilization: a modelling approach, in G.R.B ter Meulen, W.M. Stigliani, W. Salomons, E.M. Bridges, and A.C. Imeson, Eds.. Chemical Time Bombs. Proceedings European State-of-the-Art Conference on Delayed Effects of Chemicals in Soils and Sediments, Veldhoven, Foundation for Ecodevelopment, Hoofddorp, the Netherlands, pp. 107–118. Kratz, W. and R. Pöhhacker. 1994. The development of soil ecotoxicity test systems with lumbricids to assess sublethal and lethal effects, in M. Donkers, H. Eijsackers, and F. Heimbach, Eds., Ecotoxicology of Soil Organisms, Lewis Publishers, Chelsea, MI, pp. 263–272. Kula, C. 1994. A prolonged laboratory test on sublethal effects of pesticides on Eisenia fetida, in M. Donkers, H. Eijsackers, and F. Heimbach, Eds., Ecotoxicology of Soil Organisms, Lewis Publishers, Chelsea, MI, pp. 257–262. Kula, H. 1992. Measuring effects of pesticides on earthworms in the field: test design and sampling methods, in P.W. Greig Smith, H. Becker, P.J. Edwards, and F. Heimbach, Eds., Ecotoxicology of Earthworms, Intercept, Andover, U.K., pp. 90–99. Lanno, R.P., G.L. Stephenson, and C.D. Wren. 1997. The use of acute lethality thresholds in assessing the toxicity of four chemicals to the earthworm Lumbricus terrestris in three natural soils, Soil Biol. Biochem., 29:689–692. Larink, O. and H. Kula. 1994. Development and Standardization of Acute and Sublethal Laboratory Test Methods with Different Earthworm Species, SECOFASE 2nd Technical Report, NERL, Silkeborg, Denmark, pp. 19–24. Lee, K.E. 1985. Earthworms, Their Ecology and Relationships with Soils and Land Use, Academic Press, Nort Ryde, Australia. Løkke, H. and C.A.M. Van Gestel. 1998. Handbook of Soil Invertebrate Toxicity Tests, Wiley, Chichester, U.K., p. 279. Loonen, H. 1994. Bioavailability of Chlorinated Dioxins and Furans in the Aquatic Environment, Ph.D. thesis, University of Amsterdam, the Netherlands. Luttik, R., C.A.F.M. Romijn, and J.H. Canton. 1993. Presentation of a general algorithm to include secondary poisoning in effect assessment, Sci. Total Environ., Suppl., 1993:1491–1500. Ma, W.C. 1993. Speciation of Heavy Metals in Relation to Their Availability to Soil Macrobiota (Earthworms), Final Report Project C4-12, NISRP, Wageningen, the Netherlands. Ma, W.C. 1994. Methodological principles of using small mammals for ecological hazard assessment of chemical soil pollution, with examples on cadmium and lead, in M.H. Donkers, H. Eijsackers, and F. Heimbach, Eds., Ecotoxicology of Soil Organisms, Lewis Publishers, Boca Raton, FL, pp. 357–372. Ma, W.C. and J. Bodt. 1993. Difference in toxicity of the insecticide chlorpyriphos to six species of earthworms (Oligochaeta, Lumbricidae) in standardized soil tests, Bull. Environ. Contam. Toxicol., 50:864–870. Ma, W.C. and H. Eijsackers. 1989. The influence of substrate toxicity on soil fauna return in reclaimed land, in J.D. Majer, Ed., Animals in Primary Succession: The Role of Fauna in Reclaimed Land, Cambridge University Press, Cambridge, U.K., pp. 1–23. Ma, W.C., H. Siepel, and J.H. Faber. 1997. Soil Contamination in River Floodplains: A Threat to the Terrestrial Fauna of Macroinvertebrates? IBN report; published as EHR Report 72, RIZA, Leleystad, the Netherlands.

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Ma, W.C. and H. Van der Voet. 1993. A risk-assessment model for toxic exposure of small mammalian carnivores to cadmium in contaminated natural environments, Sci. Total Environ., Suppl. 1993:1701–1714. Marinissen, J.C.Y. and P.C. de Ruiter. 1993. Contribution of earthworms to carbon and nitrogen cycling in agro-ecosystems, Agric. Ecosyst. Environ., 47:59–74. Marinussen, M.P.J.C. and S.E.A.T.M. Van der Zee. 1997. Cu uptake by L. rubellus as affected by total amount of Cu in soil, soil pH and moisture and soil heterogeneity, Soil Biol. Biochem., 29:641–648. Mather, G. and O. Christensen. 1988. Surface movements of earthworms in agricultural land, Pedobiologia, 32:399–405. Mather, J.G. and O. Christensen. 1998. Earthworm surface migration in the field: influence of pesticides using benomyl as test chemical, in S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., Advances in Earthworm Ecotoxicology, SETAC Press, Pensacola, FL, pp. 327–340. Menzie, C.A., W. van der Schalie, and R. Landy. 1993. Lessons Learned from Ecological Risk Assessment Studies in the U.S, abstract book, SETAC, Lisbon, Portugal, p. 41. Morgan, A.J. and M. O’Reilly. 2002. The manganese relationships of epigeic and endogeic earthworm species living in soils associated with disused manganese mines, in Proceedings Sixth International Earthworm Ecology Symposium, Cardiff, in press. Morgan, J.E. and A.J. Morgan. 1993. Seasonal changes in the tissue-metal (Cd, Zn and Pb) concentration in two ecophysiologically dissimilar earthworm species — pollution-monitoring implications, Environ. Pollut., 82:1–7. Nahmani, J., P. Lavelle, and F. van Oort. 2002. Effects of polymetallic pollution on earthworm community (Mortagen, North of France), in Proceedings Sixth International Earthworm Ecology Symposium, Cardiff, in press. Neuhauser, E.F., P.R. Durkin, M.R. Malecki, and M. Anatra. 1986. Comparative toxicity of ten organic chemicals to four earthworm species, Comp. Biochem. Physiol., 83C:197–216. Nieboer, E. and D.H.S. Richardson. 1980. The replacement of the nondescript term “heavy metals” by a biologically and chemically significant classification of metal ions, Environ. Pollut., 1B:3–26. Noppert, F, J.W. Dogger, F. Balk, and A.J.M. Smits. 1994. Secondary poisoning in terrestrial food chain; a probabilistic approach, in H.J.P. Eijsackers and T. Hamers, Eds., Integrated Soil and Sediment Research: A Basis for Proper Protection, Kluwer, Dordrecht, the Netherlands, pp. 303–307. Posthuma, L., J. Notenboom, A.C. de Groot, and W.J.G.M. Peijnenburg. 1998. Soil acidity as major determinant of zinc partitioning and zinc uptake in two oligochaete worms exposed in contaminated field soils, in S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., Advances in Earthworm Ecotoxicology, SETAC Press, Pensacola, FL, pp. 111–127. Posthuma, L., L. Weltje, F. Mogo, and R. Baerselman. 1993. Combinatietoxiciteit van Zware Metalen bij de Regenworm Eisenia andrei, poster presented at the National Symposium NISRP 1993, Wageningen, the Netherlands. Reinecke, A.J. and S.A. Reinecke. 2002. Avoidance response in earthworms for assessing environmental contamination, in Proceedings Sixth International Earthworm Ecology Symposium, Cardiff, in press. Reinecke, S.A. and A.J. Reinecke. 1997. The influence of cadmium, lead and manganese on spermatozoa and reproduction of Eisenia fetida (Oligochaeta), Soil Biol. Biochem., 29:737–742. Romijn, C.A.F.M., R. Luttik, W. Slooff, and J.H. Canton. 1991. Presentation and Analysis of a General Algorithm for Effect-Assessment on Secondary Poisoning. Terrestrial Food Chains, RIVM Report 679102007, RIVM, Bilthoven, the Netherlands. Rozen, A. and L. Mazur. 1997. Influence of different levels of traffic pollution on haemoglobin content in earthworm Lumbricus terrestris, Soil Biol. Biochem., 29:705–712. Rundgren, S. 1995. Earthworms and soil remediation. Liming of acidic coniferous forest soils in Southern Sweden, Pedobiologia, 38:519–529. Scott-Fordsman, J.J. and J.M. Weeks. 1998. Review of selected biomarkers in earthworms, in S. Sheppard, J. Bembridge, M. Holmstrup, and L. Posthuma, Eds., Advances in Earthworm Ecotoxicology, SETAC Press, Pensacola, FL, pp. 173–189. Slimak, K.M. 1997. Avoidance response as a sublethal effect of pesticides on Lumbricus terrestris (Oligochaeta), Soil Biol. Biochem., 29:713–716. Springett, J. and R.A.J. Gray. 1992. Effect of repeated low doses of biocides on the earthworm Aporrectodea caliginosa in laboratory cultures, Soil Biol. Biochem., 24:1739–1744.

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Spurgeon, D.J. and S.P. Hopkin. 1999. Seasonal variation in the abundance, biomass and biodiversity of earthworms in soils contaminated with metal emissions from a primary smelter works, J. Appl. Ecol., 36:173–183. Spurgeon, D.J., S.P. Hopkin, and D.T. Jones. 1994. Effects of cadmium, copper, lead and zinc on growth, reproduction and survival of the earthworm Eisenia fetida (Savigny): assessing the environmental impact of point-source metal contamination in terrestrial ecosystems, Environ. Pollut., 84:123–130. Stenersen, J., E. Brekke, and F. Engelstadt. 1992. Earthworms for toxicity testing; species differences in response towards cholinesterase inhibiting insecticides, Soil Biol. Biochem., 24:1761–1764. Stephenson, G.L., C.D. Wren, I.C.J. Middelraad, and J. Warner. 1997. The exposure of earth-worms (Lumbricus terrestris) to diazinon and risk assessment to passerine birds, Soil Biol. Biochem., 29:717–720. Tamis, W.L.M., H.A. Udo de Haes, and A.J. Schouten. 1994. Ecologically recovery of some thermally and biologically cleaned field soils, in H.J.P. Eijsackers and T. Hamers, Eds., Integrated Soil and Sediment Research: A Basis for Proper Protection, Kluwer, Dordrecht, the Netherlands, pp. 341–344. Tarrant, K.A., S.A. Field, S.D. Langton, and A.D.M. Hart. 1997. Effects on earthworm populations of reducing pesticide use, Soil Biol. Biochem., 29:657–662. Tomlin, A.D., R. Protz, R.R. Martin, D.C. MacCabe, and R.J. Lagace. 1993. Relationships amongst organic matter content, heavy metal concentrations, earthworm activity and soil microfabric on a sewage sludge disposal site, Geoderma, 57:89–103. Traas, Th. and T. Aldenberg. 1992. CATS-1: A Model for Predicting Contaminant Accumulation in a Meadow Ecosystem. The Case of Cadmium, RIVM Report 719103001, RIVM, Bilthoven, the Netherlands. Van Gestel, C.A.M. 1992. Earthworms in Ecotoxicology, Ph.D. thesis, University of Utrecht, the Netherlands. Van Gestel C.A.M., E.M. Dirven-van Breemen, and R. Baerselman. 1993a. Accumulation and elimination of cadmium, chromium and zinc and effects on growth and reproduction in Eisenia andrei (Oligochaeta, Lumbricidae), Sci. Total Environ., Suppl. 1993:585–597. Van Gestel, C.A.M., E.M. Dirven-van Breemen, and J.W. Kamerman. 1993b. The influence of soil clean up on the availability of heavy metals for earthworms and plants, in H.J.P. Eijsackers and T. Hamers, Eds., Integrated Soil and Sediment Research: A Basis for Proper Protection, Kluwer, Dordrecht, the Netherlands, pp. 345–348. Van Gestel, C.A.M., L. Hezen, E.M. Dirven-van Breemen, and J.W. Kamerman. 2002. Influence of soil remediation techniques on the bioavailability of heavy metals, in G.I. Sunahara, A.Y. Renoux, C. Thellen, C.L. Gaudet, and A. Pilon, Eds., Environmental Analysis of Contaminated Sites, Wiley, Chichester, U.K., pp. 361–388. Van Gestel, C.A.M., C.D. Léon, and N.M. Van Straalen. 1997. Evaluation of soil fauna ecotoxicity tests regarding their use in risk assessment, in J. Tarradellas, G. Bitton, and D. Rossel, Eds., Soil Ecotoxicology, CRC Press, Boca Raton, FL, pp. 291–317. Van Gestel, C.A.M. and W.-C. Ma. 1993. Development of QSAR’s in soil ecotoxicology: earthworm toxicity and soil sorption of chlorophenols, chlorobenzenes and chloroanilines, Water Air Soil Pollut., 69:265–276. Van Gestel, C.A.M., W.A. van Dis, E.M. van Breemen, and P.M. Sparenburg. 1989. Development of a standardized toxicity test with the earthworm species Eisenia fetia andrei using copper, pentachlorophenol, and 2,4-dichloroaniline, Ecotoxicol. Environ. Saf., 18:305–312. Van Gestel, C.A.M. and N.M. van Straalen. 1994. Ecotoxicological test systems for terrestrial invertebrates, in M. Donkers, H. Eijsackers, and F. Heimbach, Eds., Ecotoxicology of Soil Organisms, Lewis Publishers, Chelsea, MI, pp. 205–228. Van Gestel, C.A.M. and J.M. Weeks. 2004. Future recommendations of the Third International Workshop on Earthworm Ecotoxicology, Aarhus, Denmark (August 2001), Ecotoxicol. Environ. Safety, 57:100–119. Vijver, M.G., J.P.M. Vink, C.J.H. Miermans, and C.A.M. Van Gestel. 2003. Oral sealing using a glue; a new method to distinguish between intestinal and dermal uptake in earthworms, Soil Biol. Biochem., 35:125–132. Wentsel, R.S. and M.A. Guelta. 1988. Avoidance of brass powder-contaminated soil by the earthworm Lumbricus terrestris, Environ. Toxicol. Chem., 7:241–243. Yeardly, R.B., J.M. Lazorchah, and L.C. Gast. 1996. The potential of an earthworm avoidance test for evaluation of hazardous waste sites, Environ. Toxicol. Chem., 15:1532–1537.

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Part IX Earthworms in Waste Management

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Use of Earthworms in the 18 The Breakdown of Organic Wastes to Produce Vermicomposts and Animal Feed Protein Clive A. Edwards and Norman Q. Arancon Soil Ecology Laboratory, The Ohio State University, Columbus, OH, U.S.A.

CONTENTS Introduction ....................................................................................................................................346 Breakdown of Sewage Wastes by Earthworms .............................................................................347 Breakdown of Animal, Vegetable, and Urban Industrial Organic Wastes by Earthworms ..........349 Species of Earthworms Suitable for Processing Organic Wastes .................................................351 Biology and Ecology of Suitable Earthworm Species...........................................................351 Food and Environmental Requirements of Different Earthworm Species that Process Organic Wastes ......................................................................................................353 Eisenia fetida (Savigny) and Eisenia andrei, Bouché..................................................354 Eudrilus eugeniae (Kinberg).........................................................................................354 Perionyx excavatus (Perrier) .........................................................................................355 Dendrobaena veneta (Rosa)..........................................................................................355 Polypheretima elongata (Erseus) ..................................................................................355 Lumbricus rubellus (Hoffmeister) ................................................................................355 Vermicomposts as Plant Growth Media and Soil Amendments ...................................................355 Characteristics of Vermicomposts Produced by Earthworms from Organic Wastes.............355 Growth of Plants in Vermicomposts.......................................................................................357 Effects of Vermicomposts on Plants: The Incidence of Plant Pathogens, Plant-Parasitic Nematodes, and Arthropod Pests ..............................................................................................362 Plant Diseases .........................................................................................................................362 Insect Pests..............................................................................................................................363 Plant-Parasitic Nematodes ......................................................................................................364 Earthworms as a Source of Protein for Animal Feed ...................................................................365 Nutrient Value of Earthworms as Animal Feed .....................................................................365 Production of Earthworms for Animal Feed in Animal, Vegetable, and Industrial Wastes .................................................................................................................365 Practical Production of Earthworm Feed Protein for Animals..............................................366 The Value of Worm Protein as Feed for Fish, Poultry, and Pigs ..........................................366 Fish Feeding Trials........................................................................................................367 Chicken Feeding Trials .................................................................................................367 Pig Feeding Trials .........................................................................................................367

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Economic Potential of the Production of Earthworm Protein for Animal Feed ...................367 Methods of Processing Organic Wastes with Earthworms ...........................................................368 Low-Cost Floor Beds or Windrows........................................................................................368 Containers or Box Systems ....................................................................................................368 Domestic Systems...................................................................................................................369 Earthworm Toilets...................................................................................................................369 Wedge Vermicomposting Systems..........................................................................................369 Gantry-Fed Vermicomposting Beds........................................................................................369 Continuous Flow Automated Vermicomposting Reactors .....................................................369 Complete Recycling Vermicomposting System .....................................................................370 Commercialization and Economics of Vermicomposting Systems .......................................370 References ......................................................................................................................................371

INTRODUCTION The importance of earthworms in the breakdown of organic matter and the release of the nutrients that it contains has been known for a long time (Darwin 1881). It has been demonstrated clearly that some species of earthworms are specialized to live specifically in decaying organic matter and can degrade it into fine particulate materials, rich in available nutrients, with considerable commercial potential as plant growth media soil amendments (Edwards and Bohlen 1996). For instance, earthworms are able to process sewage sludges and solids from wastewater (Neuhauser et al. 1988); brewery wastes (Butt 1993); processed potato wastes (Edwards 1983); waste from the paper industries (Butt 1993); wastes from supermarkets and restaurants (Edwards 1995a,b); animal wastes from poultry, pigs, cattle, sheep, goats, horses, and rabbits (Edwards et al. 1985; Edwards 1988); as well as horticultural residues from dead plants, yard wastes (Edwards 1995a,b), and wastes from the mushroom industry (Edwards 1988). For many years, certain species of earthworms have been bred for sale as fish bait in a wide diversity of organic wastes. Since 1978, interest in possible ways of using similar methods for processing organic wastes using earthworms to produce valuable soil additives and protein for animal feed has been expanding rapidly. This has resulted in a series of conferences aimed at reviewing and promoting such processes and systems. The first of these conferences, Utilization of Soil Organisms in Sludge Management (Hartenstein 1978), which was held in Syracuse, NY, focused on the processing of sewage sludges and biosolids by earthworms, as did the second conference, Workshop on the Role of Earthworms in the Stabilization of Organic Residues (Appelhof 1981), which was held in Kalamazoo, MI. Both conferences were attended by leading earthworm scientists. These conferences were followed by the International Symposium on Agricultural Prospects in Earthworm Farming (Tomati and Grappelli 1984), which was organized in Rome, Italy. The largest, Symposium on the Use of Earthworms in Waste Management and Environmental Management (Edwards and Neuhauser 1988), was in Cambridge, U.K., and had about 400 participants. In addition, sessions relating to earthworms and waste management have been held at the International Earthworm Symposia in Bologna, Italy, in 1985 (Pagliai and Omodeo 1987); Avignon, France, in 1990 (Kretzschmar 1992); Columbus, OH, in 1994 (Edwards 1997); Vigo, Spain, in 1998 (Diaz Cosin et al. 1999); and Cardiff, Wales, in 2002 (A.J. Morgan 2003). These are attended by increasingly larger numbers of earthworm scientists. Research into vermicomposting and commercial projects has been developed in many countries, including England, France, the Netherlands, Germany, Italy, Spain, Poland, the United States, Cuba, Mexico, Japan, the Philippines, India and other parts of Southeast Asia, Australia, New Zealand, Cuba, the Bahamas, and many countries in South America. Research in the United States at the State University of New York (SUNY) on earthworms in waste breakdown concentrated initially on the utilization of sewage sludges and solids to produce vermicomposts (Neuhauser et al. 1988) © 2004 by CRC Press LLC

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but later expanded into producing vermicomposts from food, brewery, paper, and animal wastes. Research in the U.K. at Rothamsted Experimental Station, Harpenden, focused initially on processing animal and vegetable wastes (Edwards 1988) but now has expanded to vermicomposting other organic wastes. The current leading research programs into vermiculture and vermicomposting are at the Soil Ecology Laboratory at The Ohio State University (OSU) in Columbus; at the University of Vigo in Spain, led by Jorge Dominguez (see Chapter 20 this volume); at the University of Agricultural Sciences in Bangalore, India, led by Dr. Radha Kale (see Chapter 19 this volume); and at the Instituto de Ecologia, Mexico, led by Dr. Isabelle Barois and Dr. Aranda. As public interest in vermiculture and vermicomposting increases, so does interest by farmers and commercial organizations, and U.S. journals such as Biocycle, Compost Science and Utilization, and Bioresource Technology are beginning to publish many articles on vermiculture; there are sessions at waste management and composting meetings that frequently address vermicomposting issues. Vermiculture newsletters, such as “Casting Call” and “Worm Digest,” published in the United States proliferate and attract a broad readership. There is also considerable interest in vermicomposting domestic organic garbage and food waste from schools in the United States, which is promoted by Mary Appelhof’s best-selling book, Worms Eat My Garbage (1997). At the same time, earthworm growers associations have proliferated in New Zealand, Australia, the United States, and Europe, as have Web sites, which provide ready access to vermiculture information and publication (Edwards and Steele 1997).

BREAKDOWN OF SEWAGE WASTES BY EARTHWORMS Sewage sludges and biosolids were two of the first organic wastes to attract interest and funding for research into vermiculture at SUNY Syracuse in the late 1970s (Hartenstein 1978). Quite early, they showed, on a laboratory scale, that aerobic sewage sludges can be readily ingested by earthworms such as Eisenia fetida and egested as finely divided casts; in the process, the sludge is decomposed and stabilized (i.e., rendered innocuous) much faster than non-earthwormingested sludge, probably because of the dramatic increases of microbial populations in the casts resulting from earthworm activity. During this process, relative to non-earthworm-ingested sludge, objectionable odors disappear from the wastes very quickly, and there is a rapid reduction in populations of human pathogenic microorganisms such as Salmonella enteriditis, Escherichia coli, and other Enterobacteriaceae, human viruses, and even helminth ova (Eastman et al. 2001), which has been confirmed by other workers. Although most of the sludges produced in sewage plants are anaerobic and, when fresh, can be toxic to E. fetida, when they are dewatered after becoming aerobic they are readily acceptable to this earthworm species and others (see Chapter 20 this volume). It has been found that mixing sewage sludge with other bulking materials (e.g., garden wastes, paper pulp sludge, or other lignin-rich wastes) before composting the mixtures using earthworms can accelerate the rates of decomposition. During passage through the earthworm gut, there is maceration and mixing of such materials and finely divided materials with high microbial activity in the casts (Dominguez et al. 1999). There are many possibilities for the utilization of garden, paper, and timber mill wastes and other materials as bulking agents for simultaneous disposal of these materials together with municipal sewage sludges and conversion into vermicomposts. The use of earthworms in organic waste or sludge management has been termed vermicomposting or vermistabilization Loehr et al. 1984; Neuhauser et al. 1988), and although there is now a large database, many aspects of vermicomposting still need to be researched, evaluated, and resolved to ensure the consistent success of such a process for a wide range of wastes. Factors involved in vermicomposting sludges about which additional information is needed include the following: (1) how earthworms are affected by sludge characteristics, (2) the comparative

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ability of different species of earthworms to grow and reproduce in sludge, (3) the need for methods of preprocessing sludge to make it acceptable to earthworms, and (4) the effects of mixed earthworm species on rates of sewage sludge breakdown in a vermistabilization process or system. For earthworms to be commercially useful in stabilizing sludges, they must increase the stabilization rate, which can be demonstrated best if the presence of earthworms in sludge causes an increase in the rates of volatile solids reduction. A maximum rate of reduction of volatile solids is a major goal of any sludge stabilization system. Research at SUNY followed two complementary lines: (1) basic studies to identify fundamental factors that affect the performance of the vermistabilization process with sludges and (2) applied studies to determine design and management relationships for sludge systems (Neuhauser et al. 1988). The earthworm E. fetida has been shown to accelerate the rates of destruction of sludge volatile solids in aerobic sludges greatly. This increase in the rate of destruction of volatile solids decreases the probability of putrefaction occurring in the sludge because of creation of anaerobic conditions. The main cause of increased rates of degradation of the sludge is probably the increased aeration and turnover of the waste by the earthworms. A series of experiments determined the most desirable moisture contents for optimal procession of sludges used in vermistabilization units because both excessive and insufficient moisture can have an adverse impact on the growth of earthworms (Edwards 1988). A range of sewage sludges with different moisture contents were tested, using E. fetida, at a temperature of 25°C (Neuhauser et al. 1988). Earthworm growth at the lower and higher total solids levels (6.3 to 7.9% and 17.9 to 25.1% solids) differed statistically at the p ≤ .01 probability level from the earthworm growth rates in the middle range of solids (9.3 to 15.9% solids). This indicated that optimum earthworm growth occurred over a range of total solids in the media on a wet basis ranging from about 9 to 16% (i.e., 84 to 91% moisture contents). However, more liquid wastes than these can be used in the vermistabilization process as long as they are added slowly to the system and the liquid can drain readily from the sludge, the organics are retained, and aerobic conditions are maintained. The desirability of maintaining the total solids content of the media in the range quoted above does not mean that wetter sludge slurries cannot be processed with earthworms. In fact, in practice it might be necessary to add moisture to wastes in a vermistabilization process if the media should become too dry. It is also important to correlate rates of earthworm growth with the age of the sludge, that is, the time after the sludge was removed from the aerobic reactor and dewatered (Neuhauser et al. 1988). As sludge ages, its nutritive value to earthworms decreases rapidly, whereas the ash content of the sludge increases with time, which is a further indication of sludge stabilization. The practical feasibility of using earthworms to stabilize wastewater treatment sludges was investigated (Loehr et al. 1984). These workers concluded that a liquid sludge vermistabilization process was feasible and provided data on the rates of stabilization and the physical and chemical characteristics of the residual stabilized solids resulting from vermistabilization, a conclusion that was supported by Pincince et al. (1980). There have been various practical attempts to utilize earthworms as a main component of sewage breakdown in the United States (Pincince et al. 1980; Green and Penton 1981; Loehr et al. 1984), in the U.K. (Edwards and Neuhauser 1988), and in Italy (Tomati and Grappelli 1984). However, full-scale, long-term successful vermiprocessing of sewage sludge has not yet been achieved. The potential of commercial operations was reviewed by Hartenstein (1978), Appelhof (1981), and Edwards and Neuhauser (1988), who all concluded that a major obstacle in the systems to be overcome is the possibility of exposure of the earthworms used in vermistabilization to toxic chemicals that might enter the wastewater stream accidentally. However, the use of biosolids as a substrate for vermicomposting is well established and holds considerable potential now that it is known that human pathogens are eliminated during vermicomposting (Eastman et al. 2001).

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BREAKDOWN OF ANIMAL, VEGETABLE, AND URBAN INDUSTRIAL ORGANIC WASTES BY EARTHWORMS There has been extensive research in the U.K. into the use of earthworms to break down various animal manures, such as pig and cattle solids and slurries; wastes from laying chickens, broilers, turkeys, and ducks; horse manure; and rabbit droppings (Edwards 1983). This research at Rothamsted Experimental Station, Harpenden, Herts, was extended later to studies using earthworms to break down vegetable wastes, including those from the mushroom industry, processed potato industry, brewery industry, paper pulp, food wastes from supermarkets, and yard wastes, including grass clippings and leaves. This work began as a fundamental, interdisciplinary research program on a laboratory and greenhouse scale, from 1981 to 1985, and involved many biologists, engineers, economists, and commercial organizations. It was extended to large-scale developmental research on a Þeld scale, development of appropriate engineering technology, and eventual commercialization. This U.K. research had two main aims. The Þrst was to convert animal and vegetable wastes into useful soil amendments that could be added to agricultural land to improve soil structure and fertility or be marketed for horticultural use as a bedding plant growth medium or a component of commercial potting growth media. The second was to assess the potential of harvesting earthworms from the earthworm-worked wastes and processing them into a protein supplement to feed Þsh, poultry, or pigs. To accomplish these aims, a complex network of interdisciplinary collaborative research was initiated that involved six research stations, six university or college departments, and eight commercial organizations. The following main areas of research and development were involved: 1. A laboratory screening program into the suitability of Þve different earthworm species in processing 10 different organic waste materials. The aim was to assess their biological and economic potentials and study the biology and ecology of these worms in different organic wastes (Edwards 1988). 2. Studies to assess the source of nutrition of those species of earthworms that live in organic wastes; to identify the relative importance of bacteria, protozoa, fungi, and nematodes in their diet; and to conÞrm which microorganisms are essential to their survival (Edwards and Fletcher 1988; Morgan 1988). 3. Evaluation of the rates of conversion of different organic wastes into earthworm biomass in relation to type of organic wastes, earthworm stocking rates, and effects of environmental factors, especially moisture and temperature (Edwards et al. 1985) (Figure 18.1).

1.0 0.8 Total biomass 0.6 (g) 0.4 0.2

1:64 1:128

1:16

1:8

1:4

1:32 Stocking rate: weight worm/weight waste

FIGURE 18.1 Optimum stocking rates with E. fetida for maximum biomass production.

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4. Methods of harvesting and processing of earthworms into animal feed protein, fish, poultry, and pig animal feeding trials and use of earthworms in toxicological tests (Edwards 1985; Edwards and Niederer 1988; Edwards and Bohlen 1992). 5. Development and testing of a range of technologies and systems of production of earthworm protein and vermicomposts (Edwards 1985, 1988, 1995a,b). 6. Studies into the production of plant growth media processed by earthworms from organic wastes, greenhouse plant growth trials, and their utilization on field crops. Amendments of these growth media produce maximum plant growth (Edwards 1983; Edwards et al. 1985; Edwards and Burrows 1988). Five main earthworm species were identified as potentially useful species to break down organic wastes. These were E. fetida (or the closely-related Eisenia andrei), Dendrobaena veneta, and Lumbricus rubellus from temperate regions and Eudrilus eugeniae and Perionyx excavatus from the tropics. The survival, growth, mortality, and reproduction of these species were studied in detail in the laboratory, in a range of organic wastes, including pig, cattle, duck, turkey, poultry, potato, brewery, paper, and activated sewage sludge. All the species tested could grow and survive in a wide range of different organic wastes, but some were much more prolific, others grew rapidly, and yet others attained a large biomass quickly; these were all characters contributing in different ways to the practical usefulness of the earthworms in producing vermicomposts or animal feed protein. However, there were many species-specific differences in the biology and ecology of these earthworms. Most organic wastes can be broken down by these species of earthworms, but some organic wastes have to be pretreated in various ways to make them acceptable to the earthworms, and not all organic wastes will grow earthworms equally well. The characteristics of the different wastes that have been tested are as follows: 1. Cattle manure solids are the easiest animal wastes in which to grow earthworms successfully. Except when they are very fresh, they usually contain no materials unfavorable to the growth of earthworms. Solids can be used, but it is much easier to produce good vermicomposts in solids separated mechanically from slurries of cattle manures, which are rich in nutrients for plant growth; moreover, the liquids can be added back to the solids at a later stage in vermicompost production. 2. Horse manure is an excellent material for growing earthworms and needs very little modification other than maintenance of good environmental factors in the waste. However, the earthworms do not grow as rapidly in horse manure as in pig or cattle wastes. 3. Pig manure solids are probably the most productive organic waste for growing earthworms. If the waste is in the form of a slurry, it is better if the solids are separated in mechanical presses or by sedimentation. Pig wastes tend to contain relatively large amounts of ammonia and inorganic salts, and unless these are washed out, the waste may have to be composted thermophilically for about 2 weeks or longer before to inoculation with earthworms. Pig wastes sometimes have a content of heavy metals, particularly copper (Edwards 1988; Wong and Griffiths 1991). The processed pig waste vermicomposts are high in nutrients for plant growth. 4. Poultry wastes, including chicken, duck, and turkey manures, contain significant amounts of inorganic salts, urea, and ammonia, which may kill earthworms in freshly deposited wastes. However, after removal of these materials through composting, washing, or aging, earthworms grow well in them, and the vermicomposts produced are high in nutrients for the growth of greenhouse or field plants. 5. Potato wastes, usually in the form of peel from the processed potato and frozen potato industries, make ideal growth media for earthworms and need few modifications in terms of moisture content or other kinds of preprocessing (Edwards et al. 1985). © 2004 by CRC Press LLC

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6. Food wastes or restaurant wastes are readily available wastes that are expensive to dispose in landfills or elsewhere and provide a good medium for growing earthworms if some form of bulk material such as paper waste or compost is mixed with them to lower the moisture contents. 7. Paper pulp solids are produced by mechanical separation, pressing, or sedimentation of solids from the production washings. These solids are really excellent material for any growth of earthworms, and there is no need for preprocessing or additives. They are particularly suitable for outdoor windrow processing systems because they form a crust over the bed that minimizes water loss at the same time it allows water to be added. 8. Brewery wastes need no modification, in terms of moisture content, to grow earthworms. Earthworms can process brewery wastes very quickly and grow and multiply rapidly in them. The vermicompost produced from brewery wastes has a good structure and nutrient content. 9. Spent mushroom compost is a good medium for growing earthworms, which are able to break down the straw it contains into small fragments and produce a finely structured material. However, the vermicompost produced may be low in plant nutrients. 10. Urban garden wastes, including grass clippings and tree leaves, are good growth media for earthworms, particularly when they are first macerated and thoroughly mixed before use (Huhta and Haimi 1988; Edwards 1995a,b). However, their production and availability tends to be seasonal, which makes production of a standardized vermicompost year round difficult.

SPECIES OF EARTHWORMS SUITABLE FOR PROCESSING ORGANIC WASTES BIOLOGY

AND

ECOLOGY

OF

SUITABLE EARTHWORM SPECIES

Many species of earthworms have potential for use in organic waste processing, but relatively few have been used on a widespread scale and researched adequately. The species used most commonly include Eisenia fetida and Eisenia andrei (brandling or tiger worms), Eudrilus eugeniae (African night crawler), L. rubellus (red worm), Dendrobaena veneta, Perionyx excavatus, and Perionyx hawayana. The growth of E. fetida, E. eugeniae, P. excavatus, and P. hawayana in sewage sludge has been studied in detail (Neuhauser et al. 1988). These researchers concluded that all these species have a range of optimum temperatures between 15 and 25°C for growth in sewage sludge, and all four species produced most cocoons at 25°C. Over a period of 20 weeks, P. excavatus had the slowest rate of increase in weight, and E. fetida grew slightly slower. Both P. hawayana and E. eugeniae reached their peak biomass in about 10 to 12 weeks, and E. eugeniae, which grew much faster, began to lose weight after 14 weeks. D. veneta increased most in weight and took 16 weeks to achieve maximum weight. P. excavatus produced the largest number of cocoons, and D. veneta, which produced the lowest number of cocoons, did not start producing cocoons for 10 weeks. The other three species, E. fetida, E. eugeniae, and P. hawayana, all produced similar numbers of cocoons, with peak cocoon production occurring after 10 weeks. It is also important to evaluate the number of live young earthworms that emerge from the cocoons of each species. Cocoons from five species of earthworms, D. veneta, E. fetida, E. eugeniae, P. excavatus, and P. hawayana, were collected and allowed to hatch (Edwards 1988). Individual cocoons were kept in organic waste under nonstressed conditions at 25°C and were checked twice per week to determine the number of cocoons that had hatched and the number of earthworm hatchlings that were produced per cocoon. They concluded from their data that E. fetida produced 6 cocoons per earthworm per week (19 young earthworms), D. veneta produced 5 cocoons (19 young earthworms), E. eugeniae produced 11 cocoons (20 young earthworms), P. excavatus © 2004 by CRC Press LLC

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TABLE 18.1 Maximum Reproduction Rate of Earthworms in Animal and Vegetable Wastes: Cocoon Production Species

No. of Cocoons

% Hatch

No. of Hatchlings

Net Reproductive Rates per Week

Eisenia fetida Eudrilus eugeniae Perionyx excavatus Dendrobaena veneta

3.8 3.6 19.5 1.6

83.2 81.0 90.7 81.2

3.3 2.3 1.1 1.1

10.4 6.7 19.4 1.4

TABLE 18.2 Productivity of Earthworms in Animal and Vegetable Waste: Length of Life Cycle Species Eisenia fetida Eudrilus eugeniae Perionyx excavatus Dendrobaena veneta

Time for Cocoons to Hatch (days) 32–73 12–27 16–21 40–126

Time to Sexual Maturity (days)

Time Egg to Maturity (days)

53–76 32–95 28–56 57–86

85–149 43–122 44–71 97–14

produced 24 cocoons (13 young earthworms), and P. hawayana produced 10 cocoons (9.5 young earthworms) per parent earthworm. These researchers also studied the rates of growth of different multispecific combinations of these species in polyculture. Although the total earthworm biomass tended to be greater in polyculture, the results were not clear cut because different species eventually became dominant in different wastes and climatic conditions. Edwards (1988) studied the life cycles and optimal conditions for growth and survival of E. fetida, D. veneta, E. eugeniae, and P. excavatus in animal and vegetable wastes (Table 18.1 and Table 18.2). Each of the four earthworm species differed considerably in terms of response to and tolerance of different temperatures. The optimum temperature for growth of E. fetida was 25°C, with a temperature tolerance from 0 to 35°C. D. veneta had a lower temperature optimum and was less tolerant of extreme temperatures. The optimum temperatures for E. eugeniae and P. excavatus were also about 25°C, but they died at temperatures below 9°C and above 30°C. The optimum temperatures for cocoon production for all species were much lower than for growth (Figure 18.2). These four species also differed in their optimum moisture requirements from those of E. fetida, but the differences were not great (Figure 18.3). The range over which the earthworms grew optimally was quite narrow, with optimal growth at 80 to 55% moisture content, with considerable decreases in growth at moisture contents of 70 and 90%. However, D. veneta was able to withstand a much wider range of moisture contents than the other species, such as P. excavatus. All four species of earthworms were very sensitive to ammonia and did not survive long in organic wastes containing much ammonia (e.g., fresh poultry litter). They also died in wastes containing large quantities of inorganic salts. Laboratory experiments showed that both ammonia and inorganic salts have very sharp cutoff points between toxic and nontoxic levels (i.e., <0.5 mg per ram ammonia and <0.5% salts). However, organic wastes that have too much ammonia became acceptable after the ammonia was removed by a period of composting or when both excessive

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1.0

x

x

353

25°C x 20°C 15°C

0.8 Wt (mg)

x

0.6

0.4

x

0.2

x

+

+

+

+

+

10°C

+ 20

40 Days

60

80

FIGURE 18.2 Growth of Eisenia fetida at different temperatures.

x +

0.75

x

x 30°C + 25°C 20°C 15°C

+ Wt (mg)

0.50

x

x +

+ x +

0.25

90

80

70

60

Moisture content of waste (%)

FIGURE 18.3 Growth of Eisenia fetida at different moisture contents.

ammonia and salts were washed out of the waste. Earthworms were relatively tolerant regarding pH, but when given a choice in a pH gradient, they moved toward the more acid material, with an apparent pH preference of 5.0. The optimal conditions for breeding E. fetida are summarized in Table 18.3. These do not differ much from those suitable for the other species.

FOOD AND ENVIRONMENTAL REQUIREMENTS THAT PROCESS ORGANIC WASTES

OF

DIFFERENT EARTHWORM SPECIES

Different earthworm species have quite different environmental requirements for optimal development, growth, and productivity in organic wastes. The characteristics of the more commonly used earthworm species are summarized in this section.

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TABLE 18.3 Optimal Conditions for Breeding Eisenia fetida in Animal and Vegetable Waste Condition

Requirements

Temperature Moisture content Oxygen requirement Ammonia content of waste Salt content of waste pH

15–20°C (limits 4–30°C) 80–90% (limits 60–90%) Aerobicity Low <0.5 mg/g Low <0.5% >5.0 and <9.0

Eisenia fetida (Savigny) and Eisenia andrei, Bouché The species most commonly used for breaking down organic wastes is E. fetida or the very closely related species E. andrei. There are a number of reasons why these species are preferred in vermicomposting all over the world. They are peregrine species that are very common, and many organic wastes became colonized naturally by these species. They have a wide temperature tolerance and can live in organic wastes with a range of moisture contents. They are tough earthworms, readily handled, and in mixed species cultures they usually becomes dominant, so that even when vermicomposting systems begin with other species, they often end dominated by E. fetida (or E. andrei). Although they are so similar, E. fetida and E. andrei (the “tiger worm” and the closely related “red tiger worm”) have been distinguished as separate species, but their morphological characteristics and overall reproductive performances and environmental requirements do not differ significantly (Reinecke and Viljeon 1990). These two species are widely distributed throughout the temperate regions of the world and are the species most commonly used in commercial vermiculture and organic waste reduction systems. Graff (1974), Watanabe and Tsukamoto (1976), Tsukamoto and Watanabe (1977), Hartenstein (1978), Kaplan et al. (1980), Edwards (1988), and Neuhauser et al. (1988) all investigated in detail the productivity, growth, and population biology of E. fetida when the worms were fed on animal manures, or sewage sludge, so there is much critical information on the requirements of this species. In surveys of commercial earthworm farms in the United States and Europe by Edwards (1988) and in Australia, the earthworms, sold under the name L. rubellus, were all E. fetida or E. andrei. Data on the biology, ecology, and environmental requirement of E. fetida are summarized in the previous section. Eudrilus eugeniae (Kinberg) This is a large earthworm native to Africa but commonly found in many other countries, including the United States. It is commonly known as the African night crawler, grows extremely rapidly, and is quite prolific. It is cultured extensively for fish bait in the United States, and under optimum conditions it would seem to be an ideal species for earthworm protein production. Its main disadvantages are its relatively poor temperature tolerance and poor handling capabilities because it is easily damaged and can be difficult to harvest. Eudrilus eugeniae has very high rates of reproduction (Kale and Bano 1991; Edwards 1988) and is capable of decomposing large quantities of organic wastes rapidly (Edwards 1988; Neuhauser et al. 1988; Kale and Bano 1991). However, E. eugeniae has a preference for higher temperatures and cannot tolerate extended periods below 16°C (Viljoen and Reinecke 1992); it does not survive for long below 10°C. Its use in outdoor vermiculture may therefore be limited to tropical and subtropical regions, unless lower winter

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temperatures in the wastes are avoided or controlled. This species is used extensively in the tropics, especially India. Its detailed biology has been studied by Dominguez et al. (2001). Perionyx excavatus (Perrier) This tropical species of earthworm is extremely prolific; it is almost as easy to handle as E. fetida and very easy to harvest. Its main drawback for use under temperate conditions is it inability to withstand temperature conditions below 5°C for long periods, but for tropical conditions it is an ideal species. It has an extremely high reproductive rate (Kale et al. 1982; Edwards et al. 1988; Neuhauser et al. 1988) (Table 18.1 and Table 18.2). It is a very common species in Asia and is used extensively in vermiculture in the Philippines, Australia, and India. Dendrobaena veneta (Rosa) Dendrobaena veneta is a large earthworm with good potential for use in vermiculture, and unlike E. fetida and P. excavatius, it can also survive in soil but is not very prolific although it grows quite rapidly (Edwards 1988). However, it is a suitable species for organic matter breakdown and is specifically useful for transfer and use in agricultural soils to improve soil fertility. Polypheretima elongata (Erseus) Polypheretima elongata has been tested for use in the breakdown of organic solids, including municipal and slaughterhouse wastes; human, poultry, and dairy manures; and mushroom compost in India. A project in India using this species claimed to have a commercially viable facility for the “vermistabilization” of 8 tons of solid wastes per day. They have developed a “vermifilter,” packed with vermicompost and live earthworms, that produces reusable water from sewage sludges, manure slurries, and organic wastewaters from food processing. Polypheretima elongata appears to be restricted to tropical regions and may not survive temperate winters. Lumbricus rubellus (Hoffmeister) Lumbricus rubellus is a species of earthworm common in moister soils as well as in organic matter, particularly those to which animal manure or sewage solids have been applied (Cotton and Curry 1980). It can be used for organic waste breakdown, but it is a relatively slow-growing species, and further research into its potential in vermicomposting is needed before its major adoption for this purpose. It has potential for breeding in organic wastes and transfer to soils to improve soil fertility.

VERMICOMPOSTS AS PLANT GROWTH MEDIA AND SOIL AMENDMENTS Earthworm composts (vermicomposts) can be produced from almost any kind of organic wastes with suitable preprocessing and controlled vermicomposting conditions. Vermicomposts grow plants extremely well, and they can be used as structural additives or amendments for poorer soils to provide nutrients and minimize soil erosion.

CHARACTERISTICS OF VERMICOMPOSTS PRODUCED ORGANIC WASTES

BY

EARTHWORMS

FROM

The f i nal physical structure of the plant growth media or vermicomposts produced from organic wastes depends very much on the type of parent waste from which they were produced. Some organic wastes, such as cattle, pig, and poultry manures from indoor breeding systems, as well as

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TABLE 18.4 Major Plant Nutrient Elements in Earthworm-Processed Animal Waste Element Content (% dry weight) Waste Material

N

P

K

Ca

Mg

Mn

Separated cattle solids Separated pig solids Cattle solids on straw Pig solids on straw Duck solids on straw Chicken solids on shavings Commercial plant growth medium

2.20 2.60 2.50 3.00 2.60 1.80 1.80

0.40 1.70 0.50 1.60 2.90 2.70 0.21

0.90 1.40 2.50 2.40 1.70 2.10 0.48

1.20 3.40 1.55 4.00 9.50 4.80 0.94

0.25 0.55 0.30 0.60 1.00 0.70 2.20

0.02 0.03 0.05 0.05 0.10 0.08 0.92

spent mushroom compost, all contain straw, which takes considerably longer for the earthworms to fragment than the more particulate materials, such as separated animal solids and slurries, brewery wastes, paper pulp, and similar materials. However, the final product from most organic wastes is usually a finely divided, well-stabilized and humified, peatlike material with excellent structure, porosity, aeration, drainage, and moisture-holding capacity and a low C:N ratio (Edwards 1983). Structurally, it has the appearance and many of the characteristics of peat and contains plant nutrients usually in adequate amounts for plant growth. The chemical nutrient contents of vermicomposts differ greatly, depending on the parent material from which they are processed. However, when their nutrient content is compared with that of a commercial plant growth medium, if they are processed from animal wastes, they usually contain most of the necessary mineral elements for plants, although there may sometimes be a deficiency of magnesium, which can be remedied by addition of magnesium sulfate (Table 18.4). An important feature of vermicomposts is that, during the processing of the various organic wastes by earthworms, many of the nutrients that they contain are changed to forms that are more readily taken up by plants, such as nitrate or ammonium nitrogen, exchangeable phosphorus, and soluble potassium, calcium, and magnesium (Table 18.5). Moreover, many

TABLE 18.5 Effect of Earthworm Activity on Nutrients in Organic Wastes Organic Waste Cattle waste Unworked Worm worked Pig waste Unworked Worm worked Potato waste Unworked Worm worked d.m., dry matter.

© 2004 by CRC Press LLC

Nitrate Nitrogen (ppm)

Readily Soluble P (% d.m.)

8.8 259.4

Exchangeable (% d.m.) K

Ca

Mg

0.11 0.18

0.19 0.41

0.35 0.59

0.05 0.08

31.6 110.3

1.05 1.64

1.49 1.76

1.56 2.27

0.45 0.72

74.6 1428.0

0.19 0.22

1.94 3.09

0.91 1.37

0.24 0.34

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vermicomposts produced from animal wastes tend to be on the alkaline side of neutral (pH >7.0), whereas most plants prefer a growth medium slightly on the acid side of neutral (e.g., with a pH of 6.0). The processing of organic wastes by earthworms does not change the pH of the material appreciably, so earthworm-worked wastes benefit from some acidification or from mixing with an acid medium such as peat.

GROWTH

OF

PLANTS

IN

VERMICOMPOSTS

There is very well-substantiated evidence that vermicomposts can promote the growth of plants significantly. For instance, Fosgate and Babb (1972) grew earthworms in cattle wastes and reported that the vermicomposts produced were equal to greenhouse potting mixes for production of flowering plants. Reddy (1988) reported increased growth of Vinca rosea and Oryza sativa after addition of cast material from Pheretima alexandri to soils. Buchanan et al. (1988) suggested that most vermicomposts had more available nutrients than the organic wastes from which they were processed. Edwards (1988) reported that most samples of vermicomposts had relatively high levels of available nitrogen. Handreck (1986) reviewed the utilization of vermicomposts as horticultural potting media and concluded that, although they could supply most of the trace element needs of plants, many vermicomposts may not have sufficient nitrogen to supply all the needs of the plants. However, it seems difficult to justify this conclusion because most organic wastes have excess amounts of nutrients, and usually only a small proportion of these is lost during efficient and rapid vermicomposting (Edwards et al. 1985). It seems probable that the vermicomposts that Handreck tested had been produced very slowly and had lost most of their nutrients through either volatilization or leaching. In extensive greenhouse trials testing the growth of a wide range of plants in a variety of earthworm-worked wastes (Edwards and Burrows 1988), most plants germinated earlier and grew better in vermicomposts than in commercial plant growth media. A wide range of plants were grown successfully in both undiluted worm-worked wastes as well as in a range of dilutions and mixes with bulking materials, including 3:1 or 1:1 ratios of earthworm-worked organic wastes to peat, pine bark, or a Kettering loam soil. Plants that were tested for their rates of growth in vermicomposts by Edwards and Burrows (1988) included vegetables such as aubergines (eggplant), cabbages, peppers, cucumbers, lettuces, radishes, and tomatoes; bedding plants such as Alyssum, Antirrhinum, Aster, Campanula. Calceolaria, Cineraria, Coleus, French marigold, plumose asparagus, sweet peas, Polyanthus, and Salvia (Edwards and Burrows 1988); and ornamental shrubs such as Eleagnus pungens, Cotoneaster conspicua, Pyracantha sp.: Viburnum bodnantense, Chaemaecyparis lavisonia, and Juniperus communis (Scott 1988). After processing, vermicomposts usually have about 75% moisture content and may need some drying before use. If the waste is likely to contain human pathogens or weed seeds, a brief thermophilic precomposting for 3 to 4 days may be advisable as a precaution. Usually, a magnesium sulfate supplement is necessary to rectify magnesium deficiencies that may occur in many animal wastes, and the pH can be adjusted in some way, such as by adding acid peat, to bring the medium to a pH of about 6.0. Such vermicomposts seemed to accelerate the germination and subsequent growth and yields of seedlings of most species of plants tested. A wide range of tests of seedling emergence of pea, lettuce, wheat, cabbage, tomato, and radish were made in small pots and trays using the standard European Economic Community recommended seedling-emergence test. The germination of tomatoes, cabbage, and radish seedlings tended to be significantly better in vermicomposts than in commercial plant growth media and much better than in composted animal wastes with no earthworms. Similarly, the early growth of seedlings of ornamentals up to the stage when they were transplanted into larger pots or outdoors was better in the animal waste vermicomposts mixed with peat than in a commercial plant growth medium.

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The growth of eggplants, dahlias, coleus, ornamental peppers, and polyanthus was also very good in mixtures of vermicomposts with peat or other bulking material and was usually better than in commercial plant growth media. The effects of dilution with a commercial plant growth medium of vermicomposts on the growth of three bush ornamentals at a range of levels of addition were quite dramatic (Scott 1988). When a 50/50 mixture of pig and cattle vermicompost was diluted at a range of levels, ranging from 5 to 10%, with a commercial plant growth medium, all the mixtures, even at the lowest dilutions, improved the growth of Chaemaecyparis lawsoniana, Pyracantha sp., and Viburnum bodnantense, even more than in the recommended commercial medium. Plants grown in all the dilutions and different mixtures tended to grow better than those in the 100% vermicomposts, which had a tendency to dry out more rapidly than the different mixtures. These results, for which even small amounts of vermicomposts had significant effects on plant growth, indicated that the response could not be based only on the nutrient content of the vermicomposts. Vermicomposts were also used as “blocking” materials to grow seedlings for transplanting into the field. Cabbages were grown in machine-compressed blocks made either from a commercial seedling medium or from cattle and pig vermicomposts. The seedlings were transplanted into the field in the same blocks, and their subsequent growth was followed and measured. Those seedlings grown in blocks of vermicomposts produced cabbages almost twice as large at harvest than those grown in the commercial blocking material (Edwards and Burrows 1988). There is an increasing body of evidence from current work in the Soil Ecology Laboratory at OSU that the substitution of vermicomposts into soilless bedding plant media (such as Metro-Mix 360) can increase the germination, growth, flowering, and fruiting of a wide range of greenhouse vegetables and ornamentals, such as tomatoes (Figure 18.4) (Atiyeh et al. 1999, 2000d); peppers (Arancon et al. 2004a); vegetable seedlings (Atiyeh et al. 2000a); marigolds (Atiyeh et al. 2000e, 2001); and other vegetables and ornamentals (Atiyeh et al. 2000b,c), and that amendments to field crops with low application rates of vermicomposts can increase the growth of vegetables such as peppers (Arancon et al. 2004b), fruits such as strawberries (Arancon et al. 2004c) and grapes (Arancon et al. 2004g), and ornamentals such as petunias. All these increases in crop growth in both greenhouse and field experiments were independent of nutrient supplies, which were equalized

Marketable yields (kg/plant)

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FIGURE 18.4 Marketable tomato fruit yields (mean ± SE) produced in standard commercial medium (MetroMix 360) substituted with different concentrations of pig manure vermicompost. Columns with * are significantly different from Metro-Mix 360 control (O% vermicompost) at p ≤ .05 probability. (From Atiyeh et al. 2000. With permission.) © 2004 by CRC Press LLC

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initially between treatments. In all the recent greenhouse and field experiments at OSU, increases in plant growth in response to vermicomposts were in response to quite small application rates, were significant and consistent, and were independent of nutrient supply, so there is a need to identify the mechanism of these effects. A very substantial body of evidence has accumulated demonstrating that microorganisms, including bacteria, fungi, yeasts, actinomycetes, and algae, are capable of producing plant hormones and plant-growth regulating (PGR) substances such as auxins, gibberellins, cytokinins, ethylene, and abscisic acid in appreciable quantities (Arshad and Frankenberger 1993; Frankenberger and Arshad 1995). Many of the microorganisms common in the rhizospheres of plants can produce such PGR substances; for instance, Barea et al. (1976) reported that, of 50 bacterial isolates obtained from the rhizospheres of various plants, 86% could produce auxins, 58% gibberellins, and 90% kinetinlike substances. There have been many studies of the production of PGR substances by mixed microbial populations in soil, but there are relatively few investigations into their availability from soils to plants and their persistence and fate in soils or documentation of their effects on plant growth (Arshad and Frankenberger 1993; Frankenberger and Arshad 1995). A few workers have shown that PGR substances can be taken up by plants from soil in sufficient amounts to influence plant growth. For instance, it was shown that auxins produced by Azospirillum brasilense could affect the growth of graminaceous plants (Kucey 1983). There is increasing evidence that microbially produced gibberellins can influence plant growth and development (Mahmoud et al. 1984; Arshad and Frankenberger 1993). Increased vigor of seedlings has been attributed to microbial production of cytokinins by Arthrobacter and Bacillus spp. in soils (Inbal and Feldman 1982; Jagnow 1987). It has been suggested that earthworms may be important agents that influence the enhanced production of PGR substances through promotion of greatly increased microbial activity during vermicomposting (Nielson 1965; Springett and Syers 1979; Graff and Makeschin 1980; Tomati et al. 1983, 1987, 1988, 1990; Krishnamoorthy and Vajranabhiah 1986); Dell’Agnola and Nardi 1987; Grappelli et al. 1987; Edwards and Burrows 1988; Nardi et al. 1988; Tomati and Galli 1995; Edwards 1998). Because earthworms can increase microbial activity so dramatically, sometimes by orders of magnitude, it is not unreasonable to conclude that earthworm activity might increase the rates of production of PGRs by soil microorganisms significantly to levels that could influence plant growth. The first suggestion that earthworms might produce plant growth regulators was by Gavrilov (1963). This was supported by the first report of the presence of PGR substances in the tissues of Aporrectodea caliginosa, L. rubellus and E. fetida by Nielson (1965), who extracted indole substances from earthworms and reported increases in the growth of peas caused by them. He also extracted a substance that stimulated plant growth from Aporrectodea longa, Lumbricus terrestris, and Dendrobaena rubidus, but his experiments did not exclude the possibility that the PGRs he obtained came from microorganisms living in the earthworm guts and tissues. Graff and Makeschin (1980) tested the effects of substances that had been produced by L. terrestris, A. caliginosa, and E. fetida on the dry matter production of ryegrass. They added liquid eluates from pots containing earthworms to pots containing no earthworms and concluded that plant growth-influencing (PGI) substances were released into the soil by all three species, but they did not speculate further on the nature of these substances. Tomati et al. (1983, 1987, 1988, 1990), Grappelli et al. (1987), and Tomati and Galli (1995) tested vermicomposts produced from organic wastes by the action of earthworms as media for growing ornamental plants and mushrooms. They concluded that the plant growth increases that they obtained in all their experiments were too great to be explained purely on the basis of the nutrient content of the vermicomposts. Moreover, the plant growth changes recorded included stimulation of rooting, dwarfing, earlier flowering, and lengthening of internodes. They compared the growth of Petunia, Begonia, and Coleus species after adding aqueous extracts from vermicompost, with addition of auxins, gibberellins, and cytokinins to soil and concluded there was excellent © 2004 by CRC Press LLC

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evidence of hormonal effects produced by earthworm activity, which was supported by the high levels of cytokinins and auxins they found in the vermicomposts. Krishnamoorthy and Vajranabhiah (1986) showed, in laboratory experiments involving large earthworm populations, that seven species of earthworms could promote the production of cytokinins and auxins in organic wastes very dramatically. They also demonstrated significant positive correlations (r = 0.97) between earthworm populations and the amounts of cytokinins and auxins occurring in 10 different field soils, and they concluded that earthworm activity was linked strongly with PGR production. They reported that auxins and cytokinins produced through earthworm activity could persist in soils for up to 10 weeks but degrade in a few days if exposed to sunlight. The biological activities of humic substances have been investigated extensively (McCarthy et al. 1990; Hayes and Wilson 1997). Studies of the effects of humic substances on plant growth under conditions of adequate mineral nutrition have resulted consistently in positive plant growth effects (Y. Chen and Aviad 1990; Hayes and Wilson 1997). For instance, humic substances increased the dry matter yields of corn and oat seedlings (Albuzio et al. 1994); numbers and lengths of tobacco roots (Mylonas and McCants 1980); dry weights of shoots, roots, and nodules of soybean, peanut, and clover plants (Tan and Tantiwiramanond 1983); and vegetative growth of chicory plants (Valdrighi et al. 1996) and induced shoot and root formation in tropical crops grown in tissue culture (Goenadi and Sudharama 1995). Vermicomposts originating from animal manure, food wastes, sewage sludges, or paper-mill sludges have been reported to contain high levels of humic substances (Canellas et al. 2000; Atiyeh et al. 2002; Arancon et al. 2004f). The biological activities and effects of humic substances derived from earthworm feces on plants have been investigated (Dell’Agnola and Nardi 1987; Nardi et al. 1988; Muscolo et al. 1993). For instance, Dell’Agnola and Nardi (1987) reported hormonelike effects of depolycondensed humic fractions obtained from the feces of the earthworms Allolobophora rosea and Allolobophora caliginosa. Treating carrot cells with humic substances obtained from the feces of the earthworm A. rosea increased cell growth and induced morphological changes similar to those induced by auxins (Muscolo et al. 1996). Work in the Soil Ecology Laboratory at OSU indicates that it seems that vermicomposts, which consist of an amalgamate of humified earthworm feces and organic matter, can stimulate plant growth beyond that produced by mineral nutrients because of the direct or indirect effects of the humic substances in the vermicomposts acting as PGRs; this has been confirmed (Atiyeh et al. 2000a,b,c,d,e; Arancon et al. 2004f). In this work at OSU, typical growth responses after treating plants such as tomatoes with humic substances were for increased growth correlated with increasing concentrations of humic substances but usually with a decrease in growth at higher concentrations of the humic materials (Figure 18.5). This stimulatory effect of humic substances on plant growth at low concentrations has been explained by various theories, the most convincing of which hypothesizes a “direct” hormonal action on the plants together with an indirect action on the metabolism of soil microorganisms, the dynamics of soil nutrients, and soil physical conditions (Cacco and Dell’Angola 1984; Nardi et al. 1988; Casenave de Sanfilippo et al. 1990; Y. Chen and Aviad 1990; Muscolo et al. 1993, 1996, 1999; Albuzio et al. 1994). Laboratory and greenhouse research at OSU has provided new evidence that the effects of earthworm activity on organic matter to produce vermicomposts can lead to the production of water-extractable and base-extractable plant PGI substances in vermicomposts in quantities that could significantly influence plant germination, growth, flowering, and yields of greenhouse crops. For instance, in bioassays, the leaf development of radish seedlings grown in a full Hoagland nutrient solution was compared with that in complete nutrient solutions amended with 2 or 5% aqueous extracts of vermicomposts. The extracts increased the leaf area significantly, suggesting a non–nutrient-mediated plant growth response. In aqueous extracts of vermicomposted cattle waste, in fractions separated by high-performance liquid chromatography (HPLC) and then analyzed by gas chromatography-mass spectrometry (GC-MS), indole acetic acid (IAA) was reported conclusively, and smaller amounts of gibberellins and cytokinins were present. © 2004 by CRC Press LLC

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a

Leaf area (cm)

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ab abc

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361

d

cd

abcd

abc

abcd bcd

cd

190 170 150 0

50

100

150

200

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500 1000 2000 4000

Humate concentration (mg/kg) FIGURE 18.5 Leaf area of tomatoes (mean ± SE) grown in commercial medium (Metro-Mix 360) substituted with humic acids extracted from vermicomposts. Columns with the same letter or letters are not significantly different at p ≤ .05 probability. (From Atiyeh et al. 2002. With permission.)

In laboratory/greenhouse bioassays at OSU, it was shown that auxins were present in vermicomposts in significant amounts using a Coleus bioassay. It was also shown that gibberellins were present in relatively small quantities by using a dwarf 5 maize bioassay (Sembdner et al. 1976), and that cytokinins were present in small amounts by using a cucumber bioassay (Hahn and Bopp 1968). Chemical analyses for IAA in vermicomposts by using HPLC and GC-MS confirmed these findings. The tomato seedlings responded positively to IAA and gibberellic acid 3 (GA3) and negatively to a single application of kinetin but positively to a second application of this PGR. These experiments demonstrated clearly that tomato plants could take up PGRs, including those produced in vermicomposts, from soil through their roots in quantities sufficient to influence their growth. Humic acids were extracted from vermicomposts, and a range of doses of these substances was added to tomato seedlings provided all needed nutrients. In greenhouse experiments, in response to some humate doses, very significant increases in plant growth occurred (Figure 18.5). The humic acids were extracted from pig-manure–based vermicomposts using the classic alkali/acid fractionation procedure (Valdrighi et al. 1996). The dry yield of humates was 4 g kg−1 of vermicomposts. The incorporation of 150, 200, 250, and 500 mg.kg–1 of humates from pig manure vermicompost into the soil less medium Metro-Mix 360 significantly increased the heights and leaf areas of tomato seedlings grown in these mixtures compared with those grown in the Metro-Mix 360 controls with no humates added. The greatest plant heights occurred in potting mixtures containing 200 mg/kg humates, whereas the greatest leaf areas occurred in potting mixtures containing 500 mg/kg humates. The dry weights of shoots of tomato seedlings grown in mixtures containing 200, 250, and 500 kg/mg humates were 47.0, 37.4, and 43.4%, respectively, greater than those of seedlings grown in Metro-Mix 360 controls. The dry weights of roots of tomato seedlings grown in media with 250, 500, and 1000 mg/kg of humates from pig manure vermicompost added were 77.5, 79.3, and 72.1%, respectively, more than those of seedlings grown in the controls with no humates. These effects of humates on plant growth all occurred when the plants were supplied with all their required nutrients (Atiyeh et al. 2002). In later experiments, humates extracted from cattle, food, and paper waste vermicomposts produced similar growth increases on peppers and strawberries (Arancon et al. 2004b). We hypothesized (Atiyeh et al. 2002; Arancon et al. 2004f) that plant growth hormones may be very transient in soils because they are very water soluble and degrade rapidly in sunlight. However, if they became adsorbed on humic acids, which are extremely stable, they would persist much longer in soils and continue to influence plant growth. This theory was confirmed by Canellas © 2004 by CRC Press LLC

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et al. (2000), who demonstrated that there were exchangeable auxin groups in the macrostructure of humic acids extracted from vermicomposts. These workers also showed that these complexes influence lateral root development of maize. This research provided valuable clues why vermicomposts influence plant germination, growth, flowering, and yields so dramatically over and above their content of readily available nutrients and make positive contributions to soil structure and fertility. Thus, there is increasing evidence of the various ways in which components of vermicomposts can increase the germination, growth, flowering and fruiting of a wide range of crops as discussed in this chapter. This also has implications for organic farming, for if earthworms can promote the activity and effects of PGRs in organic wastes, it may also be true that in soils to which organic matter is added, the production of PGRs by microorganisms may be increased by soil-inhabiting earthworm activity.

EFFECTS OF VERMICOMPOSTS ON PLANTS: THE INCIDENCE OF PLANT PATHOGENS, PLANT-PARASITIC NEMATODES, AND ARTHROPOD PESTS The suppression of plant pathogens by organic matter and thermophilic composts (Hoitink and Grebus 1997) and plant-parasitic nematodes by various forms of organic matter is well documented (Akhtar and Malik 2000). There are many unsubstantiated reports in the popular organic literature of the control of pests by organic matter application. However, only recently has the potential of vermicomposts in the suppression of pests been explored.

PLANT DISEASES There is extremely extensive literature on the suppression of plant diseases by organic amendments (Ara et al. 1996; Arafa and Mohammed 1999; Bettiol et al. 1997, 2000; Blok et al. 2000; Chen et al. 1987; Cotxaterra et al. 2001; Dixon et al. 1998; Diyora and Khondar 1995; Dutta and Hegde 1995; Ehteshamul et al. 1998; Fikre et al. 2001; Goldstein 1998; Goudar et al. 1998; Harender et al. 1997; Hoitink and Kuter 1986; Hoitink et al. 1997; Kannangowa et al. 2000; Karthikeyan and Karunanithi 1996; Lazarovits et al. 2000; Lima et al. 1997; Nam et al. 1988; Panneerselvam and Saravanamuthu 1996; Pitt et al. 1998; Raguchander et al. 1998; Rajan and Sarma 2000; Ramamoorthy et al. 2000; Sanudo and Molina-Valero 1995; Schiau et al. 1999; Somasekhara et al. 2000; Velandia et al. 1998) and traditional thermophilic composts (Hoitink et al. 1997; Huelsman and Edwards 1998). Various mechanisms have been suggested for this suppression, but most of these are based on some form of microbial antagonism. Traditional composting is a thermophilic process that promotes microbial activity selectively, whereas vermicomposting is a nonthermophilic method and promotes greatly increased activity by a wide range and diversity of microorganisms. We have considerable evidence from our research at OSU of much greater microbial activity and biodiversity in vermicomposts than in thermophilic composts. Our laboratory and field research provide evidence that vermicomposts may have an even greater potential for disease suppression than traditional thermophilic composts. For instance, general observational evidence of decreases in plant disease incidence and of pathogen suppression were recorded in studies involving 28 species of crop plants grown in vermicomposts (Edwards and Burrows 1988; Scott 1988). Nakamura (1996) reported suppression of Plasmodiophora brassicae, Phytophthora nicotianae (tomato late blight), and Fusarium lycopersici (tomato fusarium wilt) by vermicomposts. Szczech (1999) and Szczech et al. (2002) reported vermicompost suppression of Fusarium lycopersici and Phytophthora nicotianae on tomatoes. Rodríguez et al. (2000) demonstrated general suppression of fungal diseases of gerbera plants such as Rhizoctonia solani, Phytophthora drechsleri, and Fusarium oxysporum by the incorporation of vermicompost into the growth media. © 2004 by CRC Press LLC

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Food waste 5.0 t/ha

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Food waste 10 t/ha

FIGURE 18.6 Verticillium wilt damage to strawberries grown in field soils treated with vermicomposts produced from food waste and paper waste vermicomposts. Columns with the same letter or letters are not significantly different at p ≤ .05.

Orlikowski (1999) described sporulation reduction of the pathogen Phytophthora cryptogea after treatment with vermicomposts. Studies by Nakasone et al. (1999) showed that aqueous extracts of vermicomposts inhibited the mycelial growth of Baptista cinerea, Sclerotinia sclerotiorum, Sclerotium rolfsii, Rhizocronia solani, and Fusarium oxysporum. In greenhouse experiments in the Soil Ecology Laboratory at OSU, there was significant suppression of Pythium and Rhizoctonia resulting from substituting low rates (10 to 30%) of vermicompost into horticultural bedding mixtures (Chaoui et al. 2002). Suppression of diseases of field crops was achieved with low application rates of vermicomposts. The diseases suppressed in the field were Verticillium wilt on strawberries (Figure 18.6) and Phomopsis and powdery mildew (Sphaerotheca fulginae) on grapes. Two mechanisms of pathogen suppression have been described: one is based on microbial competition, antibiosis, hyperparasitism, and possibly systemic plant resistance (Hoitink and Grebus 1997). The second method of suppression of diseases such as Rhizoctonia has only a narrow range of microorganisms facilitating the suppression, termed specific suppression (Hoitink et al. 1997). It seems likely that these two mechanisms of suppression also apply to vermicomposts, but probably general suppression is much more common for vermicomposts because vermicomposting increases the biodiversity of microorganisms greatly whether pathogenic or beneficial.

INSECT PESTS There are some reports in the literature demonstrating that field applications of various types of organic matter and traditional thermophilic composts can suppress attacks by insect pests such as aphids and scales (Costello and Altieri 1995; Eigenbrode and Pimentel 1998; Morales et al. 2001; Phelan et al. 1996; Sudhakar et al. 1998; Yardim and Edwards 1999). There have been scattered reports of the suppression of insect pest attacks on plants by vermicompost amendments. Biradar et al. (1998) reported a clear correlation between the amounts of vermicomposts in the medium in which Leucaena leucocephala was grown and the degree of infestation by the psyllid Heteropsylla cubana. Rao et al. (2001) reported lower overall pest densities of ground nut leaf miner (Aproaerema modicella) in plots treated with vermicomposts. Ramesh (2000) described decreased attacks by sucking pests in response to vermicomposts. Rao (2002) reported very large decreases in attacks by the jassid (Empoasca verri) and the aphid (Aphis craccivora) and changed predator populations © 2004 by CRC Press LLC

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in response to field applications of vermicomposts. Such reports, although not numerous, provide an adequate basis to justify further research into this subject. Greenhouse experiments at OSU have demonstrated significant suppression of populations of aphids (Myzus persicae), mealy bugs (Pseudococcus), and caterpillars (Pieris brassicae) by substituting low rates of vermicomposts into a soilless plant growth medium (Metro-Mix 360) for tomatoes, peppers, and cabbages (Arancon et al. 2004d). The possible mechanisms of arthropod pest suppression by organic matter, composts, and vermicomposts are still speculative, but changes in the nutrient characteristics and balances of plants in response to vermicomposts compared with inorganic fertilizers and possibly the phenol contents of plant leaves have been suggested mechanisms because organic nitrogen is released more slowly from organic amendments such as vermicomposts than from inorganic fertilizers. This would make plants less acceptable to arthropod attacks (Patriquin et al. 1995).

PLANT-PARASITIC NEMATODES There is very extensive scientific literature demonstrating that additions of organic matter to soils may sometimes decrease populations of plant-parasitic nematodes appreciably (Addabdo 1995). Akhtar (2000) reviewed 212 scientific papers that discussed the effects of various organic amendments including composts (Chen and Zuckerman 2000; Gupta and Kamar 1997; McSorley and Gallaher 1995; Miller 2001; Zambolin et al. 1996) on plant-parasitic nematode populations. There have been a few reports in the scientific literature of vermicomposts suppressing populations of plant-parasitic nematodes. Swathi et al. (1998) demonstrated that 1.0 kg m−2 of vermicompost suppressed attacks of Meloidogyne incognita in tobacco plants. Morra et al. (1998) reported partial control of Meloidogyne incognita by vermicompost amendments to soils in a tomato-zucchini rotation. Ribeiro et al. (1998) reported that vermicomposts decreased the numbers of galls and egg masses of Meloidogyne javanica. Arancon et al. (2002, 2004e) reported significant suppression of plant-parasitic nematodes by field applications of vermicomposts, ranging from 2 to 8 kg/ha, applied to tomato (Figure 18.7), pepper, strawberry, and grape crops (Figure 18.8). Suppression of plant-parasitic nematodes by field applications of paper waste, food waste, and cattle manure vermicomposts was also reported by the same group of workers at OSU (Arancon et al. 2002, 2004e). There are a number of possible mechanisms that may explain decreases in populations of plantparasitic nematodes by vermicomposts. Predatory-prey interactions, which decrease populations of plant-parasitic nematodes, may also provide one feasible explanation of the mechanism. For

Numbers/20-g soil sample

16 14 12 10 8 6 4 2 0

Inorganic Paper Paper Cattle Cattle Food Food Compost fertilizer waste waste manure waste waste waste 20 t/ha 20 t/ha 10 t/ha 20 t/ha 10 t/ha 20 t/ha 10 t/ha

FIGURE 18.7 Plant-parasitic nematodes (mean ± SE) in field soils planted with tomatoes and amended with vermicomposts.

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instance, according to Bilgrami (1996), the mite Hypoaspis calcuttaensis preys voraciously on plant-parasitic nematodes, predaceous nematodes, and saprophagous nematodes. Perhaps vermicomposts increase numbers of omnivorous nematodes or arthropods that prey selectively on plantparasitic nematodes. Vermicomposts might also promote the growth of nematode-trapping fungi and species of fungi that attack and destroy nematode cysts and affect, either directly or indirectly, populations of plant-parasitic nematodes (Kerry 1988). Alternatively, rhizobacteria can colonize roots and kill plant-parasitic nematodes by producing enzymes and toxins that are toxic to them (Siddiqui and Mahmood 1999). As well as such biotic interactions, abiotic factors provided by vermicomposts might reduce populations of plant-parasitic nematodes. For example, vermicomposts may contain compounds that might affect the survival of nematodes. For instance, nematodes can be killed by the release of toxic substances such as hydrogen sulfide, ammonia, and nitrates during vermicomposting (Rodriguez-Kabana 1986).

EARTHWORMS AS A SOURCE OF PROTEIN FOR ANIMAL FEED Many mammals, birds, and fish prey on earthworms in nature. It was first suggested that earthworms contain sufficient protein to be considered as animal food by Lawrence and Millar (1945), and this potential of earthworms as animal feed has been confirmed by full analyses of the body tissues of earthworms, which show the kinds of amino acids that they contain and the nature of the other chemical body constituents. The first successful animal feeding trials on earthworms were by Sabine (1978), but subsequently there have been a number of other such trials by various workers (Edwards and Niederer 1988).

NUTRIENT VALUE

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The first analyses of the constituents of the tissues of different species of earthworms were by Lawrence and Millar (1945) and McInroy (1971), and there have been various other analyses since (Schulz and Graff 1977; Sabine 1978; Yoshida and Hoshii 1978; Mekada et al. 1979; Taboga 1980; Graff 1982; Edwards 1985; Edwards and Niederer 1988). Some of these show clearly that the essential amino acid spectrum for earthworm tissues, as reported by these different authors, compares well with those from other currently used sources of animal feed protein, and that the mean amounts of essential amino acids recorded are very adequate for a good animal feed. In addition, earthworm tissues contain a preponderance of long-chain fatty acids, many of which cannot be synthesized by nonruminant animals and an adequate mineral content. They have an excellent range of vitamins and are rich in niacin, which is a valuable component of animal feeds, and are an unusual source of vitamin B12. The overall nutrient spectrum of worm tissues shows an excellent potential as a protein supplement to feed for fish, poultry, pigs, or domestic animals.

PRODUCTION OF EARTHWORMS INDUSTRIAL WASTES

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IN

ANIMAL, VEGETABLE,

AND

The growth patterns of individual earthworms or whole earthworm populations in organic wastes follow classical sigmoid growth curves, with a rapid initial growth phase followed by a steadier phase and subsequent leveling. The maximum protein production per unit time can be achieved by inoculating relatively large volumes of animal wastes with small numbers of young worms to take maximum advantage of the initial fast phase of population growth. Dry matter conversion ratios of waste to earthworm biomass, which range from 10% for cattle and pig waste (Figure 18.1) to 2% for duck waste, have been achieved readily in the laboratory, although rather lower conversion rates have been attained in the field because of the difficulty of maintaining optimal conditions.

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PRACTICAL PRODUCTION

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EARTHWORM FEED PROTEIN

FOR

ANIMALS

The efficient production of earthworm protein depends mainly on detailed knowledge and management of the population dynamics of the appropriate earthworm species, maintenance of optimal environmental conditions, and the availability of efficient production and processing methods and particularly on the development and use of systems of harvesting earthworms from organic wastes that involve relatively little labor input. Earthworms that process organic wastes grow best at relatively high moisture levels (80 to 90% moisture content); this raises subsequent harvesting problems because it is not easy to separate earthworms mechanically from finely divided organic matter at high moisture contents, and some drying of the vermicompost before harvesting is usually necessary. There are various methods of separating earthworms from fully worked organic wastes, but they tend to be labor intensive, and an improved separation method was developed at Rothamsted and the National Institute for Agricultural Engineering, Silsoe, U.K. (Price and Phillips 1990). The efficiency of this machinery in terms of percentage recovery of earthworms is very high. The machine described will separate the worms from about 1 ton of waste per hour and can be automated and scaled up in size to increase its output and efficiency. After earthworms are collected from the separating machine, they may have small particles of waste attached to their bodies and are likely to contain organic waste in their guts. The earthworms must be washed thoroughly and left standing in water for several hours to evacuate the residual waste particles from their guts completely. A range of different methods of processing the earthworms into materials suitable for animal feed has been developed (Edwards and Niederer 1988). Two of these methods produced a moist paste product, and the other four produced dry meals; all the products were acceptable formulations for particular types of animal feeds. The ultimate choice of a method of processing depends on (1) the species of animal to be fed, (2) the type of animal feed required, (3) minimal loss of dry matter allowed, (4) minimal loss of nutrient value allowed, and (5) the costs of production. The following were the methods developed: 1. 2. 3. 4. 5. 6.

Incorporation of earthworms with molasses Ensiling earthworms with formic acid Air-drying earthworms at room temperatures Freeze-drying earthworms Oven-drying earthworms at 95°C Acetone immersion of earthworms followed by oven-drying at 95°C

The type of processing method used affected the amounts of total and essential amino acids in the feeds very little; however, the lysine content was decreased slightly by ensiling with molasses using formic acid and by freeze-drying compared with the other methods. The dry weight matter yields differed slightly among methods. Clearly, a stable protein feed can be produced by any of the methods listed, and the choice of method must depend mainly on the use to which the protein is to be put, the animal that is to be fed, and the cost of the processing method in relation to the feed value of the protein.

THE VALUE

OF

WORM PROTEIN

AS

FEED

FOR

FISH, POULTRY,

AND

PIGS

The main outlets suggested for utilization of earthworms as animal feed protein have been in fish farming and poultry and breeder pig feeds. All have been tested experimentally (Sabine 1978; Tacon et al. 1983; Edwards 1982, 1985; Edwards and Niederer 1988).

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Fish Feeding Trials The first use of earthworms as a protein source in fish feed was reported by Tacon et al. (1983), who studied the growth of trout fed on E. fetida, A. longa, and L. terrestris as a total feed compared with that of fish fed on a commercial formula. The fish fed frozen A. longa and L. terrestris grew as well as or better than fish fed on commercial trout pellets. However, trout did not grow so well on a whole diet of freeze-dried E. fetida, although they grew almost as well on E. fetida blanched in boiling water before freezing. Protein makes up only 15 to 30% of commercial fish diets, and dried earthworm meal derived from E. fetida, even when it had not been blanched satisfactorily, replaced the fishmeal component of formulated trout pellets at the normal levels of inclusion between 5 and 30%. Clearly, earthworms have good potential as both complete feed or protein supplement for trout or other fish. Guerrero (1983) reported that Tilapia fish grew better on diets containing earthworm protein supplements from P. excavatus than those provided with other fish meal supplements. Velasquez et al. (1991) reported that earthworm meal produced from E. fetida produced satisfactory growth of rainbow trout with a significant increase in lipid content. Chicken Feeding Trials The first reports of the growth of chickens on a diet with earthworm protein were by Harwood and Sabine (1978). They found no significant difference in growth of chickens fed on earthworms and chickens fed on a meat meal protein supplement. Similar results were reported by Taboga (1980) and Mekada et al. (1979). Jin-you et al. (1982) reported that chickens fed on earthworms gained weight faster than those given other diets (including fish meal), had more breast muscle, and consumed less food. Freeze-dried earthworm protein was used by Fisher (1988), and chickens fed this protein supplement grew well, gained weight per unit of food well, and had an excellent nitrogen retention when fed diets containing levels of earthworm meal from 72 to 215 g/kg. Pig Feeding Trials In feeding trials with both starter and grower pigs (Sabine 1978), young pigs fed an earthworm protein supplement grew equally well and had similar feed conversion ratios to those grown on regular commercial protein feeds. Jin-you et al. (1982) also reported that piglets grew better on diets with earthworm protein supplements than on other protein supplements, weaning was accelerated, estrus in sows was earlier, and there was increased disease resistance and a decreased incidence of white diarrhea. Good growth of young pigs on earthworm protein was also reported by Edwards and Niederer (1988).

ECONOMIC POTENTIAL ANIMAL FEED

OF THE

PRODUCTION

OF

EARTHWORM PROTEIN

FOR

A detailed study of the economics of production of earthworm protein concluded that the earthworm meal must be produced at an economic price, although the value of the vermicompost produced can also be taken into account as complementary income. Earthworms can be produced economically with relatively low labor inputs. The labor-intensive part of earthworm protein production is the harvesting process, and this remains the main barrier to successful commercial production of earthworm protein. More efficient methods of harvesting have been tested and hold considerable promise for the future.

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In a computer analysis of the economic value of earthworm protein based on its amino acid, fatty acid, mineral, and vitamin contents, it can be seen that it is extremely valuable as feed for certain animals, particularly eels and young turkeys, and it has the same value for fish, pig, and poultry feed as fish meal or meat.

METHODS OF PROCESSING ORGANIC WASTES WITH EARTHWORMS Processing systems of earthworms range from very simple methods involving low technology such as windrows, waste heaps, or containers, through moderately complex to completely automated continuous flow reactors (Jensen 1993; Edwards 1995a,b). The basic principle of all successful processing systems is to add the wastes at frequent intervals in small, thin layers to the surface of the system and allow the earthworms to move into this and process successive aerobic layers of wastes. The earthworms will always move up and concentrate themselves in the upper 15 cm of waste and continue to move upward as each successive waste layer is added. Many of the operations involved in vermicomposting can be mechanized; a suitable balance is needed between the costs of mechanization and the savings in labor that result. The key to combining maximum productivity of vermicompost with the greatest rates of earthworm growth is to maintain aerobicity and optimal moisture and temperature conditions in the waste and to avoid wastes with excessive amounts of ammonia and salts. The addition of organic wastes in thin layers avoids overheating through thermophilic composting, but enough usually occurs to maintain suitable temperatures for earthworm growth during cold winter periods. Hence, for year-round production to maintain a reasonable temperature in temperate climates, the processing should always be done under cover, although heating is not usually necessary if the waste additions are managed well with addition of thicker layers during cold periods to provide some thermophilic composting.

LOW-COST FLOOR BEDS

OR

WINDROWS

Outdoor windrows or beds, either heaped or with low simple walls, are the most common type of process generally used. The size of such beds is flexible, but the width of the beds should not exceed 8 ft (2.4 m), which allows the entire bed to be inspected easily. Because there is no need to walk on the bed, many suitable surface coverings and construction materials can be used. The length is less important and depends on the area available. They should not be laid on soil because soil particles would be picked up with the processed vermicompost. Concrete areas are ideal for earthworm processing systems because they provide a firm surface for tractor operations. However, it is essential for precautions to be taken to prevent too much water from entering the beds and to allow excess water to drain away from the bed. Often, the wastes in such floor beds are covered with permeable material such as canvas or bamboo sheets, which are removed only for watering and addition of new waste materials. Windrows and floor beds process organic wastes relatively slowly, often taking 6 to 12 months. During this period, there may be losses of plant nutrients through volatilization or leaching. The major drawback to windrows is the difficulty in harvesting the vermicompost and the need for a trommel or other separation stage to recover earthworms from the vermicompost before it is used. Although the initial capital outlay, other than land, is low, large areas of land are needed, labor costs are high, and the rate of processing is slow.

CONTAINERS

OR

BOX SYSTEMS

Edwards (1988) discussed methods of batch vermicomposting in large or small stacked boxes or containers and suggested that most of the methods tested were too labor intensive because batches had to be moved to add more wastes. It is difficult to access them or add water to them because they are usually stacked one above the other. There have been attempts to develop improved batch © 2004 by CRC Press LLC

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systems and modular container systems. However, they all have the same disadvantage as windrows, that is, the need for a labor-intensive separation of waste from the earthworm stage.

DOMESTIC SYSTEMS Small-scale systems of vermicomposting for disposal of household wastes have been used extensively in homes and schools (Appelhof 1997). They range from simple containers with perforated lids for aeration to more sophisticated commercially produced stacking systems of different sizes and complexity, including circular stacking systems such as the Can-O-Worms and the Worm Wigwam System, a rectangular stacking system called Worm Factory, and systems such as the Eliminator, which has a breaker bar and collection drawer at the base (Worm Digest). These systems have attracted the interest of some urban waste authorities, who have encouraged home owners to use them.

EARTHWORM TOILETS Various systems of toilets based on earthworms as an alternative to septic systems have been developed. The most successful is the Dowmus Composting Toilet. This system is well engineered, completely odor free, and built in Australia. It needs little maintenance and has been adopted for use in some state parks (Windust 1994).

WEDGE VERMICOMPOSTING SYSTEMS A wedge system was designed by Edwards and colleagues in the U.K. It is based on adding successive thin layers (5 to 10 cm) of organic waste at a 45° angle from a vertical removable barrier (Phillips 1988; Fletcher 1991). The wedge system can be any width or length but is limited in height to about 1.2 to 1.5 m for ease of loading. It should be situated on concrete or some other solid surface. The system starts with an initial layer of partially vermicomposted biosolids or other organic waste containing 9 kg (wet weight) of E. fetida (or other species) per square meter to a depth of about 15 cm. The surface is kept moist to a depth of 15 cm (80% moisture content) by a fine water spray twice daily. The earthworms move rapidly from the older layers of fully processed organic waste into the fresh material at the wedge surface so that the entire earthworm population is always concentrated in the top 15 cm below the leading surface. At convenient intervals (e.g., 1 to 2 months), the removable barrier is taken away and replaced about 60 cm behind the leading face of the wedge, so that no earthworms are removed when the waste is collected. All the processed waste behind this barrier can be removed with front loader machinery and collected free of earthworms for subsequent drying to 35 to 45% moisture, sieving, and packaging. Processing of wastes in a wedge system takes about 3 to 4 months.

GANTRY-FED VERMICOMPOSTING BEDS As discussed, an important principle in improving the efficiency of processing organic wastes by earthworms is to be able to add the wastes at frequent intervals to beds as deep as 1 m in thin layers 1 to 2 cm thick. This can be done most readily by adding the wastes by an overhead gantry running on wheels on the walls of the beds. This gradual addition of waste maximizes rates of waste processing, minimizes the generation of heat through composting, and ensures that earthworms are continually processing the fresh wastes near the surface.

CONTINUOUS FLOW AUTOMATED VERMICOMPOSTING REACTORS Continuous flow automated vermicomposting reactor systems were designed at the National Institute for Agricultural Engineering, Silsoe, Beds, U.K., by Phillips, Price, Edwards, and Fletcher

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(Price and Phillips 1990; Fletcher 1991; Edwards 1995a,b). These systems use large containers raised on legs above the ground (Figure 18.3). This allows organic wastes to be added in thin layers to the surface from mobile gantries at 1- to 2-day intervals, and the vermicomposts can be collected mechanically through mesh floors at the bottom using manual power or electrically driven breaker bars, which travel up and down the length of the system on a winch. Waste released on the floor can be brought from under the reactor to one end by hydraulically driven flap scraper systems of the kind used to collect manure from dairy cows in barns. Such reactors can range from mediumtechnology systems using manually operated loading and waste collection systems to large (40 m long, 2.8 m wide, 1 m deep or 1 m long legs) processes that are completely automated, electrically hydraulically driven, continuous flow reactors, which have operated successfully in the U.K., United States, and Australia for several years (Edwards 1998). The earthworm populations in such reactors tend to reach an equilibrium biomass of about 9 kg per m2. Such reactors can fully process the whole 1-m depth of suitable organic wastes they contain in about 30 to 45 days (Edwards 1995b, 1998). Economic studies have shown that such reactors have a much greater economic potential to produce high-grade plant growth media with few losses very quickly and much more efficiently than do windrows or ground beds.

COMPLETE RECYCLING VERMICOMPOSTING SYSTEM A complete, vermiculture-based, urban waste recycling system has been developed in France. This involves putting waste through a selector that breaks up plastic wastes and removes them, followed by manual sorting, sorting of rolling objects such as bottles, and separation of ferrous metal objects with magnets. The waste is then transported to a thermophilic compost system for 30 days, followed by vermicomposting in a very deep continuous flow system for about 60 days before removing earthworms, storing, and packaging. This system can turn as much as 27% of the total urban waste stream into vermicompost. This was sold and added to the commercial potential of the recycling considerably.

COMMERCIALIZATION

AND

ECONOMICS

OF

VERMICOMPOSTING SYSTEMS

The use of organic wastes to grow earthworms is an extensive cottage industry in the United States and other parts of the world. Many of these small-scale producers market the castings they produce for growing plants. Most operations in the United States are based primarily on windrow systems, which have many economic and environmental drawbacks, as discussed here. They are ground based and require large areas of land, with potential for groundwater pollution with nutrients and other contaminants because they are watered regularly and usually have no protection against leaching. The process is slow, taking 4 to 12 months to complete. The harvesting of the vermicompost is laborious and time consuming because the earthworms in the waste have to be separated, usually by a screening process, before marketing. Although the initial capital outlay, other than land, is low, its labor costs are high at all stages of operation. The wedge system designed by Edwards and colleagues described in this section has been used by a number of organizations. Although it uses an innovative but relatively inexpensive technology and requires less equipment, it overcomes many of the labor, economic, and environmental drawbacks associated with windrows. In particular, it uses less land, there is no leaching into groundwater, and there is no need to separate earthworms from the vermicompost. The processing time is shorter (3 to 4 months). The automated continuous flow reactor system designed by Edwards and colleagues (Edwards 1995b, 1998) and used by the Oregon Soil Corporation since 1992 has totally different environmental, operating, and economic characteristics. The equipment has to be under cover to maintain controlled environmental conditions, and the waste compartment is raised above the floor and is maintained at 80% moisture content and 20 to 32°C with no leaching. The retention or processing

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time of wastes in the reactor is 30 to 45 days. The economics of automated reactors are totally different from those of other systems because they involve a need for high initial capital inputs for reactors, which can process up to 3 tons of waste per day ($35,000 to $50,000) each, as well as ancillary loading and transport equipment such as moving belts, macerators, and loaders for operation. However, its labor and running costs are extremely low, and reactors reach equilibrium and can be run trouble free for a number of years. The capital expenditure can be recovered in 1 to 2 years. A number of relatively expensive systems based on this system have been marketed, but are less attractive economically.

REFERENCES Addabdo, T.D. 1995. The nematicidal effect of organic amendments: a review of the literature 1982–1994, Nematol. Mediterranea, 23:299–305. Akhtar, M. 2000. Approaches to biological control of nematode pests by natural products and enemies, J. Crop Prod., 3:367–395. Akhtar, M. and Malik, A. 2000. Role of organic amendments and soil organisms in the biological control of plant parasitic nematodes: a review, Biores. Technol., 74:35–47. Albuzio, A., Concheri, G., Nardi, S., and Dell’Agnola, G. 1994. Effect of humic fractions of different molecular size on the development of oat seedlings grown in varied nutritional conditions, in Humic Substances in the Global Environment and Implications on Human Health, N. Senesi and T.M. Miano, Eds., Elsevier Science, New York, pp. 199–204. Appelhof, M. 1981. Workshop on the Role of Earthworms in the Stabilization of Organic Residues, Vol. 1, Proceedings, Beech Leaf Press, Kalamazoo, MI. Appelhof, M. 1997. Worms Eat My Garbage, Flower Press, Kalamazoo, MI. Ara, J, Ehteshamul, H.S., Sultana, V., Qasim, R., and Ghaffar, F. 1996. Effect of Sargassum seaweed and microbial antagonists in the control of root rot disease of sunflower, Pakistan J. Bot., 28:219–223. Arafa, M.K. and Mohamed, E.I. 1999. Soybean seed borne fungi and their control. 2. Effect of soil amendments on the incidence of Fusarium root rot and chlamydospores germination, Egyptian J. Agric. Res., 77:97–111. Arancon, N.Q., Atiyeh, R.M., Edwards, C.A., and Metzger, J.D. 2004a. Effects of vermicomposts produced from food waste on greenhouse peppers, Biores. Technol., in press. Arancon, N.Q., Edwards, C.A., Bierman, P., Metzger, J.D., Lee, S., and Welch, C. 2004b. Effects of vermicomposts applied to tomatoes and peppers grown in the field and strawberries under high plastic tunnels, Pedobiologia, in press. Arancon, N.Q., Edwards, C.A., Bierman, P., Welch, C., and Metzger, J.D. 2004c. The influence of vermicompost applications to strawberries: Part 1. Effects on growth and yield, Biores. Technol., in press. Arancon, N.Q., Galvis, P., and Edwards, C.A. 2004d. Suppression of insect pest populations and plant damage by vermicomposts, Biores. Technol., in press. Arancon, N.Q., Galvis, P., Edwards, C.A., and Yardim, E. 2004e. The trophic diversity of nematode communities in soils treated with vermicomposts Pedobiologia, in press. Arancon, N.Q., Lee, S., Edwards, C.A., and Atiyeh, R.M. 2004f. Effects of humic acids derived from cattle, food and paper-waste vermicomposts on growth of greenhouse plants, Pedobiologia, in press. Arancon, N.Q., Edwards, C.A., Galvis, P. 2004g. Effects of food waste vermicompost on yield and quality of Seyval grapes, Biores. Technol., in press. Arancon, N.Q., Edwards, C.A., and Lee, S. 2002. Management of plant parasitic nematode populations by use of vermicomposts, Proc. Brighton Crop Prot. Conf. Pests Dis., 8B-2:705–716. Arshad, M. and Frankenberger, W.T., Jr. 1993. Microbial production of plant growth regulators, in Soil Microbial Ecology: Applications in Agricultural and Environmental Management, F.B. Metting, Jr., Ed., Marcel Dekker, New York, pp. 307–347. Atiyeh, R.M., Arancon, N.Q., Edwards, C.A., and Metzger, J.D. 2000a. Influence of earthworm-processed pig manure on the growth and yield of green house tomatoes, Biores. Technol., 75:175–180. Atiyeh, R.M., Arancon, N.Q., Edwards, C.A., and Metzger, J.D. 2001. The influence of earthworm-processed pig manure on the growth and productivity of marigolds, Biores. Technol., 81:103–108.

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Atiyeh, R.M., Dominguez, J., Subler, S., and Edwards, C.A. 2000b. Biochemical changes in cow manure processed by earthworms (Eisenia rapez) and their effects on plant-growth, Pedobiologia, 44:709–724. Atiyeh, R.M., Edwards, C.A., Subler, S., and Metzger, J.D. 2000c. Earthworm processed organic wastes as components of horticultural potting media for growing marigolds and vegetable seedlings, Compost Sci. Util., 8:215–223. Atiyeh, R.M., Lee, S., Edwards, C.A., Arancon, N.Q., and Metzger, J.D. 2002. The influence of humic acids derived from organic wastes on plant growth, Biores. Technol., 84:7–14. Atiyeh, R.M., Subler, S., Edwards, C.A., Bachman, G., Metzger, J.D., and Shuster, W. 2000d. Effects of vermicomposts and compost on plant growth in horticultural container media and soil, Pedobiologia, 44:579–590. Atiyeh, R.M., Subler, S., Edwards, C.A., and Metzger, J.D. 1999. Growth of tomato plants in horticultural media amended with vermicompost, Pedobiologia, 43:724–728. Barea, J.M., Navarro, E., and Montana, E. 1976. Production of plant growth regulators by rhizosphere phosphate-solubilizing bacteria, J. Appl. Bacteriol., 40:129–134. Bettiol, W., Migheli, Q., and Garibaldi, A. 1997. Control, with organic matter, of cucumber damping off caused by Pythium ultimum Trow, Pesquisa Agropecuaria Bras., 32:57–61. Bettiol, W., Migheli, Q., and Garibaldi, A. 2000. Control of Pythium damping-off of cucumber with composted cattle manure, Fitopatol Bras., 1:84–87. Bilgrami, A.L. 1996. Evaluation of the predation abilities of the mite Hypoaspis calcuttaensis, predaceous on plant and soil nematodes, Fund. Appl. Nematol., 20:96–98. Biradar, A.P., Sunita, N.D., Teggelli, R.G., and Devaranavadgi, S.B. 1998. Effects of vermicompost on the incidence of subabul psyllid, Ins. Environ., 4:55–56. Blok, W.J., Lamers, J.G., Termoshuizen, A.J., and Bollen, G.J. 2000. Control of soil-borne plant pathogens by incorporating fresh organic amendments followed by tarping, Phytopathology, 90:253–259. Buchanan, M.A., Russell, E., and Block, S.E. 1988. Chemical characterization and nitrogen mineralization potentials of vermicomposts derived from differing organic wastes, in Earthworms in Environmental and Waste Management, C.A. Edwards and E.F. Neuhauser, Eds., SPB Academic Publishing, The Hague, the Netherlands, pp. 231–240. Butt, K.R. 1993. Utilization of solid paper mill sludge and spent brewery yeast as a feed for soil-dwelling earthworms, Biores. Technol., 44:105–107. Cacco, G. and Dell’Agnola, G. 1984. Plant growth regulator activity of soluble humic complexes, Can. J. Soil Sci., 64:225–228. Canellas, L.P., Olivares, F.L., Okorokova, A.L., and Facanha, A.R. 2000. Humic acids isolated from earthworm compost enhance root elongation, lateral root emergence, and plasma H+-ATPase activity in maize roots, Plant Physiol., 130:1951–1957. Casenave de Sanfilippo, E., Arguello, J.A., Abdala, G., and Orioli, G.A. 1990. Content of auxin-, inhibitorand gibberellin-like substances in humic acids, Biol. Plant, 32:346–351. Chaoui, H., Edwards, C.A., Brickner, A., Lee, S., and Arancon, N.Q. 2002. Suppression of the plant parasitic diseases: Pythium (damping off), Rhizoctonia (root rot) and Verticillium (wilt) by vermicompost, Proc. Brighton Crop Prot. Conf. Pests Dis., 8B-3:711–716. Chen, J., Abawi, G.S., and Zuckerman, B.M. 2000. Efficacy of Bacillus thuringiensis, Paecilomyces marquandii, and Streptomyces costaricanus with and without organic amendment against Meloidogyne hapla infecting lettuce, J. Nematol., 32:70–77. Chen, W., Hoitink, H.A., Schmitthenner, A.F., and Touvinen, O.H. 1987. The role of microbial activity in suppression of damping off caused by Pythium ultimum, Phytopathology, 78:314–322. Chen, Y. and Aviad, T. 1990. Effects of humic substances on plant growth, in Humic Substances in Soil and Crop Sciences: Selected Readings, P. MacCarthy, C.E. Clapp, R.L. Malcolm, and P.R. Bloom, Eds., ASA and SSSA, Madison, WI, pp. 161–186. Costello, M.J. and Altieri, M.A. 1995. Abundance, growth rate and parasitism of Brevicoryne brassicae and Myzus persicae (Homoptera: Aphididae) on broccoli grown in living mulches, Agric. Ecosyst. Environ., 52:2187–196. Cotton, D.C.F. and Curry, J.P. 1980. Effects of cattle and pig slurry fertilizers on earthworms (Oligochaeta, Lumbricidae) in grassland managed for sludge production, Pedobiologia, 20, 181–188.

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Cotxarrera, L., Trillas-Gay, M.I., Steinberg, C., and Alabouvette, C. 2001. Use of sewage sludge compost and Trichoderma asperellum isolates to suppress Fusarium wilt of tomato, Soil Biol. Biochem., 2002:34467–476. Darwin, C. 1881. The Formation of Vegetable Mould Through the Action of Worms, with Observations of Their Habits, Murray, London. Dell’Agnola, G. and Nardi S. 1987. Hormone-like effect and enhanced nitrate uptake induced by depolycondensed humic fractions obtained from Allolobophora rosea and A. caliginosa feces, Biol. Fertil. Soils, 4:115–118. Diaz Cosin, D.J., Jesus, J.B., Trigo, D., and Garvin, M.H, Eds. 1999. Proceedings of Sixth International Symposium of Earthworm Ecology, Pedobiologia, 43:481–908. Dixon, G.R., Walsh, U.F., and Szmidt, R.A. 1998. Suppression of plant pathogens by organic extracts a review, Acta Hortic., 469:383–390. Diyora, P.K. and Khandar, R.R. 1995. Management of wilt of cumin (Cuminum cyminum L.) by organic amendments, J. Spices Aromatic Crops, 4. Dkhar, M.C. and Mishra, R.R. 1986. Microflora in earthworm casts, J. Soil Biol. Ecol., 6:24–31. Dominguez, J. and Edwards, C.A. 2004. Vermicomposting organic waste: a review, in Soil Animals and Sustainable Development, in press. Dominguez, J., Edwards, C.A., and Ashby, J. 2001. The biology and population dynamics of Eudrilus eugeniae in cattle waste solids, Pedobiologia, 45:341–353. Dominguez, J., Edwards, C.A., and Webster, M. 1999. Vermicomposting of sewage sludge: effect of bulking materials on the growth and reproduction of the earthworm Eisenia andrei, Pedobiologia, 193(4):372–379. Dutta, P.K. and Hegde, R.K. 1995. Effect of organic amendments on the suppression of Phytophthora palmivora (Butler) Butler causing black pepper wilt, Plant Health, 1:56–60. Eastman, B.R., Kane, P.N., Edwards, C.A., Trytek, L., and Gunadi, B. 2001. The effectiveness of vermiculture in human pathogen reduction for USEPA class A stabilization, Compost Sci. Util., 9:38–49. Edwards, C.A. 1983. Production of earthworm protein for animal feed from potato waste, in Upgrading Waste for Feed and Food, D.A. Ledward, A.J. Taylor, and R. Laurio, Eds., Butterworths, London, 153–162. Edwards, C.A. 1983. Earthworms, organic wastes and food, Span. Shell Chemical Co., 26:106–108. Edwards, C.A. 1985. Production of feed protein from animal wastes by earthworms, Philos. Trans. R. Soc. London, 310:299–307. Edwards, C.A. 1988. Breakdown of animal, vegetable, and industrial organic wastes by earthworms, in Earthworms in Waste and Environmental Management, Edwards, C.A. and Newhauser, E.F., Eds., SPB Publ., The Hague, the Netherlands, pp. 21–23. Edwards, C.A. 1995a. The commercial and environmental potential of vermicomposting: a historical overview, Biocycle, June: 62–63. Edwards, C.A. 1995b. A historical overview of vermicomposting, Biocycle, June: 56–58. Edwards, C.A. 1998. Earthworm Ecology, Lewis Publishers, Boca Raton, FL. Edwards, C.A. and Bohlen, P.J. 1996. The Biology and Ecology of Earthworms, Chapman & Hall, London. Edwards, C.A., Bohlen, P.J., Linden, D.R., and Subler, S. 1995. Earthworms in agroecosystems, in Ecology of Earthworms in Forest Rangeland and Crop Ecosystems in North America, P.F. Hendrix, Ed., Lewis Publishers, Chelsea, MI, pp. 185–213. Edwards, C.A. and Bohlen, P.J. 1992. The effects of toxic chemicals on earthworms, Rev. Environ. Contam. Toxicol., 125:23–99. Edwards, C.A. and Burrows, I. 1988. The potential of earthworm composts as plant growth media, in Earthworms in Environmental and Waste Management, C.A. Edwards and E.F. Neuhauser, Eds., SPB Academic Publishing, The Hague, the Netherlands, pp. 211–220. Edwards, C.A., Burrows, I., Fletcher, K.E., and Jones, B.A. 1985. The use of earthworms for composting farm wastes, in Composting of Agricultural and Other Wastes, J.K.R. Gasser, Ed., Elsevier, Amsterdam, the Netherlands, pp. 229–242. Edwards, C.A. and Fletcher, K.E. 1988. Interactions between earthworms and microorganisms in organicmatter breakdown, Agric. Ecosyst. Environ., 24:235–247. Edwards, C.A. and Neuhauser, E.F., Eds. 1988. Earthworms in Waste and Environmental Management, SPB Academic Publishing, The Hague, the Netherlands.

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Use of Earthworms: 19 The Nature’s Gift for Utilization of Organic Wastes in Asia Radha D. Kale Department of Zoology, University of Agricultural Sciences, Bangalore, India

CONTENTS A Spectrum of Rural and Urban Solid Waste Generation ............................................................383 Status of Agriculture in Karnataka State, India ............................................................................384 Earthworms as Biodegraders of Organic Waste Biomass.............................................................385 Solid Organic Waste Utilization for Compost Production Using Earthworms ............................386 Urban Solid Organic Waste Management .....................................................................................387 Vermicomposting Operations by Agroindustries...........................................................................388 Development of Vermicomposting Technology in Rural Areas....................................................389 Humus Produced by Earthworm Activity: Organic Waste to Soil ...............................................391 The Status of Vermiculture and Vermicomposting at the Turn of the Century............................393 Advantages of Vermicomposts.......................................................................................................395 Earthworm/Microbial Interactions .................................................................................................396 References ......................................................................................................................................396

Earthworms are soil macroinvertebrates well known for their contributions to soil formation and turnover and with a widespread global distribution. The role of earthworms in the breakdown of organic debris on the soil surface and in soil turnover was first highlighted by Darwin (1881). Since then, it has taken almost a century to appreciate the important contributions of earthworms to curbing organic pollution and improving topsoil. This realization, although a long time coming, has awakened the global human population to the potential for utilizing earthworms for ecological benefits. Their potential as a biological tool is much better understood; selected species of earthworms can facilitate organic farming and make sustainable development a reality. This concept is gaining priority status as the quality of the soil is increasingly related to its capacity to accept, store, and recycle nutrients and water. Soil improvement is urgently needed to maintain economic crop yields and improve environmental quality (Haynes 1997). Many scientists have realized the possibility of utilizing earthworms to break down organic wastes, which often have been the causative agents of organic pollution (Edwards 1988; Edwards and Neuhauser 1988; Dominguez and Edwards 1997; Edwards 1997). By regulating moisture levels in wastes and mixing the ingredients in ratios that can be accepted by the earthworms, coffee pulp (Arellano et al. 1994), sugar factory waste (Kale et al. 1994), and pig solids (Dominguez and Edwards 1997) can all be converted into good-quality soil additives, along with the biomass production of earthworms.

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Reports by Vinceslas-Apka and Loquet (1997) and Fredrickson et al. (1997) showed that the involvement of earthworms in a composting process to produce vermicomposts decreases the time of stabilization of the wastes and produces an efficient organic material with nutrients and energy reserves. Garcia et al. (1994) and Elvira et al. (1994) showed that sludges from both agri-based industries and domestic sewage plants, which are not accepted as soil additives directly onto fields, can be an excellent food source for vermicomposting earthworm species, with suitable organic amendments such as plant litter or animal wastes used to produce earthworm biomass and obtain better quality soil additives. In the case of those earthworms used for vermicomposting of organic wastes, the relationship between biomass production and earthworm population increase is of utmost importance. Meyer and Bouwman (1997) provided an explanation for the possibility of not getting the population increases they expected in these earthworms. They suggested that anisopary, which is a physiological behavioral pattern of earthworms in which some percentage of the population remains only as sperm donors and fails to produce cocoons, as well as various extrinsic factors such as the available food sources and climatic factors can also contribute to changing their reproductive potential. Similarly, the levels of hazardous heavy metals in the organic wastes fed to earthworms can affect their spermatogenesis and spermateleosis. This was made clear in the ultrastructural studies carried out by Reinecke and Reinecke (1997) (see Chapter 16, this volume), which were based on exposing the earthworms to different concentrations of cadmium, lead, and manganese mixed with cattle dung. These hazardous materials are taken up into earthworm tissues and may at some point enter the upper trophic levels of the food chain, as was demonstrated in studies on earthworm populations at a hazardous waste site (Stair et al. 1994). In developing countries, where the current trend is to convert to more ecofriendly organic farming, Jambhekar (1994) demonstrated improvements in the fertility status of the soil in vineyards because of earthworms and increased yields of grapes, which is very encouraging. Similarly, Mba (1994) highlighted the effects of vermicomposts on a ground cover crop Telfaria occidentalis and their influence in increasing earthworm activity in the soil. Encouraging the endogeic species populations of earthworms by amending soil with organic manures and ground cover crops can contribute significantly to soil fertility in tropical countries. It is true that clear-cut demarcations between the role of epigeic and endogeic species of earthworms and their limitations have to be made, as was summarized by Buckerfield (1994). If there is any confusion, it has to be clarified by highlighting the importance of endogeic species of earthworms and the relative inability of epigeic species of earthworm to overcome adverse physical conditions in nature. Once this is resolved, it should be possible to designate the species of earthworm that can work best on particular organic wastes and in particular climate conditions. Mitchell (1997) discussed composting using epigeic earthworms to overcome organic pollution. In most of the presentations at the Fifth International Symposium on Earthworm Ecology (Edwards 1998), Eisenia fetida and Eisenia andrei were the species chosen for vermicomposting. Although there is information about the biology of other earthworm species like Perionyx excavatus or Eudrilus eugeniae, more is needed (Salazar et al. 1994). Fayolle et al. (1997) reported the possibility of rearing Dendrobaena veneta on organic wastes, such as horse manure and paper sludges. Studies of the life cycle of this species on other kinds of organic wastes would increase the number of earthworms available to use in organic waste processes. In tropical and subtropical conditions, E. eugeniae and P. excavatus are the best earthworms for vermicomposting. Work is still in progress to test the possibility of using other species from forestlands in vermicomposting. The use of epigeic species of earthworms to combat organic pollution and produce high-quality organic soil amendments and earthworm biomass is progressing at a slow pace in most Asian scientific laboratories compared with workers in the United States, who have been successful in promoting the technology of vermicomposting. Appelhof (1994) and Bisesi and Appelhof (1994) initiated vermicomposting food waste programs in schools and have prepared worksheets and manuals to increase interest and awareness among youngsters. White (1994) initiated the publication © 2004 by CRC Press LLC

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of a quarterly newsletter, Worm Digest, in which he focuses on earthworms as the answer to solving waste disposal problems and rejuvenating neglected land. From my personal research experience, it is a challenge for scientists to bridge the gap between the laboratory and the field. The whole human race will benefit if vermicomposting technology is accepted and adopted. The activities in the small state of Karnataka, India, described here show a general acceptance of vermicomposting by agriculturists, urban residents, and industrialists. Papers presented at the Fourth (Kretzschmar 1992), Fifth (Edwards 1998), and Sixth (Diaz Cosin et al. 1999) International Symposia on Earthworm Ecology covered many aspects of vermicomposting and the use of vermicomposts. The changing scenario regarding use of earthworms in organic waste management, especially the voluminous waste from agri-based industries, is a welcome feature. It is incredible to see the efforts made to tap the talents of resourceful schoolchildren and initiate them into using earthworms to handle problems of urban soil organic wastes. Although E. fetida predominates as the preferred vermicomposting earthworm, there are various attempts to test the utilization of other earthworm species to encourage the use of locally available species or the species that can adapt best to the different zoogeographic regions and climates. An aspect not usually covered is the implementation of the benefits of vermicomposting technology in developmental programs. The efforts of researchers to involve the public through various media have to be strengthened further by increased publication of scientific data to confirm earthworms as one of the best means of minimizing organic waste pollution and decreasing indiscriminate use of inorganic fertilizers. Proper technologies and monitoring of the production of earthworm biomass can also add to animal protein substitutes that are available to animal feed production units. Available areas of the countryside and agricultural land are decreasing rapidly in many countries as they move toward industrialization and away from an agrarian emphasis. Increased poverty, brought about by declines in available fertile land to grow food grains, is often accompanied by greed to use the land for cash crops. Under such circumstances, the present problem depends on improving the fertility of available land to overcome problems of hunger and starvation. From these points of view, progress achieved in academic institutions, its adoption by farming communities and agro-based industries, and its commercial viability in this subtropical part of India are summarized here with the aim of furthering research on organic waste management using earthworms.

A SPECTRUM OF RURAL AND URBAN SOLID WASTE GENERATION The Indian subcontinent, with its varied physiographic and agroclimatic zones and cultural practices, produces a great deal of agricultural and human wastes. Depending on the major crop of the zone, agroindustries are established in different townships. The kinds of waste generated in these units are extremely diverse. They are of plant or animal origin and can undergo degradation to release plant nutrients from their bound complex forms. Because the complex structural composition of these substances resists easy breakdown, decomposition is usually a slow process. This has resulted in accumulations of these materials without proper utilization. Stubbles, weeds, and crop litter are the main biomass that comprises major farm wastes. The organic wastes released from dairy farms can be added. Table 19.1 lists the major farm wastes and agroindustrial wastes that are posing disposal problems. The rapid growth of cities has resulted in ever-increasing accumulation of organic wastes. Most farm produce enters cities to cater to the food needs of the urban population. Thus, most agricultural produce is moved away from its origin. The unutilized and unwanted part of organic crop residues is often put into landfills. Organic materials that could have returned to the biogeochemical cycles

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TABLE 19.1 Wastes Tested for Vermicomposting Sl. No.

Source of Waste Generation

I.

Agricultural wastes Agricultural fields Plantations Animal wastes Urban solid waste

1. 2. 3. II. III. 1. 2. 3. 4. 5. 6. 7.

Agroindustries wastes Food processing units Vegetable oil refineries Sugar factories Breweries and distilleries Seed production units Aromatic oil Extraction units Coir industries

Utilizable Waste for Vermicomposting

Stubble, weeds, husk, straw, and farm-yard manure Stems, leaf matter, fruit rind, pulp, and stubble Dung, urine, and biogas slurry Kitchen waste from households and restaurants, waste from market yards and places of worship, and sludge from sewage treatment plants Peel, rind, and unused pulp of fruits and vegetables Pressmud and seed husk Pressmud, fine bagasse, and boiler ash Spent wash, barley waste, and yeast sludge Core of fruits, paper, and date-expired seeds Stems, leaves, and flowers after extraction of oil Coir pith

in farmland to enrich the nutrient status of soil have been turned into pollutants in the city. In India, domestic waste is mostly of an organic nature and contributes 70 to 80% of total urban solid wastes. Each household of four family members generates 0.5 to 0.75 kg kitchen waste per day (Kale and Sunita 1993).

STATUS OF AGRICULTURE IN KARNATAKA STATE, INDIA Food production is a primary concern of any country. When productive lands are scarce and human population growth is increasing at an alarming rate, maximum productivity must be achieved in countries such as India from the available land. A little more than 71% of the population of Karnataka state lives in rural areas. Among this rural population, 94% depend on agriculture for a living. Reports from the Directorate of Agriculture, Economics, and Statistics and Economic Survey show that land utilization for agriculture over the state averages 1.2 ha per individual. In this area, nearly 56.2% of land is used for sowing annual crops, and 8.8% is cultivated more than once annually. Only 12% of agricultural land is irrigated. For every ton of food grains produced, an average consumption of chemical fertilizers, as an energy subsidy, is 50 kg N, 30 kg P, and 90 kg K. Inorganic fertilizer utilization has more than doubled in the last decade and so has the use of pesticides for pest control increased. These increases in chemical applications have not boosted agricultural yields significantly, except for those of sugarcane. For the rest of the crops, production has either remained the same or has declined over the last 5 years. It is costlier to maintain crop plants in hot climates than in temperate areas because of the high temperatures and water stresses (Best 1962). A mere 3% of the agricultural community in Karnataka owns more than 10 ha of land, and the rest of the landholders are marginal-to-medium farmers with less than 10 ha of land. The current strategy of a tenfold increase in use of fertilizers, pesticides, and machinery to provide a mere twofold increase in agricultural production is economically unviable under the existing situations in the state (Bennett and Robinson 1967). The levels of pollution caused by different sources have been emphasized. Factory production units of fertilizers are releasing various hazardous by-products that can pollute air, water, and soil if proper preventive measures are not taken. Similarly, sludges from sewage treatment plants and other organic degradable refuse are causing harm to the ecosystem. The same organic refuses can

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be utilized for the agricultural system if regulated properly. To utilize the nutrients in these materials, detritus consumption on agricultural land has to be activated. To meet the needs of short-duration crops, which have been adopted in the present-day agricultural systems, more and more chemical fertilizers are currently used. The repeated use of chemicals in tropical lands is causing increasingly adverse effects on soil properties, productivity, and fertility. However, this alarming situation has produced a reverse in this strategy of agricultural practice by a small percentage of agriculturists in the state. Their practices involve selecting varieties of plants that are most resistant to pest attack and have a low demand for inorganic nutrient inputs and practicing multiple cropping systems. This also includes the production and use of biological insecticides and biomass as organic soil amendments. By tapping organic detritus as a resource, it may be possible to rejuvenate the topsoil and increase production. In this organic detritus group, earthworms can be managed to feed on varieties of organic wastes, thereby serving as tools to hasten the process of organic degradation and release of nutrients and improving the nutrient status of the tropical soils. The propagation of earthworms for humus production under seminatural conditions is developing rapidly in the state and is revolutionizing its agriculture. From the situation that was reported in 1991 to 1992 regarding the agricultural scenario in the state, the latest census of 1998 to 2000, the available land for agriculture has decreased by 5.9% and land used more than once for agriculture per annum by 3.6% for the expected population increase of 21%. Interestingly, because of the awareness created about the importance of organic manures for sustainable agriculture, in this decade there has been a 23% decrease in inorganic fertilizer consumption.

EARTHWORMS AS BIODEGRADERS OF ORGANIC WASTE BIOMASS The loss of topsoil because of practices associated with the indiscriminate use of chemicals is a major cause of the decline in productivity status of our tropical soils. The realization of this has begun to increase the use of organic wastes on agricultural fields. However, there is a decline in the availability of cattle dung, and a current need is to obtain the required organic soil additives from underused available plant biomass residues in minimum time. Various methods have come into practice in addition to the traditional practice of pit composting. Engineering skills are used to provide better aeration of composting materials to minimize the time of composting. Selected groups of fungi are often used for composting of lignocellulose-based materials. Much of the importance given to earthworm activity in temperate soils is lacking in tropical regions, although earthworm species diversity and richness is not much less than in temperate regions (Bano and Kale 1991). Food niches and strategies that have developed among different species of earthworms have helped broaden their use in organic waste breakdown. Epigeic earthworm species, which show a greater affinity for nitrogen-rich organic matter, live in an unstable environment. They resemble the seral stages of ecological development, with a smaller body size, higher metabolic rate, higher fecundity, and shorter life cycle. In natural conditions, their survival depends on environmental conditions and degree of biotic pressure from predators. These earthworms form the natural components of the tropical forest floor community. With the loss of natural forests, they have adopted to the agricultural plantations. Increased yields from the use of chemical fertilizers have led the farming community to neglect the importance of organic manures and additives. Many farmers have been unaware that organic manures are essential to restore the cycle of events in the soil to keep it productive. Chemicals applied to the soil short-circuited this cycle of events and deprived the soil organisms of their energy sources, which restore the health of the soil through their activity. When earthworms are isolated to work on waste organic matter, their major contribution is in fragmenting the organic matter and making it more available to microorganisms. The microenvironment they provide for the establishment and multiplication of microorganisms is of utmost

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importance in the vermicomposting process. Their excretions of ammonia and urea add to the nitrogen content. Other organic acids and mucus secretions provide nourishment for microbial populations. With the organic amendments to the fields and use of minimum tillage practices, the soil organisms become more important.

SOLID ORGANIC WASTE UTILIZATION FOR COMPOST PRODUCTION USING EARTHWORMS We tried vermicomposting animal waste under laboratory conditions using P. excavatus (Kale et al. 1982), a species that is distributed all over India. These earthworms can be found congregated near cattle sheds and in biogas slurry pits. Plantations with a good litter cover also form a habitat for these worms. Similarly, Perionyx sansibaricus in Kerala and P. pallus in Maharashtra have been tested for their potential use in organic waste degradation. Apart from these species, polyhumic worms such as Lampito mauritii have also been tried for vermicomposting. As in temperate countries, E. eugeniae and E. fetida are also gaining importance as the most suitable species for humus production. Eudrilus eugeniae, a tropical species, is establishing well in southern peninsular India, whereas E. fetida establishes best in the more temperate part of the country. Eudrilus eugeniae is preferred for the degradation of agricultural waste, and E. fetida is used for urban waste degradation at the community level and in households. The systems for utilizing these earthworms were developed at different stages in the laboratory. Methods were aimed at minimizing the time of vermicomposting, improving the quality of the vermicompost, and minimizing the cost of labor and making it an economic venture for the spread of the technology to the rural areas. The work started with using organic matter that breaks down slowly to form the first tier in a vermicomposting container, followed by a layer of finely sieved sand and garden soil. Populations of E. eugeniae, E. fetida, or P. excavatus are introduced into these prepared containers. Soft organic wastes, like many animal wastes or vegetable wastes, were spread on the surface. In the initial stages, earthworms fed on the surface organic layer and produced casts. The feed mix was added as needed in layers after collecting the casts. Every 3 months, the contents of the culture containers were emptied; by this time, the entire organic matter, including stubbles from agriculture fields, sugarcane trash, or coir waste, had decomposed. This was shown by the reduction in levels of residual cellulose and lignin in the earthworm-worked materials (Kale et al. 1991). Because this procedure was not viable for the farming community, an alternative and a much simpler method was evolved. In older agricultural practices, when the available land area for food production per individual was larger, farmers were in the habit of reserving a part of their land for growing certain shrubs and grazing cattle. These shrubs were pruned periodically to mix the green matter into the soil to serve as green manure. Most of these plants had a high level of nitrogen. Based on this practice, green leaves from different plants, weeds, hedge cuttings, and dry leaves were mixed with cow dung slurry and left in the earthworm culture containers for 2 weeks. During this period, two turns were given to the mix to aerate the material. After 2 weeks of conditioning, the earthworms were released to the waste surface, where they fed on the surface organic materials and moved down as most of the material was converted into earthworm casts. Occasionally, water was sprinkled on the surface. Earthworms fed on the soft material at the beginning, and the hard and unfed materials were mixed again with fresh material. The collection and preparation of organic mixes are simple and time saving; farmers further simplified the method and have developed their own techniques for vermicomposting the wastes listed in Table 19.1. To minimize nutrient losses from the wastes during vermicomposting and to provide a vent for released carbon dioxide or other gases during the decomposition process, a 2.5-cm thick mud pack was placed on the surface of the decomposing material, and polyvinyl chloride tubes — or even the hollow shoots of the plants 3 cm in diameter with a series of holes were inserted into the packed

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TABLE 19.2 Range of Nutrients in Vermicompost Organic carbon (%) Total nitrogen (%) Available phosphorus (%) Available potassium (%) Available sodium (%) Calcium and magnesium (meq/100 g) Copper (ppm) Iron (ppm) Zinc (ppm) Available sulfur (ppm)

9.15–17.98 0.50–1.50 0.10–0.30 0.15–0.56 0.06–0.30 22.67–47.60 2.00–9.50 2.00–9.30 5.70–11.50 128.00–548.00

material 30 cm apart. After 2 weeks, earthworms were released into the vermicompost pits through temporary holes made on the mudpack. By this method, moisture loss from the waste was minimized. Thus, the light and oxygen kept the earthworms active beneath the surface of the mudpack in the organic wastes. This increases the biomass of the earthworms and the nutrient status of the compost produced in a short interval of time (Table 19.2). The species E. eugeniae, E. fetida, and P. excavatus all work well under these conditions, whereas Dichogaster curgensis, which needs a soil base, failed to survive. The large body size and low population-carrying capacity of E. eugeniae led us to depend more on E. fetida, which tolerates high population density pressure in its use for urban solid organic wastes. Neuhauser et al. (1979, 1980) showed the sensitivity of E. eugeniae to population density pressure. The carrying capacity of E. eugeniae on organic wastes under laboratory conditions was studied and was measured as 0.015 g/cc (Kale and Bano 1991).

URBAN SOLID ORGANIC WASTE MANAGEMENT The practice of waste segregation in India is growing both at the community level and in individual households. The organic wastes are segregated in the home before handing over to garbage collectors. Only the compostable waste is dumped into pits, constructed in public parks of residential areas, for vermicomposting. The pits are covered with metallic mesh to keep away predators. A low level and sloped roofs protect the pits from inundation during heavy rains and from direct sunshine. Care is taken to maintain the aesthetic look of the parks while constructing the pits. Similar vermicomposting activities are done in houses in small tubs or even in knitted sacks. The daily garbage production in Bangalore (Karnataka) averages 2000 tons. In residential localities, 70 to 80% of the solid waste is biodegradable organic waste. Table 19.3 provides information on the composting of such urban wastes in a residential area. The organic matter can result in air pollution and groundwater pollution, both recognizable by the obnoxious odors, at the dump sites and by released leachates. As an alternative, by collecting the organic wastes into impermeable pits under aerobic conditions with selected species of earthworms, the entire material is converted in a short time into good vermicomposts. The organic wastes that cause pollution and are a source of epidemiological problems now become a source of income for those people who have begun vermicomposting, and this vermicompost provides a high nutrient source that is very beneficial to plants and evokes biological activity in impoverished soils. The Karnataka Compost Development Corporation (KCDC), in association with Bangalore City Municipality, is converting a portion of the city garbage into vermicompost (150 to 200 tons/day) and the same amount of garbage into enriched compost. They have a clientele convinced about the high quality of the product. Their success has made them consultants to municipalities of other states in the country.

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TABLE 19.3 Composting of Household Waste Using Three Species of Earthworms: Eudrilus eugeniae, Eisenia fetida, and Perionyx excavatus

Earthworm Species I set Eudrilus eugeniae Eisenia fetida Perionyx excavatus II set Eudrilus eugeniae Eisenia fetida Perionyx excavatus

Organic Waste (kg)

Time Taken for Composting (days)

Compost Recovery (kg)

Initial Biomass (g)

Final Biomass of Worms (g)

Population Increase (no. of young ones)

200 200 200

60 90 90

160 160 160

300 150 150

1500 350 205

32,000 30,000 6,000

400 400 400

60 100 100

250 250 250

1500 350 205

1800 500 280

100,000 80,000 20,000

Note: Initially, 1000 nonclitellate worms of each species were introduced into different pits. The population of Eudrilus eugeniae reached carrying capacity by the end of 5 months. Source: From Kale and Sunita 1993.

As the interest in processing kitchen wastes is increasing, several designs of vermicomposting vessels have evolved to suit different purposes. One of the containers is a cylindrical drum erected horizontally on a stand. Mesh covers on either side of the drum provide ventilation. The capacity of the drum is about 5 kg. Rotation at an interval ranging from 48 to 72 h (maximum five rotations) provides good aeration in the medium. Such a closed and compact system can be maintained easily in any household without occupying much space. Two drums of this capacity are ideal for a typical Indian household with four to five family members. Different cities in the country are planning to devise methods to utilize such vermicomposting technology.

VERMICOMPOSTING OPERATIONS BY AGROINDUSTRIES An average pressmud (a by-product of the sugar processing industry) contains the elements added to clarify juice and sediments, solids like bagacillo, the smaller fibers of cane sugar, and particulates and mud. Production of pressmud per ton of sugarcane is 35 kg. A sugar factory that has taken up vermicomposting has a unit capacity for crushing 2000 tons of sugarcane per day. The resulting pressmud released into lagoons averages 2000 tons per month. The other organic wastes in lesser quantities are the fine baggase and ash from boilers. These are currently occupying large areas of land around the factory units as dump sites. The pressmud, as such, can be a good soil additive for its nutrient status, but the farmers are not willing to apply this to their lands for fear of developing complications in the fields. The pressmud has no other use and becomes a water and air pollutant if no alternative way is available for its safe disposal; vermicomposting is one of the only alternatives. Khoday’s group of companies is converting this waste into vermicomposts together with the organic wastes from their dairy farms and crop residues from their plantations. They started with 5000 E. eugeniae in June 1993, multiplied them in cement cisterns, and within 1 year they had a production level of 20 tons of vermicompost per month. This prompted them to develop a unit to convert all the waste from the sugar factory, the sludge from the distillery, and coir pith from coir industry into vermicomposts by investing Rs. 25,000,000 ($710,000 U.S.) for a single production unit. By the year 2000, many other distilleries and sugar factories in the state and in neighboring states have realized the importance of vermicomposting the effluents and have started units on their premises. The Central © 2004 by CRC Press LLC

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Pollution Control Board has given permission to distilleries and breweries to use effluents in vermicomposting as one of the means to curb groundwater pollution and pollution of other water bodies in the vicinity. An aromatic oil extraction unit near Bangalore city had a problem in disposal of flower and plant wastes after extracting the oils using organic solvents. The quantity of organic wastes generated was very high relative to the quantity of the oil extracted. The materials disposed in the dump sites caused bad odors and were not properly decomposed, even after 1 year. Now, the unit has tested the efficiency of earthworms at converting this material into vermicomposts. Encouraging results, both in terms of composting time and the nutrient value of the recovered compost, have prompted this unit to enter into much larger scale production. After achieving the success in the central unit at the factory premises, it has decentralized the activity by providing the know-how to the farmers of the neighboring villages. Now, farmers are collecting the organic wastes from this site and are using them for producing vermicomposts to use in their fields. Thus, the factory has found the way to shift the waste generated from their site of production for better cause. A public sector machine tool factory has taken up the vermicomposting of sludges from a domestic sewage plant. The agriculturists from the surrounding areas are not prepared to use the sludge directly on the fields, and it is not used for the gardens in the factory. Instead, it is currently heaped up near the treatment plant and wasted. Now, the factory is mixing other plant residues from pruning with the sludge and converting the sludges successfully into vermicompost. The competition and awareness of the potential of using organic wastes for vermicomposting will maintain the cost of the wastes at reasonable levels to meet the needs of the users. For big units, this is a means to dispose of the unwanted wastes. It also provides an additional income to the units and creates job opportunity for the unemployed. The organic waste has turned into profits with the use of earthworms in these units. A vegetable seed production and tissue culture division in Bangalore produces organic wastes in the form of ripened rind of fruit after the seeds are separated. Because they are producing a large number of plants by adopting tissue culture techniques, from agar to cotton buds and paper, lots of organic waste is generated. The unsold seeds after the expiration date have to be destroyed. This unit is now using these materials for vermicomposting. As the nature of the waste generated is different kinds, vermicompost is produced using different suitable combinations of these wastes.

DEVELOPMENT OF VERMICOMPOSTING TECHNOLOGY IN RURAL AREAS The green revolution in India encouraged the indiscriminate use of fertilizers to obtain two to three crop yields per year using irrigation and pesticides. As a result, tropical soils, which are prone to loss of nutrients and depletion of carbon levels, are becoming unproductive. There is concern about sustaining productivity rather than enjoying the high, but short-lived, yields and financial returns. High temperatures during most of the year and the unequal and unpredictable rainfall provide much less scope for the activity of soil organisms. Only termites and ants can be active in such situations. Moisture-sensitive organisms such as earthworms are active for only very short periods. Low moisture contents in the rain-fed soil, regular plowing, little available organic matter in the land, and heavy pesticide applications have further decreased their populations. All these factors contribute to unawareness among farmers about biological processes in soil and the importance of earthworms in the processes. The overall populations of soil organisms that contribute to the formation of topsoil by fragmenting the organic material and mixing into the soil strata are declining. The productivity status of the soil is showing the same downward trend. At this juncture, because of a lack of availability of manure at the right time in required quantities, many farmers are looking for different resources to build up the topsoil. It is essential to restore the carbon level in tropical soils, in which the depletion of carbon takes place at a rapid rate.

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A technology using suitable species of earthworms for degradation of any organic wastes of plant and animal origin for a minimum of 6 weeks was released for the benefit of farmers in 1984 from the University of Agricultural Sciences, Bangalore. Nevertheless, few farmers came forward to make use of the technology because of the heavy subsidy at that time on chemical fertilizers. The ill effects of using only inorganic chemicals were not realized at that time. Applying fertilizers, pesticides, and herbicides was timesaving and less labor intensive. By 1990, farmers realized that the subsidy on inorganic chemicals was not an everlasting feature. To get the same yields from their fields, they had to continually increase the quantities of the inorganic fertilizers used. Pest populations developed resistance to pesticides, which instead of bringing down the incidence of pest attack started to destroy the crops. Because of this, farming has become uneconomical. The nightmare of losing soil fertility combined with uncontrollable pest attacks has made Indian farmers look for alternative methods of production available through a more sustainable agriculture. From 1990, vermicompost production using available organic waste biomass and epigeic earthworms like E. eugeniae has advanced rapidly. Table 19.4 provides information on the spread of this technology in the area from February 1983 to 2000. Direct dialogue between the farmers and representatives of various organizations had led them to try out different degradable materials and improvise methods to suit the local conditions. Farmers are accepting this as a highly viable technology and are convinced that, with minimum expense, they can produce better quality organic soil additives. The feedback from farmers has revealed that the farmers are more optimistic because of this technology and innovation. In some of the districts of this state, almost all the farmers know about this new technology, and as the well-known slogan goes, a process of “each one teach one” has started. The message of protecting the soil for future generations to enjoy is spreading and boosting the morale of farmers. The art of mass rearing and maintenance of earthworm cultures, and the tapping of organic wastes for their maintenance, has provided good scope for developing

TABLE 19.4 Dissemination of Vermicomposting Technology from UAS Bangalore from 1983 to 1994 Year

Within Karnataka

1983 1984 1985 1986 1987 1988 1989 1990 1991 1992

2 1 1 0 1 0 0 4 51 290

1993 1994 (March) 2000

361 116 ∞

Outside Karnataka MAH/1 TN/1 AP/1 MAH/1 TN/1 KER/1 MAH/1 0 AP/1 MAH/1 TN/1 KER/1 PUN/1 TN/3 MAH/1 KER/4 AP/4 MAH/3 TN/6 KER/10 MAH/5 GUJ/2 AP/2 PUN/1 MAN/1 DEL/1 HP/1 MP/1 RAJ/1 OR/3 BEN/1 UP/1 HAR/1 TN/22 KER/17 AP/19 OR/4 GUJ/2 PUN/1 MAH/10 DEL/2 BEN/1 SIK/1 POND/1 AP/3 MAH/1 TN/3 KER/1 All the states in the country and Sri Lanka and Bangladesh

AP, Andhra Pradesh; BEN, Bengal; DEL, Delhi; GUJ, Gujarat; HAR, Haryana; HP, Himachal Pradesh; KER, Kerala; MAH, Maharashtra; MAN, Manipur; MP, Madhya Pradesh; POND, Pondichery; PUN, Punjab; RAJ, Rajasthan; SIK, Sikkim; TN, Tamil Nadu; UAS, University of Agricultural Sciences; UP, Uttar Pradesh.

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vermicomposting as a cottage industry in this country, where there is no dearth of organic wastes or labor. The tapping of a resourceful technology is of utmost importance for the present day because “soil is the placenta of life.”

HUMUS PRODUCED BY EARTHWORM ACTIVITY: ORGANIC WASTE TO SOIL Joshi and Kelkar (1952) reported higher electrical conductivity in earthworm casts, which denotes an increase in the levels of soluble salts over the surrounding soil. They showed that casts have greater nitrifying power than the soil. Earthworm casts, as aggregates, make the site more aerobic. The stability of the casts depends on the concentration and type of organic matter present and the bacterial and fungal polysaccharides (Dutt 1948; Bhandari et al. 1967). Earthworm casts are better sources of organic manure than other anaerobically degraded composts because of the factors discussed in this section. Casts are loosely packed granular aggregates of semidigested organic matter that provide energy for the establishment and multiplication of various microorganisms. Some of the microorganisms associated with the casts remove the bad odor from decomposing materials (Watanabe et al. 1982). The casts also form a suitable base for free-living beneficial microorganisms with activity that is essential for the release of nutrients to plants (Atlavinyte and Daciulyte 1969; Atlavinyte and Vanagas 1982; Ross and Cairns 1982). The establishment of a microenvironment takes place in the presence of earthworms in the given medium (Kale et al. 1987). Especially in tropical countries, earthworms cannot remain active throughout the whole year (Roy 1957; Reddy and Alfred 1978; Chauhan 1980). Nutrients from the soil leach out at a rapid rate under ambient soil and environmental conditions in tropical countries, particularly during monsoons. Under such circumstances, the application of earthworm casts or vermicomposts to fields can improve the physicochemical and biological properties of the soil (Kale et al. 1992). The biochemical activity of established microorganisms and earthworm exudates has a stimulatory effect on plant growth (Ross and Cairns 1982). The presence of earthworms in culture pots improved the germination, growth, and yield of a barley crop (Atlavinyte and Zimkuviene 1985). A higher level of vitamin B12 in the medium because of earthworm activity was reported by Atlavinyte and Daciulyte (1969). Nielson (1965) isolated indole acetic acid (IAA)–like substances from homogenates of different species of earthworms. Springett and Syers (1979) reported increases in net crop production after application of earthworm casts. Similar information is available on improvements in the growth and yields of crops influenced by earthworm exudates (Graff and Makeschin 1980). Tomati et al. (1990) reported on the presence of growth regulator substances in earthwormworked soils. Increases in the protein synthesis of Agaricus bisporus and radish Raphanus sativam was recorded in plants grown in the presence of earthworm casts (Galli et al. 1990; Tomati et al. 1990). Increase in the rates of uptake of nutrients with the increase in symbiotic microbial associations in cereal and ornamental plants after using vermicompost as a source of organic manure were observed (Kale et al. 1987, 1992). Similarly, in a perennial crop such as mulberry, there was no difference in the uptake of nutrients in plants grown with vermicomposts or chemical fertilizers. Although the NPK concentrations applied to land in the form of chemicals was higher than the level of the same components in vermicompost, the level that entered the plant tissues remained the same irrespective of the concentrations that entered the soil (Gunatrilakara and Ravignanami 1996). Karuna et al. (1999) observed improved growth and increased flower blooms in tissuecultured crinkle red variety of Anthurium andreanum Lind. on using earthworm body fluid (vermiwash) as a spray against the urea-sprayed plants.

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TABLE 19.5 Effect of Vermicompost on Crop Growth and Yield of the Tested Varieties Serial No.

Station

Crop Variety

Parameters

Ref.

1

Wetland at Bangalore North

2

Yield unaffected on reducing Bano et al. the chemical dosage by (1994) 50% of recommended dose when applied with vermicompost Some flowering plants Bano et al. needed minimum fertilizers (1994) along with vermicompost; no need for additional fertilizers when vermicompost was used in some cases Vegetative growth and seed No change in vegetative Kale et al. Botanical garden, Two varieties of Helianthus annuus: yield characters with different (1991) GKVK, EC 68415 and combinations of fertilizers; Bangalore Morden vermicompost with half the recommended dose of chemicals increased the yield Farmland at Arachis hypogaea and Same as above Same as above Bangalore Glycine max South

3

4

5

Summer variety paddy (1) Nutrient uptake in (4- to 5-mo crop) plants; (2) nutrient levels in soil after harvest; (3) microbial load in soil before and after harvest; (4) Mycorrhizal association with roots Vegetable garden Solanum melongena, Yield with irrigation Lycopersicon esculentum, facility in Raphanus sativus, Bangalore and Daucus carota North Botanical garden, Different ornamental Vegetative growth and GKVK, plants flower yield Bangalore

Inferences

Improved uptake of Kale et al. nutrients; increased level of (1992) N, P, and microbial load; higher level of symbiotic association

Note: To study treatment effects on various crops, the recommended dosages of fertilizers given in the package of practices of University of Agricultural Sciences, Bangalore, were taken as a control to compare the treatments using vermicompost.

Table 19.5 provides information on the yield data of different crops for which vermicomposts had been used as source of organic matter. These results authenticate that when earthworms work on organic matter, they improve the quality of the humus produced by their activities. The production of degradable organic wastes and the problems of its disposal are global issues. To protect the topsoil, to restore the sustainability of productive soils, and to rejuvenate the degraded soils are major concerns at the international level (Blussee 1994). Provision of a suitable environment in the soil by amending it with good-quality organic soil additives enhances the development of resistance in plants to pests and diseases. Minimizing the use of fertilizers and pesticides or herbicides brings down the level of air, water, and soil pollution. When farmers are asked to use more manure or soil additives to revive the productivity status of their soils, where do they get the required quantity of materials needed to feed the soil? No doubt, the crop varieties used require heavy energy subsidies to give the expected yields. By minimizing the time of humification for organic materials, by optimizing their production, and by evolving the methods to minimize the loss of nutrients during the course of decomposition, the fantasy becomes fact. Earthworms can

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TABLE 19.6 Response to Media Use by University of Agricultural Sciences (UAS), Bangalore, for Dissemination of Technology and the Range of Vermicompost Production by Different Groups on Random Data Collection Media Response Different Media

Average Vermicompost Production (kg/mo)

Different Categories Response (%)

Activity Groups

%

Range

%

<100 100–200 200–500 500–1,000 1,000–2,000 2,000–3,000 3,000–5,000 5,000–10,000 10,000–50,000

35 23 12 12 10 3 2 2 1

Newspapers, magazines, and books

48

Individuals from rural areas

55

Radio and television Workshops and exhibitions From the Centre (UAS) Other than UAS

7 10 14 20

Individuals from urban areas Nongovernment organizations Entrepreneurs Industrial units

15 10 10 5

serve as practical tools to facilitate these functions. Just as the truth lies in their serving as “nature’s plowman,” as Aristotle suggested, they act as nature’s gift to produce good humus, which is a most precious material to fulfill the need of crops. Table 19.6 and Table 19.7 provide information on the feedback received from randomly selected individuals who are using earthworms to produce vermicomposts. Information on the quality of compost produced by them, its utility, and their opinion on the use of manure for different crops is furnished to summarize the success story of vermicomposting in the state of Karnataka.

THE STATUS OF VERMICULTURE AND VERMICOMPOSTING AT THE TURN OF THE CENTURY The trend in the selection of earthworm species for vermicomposting in different places has remained the same, and the most accepted species is E. fetida. Blakemore (2000) reported the use of Anisochaeta buckerfieldi and another species of Anisochaeta for vermicomposting in Australia. He suggested that research should be targeted at studies on the taxonomy and behavior of earthworms to understand their adaptability to different organic wastes and to different environmental conditions. This could provide opportunity to maintain polycultures of earthworms rather than the current dependency on E. fetida for organic waste management. Interest in producing vermicomposts from available wastes is increasing, and organic waste availability ranges from individual homes to small farms to communities in cities. Based on this, much research has gone into this aspect to provide different types of bins and vermicomposting containers for users. The bins are marketed under different brand names. The concept of safe and scientific handling of decomposable organic wastes has become an element of awareness programs. At present, in India, when compared with the previous decade, vermicomposting has become a common part of the activities in villages, in agro-based industries (especially the breweries and distilleries), and in residential areas in cities. This has been achieved through extensive outreach programs, such as organized meetings by government and nongovernmental organizations, exhibitions, popular articles in print media, and field

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TABLE 19.7 Feedback Received from Farmers on the Effectiveness of Vermicompost on Different Crops I. Cereal Crop Variety

II. Pulses

III. Oil Seeds

Opinion

Crop Variety

Opinion

Crop Variety

Opinion

1. Jowar

+++

1. Pisum sativum (garden pea)

+++

++++

2. Oryza sativa (rice)

++++

+++

3. Zea mays (maize)

+++

4. Eleusine coracana (ragi)

+++

2. Phaseolus mungo (blackgram) 3. Cajanus cajan (Tuvar Dal/pigeongram) 4. Dolichos lablab (country bean)

1. Helianthus annuum (sunflower) 2. Arachis hypogaea (groundnut) 3. Glycine max (soybean)

IV. Spices Crop Variety

+++ +++

V. Vegetables Opinion

1. Cardamom

++++

2. Piper nigrum (pepper) 3. Murrya paniculata (curry leaf plant) 4. Curcuma longa (turmeric) 5. Cinnamomum zeylanicum (cinnamon) 6. Clove 7. Vanilla

Crop Variety

+++ ++++ +++

4. Brassica compestris (mustard) VI. Fruits

Opinion

Crop Variety

Opinion

++++

1. Mangifera indica (mango) 2. Musca paradisiacal (banana) 3. Achoros sapota (sapota)

+++

++++

1. Brassica oleracea var. Capitala (cabbage) 2. Raphanus sativus (radish)

++++

++++

3. Daucus carrota (carrot)

++++

+++

4. Solanum tuberosum (potato)

++++

+++

5. Lycopersicum esculentum (romato)

++++

+++ ++++

6. Capsicum annuum (chillies) 7. Cucurbita pepo (pumpkin)

++++ ++++

8. Luffa acutangula (ribbed gourd) 9. Cluster beans

++++

10. Cucumin sativus (cucumber) 11. Ipomoea batatas (sweet potato) 12. Phaseolus vulgaris (French beans) 13. Abelmoschus esculentus (lady’s finger) 14. Solanum melongena (Brinjal) 15. Snakegourd

+++

+++

++++ +++

4. Citrullus lanatus (watermelon) 5. Citrus spp. (lemon)

++++

6. Vitis vinifera (grapes) 7. Artocarpus heterophyllus (jack tree) 8. Ber

++++ +++

9. Punica granatum (pomegranate) 10. Annona squamosa (custard apple)

+++

++++ ++++ ++++

+++ ++++

11. Tamarindus indica (tamarind)

+++

+++ ++++ +++ (continued)

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TABLE 19.7 Feedback Received from Farmers on the Effectiveness of Vermicompost on Different Crops (Continued) VII. Ornamental Plants Crop Variety

VIII. Cash Crops

Opinion

Crop Variety

IX. Plantation Crops Opinion

Crop Variety

Opinion

1. Cocos nucifera (coconut) 2. Areca catechu (arecanut) 3. Teetona grandis (teak)

++++

1. Roses

++++

1. Theobroma cocao (cocoa)

++

2. Pyrethrum sp. (Chrysanthemum) 3. Vanda roxburghi (Orchids, vanilla) 4. Impatiens balsamina (Balsam) 5. Tagetes patula (Marigold) 6. Pimpinella anisa (Lady’s lace) 7. Polianthes tuberosa (tuberose) 8. Celotia 9. Zinnia 10. Anthurium 11. Carnation

++++

2. Coffea arabica (coffee)

+++

++++

3. Thea sinensis (tea)

++

+++

4. Morus alba (mulberry)

++++

++++

++++

++++

5. Saccharum officinarum (sugarcane) 6. Beetle leaf

++++

++++

7. Gossypium hirsutum (cotton)

+++

++++ +++

+++ +++ ++++ ++++

++++, excellent; +++, very good; ++, good.

demonstrations. Interactions between scientific and government agencies has resulted in considering vermicomposting as a need-based activity in the developmental programs. The Khadi Village Industries Commission (KVIC), which promotes and supports the rural development programs in India, is supporting the farmers financially to establish vermicomposting units in villages. The commission has also taken responsibility to set standards for the recovered vermicomposts and market the product through its sale counters. Similarly, the Directorate of Horticulture, Sericulture, Coir Board, and Spices Board are financially helping farmers establish units on their lands. Thus, much awareness has been created among the farming community regarding the use of organic matter to restore the soil productivity. This has contributed, in a developing country like India, to improve the soil productivity through organic amendments and organic solid waste management for environmental protection. The program has linked the urban and rural areas from two different angles to reach a point of developmental progress. Vermiculture and vermicomposting are included in the curriculum of higher secondary education as an introductory subject and at higher levels as an optional subject to make an in-depth study of the subject. The agricultural universities and other related institutions in the country are actively carrying out research on earthworms, vermicomposting, and its applications.

ADVANTAGES OF VERMICOMPOSTS The vermicomposts recovered are homogeneous and characterized by a strong aggregate structure. The earthworm secretions and products of microorganisms harbored in vermicompost can act as plant growth stimulators. Improved active nodulation in legumes (Kale 1997) and increased symbiotic mycorrhizal association with roots has been reported (Kale et al. 1987; Harinikumar et al. © 2004 by CRC Press LLC

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1991). An increase in total microbial activity takes place in fields receiving vermicompost (Kale et al. 1992; Nair et al. 1997). Vermicomposts have been used as a rooting medium and for establishment of saplings in nurseries (Vadiraj et al. 1993). Anand et al. (1995) reported improved germination and seedling growth of tomatoes in a medium treated with vermicompost extracts. Vermicompost extracts and earthworm body fluids (vermiwash) have been used as media for root induction in two varieties of carnation. The responses of cuttings to these fluids were better than observed with synthetic growth promoter indole butyric acid (IBA) solution (Ashwini, personal communication 2003).

EARTHWORM/MICROBIAL INTERACTIONS Another field of interest in vermiculture is microbial and earthworm interactions. The dietary composition (organic waste used) contributes to the nature of the microflora associated with earthworm activity. Addition of neem cake (2%) to organic wastes before feeding them to earthworms increased N-fixing microbial populations in the recovered vermicompost (Kale et al. 1986). A change in microbial populations was observed in earthworm-worked pressmud (Parthasarathy et al. 2001). Rajani (2001) used microbial density and enzyme activity as tools to measure the effectiveness of the process of vermicomposting. It is essential to carry out in-depth studies on the functional role of microorganism associated with earthworms on crop responses at different stages of development. An increase in the actinomycetes population has been observed in the gut region of earthworms. Different cellulolytic and P-solubilizing organisms belonging to this group have been studied. Stimulatory effects on plant growth by some of these isolates were tested in pot cultures of tomato and fingermillet (Raghavendra Rao 2001). Thus, microbial earthworm associations are an essential contributory factor in quality determination of vermicomposts. Incorporation of 12.5% rock phosphate to agricultural wastes (stubbles) with 10% cow dung and introduction of earthworms after 45 to 60 days of initial decomposition enhanced the population growth of phosphate-solubilizing microorganisms by 90 days. The response of French beans to applications of the derived vermicompost was equivalent to the control, which received the recommended chemical fertilizers at level of 7.5 to 10 t/ha (Trimurthy 2002). In general, research on vermiculture and vermicomposting is leading toward appreciation of some of the finer aspects that add to other benefits derived from vermiculture. With the transfer of the technology from laboratory to land and the role of nongovernmental organizations in liaisons between academic institutions and private entrepreneurs, vermiculture and vermicompost production is very effective in reviving organic farming in the country.

REFERENCES Anand, J.A., M.D.P. Wilson, and R.D. Kale. 1995. Effect of “Vermiwash” on seed germination and seedling growth, J. Soil Biol. Ecol., 15:90–95. Appelhof, M. 1994. Vermicomposting School Lunchroom Waste Utilizing Eisenia fetida, abstract presented at ISSE5, Columbus, OH, U.S.A. Arellano, R.P., I. Barois, and E. Arand. 1994. Earthworm Carrying Capacity for Coffee Pulp Using Eisenia andrei and Perionyx excavatus, abstract presented at ISSE5, Columbus, OH, U.S.A. Atlavinyte, O. and J. Daciulyte. 1969. The effects of earthworms on the accumulation of vitamin B12 in soil, Pedobiologia, 9:165–170. Atlavinyte, O. and J. Vanagas. 1982. The effect of earthworms on the quality of barley and rye grain, Pedobiologia, 23:256–262. Atlavinyte, O. and A. Zimkuviene. 1985. The effect of earthworms on the barley in the soil of various density, Pedobiologia, 28:305–310.

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Bano, K. and R.D. Kale. 1991. Earthworm fauna of southern Karnataka, India, in Advances in Management and Conservation of Soil Fauna, G.K. Veeresh, D. Rajagopal, and C.A. Viraktamath, Eds., Oxford and IBH, New Delhi, India, pp. 627–634. Bano, K., R.D. Kale, and G.P. Satyarathi. 1994. Vermicompost as fertilizer for vegetable crops, in Soil Organisms and Substainability, D. Rajagopal, R.D. Kale, and K. Bano, Eds., ISSBE, Bangalore, India, pp. 157–164. Bennett, I.L. and H.L. Robinson. 1967. The World Food Problem. A Report of the President’s Science Advisory Committee, Panel on the World Food Supply, Superintendent of Documents, Washington, D.C. Best, R. 1962. Production factors in the tropics, Neth. J. Agric. Sci., 10(5, Special Issue):347–353. Bhandari, G.S., N.S. Randhawa, and M.S. Maskina. 1967. Polysaccharide content of earthworm casts, Curr. Sci., 36:519–520. Bisesi, M. and M. Appelhof. 1994. Qualitative Evaluation of a Worm-A-Way Composter, abstract presented at ISEE5, Columbus, OH, U.S.A. Blakemore, R.J. 2000. Vermicology — I. Ecological Considerations of the Earthworm Species Used in Vermiculture, in Vermillennium Abstracts, Flower Field Ent., Kalamazoo, MI (September 16–22, 2000), Abstract S1-5. Blussee, P.A. 1994. Start and Development of Zeeland Wormicultures Frusan, Netherlands, abstract presented at ISEE5, Columbus, OH, U.S.A. Buckerfield, J.C. 1994. Appropriate Earthworms for Agriculture and Vermiculture in Australia, paper presented at ISEE5. Chauhan, T.P.S. 1980. Seasonal changes in activities of some tropical earthworms, Comp. Physiol. Ecol., 5:288–298. Darwin, C., 1881. The Formation of Vegetable Mould through the Action of Worms with Observations of Their Habits, Murray, London, p. 326. Diaz Cosin, D.J., J.B. Jesus, D. Trigo, and M.H. Garvin, Eds. 1999. Proceedings of Sixth International Symposium on Earthworm Ecology, Pedobiologia, 43:481–908. Dominguez, J. and C.A. Edwards. 1997. Effects of stocking rate and moisture content on the growth and maturation of Eisenia fetida (Oligochaeta) in pig manure, Soil Biol. Biochem., 29:743–746. Dutt, A.K. 1948. Earthworm and soil aggregation, J. Am. Soc. Agron., 40:407–410. Edwards, C.A. 1988. Breakdown of animal, vegetable and industrial organic wastes by earthworms, Agric. Ecosys. Environ., 24: Academic Publishers, The Hague, the Netherlands, pp. 21–31. Edwards, C.A. 1998. Proceedings of the Fifth International Symposium on Earthworm Ecology, Soil Biol. Biochem., 29:215–766. Edwards, C.A. and E.F. Neuhauser, Eds. 1988. Earthworms in Waste and Environmental Management. SPB Academic Publishers, The Hague, Netherlands. Fayolle, L., H. Michaud, D. Cluzeau, and J. Stawiecki. 1997. Influence of temperature and feeding patterns on the life cycle of the earthworm Dendrobaena veneta (Oligochaeta), Soil Biol. Biochem., 29:747–750. Frederickson, J., K.R. Butt, R.M. Morris, and C. Daniel. 1997. Combining vermiculture with traditional green waste composting systems, Soil Biol. Biochem., 29:725–730. Galli, E., V. Tomati, A. Grappelli, and G. de Lena. 1990. Effect of earthworm cast on protein synthesis in Agaricus bisporus, Biol. Fertil. Soil, 9:1–2. Garcia, M., D. Otero, and S. Mato. 1994. Reproductive Behaviour of E. andrei (Oligochaeta; Lumbricidae) in Individual Cultures with Sewage Sludge, abstract presented at ISEE5, Columbus, OH, U.S.A. Graff, O. and F. Makeschin. 1980. Beeinflüssung des Ertrags von Weidelgras (Lolium multiforum) durch Ausscheidungen von Regenwürmen drei verschiendenen Arten, Pedobiologia, 20:176–180. Gunathilakaraj, S. and T. Ravignanam. 1996. Effect of vermicompost on mulberry sapling establishment, Madras Agric. J., 83:466–467. Harinikumar, K.M., D.J. Bagyaraj, and R.D. Kale. 1991. Vesicular mycorrhizal propogules in earthworm cast, in Advances in Management and Conservation of Soil Fauna, G.K. Veeresh, D. Rajagopal, and C.A. Viraktamath, Eds., Oxford and IBH, New Delhi, India, pp. 657–664. Haynes, R.J. 1997. The concept of soil quality and its applicability to sugarcane production, Proc. Ann. Congr. South Afr. Sugar Technol. Assoc., 71:9–14.

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Jambhekar, H. 1994. Vermicompost Experience in Grape Cultivation, abstract presented at ISEE 5, Columbus, OH, U.S.A. Joshi, N.V. and B.V. Kelkar. 1952. The role of earthworms in soil fertility, Indian J. Agric. Sci., 22:189–196. Kale, R.D. 1997. Earthworm — significant contributions to organic farming and sustainable agriculture, in Organic Farming and Sustainable Agriculture, G.K. Veeresh, K. Shivashankar, and M.A. Singlachar, Eds., Association for Promotion of Organic Farming, Bangalore, India, pp. 52–57. Kale, R.D. and K. Bano. 1991. Time and space relative population growth of Eudrilus eugeniae, in Advances in Management and Conservation of Soil Fauna, G.K. Veeresh, D. Rajagopal, and C.A. Viraktamath, Eds., Oxford and IBH, New Delhi, India, pp. 657–664. Kale, R.D., K. Bano, and R.V. Krishnamoorthy. 1982. Potential of Perionyx excavatus for utilization of organic wastes, Pedobiologia, 23:419–425. Kale, R.D., K. Bano, M.N. Sreenivasa, and D.J. Bagyaraj. 1987. Influence of wormcast (Vee Comp. E. UAS. 83) on the growth and mycorrhizal colonization of two ornamental plants, South Ind. Hort., 35:435–437. Kale, R.D., K. Bano, M.N. Sreenivasa, K. Vinayak, and D.J. Bagyaraj. 1991. Incidence of cellulolytic and lignolytic organisms in the earthworm worked soils, in Advances in Management and Conservation of Soil Fauna, G.K. Veeresh, D. Rajagopal, and C.A. Viraktamath, Eds., Oxford and IBH, New Delhi, India, pp. 599–604. Kale, R.D., B.C. Mallesh, K. Bano, and D.J. Bagyaraj. 1992. Influence of vermicompost application on the available macronutrients and selected microbial populations in a paddy field, Soil Biol. Biochem., 24:1317–1320. Kale, R.D., S.N. Seenappa, and C.B.J. Rao. 1994. Sugar Factory Refuse for Production of Vermicompost and Worm Biomass, paper presented at ISEE5. Kale, R.D. and N.S. Sunita. 1993. Utilization of earthworms in recycling of household refuse — a case study, in Biogas Slurry Utilization, CORT, New Delhi, India, pp. 75–79. Kale, R.D., K. Vinayaka, and D.J. Bagyaraj. 1986. Suitability of neem cake as an additive in earthworm feed and its influence on the establishment of microflora, J. Soil Biol. Ecol., 6:98–103. Kretzschmar, A. Ed. 1992. Fourth International Symposium on Earthworm Ecology, Soil Biol. Biochem., 24:1193–1774. Karuna, K., C.R. Patil, P. Narayanaswamy, and R.D. Kale. 1999. Stimulatory effect of earthworm body fluid (Vermiwash) on crinkle red variety of Anthurium andreanum Lind, Crop Res., 17:253–257. Mba, C.C., 1994. Interaction of Tolfaria occidentalis with Two Vermicomposts, paper presented at ISEE5. Meyer, W.J. and H. Bouwman. 1997. Anisopary in compost earthworm reproductive strategies (Oligochaeta), Soil Biol. Biochem., 29:731–736. Mitchell, A. 1997. Production of Eisenia fetida and vermicompost from feed-lot cattle manure, Soil Biol. Biochem., 29:763–766. Nair, S.K., A. Naseema, K.S. Meenakumari, P. Prabhakumari, and C.K. Peethambaran. 1997. Microflora associated with earthworms and vermicompost, J. Trop. Agric., 35:68–70. Neuhauser, F., R. Hartenstein, and D.L. Kaplan. 1980. Growth of earthworm Eisenia fetida in relation to population density and food rationing, Oikos, 35:93–98. Neuhauser, F., D.L. Kaplan, and R. Hartenstein. 1979. Life history of earthworm Eudrilus eugeniae, Rev. Ecol. Biol. Sol, 16:525–534. Nielson, R.L. 1965. Presence of plant growth substances in earthworms demonstrated by paper chromatography and Went Pea test, Nature (Lond.), 208:1113–1114. Parthasarathy, K., L.S. Ranganathan, and J. Zeyer. 2001. Species specific predation of fungi by Lampito mauritii (Kinb) and Eudrilus eugeniae (Kinb) reared on pressmud, Abst. Session 3. Earthworms and Vermicomposting, in Seventh National Symposium on Soil Biology and Ecology, C.A. Virakthamath, Ed., Indian Society of Soil Biology and Ecology, India, Ceske, Czech Republic, pp. 67. Raghavendra Rao, B. 2001. Assessment of Microbial and Biochemical Quality of Urban Compost and Its Impact on Soil Health, Ph.D. (Agri.) thesis, University of Agricultural Sciences, Bangalore, India. Rajani, B.S. 2001. Biodiversity of Microorganisms and Biochemical Characteristics During Composting and Vermicomposting of Urban Solid Waste, M.Sc. (Agri.) thesis, University of Agricultural Sciences, Bangalore, India.

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Reddy, M.V. and J.R.B. Alfred. 1978. Some observations on the earthworm population and biomass in subtropical pine forest soil, in Soil Biology and Ecology in India, C.A. Edwards and G.K. Veeresh, Eds., UAS Technical Series No. 22, UAS Press, Bangalore, India, pp. 78–82. Reinecke, S.A. and A.J. Reinecke. 1997. The influence of lead and manganese on spermatozoa of Eisenia fetida (Oligochaeta), Soil Biol. Biochem., 29:737–742. Ross, D.J. and A. Cairns. 1982. Effects of earthworms and ryegrass on respiratory and enzyme activities of soil, Soil. Biol. Biochem., 14:583–587. Roy, S.K. 1957. Studies on the activities of earthworms, Proc. Zool. Soc. Calcutta, 10:81–98. Salazar, T., E. Aranda, and L. Barois. 1994. Comparative Vermicompost Production by Eisenia andrei, E. fetida and perionyx excavatus Using Coffee Pulp and Carrying Capacity for Coffee Pulp Using Eisenia andrei and Perionyx excavatus, paper presented at ISEE 5, Columbus, OH, U.S.A. Springette, J.A. and J.K. Syers. 1979. The effect of earthworm casts on ryegrass seedlings, in Proceedings of 11 Australasian Conference on Grassland Invertebrate Ecology, T.K. Crosby and R.P. Pottinger, Eds., Government Printer, Wellington, New Zealand, pp. 44–47. Stair, D.M., T.W. Hensen, M.M. Suggs, T.L. Ashwood, G.W. Suter, and B.E. Sample. 1994. Sampling of Resident Earthworms to Evaluate Ecological Risk at a Hazardous Waste Site, paper presented at ISEE5. Tomati, U., E. Galli, A. Grappelli, and G. Dilena. 1990. Effect of earthworm casts on protein synthesis in radish (Raphanus sativum) and lettuce (Lactuga sativa) seedlings, Biol. Fertil. Soils, 9:1–2. Trimurthy, N. 2002. Dynamics of Phosphate Solubilizing Micro-organisms in Organic Waste Decomposition and Its Impact on Plant Growth, Ph.D. (Agri.) thesis, University of Agricultural Sciences, Bangalore, India. Vadiraj, B.A., V. Krishnakumar, M. Jayakumar, and R. Naidu. 1993. Vermicomposting and its effect on cardamom nursery, in Proceedings of IV National Symposium on Soil Biology and Ecology, D. Rajagopal, R. D. Kale, and K. Bano, Eds., ISSBE, UAS, Bangalore, India, pp. 151–156. Vinceslas-Akpa, M. and M. Loquet. 1997. Organic matter trans-formation in lingo-cellulosic waste products composted or vermicomposted (Eisenia fetida andrei) chemical analysis and 13C CPMAS NMR spectroscopy, Soil Biol. Biochem., 29:751–758. Watanabe, H., I. Hattori, Y.R. Takai, and H. Hasegava. 1982. Effect of excrement of earthworms on deodorization of ammonia, J. Tokyo Univ. Fish., 69:11–18. White, S. 1994. The Art of Small Scale Vermicomposting, abstract presented at ISEE 5, Columbus, OH, U.S.A.

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and New 20 State-of-the-Art Perspectives on Vermicomposting Research Jorge Domínguez Departamento de Ecoloxía e Bioloxía Animal, Universidade de Vigo, Spain

CONTENTS Introduction ....................................................................................................................................402 What is Vermicomposting? ............................................................................................................402 Earthworms ....................................................................................................................................403 Earthworm Life Histories ..............................................................................................................403 Earthworm Species Suitable for Vermicomposting.......................................................................404 Temperate Species ..................................................................................................................405 Eisenia fetida (Savigny, 1826) and Eisenia andrei (Bouché, 1972)............................405 Dendrobaena rubida (Savigny, 1826) ..........................................................................405 Dendrobaena veneta (Rosa, 1886) ...............................................................................406 Lumbricus rubellus (Hoffmeister, 1843) ......................................................................406 Tropical Species......................................................................................................................406 Eudrilus eugeniae (Kinberg, 1867) ..............................................................................406 Perionyx excavatus Perrier, 1872..................................................................................407 Pheretima elongata (Perrier, 1872) ..............................................................................407 Influence of Environmental Factors on Survival and Growth of Earthworms .............................407 Temperature.............................................................................................................................409 Moisture Content ....................................................................................................................409 pH............................................................................................................................................410 Aeration...................................................................................................................................410 Ammonia.................................................................................................................................410 Effects of Diet on the Growth and Reproduction of Earthworms................................................411 Ecology of Vermicomposting: A Case Study................................................................................411 pH during Vermicomposting...................................................................................................412 Carbon Mineralization during Vermicomposting ...................................................................412 Nitrogen Transformations during Vermicomposting..............................................................412 Vermicomposting and Heavy Metal Availability ...................................................................413 Humification during Vermicomposting ..................................................................................414 Stability of Organic Wastes and Maturity of the Vermicomposts .........................................414 Vermicomposting and Human Pathogen Destruction ............................................................414 Soil Food Webs in the Vermicomposting System .........................................................................414 Applications of Vermicomposting .................................................................................................416 New Perspectives in Vermicomposting Research..........................................................................417 Operation of the Process: How Vermicomposting Works......................................................417 401 © 2004 by CRC Press LLC

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Timing of the Vermicomposting Process and Longevity of Vermicomposts ........................418 Effects of Vermicomposts on Plant Growth...........................................................................418 Vermicomposts as Suppressors of Plant Diseases and Plant-Parasitic Nematodes...............420 References ......................................................................................................................................421

INTRODUCTION The importance of biological processes in the management and recycling of organic wastes has been widely recognized; this chapter deals with vermicomposting, which is one of the most efficient methods for converting solid organic materials into environmentally friendly, useful, and valuable products for crop production. Vermicomposting is an accelerated process of biooxidation and stabilization of organic wastes that involves interactions between earthworms and microorganisms. Although Darwin (1881) already drew attention to the great importance of earthworms in the breakdown of organic matter from dead plants and the release of nutrients from them, it was necessary to wait almost 100 years until this concept was taken seriously as a technology or even a field of scientific knowledge. After 2 decades of research and technical development on vermicomposting, it is still necessary to depend on a series of fundamental aspects to understand how the process works. Certain species of earthworms, the main actors in the vermicomposting process, are described briefly in terms of biology and ecology, showing how these animals can be important organic waste decomposers to produce useful materials. The different earthworm species suitable for vermicomposting organic wastes have quite different requirements for optimal development, growth, and productivity. In this chapter, the life cycles of these species and the general requirements of ideal vermicomposting species of earthworms are first reviewed. Vermicomposting is a complex biological and ecological process; to illustrate some of the important physical, chemical, and biological actions and transformations occurring during it, a case study is presented. Although earthworms are critical in the process of vermicomposting, complex interactions among the organic matter, microorganisms, earthworms, and other soil invertebrates result in the fragmentation, biooxidation, and stabilization of the organic matter. As an example, some of the interactions between earthworms and nematodes are presented. Finally, some comments are made on the applications of vermicomposting to plant growth, and some new perspectives on vermicomposting research are discussed.

WHAT IS VERMICOMPOSTING? The disposal of organic wastes from domestic, agricultural, and industrial sources into landfills and other outlets has caused increasing environmental and economic problems, and many different technologies to address this problem have been developed and tested. The growth of earthworms in organic wastes has been termed vermiculture, and the managed processing of organic wastes by earthworms to produce casts is termed vermicomposting. Vermicomposting, which involves the breakdown of organic wastes through earthworm activity, has been successful in processing sewage sludge and solids from wastewater (Neuhauser et al. 1988; Domínguez et al. 2000); materials from breweries (Butt 1993); paper wastes (Butt 1993; Elvira et al. 1995, 1997); urban residues, food wastes, and animal wastes (Allevi et al. 1987; Edwards 1988; Elvira et al. 1996a, 1997; Domínguez and Edwards 1997; Atiyeh et al. 2000a); as well as horticultural residues from processed potatoes, dead plants, and the mushroom industry (Edwards 1988). Vermicomposting is a decomposition process involving interactions between earthworms and microorganisms. Although the microorganisms are responsible for the biochemical degradation of the organic matter, earthworms are the crucial drivers of the process by fragmenting and conditioning the substrate, increasing surface area for microbiological activity, and altering its biological © 2004 by CRC Press LLC

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activity dramatically. Earthworms act as mechanical blenders, and by comminuting the organic matter, they modify its biological, physical, and chemical status, gradually reducing its C:N ratio, increasing the surface area exposed to microorganisms, and making it much more favorable for microbial activity and further decomposition. During passage of organic matter through the earthworm gut, the fragments and bacteria-rich excrements are moved, thereby homogenizing the organic material. Vermicompost, which is the end product, is a stabilized, finely divided peatlike material with a low C:N ratio and high porosity and water-holding capacity that contains most nutrients in forms that are readily taken up by the plants. These earthworm casts are rich in organic matter and have high rates of mineralization, which reflect greatly enhanced plant availability of nutrients, particularly ammonium radicals and nitrates.

EARTHWORMS Earthworms can be defined as segmented and bilaterally symmetrical invertebrates with an external gland (clitellum) producing an egg case (cocoon), a sensory lobe in front of the mouth (prostomium), with the anus at the posterior end of the animal body, and with no limbs but possessing a small number of bristles (chaetae) on each segment. They are hermaphrodites, and reproduction normally occurs through copulation and cross fertilization, after which each of the mated individuals can produce cocoons (oothecae) containing between 1 and 20 fertilized ova (although parthenogenesis is also possible). The resistant cocoons, which can survive many years, are tiny and roughly lemon shaped with specific characteristics. After an incubation period that varies according to species and climatic conditions, the cocoons hatch. The young earthworms, which are white and only a few millimeters in length after emerging from the cocoons, gain their specific adult pigmentation within a day. Assuming favorable conditions, many species can reach sexual maturity within weeks after emergence, although some species that live mainly in soil take longer. Mature individuals can be distinguished easily by the presence of the clitellum, which is a pale or dark-colored swollen band located behind the genital pores. After fertilization, the clitellum secretes the fibrous cocoon, and the clitellar gland cells produce a nutritive albuminous fluid that fills the cocoon. The earthworms can continue to grow in size after completing their sexual development but never add further segments. The number of earthworm species is enormous; according to Reynolds (1994), there are as many as 7254 species in the Oligochaeta, of which about half (3627) are terrestrial earthworms, with an average annual description of about 68 new species. For most earthworm species, the original genus and species description is the only information available, and for many species, little or nothing is known of their life cycles, distribution, ecology, and the like. Through feeding, burrowing, and casting, earthworms modify the physical, chemical, and biological properties of the organic matter. Physical properties in soils and wastes processed by earthworms include improved aggregation, stability, and porosity; soil biological and chemical properties that may be modified include nutrient cycling (mainly N and P), organic matter decomposition rates, and chemical forms of nutrients in soil and their availability to plants. They also change the soil pH, organic matter dynamics in terms of quality and quantity, microbial and invertebrate activity (including production of enzymes and plant growth regulators), and the abundance, biomass, species composition, and diversity of the microflora and fauna (Lavelle et al. 1998).

EARTHWORM LIFE HISTORIES Earthworms, as all organisms, have to distribute the energy obtained in feeding to two main compartments: the reproductive compartment and the somatic compartment. This assignment of resources to either growth or reproduction can be modified according to evolutionary answers to

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different environmental factors. These include the availability and the quality of food as well as physical and chemical factors that can affect the earthworms directly or indirectly, modifying the availability of food and other biotic factors like competition. Finally, life histories depend on the different components of the life cycle of each earthworm species. Different species of earthworms have quite different life histories, behaviors, and environmental requirements occupying different ecological niches. They have been formally classified into three major ecological categories based primarily on their feeding and burrowing strategies (Bouché 1977): epigeic, endogeic, and anecic. Epigeic species are essentially litter dwellers; they live in organic horizons in or near the surface litter and feed primarily on coarse particulate organic matter, ingesting large amounts of undecomposed litter. These species produce ephemeral burrows into the mineral soil for periods of diapause, so most of their activities and effects are limited primarily to the upper few centimeters of the soillitter interface. They are essentially “litter transformers.” They are typically small, uniformly pigmented species with high metabolic and reproductive rates, which represent adaptations to the highly variable environmental conditions at the soil surface. In habitable tropical regions, earthworms in this category can be found aboveground in microbially rich accumulations of soil and water in the axils of plants such as Bromeliaceae (Lavelle and Barois 1984). When the environmental conditions within heterotrophic decomposition systems are unsuitable or food is limited, epigeic species are difficult to find, despite their great potential for rapid reproduction. This group of epigeic species includes Lumbricus rubellus, Eisenia fetida, Eisenia andrei, Dendrobaena rubida, Eudrilus eugeniae, Perionyx excavatus, and Eiseniella tetraedra. Endogeic earthworm species live deeper in the soil profile and feed primarily on both soil and associated organic matter. They have little pigmentation, and they generally construct horizontal, deep-branching burrow systems that fill with cast material as they move through the organic-mineral layer of the soil. Earthworms of this type can burrow deep into soils, and unlike r-selected epigeic species of earthworms, they are k-selected species (Satchell 1980; Lavelle 1983) that require a much longer time to achieve their maximum weight and appear to be more tolerant of periods of starvation than are epigeic species (Lakhani and Satchell 1970). These species are apparently of no major importance in litter incorporation and decomposition because they feed on subsurface soil material; they are important in other soil formation processes, including root decomposition, soil mixing, and aeration. Species such as Allolobophora caliginosa, Aporrectoedea rosea, and Octolasion cyaneum are included in this endogeic group of species. Anecic earthworm species live in more or less permanent vertical burrow systems that may extend several meters into the soil profile. The permanent burrows of anecic earthworms create a microclimatic gradient, and the earthworms can be found at either shallow levels or deep in their burrows, depending on the prevailing soil environmental conditions. They cast at the soil surface and emerge at night to feed primarily on surface litter, manure, and other partially decomposed organic matter, which they pull down into their burrows. Some anecic species also may create heaps of cast material termed middens at the burrow entrance; these consist of a mixture of cast, soil, and partially incorporated surface litter. Characteristically, these earthworms are large in size as adults and dark in color anteriorly and dorsally; their reproduction rates are relatively slow. Anecic species of earthworms, intermediate on the r-k scale (Satchell 1980; Lavelle 1983; Lavelle and Barois 1988), are very important agents in organic matter decomposition, nutrient cycling, and soil formation, accelerating the pedological processes in soils worldwide. Lumbricus terrestris, Aporrectodea trapezoides, and Allolobophora longa are included in this ecological anecic group of earthworms.

EARTHWORM SPECIES SUITABLE FOR VERMICOMPOSTING Looking at this general ecological grouping, it is obvious that only epigeic species can be expected to be suitable for vermiculture and vermicomposting. Moreover, to consider a species suitable for © 2004 by CRC Press LLC

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use in vermicomposting, it should possess certain specific biological and ecological characteristics, that is, an ability to colonize organic wastes naturally; high rates of organic matter consumption, digestion, and assimilation; ability to tolerate a wide range of environmental factors; high reproductive rates by producing large numbers of cocoons, which should not have a long hatching time, and growth and maturation rates from hatchlings to adult individuals should be rapid; and they should be strong, resistant and survive handling. Not too many species of earthworms possess all these characteristics.

TEMPERATE SPECIES Eisenia fetida (Savigny, 1826) and Eisenia andrei (Bouché, 1972) The closely related E. fetida and E. andrei species are the ones most commonly used for the management of organic wastes by vermicomposting. There are several reasons why these two species are preferred: They are peregrine and ubiquitous with a worldwide distribution, and many organic wastes become naturally colonized by them; they have good temperature tolerance and can live in organic wastes with a range of moisture contents. They are resilient earthworms and can be handled readily; in mixed cultures with other species, they usually become dominant, so that even when systems begin with other species, they often end up with dominant Eisenia spp. The biology and ecology of E. fetida and E. andrei, when fed on animal manures or sewage sludge, have been investigated by several authors (Graff 1953, 1974; Watanabe and Tsukamoto 1976; Hartenstein et al. 1979; Kaplan et al. 1980; Edwards 1988; Reinecke and Viljoen 1990; Elvira et al. 1996a; Domínguez and Edwards 1997; Domínguez et al. 1997; Domínguez et al. 2000). Under optimal conditions, their life cycles, from freshly deposited cocoon through sexually mature clitellate earthworm and the deposition of the next generation of cocoons, range from 45 to 51 days. The time for hatchlings to reach sexual maturity ranges from 21 to 30 days. Copulation in these species, which takes place in the organic matter, has been described by various authors since 1845 and has been observed more often than for any other megadrile species. Cocoon laying begins 48 hours after copulation, and the rate of cocoon production is between 0.35 and 1.3 per day. The hatching viability is 72 to 82%, and the incubation period ranges from 18 to 26 days. The number of young earthworms hatching from each viable cocoon varies from 2.5 to 3.8 depending on temperature. Maximum life expectancy is 4.5 to 5 years, but the average life survival was 594 days at 28°C and 589 days at 18°C, although under natural conditions it may be considerably less than these figures because they have so many predators and parasites in the wild (Edwards and Bohlen 1996). Dendrobaena rubida (Savigny, 1826) Dendrobaena rubida is a temperate species of earthworm with a clear preference for organic soils, and it inhabits substrates such as decaying rooting wood and straw, pine litter, compost, and peat and is found near sewage tanks and animal manures. Although some aspects of their biology have been investigated (Evans and Guild 1948; Gates 1972; Sims and Gerard 1985; Bengtsson et al. 1986; Cluzeau and Fayolle 1989; Elvira et al. 1996b), this species is not widely used in vermicomposting systems. Dendrobaena rubida can complete its life cycle in 75 days, and its rapid maturation and high reproductive rate could make it a suitable species for vermicomposting. Compared with other vermicomposting species, D. rubida grows relatively slowly, although it reaches sexual maturity relatively quickly (54 days after hatching). Cluzeau and Fayolle (1989) reported that it was sexually mature after 44 ± 10 days. We found that the net reproductive rate for D. rubida was 2.06 hatchlings per mature earthworm−1 week−1 (Elvira et al. 1996b), although cocoon production rates by D. rubida reported in the literature of 3.22 cocoons week−1 (Cluzeau and Fayolle 1989) are usually higher than those we reported (2.31 cocoons week−1; Bengtsson et al. 1986). Gates (1972) reported that only one earthworm emerged from 75% of the cocoons of D. rubida, with 2 © 2004 by CRC Press LLC

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to 4 hatchlings emerging from the remaining cocoons. According to Cluzeau and Fayolle (1989), one of the factors that contributes to the high fertility rate of D. rubida is because its reproduction may be facultatively biparental, amphimitic, or uniparental, either by parthenogenesis (Omodeo 1952) or by self-fertilization (André and Davant 1972). Dendrobaena veneta (Rosa, 1886) Dendrobaena veneta is a large species of earthworm with considerable potential for use in vermiculture that can also survive in soil (Satchell 1983). Although it is not very prolific and does not grow very rapidly, it is used by a number of vermiculturalists (Edwards 1988; Viljoen et al. 1991). Of the species that have been considered for vermiculture, it is probably one of the least suitable species for use in organic waste processing or vermicomposting, although it may have some potential for protein production systems and for breeding for soil improvement. Dendrobaena veneta is a robust earthworm that can tolerate much wider moisture ranges than many other species and has a preference for mild temperatures (15 to 25°C). Its life cycle can be completed in 100 to 150 days, and 65 days is the average time to reach sexual maturity. Mean cocoon production has been reported as 0.28 per day, but the hatching viability is low (20%), and the mean cocoon incubation period is 42 days. The mean number of earthworms hatching from each viable cocoon was about 1.10 (Lofs-Holmin 1986; Viljoen et al. 1991, 1992; Muyima et al. 1994). Lumbricus rubellus (Hoffmeister, 1843) This Lumbricus rubellus species is found commonly in moist soils, particularly those to which animal manures or sewage solids have been applied (Cotton and Curry 1980a,b). In surveys of commercial earthworm farms in the United States, Europe, and Australia, earthworms sold under the name L. rubellus were all E. fetida or E. andrei (Edwards and Bohlen 1996). Lumbricus rubellus has a relatively long life cycle (120 to 170 days) with a slow growth rate and a long maturation time (74 to 91 days) (Cluzeau and Fayolle 1989; Elvira et al. 1996b). We estimated the net reproductive rate to be 0.35 hatchlings earthworm−1 week−1 because of the low cocoon production rate (0.54 cocoons week−1), and only one hatchling emerged from each cocoon (Elvira et al. 1996b). Other researchers have recorded higher cocoon production rates for this species, ranging from 0.49 (Cluzeau and Fayolle 1989) to 1.75 cocoons week−1 (Evans and Guild 1948). Its low maturation and reproductive rates suggest that it is not an ideal earthworm species for vermicomposting, although its size, vigor, and ability to survive in soils could make it of interest as fish bait or for land improvement. Moreover, L. rubellus is not an opportunistic species, with obligatory biparental reproduction (Sims and Gerard 1985), which contributes to its low reproductive rates.

TROPICAL SPECIES Eudrilus eugeniae (Kinberg, 1867) The E. eugeniae species of earthworm belongs to the Eudrilidae; it is a native African species that lives in both soils and organic wastes but has been bred extensively in the United States, Canada, and elsewhere for the fish-bait market, where it is commonly called the African night crawler. It is a large, robust earthworm that grows extremely rapidly, and it is relatively prolific when cultured. Under optimum conditions, it could be considered an ideal species for animal feed protein production. Its main disadvantages are a relatively narrow temperature tolerance and some sensitivity to handling. Eudrilus eugeniae can live in soils and has high reproduction rates (Bano and Kale 1988;

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Edwards 1988); it is capable of decomposing large quantities of organic wastes rapidly and incorporating them into the topsoil (Neuhauser et al. 1979, 1988; Edwards 1988; Kale and Bano 1988). The life cycle of E. eugeniae ranges from 50 to 70 days, and its life span can be 1 to 3 years. This species is more productive in terms of rates of growth than many other earthworm species and would seem to be a suitable candidate for vermicomposting systems in regions where maintaining its optimal temperature of 25°C is both feasible and economic. Although the large size of E. eugeniae makes it much easier to handle and harvest than commonly used species such as E. fetida and P. excavatus, it seems more sensitive to disturbance and handling and may occasionally migrate from breeding beds. However, it has been grown commercially for fish bait for a long time in the United States, which is evidence that is comparatively easy to rear. It is probably one of the two preferred species, together with P. excavatus, for vermiculture and vermicomposting in tropical climates (Domínguez et al. 2001) (see Chapter 19, this volume). Perionyx excavatus Perrier, 1872 Perionyx excavatus is an earthworm belonging to the Megascolecidae, commonly found over a large area of tropical Asia (Stephenson 1930; Gates 1972), although it has also been transported to Europe and North America. This is an epigeic species that lives solely in organic wastes. High moisture contents and adequate amounts of suitable organic material are required for populations to become fully established and to process organic wastes efficiently. The life cycle of P. excavatus takes 40 to 71 days from hatching to maturity. This species prefers high temperatures and may die at temperatures below 5°C. This is a prolific species that, with about 90% hatching rate and 1.1 hatchlings per cocoon, has a net reproductive rate of nearly 20 cocoons week−1 (Edwards and Bohlen 1996; Edwards et al. 1998). Pheretima elongata (Perrier, 1872) Pheretima elongata is a megascolecid earthworm species has been tested for use in vermicomposting organic solids, including municipal and slaughterhouse wastes; human, poultry, and dairy manures; and mushroom compost in India. A project in India using this species claimed it had a commercially viable processing facility for the “vermistabilization” of 8 tons of organic solid waste day−1. These workers developed a “vermifilter” (packed with vermicompost and live earthworms) that produced reusable water from sewage sludge, manure slurries, and organic wastewaters from food processing (Edwards and Bohlen 1996). Pheretima elongata appears to be restricted to tropical regions and may not survive severe winters such as those in temperate regions. Table 20.1 summarizes some aspects of the biology of these vermicomposting earthworm species. A comparison of the duration of the life cycles and the reproduction potential of the earthworm species suitable for vermicomposting is presented in Figure 20.1.

INFLUENCE OF ENVIRONMENTAL FACTORS ON SURVIVAL AND GROWTH OF EARTHWORMS Cocoon production, rates of development, and growth of earthworms are all affected critically by environmental conditions. Species of earthworms that can be used successfully in vermicomposting are relatively tolerant of the varied environmental conditions in organic wastes, so relatively simple low-management windrow or ground bed systems have been used extensively in the past to process wastes. However, it has been demonstrated clearly that these earthworm species have well-defined limits of tolerance to certain parameters, such as moisture and temperature, and that the wastes are processed much more efficiently under a relatively narrow range of favorable chemical and environmental conditions. If divergence from these limits is great, the earthworms may move to more

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Eisenia fetida Color Size of adult worms

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Dendrobaena rubida

Brown and buff bands 4 to 8 mm × 50 to 100 mm 0.55 g

Red

Reddish purple

4 to 8 × 50 to 100 mm 0.55 g

28–30

Dendrobaena veneta

Lumbricus rubellus

Drawida nepalensis

Eudrilus eugeniae

Perionyx excavatus

Reddish brown

Reddish brown

3 to 4 × 35 to 60 mm 0.25 g

Reddish and purple Reddish brown ? bands 5 to 7 × 50 to 4 × 70 to ? 80 mm 150 mm 0.92 g 0.80 g 0.82 g

5 to 7 × 80 to 190 mm 2.7–3.5 g

4 to 5 × 45 to 70 mm 0.5–0.6 g

21–28

54

65

74–91

34–42

40–49

28–42

0.35–0.5

0.35–0.5

0.20

0.28

0.07–0.25

0.15

0.42–0.51

1.2–2.7

4.85 × 2.82 mm 18–26

4.86 × 2.64 mm 18–26

3.19 × 1.97 mm 15–40

3.14 × 1.93 mm 42.1

3.50 × 2.46 mm ? 35–40 24

? 12–16

? 18

73–80

72

85

20

60–70

75–81

75–84

90

2.5–3.8

2.5–3.8

1.67

1.10

1

1.93

2–2.7

1–1.1

+ 45–51 25°C (0–35°C) 80–85% (70–90%)

+ 45–51 25°C (0–35°C) 80–85% (70–90%)

+ 75 ? ?

? 100–150 25°C (15–25°C) 75% (65–85%)

– 120–170 ? ?

+ 100–120 ? ?

– 50–70 25°C (16–30°C) 80% (70–85%)

? 40–50 25–37°C 75–85%

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Mean weight of adults Time to maturity (days) Number of cocoons day−1 Mean size of cocoons Incubation time (days) Hatching viability (%) Number of worms cocoon-1 Self-fertilization Life cycle (days) Limits and optimal Tª Limits and optimal moisture

Eisenia andrei

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TABLE 20.1 Comparison of Some Aspects of the Biology of the Earthworm Vermicomposting Species

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A

409

B 160 140

0.7

D. veneta

100

D. nepalensis

80 D. rubida

60 40

Eisenia

20 0

P. excavatus E. eugeniae

Number of cocoons mature ew–1 day–1

Time (day)

120

P. excavatus (1.1–1.4)

L. rubellus

0.6 0.5

Eisenia fetida & E. andrei

E. eugeniae

0.4 D. rubida

0.3 0.2

D. veneta D. nepalensis

0.1 0

L. rubellus

FIGURE 20.1 (A) Length of the life cycle and (B) mean cocoon production of the earthworm species suitable for vermicomposting.

suitable zones in the waste, leave the waste, or die, so that the wastes are processed only slowly. This means that processing under some form of cover is preferable.

TEMPERATURE Earthworms have fairly complex responses to changes in temperature. Neuhauser et al. (1988) studied the potential of several species of earthworms to grow in sewage sludge, and they concluded that all these species have a range of preferred temperatures for growth ranging between 15 and 25°C. In their studies, cocoon production was restricted more by temperature than by growth, and the species studied produced most of the cocoons at 25°C. Edwards (1988) studied the life cycles and optimal conditions for survival and growth of E. fetida, D. veneta, E. eugeniae, and P. excavatus. Each of these four species differed considerably in terms of response and tolerance to different temperatures. The optimum temperature for E. fetida was 25°C, and its temperature tolerance was between 0 and 35°C. Dendrobaena veneta had a rather low temperature optimum and rather less tolerance to extreme temperatures. The optimum temperatures for E. eugeniae and P. excavatus were around 25°C, but they died at temperatures below 9°C and above 30°C. Optimal temperatures for cocoon production were much lower than those most suitable for growth for all these species. Temperatures below 10°C generally resulted in reduced or little feeding activity; below 4°C, cocoon production and development of young earthworms ceased completely (Edwards and Bohlen 1996). In extreme temperature conditions, earthworms tend to hibernate and migrate to deeper layers of the windrow or soil for protection. It appears that earthworms can acclimate to temperature in autumn and survive the winter, but they cannot survive for long periods when exposed to freezing conditions. The unfavorable effect of high temperatures (above 30°C) on most species of earthworms is not entirely a direct effect because these warm temperatures also promote chemical and microbial activities in the substrate, and the increased microbial activity tends to consume the available oxygen, with negative effects on the survival of earthworms.

MOISTURE CONTENT There are strong relationships between the moisture contents in organic wastes and the growth rate of earthworms. In vermicomposting systems, the optimum range of moisture contents for most species has been reported to be between 50 and 90% (Edwards 1998). Eisenia fetida can survive in moisture ranges between 50 and 90% (Sims and Gerard 1985; Edwards 1988) but grows more rapidly between 80 and 90% in animal wastes (Edwards 1988). Reinecke and Venter (1985) reported

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that the optimum moisture content for E. fetida was above 70% in cow manure. By comparison, E. andrei cultured in pig manure grew and matured best between 65 and 90% moisture content, with 85% the optimum (Domínguez and Edwards 1997). According to Reinecke and Venter (1985), it seems likely that lowering of the growth rate because of low moisture conditions can also retard sexual development, so earthworms of the same age could develop clitella at different times under different moisture conditions. PH

Most species of epigeic earthworms are relatively tolerant to pH, but when given a choice in the pH gradient, they moved toward the more acid material, with a pH preference of 5.0. However, earthworms will avoid acid soils of pH less than 4.5, and prolonged exposure to such soils could have lethal effects (Edwards and Bohlen 1996). Minor increases in acidity caused by addition of fresh wastes to the vermicomposting bed can be neutralized by the intestinal calcium secretions of earthworms and excreted ammonia. Lime is commonly added to vermicomposts.

AERATION Earthworms have no specialized respiratory organs; they obtain oxygen by diffusion through the body wall and lose carbon dioxide by diffusion. However, earthworms are very sensitive to anaerobic conditions, and their respiration rates are depressed in low oxygen concentrations of around 55 to 65%, e.g., at oxygen levels of 0.25 its normal partial pressure (Edwards and Bohlen 1996); feeding activities might be reduced under these suboptimal conditions. Individuals of E. fetida and other species have been reported to migrate in large numbers from a water-saturated substrate in which the oxygen conditions had been depleted or in which carbon dioxide or hydrogen sulfide had accumulated. However, they can live for long periods in aerated water, such as in trickling filters in wastewater treatment plants.

AMMONIA Earthworms are very sensitive to ammonia and cannot survive in organic wastes containing high levels of this cation (e.g., fresh poultry litter). They also die in organic wastes with large quantities of inorganic salts. Both ammonia and inorganic salts have very sharp cutoff points between toxic and nontoxic (i.e., <1 mg/g of ammonia and <0.5% salts) (Edwards 1988). However, organic wastes containing large amounts of ammonia can become acceptable after its removal by a period of composting. Outside the limits of these environmental parameters, both earthworm activity and the rates of organic waste processing decrease dramatically; for maximum vermicomposting efficiency, wastes should be preconditioned to make them suitable for vermicomposting. Earthworm population density is known to affect the rates of earthworm growth and reproduction. Even when the physical-chemical characteristics of the wastes are ideal for vermicomposting, problems can develop because of overcrowding. Reinecke and Viljoen (1990), in studies with E. fetida reared in cow manure, and Domínguez and Edwards (1997), who studied the growth and reproduction of E. andrei in pig manure, reported that, when grown at different population densities, the earthworms in the crowded dishes grew more slowly and ended with a lower final body weight, although the total weight of earthworm biomass produced per unit of waste was greater in the crowded dishes. Maturation rates were also affected by the population density; earthworms of the same age developed a clitellum at different times in cultures with different earthworm stocking rates; usually, it was later in dense populations. When the environmental conditions are maintained within adequate ranges, a maximum yield of about 10 dry unit weights of earthworm biomass can be expected from an initial 100 units (dry weight) of substrate, independent of nitrogen concentration, when a minimum of about 1% or more © 2004 by CRC Press LLC

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N is present initially (Hartenstein 1983; Edwards 1988). Although this conclusion was based on laboratory experiments, a similar earthworm yield can be expected from field systems that are managed well. It is likely that ingestion of organic waste by earthworms stops when a critical level of humified material appears, rich in free-radical and noningestible content, despite the remaining abundance of oxidizable carbon (Hartenstein 1983; Hartenstein and Neuhauser 1985). This might account for the relatively low biomass of earthworms in tropical soils despite the high availability of organic carbon from vegetation and the rapid rates of soil and organic matter turnover. However, this needs further investigation.

EFFECTS OF DIET ON THE GROWTH AND REPRODUCTION OF EARTHWORMS Earthworms obtain their energy from the organic matter on which they feed, and their effects on the characteristics of this organic matter will depend on the quality of the resource and on the earthworm species. Earthworms fragment organic wastes with a grinding gizzard, aided by grit and sand, and this increases the surface area of the organic matter and promotes very high microbial activity. Moreover, the earthworms use the microorganisms for a nutrient source rather than the organic matter. Vermicomposts can be produced from almost any kind of organic waste with suitable preprocessing and controlled processing conditions. However, the growth and reproduction of earthworms depends very much on the quality of their food resources in terms of their potential to increase microbial activity. Depending on this quality, earthworms can invest more energy either in growth or in reproduction. For example, studying the effect of different residual bulking agents (e.g., paper, cardboard, grass clippings, pine needles, sawdust, and food wastes) mixed with sewage sludge (1:1 dry weight) on the growth and reproduction of E. andrei, we found that the maximum earthworm weights achieved, and the highest growth rates, occurred in the mixture with food waste added (755 ± 18 mg and 18.6 ± 0.6 mg day−1, respectively), whereas the smallest earthworm sizes and the lowest growth rate occurred in a mixture of sewage sludge with sawdust (572 ± 18 mg and 11 ± 0.7 mg day−1, respectively). However, the earthworms reproduced much faster in the paper and cardboard mixtures (2.82 ± 0.39 and 3.19 ± 0.30 cocoons earthworm−1 week−1, respectively) compared with the reproduction in the control with sewage sludge alone (0.05 ± 0.01 cocoons earthworm−1 week−1) (Figure 20.2) (Domínguez et al. 2000).

ECOLOGY OF VERMICOMPOSTING: A CASE STUDY An experiment in our laboratory at the University of Vigo in Spain studied a vermicomposting system with different mixtures of pig manure slurries and agro-forestry by-products. This research project evaluated the characteristics of the vermicomposts produced after different processing times. The vermicomposting boxes were sampled monthly during a year; numbers and total weights of earthworms and cocoons were recorded, and several physical and chemical parameters were measured. In most of the vermicomposting systems, an initial decrease in earthworm biomass was observed at the start of the experiment; this was more marked in the mixtures of pig slurry with pine bark and pine needles. Later, the earthworm populations recovered, and their biomass increased gradually to final values that were considerably greater than the initial ones. One possible cause is that, in microcosm experiments, earthworms are unable to find suitable ecological habitats and may suffer an initial stress in activity and feeding. As a consequence, the effects of earthworms on the decomposition of the organic matter were much greater during the final stages of the process when the earthworm populations were more conditioned and

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FOOD

CONTROL
GRASS CLIPPINGS


+Wei

Mat +

Gwt +






SAWDUST

GROWTH (PC1)

PINE NEEDLES

PAPER

+ Coc


CARDBOARD

REPRODUCTION (PC2)

FIGURE 20.2 Effect of different diet treatments in the plan defined by factorial axes representing growth and reproduction of the earthworm Eisenia andrei. Mat, sexual maturity; Wei, earthworm weight; Gwt, earthworm growth; Coc, cocoon production; PC1 represents 46.5% of total inertia, and PC2 represents 36.87%.

active. Transformations of the organic wastes after 2 months during this effective vermicomposting period are summarized in Figure 20.3. PH DURING

VERMICOMPOSTING

The pH of the pig slurry used in our vermicomposting experiments ranged from 8.2 to 8.7. The vermicompost obtained after 2 months was slightly acidic, with similar pH values to the parent waste without earthworms (control), proving that earthworms did not affect the pH values to any great extent (Figure 20.3A). The effects of earthworms on pH of wastes during vermicomposting is probably related to increases in the mineral nitrogen content of the substrates, changes in the ammonium-nitrate equilibrium, and accumulation of organic acids from microbial metabolism or from the production of fulvic and humic acids during decomposition.

CARBON MINERALIZATION

DURING

VERMICOMPOSTING

Similar to other invertebrates in the organic matter decomposer community, earthworms can assimilate carbon best from the more recently deposited organic matter fractions, consisting mainly of easily degradable substances. The degradation process resulted in carbon losses by mineralization, which produced a decrease in the amounts of total organic carbon and in the carbon contributions to the organic matter (Figure 20.3B). Although earthworms consume and process large amounts of organic matter, their contributions to the total heterotrophic respiration is quite low because of their poor assimilation efficiency; only when there are large active earthworm populations, as in vermicomposting systems, can they contribute appreciably to the total heterotrophic respiration.

NITROGEN TRANSFORMATIONS

DURING

VERMICOMPOSTING

Earthworms had a great impact on the nitrogen transformations in the pig manure by enhancing nitrogen mineralization, so that most mineral nitrogen was retained as nitrates. The net total nitrogen and the different nitrogen fractions decreased during vermicomposting, and important reductions in organic nitrogen content and a high nitrification rate were noted (Figure 20.3C). This implies © 2004 by CRC Press LLC

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A

B 9 Total Organic C (g)

400

8 7

pH

6 5 4 Initial

200 100 0

Available Zn (mg kg–1)

2 Nitrate-N (%)

300

Final

C

1.5 1 0.5

Initial

Final

Initial

Final

Initial

Final

D 800 600 400 200

0

0 Initial

Final

E

F FA HA

80 60 40 20 0

Initial

100 Germination index (%)

100 Humic and fulvic organic carbon (%)

413

Final

80 60 40 20 0

FIGURE 20.3 Changes in some properties of pig manure after 2 months of effective vermicomposting with the earthworm Eisenia andrei: (A) effect of the vermicomposting process on pH; (B) carbon mineralization during vermicomposting; (C) nitriÞcation during vermicomposting; (D) heavy metal availability (extractable with ammonium bicarbonate diethylene triamine pentaacidic acid (AB-DTPA) after vermicomposting; (E) humiÞcation during vermicomposting; (F) germination index of Lepidium sativum in vermicompost from pig manure.

that earthworms (E. andrei in this case) provided conditions in the manure that favored nitriÞcation, resulting in the rapid conversion of ammonium into nitrates. Similar results have been reported by Hand et al. (1988), who found that E. fetida in cow slurry increased the nitrate concentration of the substrate (Atiyeh et al. 2000a) (see Chapter 18, this volume).

VERMICOMPOSTING

AND

HEAVY METAL AVAILABILITY

It is important to follow changes in the total and available contents of heavy metals in the organic wastes during the vermicomposting process because they may cause problems in some animal manures, sewage sludges, and industrial organic wastes. In our experiments, the total amounts of heavy metals increased (by between 25 and 30%) as a consequence of the carbon losses by mineralization during vermicomposting, and the amounts of bioavailable heavy metals tended to

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decrease with a decrease because of chemical binding of between 35 and 55% in bioavailable metals in 2 months (Figure 20.3D). Similar results have been reported in other studies of both composting and vermicomposting, and this implies a lower availability of heavy metals for plants from composts or vermicomposts. During vermicomposting, heavy metals tend to form complex aggregates with the humic acids and the most polymerized organic fractions.

HUMIFICATION

DURING

VERMICOMPOSTING

Saviozzi et al. (1988) reported that organic wastes, to be compatible with their agricultural uses and to avoid adverse effects on plant growth, must be transformed into a humuslike material and become stabilized. In our case study, decreases in the carbon from fulvic acids and increases in the percentages of the carbon from humic acids were observed throughout the vermicomposting process (Figure 20.3E), so, clearly, earthworm activity accelerates the humification of organic matter. Moreover, during vermicomposting, the amounts of humic materials increased from 40 to 60%, which was more than the values obtained in a composting process using the same waste materials. Humification processes are accelerated and enhanced not only by the fragmentation and size reduction of the organic matter, but also by the greatly increased microbial activity within the intestines of the earthworms and by aeration and turnover of the organic matter through earthworm movement and feeding.

STABILITY

OF

ORGANIC WASTES

AND

MATURITY

OF THE

VERMICOMPOSTS

The stability and maturity of organic wastes, which imply a potential for the development of beneficial effects to plants when they are used as growth media, can be determined by plant germination experiments and growth bioassays (Chen and Inbar 1993). In our example, the germination percentages of Lepidium sativum indicated that the initial organic wastes were toxic to the plants, probably because of their high ammonium content, but this toxicity was removed gradually during the vermicomposting process. Moreover, the results obtained for the germination index (which combined germination percentages and coleoptile elongations) demonstrated a beneficial effect of the earthworms on germination (Figure 20.3F).

VERMICOMPOSTING

AND

HUMAN PATHOGEN DESTRUCTION

Preliminary research in our laboratory, and in the Soil Ecology Laboratory at The Ohio State University, has shown that vermicomposting involves a great reduction in populations of human pathogenic microorganisms, as in composting. It is generally accepted that the 72 hours of the thermophilic stage of the composting process eliminate pathogenic organisms, but these studies have shown that human pathogens also do not survive vermicomposting. After 60 days of vermicomposting, the amounts of fecal coliform bacteria in biosolids dropped from 39,000 MPN (most probable number)/g to 0 MPN/g. In that same time period, Salmonella sp. dropped from <3 MPN/g to <1 MPN/g. Similar results have been reported by Eastman (1999), also for fecal coliforms and Salmonella sp. and for enteric virus and helminth ova, and other authors (see Chapter 18 this volume).

SOIL FOOD WEBS IN THE VERMICOMPOSTING SYSTEM Earthworms participate in soil functions through the drilosphere, which is defined as the space of interactions among earthworms, soil or waste physical structure, and the whole microbial and invertebrate communities (Lavelle et al. 1998). As a result of organic matter digestion processes by earthworms and the creation of soil structures (see Chapter 11 this volume), the overall composition, structure, and the relative importance of the drilosphere is clearly determined by environmental conditions, soil characteristics, and the quality and amounts of the organic matter inputs. © 2004 by CRC Press LLC

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In vermicomposting, complex interactions occurring between the organic matter, microorganisms, earthworms, and other soil invertebrates result in a rapid biooxidation and stabilization of the organic matter. Most vermicomposting systems sustain complex food webs and at the same time modify the chemical forms of several nutrient elements into longer-lived organic compounds important for nutrient dynamics (Domínguez et al. 1997). Although populations of some sensitive organisms may be reduced drastically or eliminated during vermicomposting, the substrate maintains an overall increased active community of decomposer organisms, which in addition to earthworms, includes enchytraeids, nematodes, springtails, mites, protozoa, and very large populations of microorganisms. The complex food webs in the vermicomposting systems can be represented as a pyramid with primary-, secondary-, and tertiary-level consumers. The base of the pyramid, the source of energy, is composed of decaying organic matter, including plant and animal residues. In the same way as in soil, the spatial scales at which soil organisms act in a vermicomposting system are determined mainly by their size, number, and modes of operation. At the microbial microscale, there are basically bacteria (unable to move long distances except if transported by water or larger soil organisms), fungi (in which hyphal growth provides the capacity to colonize new zones), and actinomycetes. Concomitantly, still at a microscale level but gradually increasing in size and spatial influence, the micro–food web includes microinvertebrates, such as nematodes, protozoa, and rotifers, which feed primarily on microorganisms. At the mesoscale level, there are larger organisms, such as enchytraeids and mesoarthropods, that feed on decaying organic matter, microorganisms, and microinvertebrates and are important in facilitating nutrient cycling and the small-scale dispersal of microorganisms. Finally, at the macroscale level, there is the main component of the vermicomposting system, the earthworms, which feed on and disperse microorganisms. As they feed on decaying organic matter, earthworm burrowing and tunneling activities aerate the substrate and enable water, nutrients, and oxygen to filter through it; their feeding activities increase the surface area of organic matter for microorganisms to act on. As some decomposers die, more food is added to the food web for other decomposers. As organic matter passes through the earthworm’s gizzard, it is finely ground before digestion. Then, digestive microorganisms, and possibly enzymes and other fermenting substances, continue the breakdown process. The organic matter passes out of the earthworm’s body in the form of casts, or vermicomposts, which are rich in nutrients and microorganisms and are of fine quality and structure. Earthworms can exert various influences on soil microorganisms and invertebrate populations directly or indirectly via comminution, burrowing, casting, grazing, and dispersal. Not only does the physicochemical and biological status of the organic matter and soil change for the better during the course of these activities, but the characteristics of the drilosphere may also be altered dramatically (see reviews by Brown 1995 and Doube and Brown 1998; see Chapter 12, this volume). The drilosphere is the soil system influenced directly or indirectly by earthworm activities (Lavelle 1988), whether in the gut of the earthworm (internal processes) or in its burrows and casts (external processes). As a consequence, the entire soil invertebrate community plays an important role in organic matter degradation through its interactions with soil microorganisms. Because vermicomposting systems teem with an enormous biodiversity of microorganisms and invertebrates, they provide ideal sites for complete and effective inoculation of the organic wastes with complex communities of beneficial soil organisms. This may be especially important for producing bedding plant container media and for soils that have been intensively chemically managed or have become impoverished. As the understanding of soil ecology increases, the determination and analysis of the structure of decomposer food webs in organic amendments may become an important predictive tool in evaluating their potential qualities and value. An interesting example of such interactions in the food web is the effect of earthworms (E. andrei) on nematode populations during vermicomposting (Domínguez et al. 2003) (Figure 20.4). © 2004 by CRC Press LLC

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B

12000

Nematodes (No g–1 dry wt)

Nematodes (No g–1 dry wt)

A SEWAGE SEWAGE + EARTHWORMS

10000 8000 6000 4000 2000 0 0

2

4

6

8 10 12 Time (weeks)

14

16

MANURE MANURE + EARTHWORMS

200 180 160 140 120 100 80 60 40 20 0

F

p

Earthworm 7.9 0.023 Time 5.5 0.000 Ew × Time 3.8 0.003

0

2

4

6 8 10 12 Time (weeks)

14

16

FIGURE 20.4 (A) Bacterivore and (B) fungivore nematode abundance (mean ± SE) in presence and absence of the earthworm Eisenia andrei during vermicomposting of cow manure and sewage sludge. (From Domínguez et al. 2003.)

The density of bacterivorous nematodes increased with time during the first 6 weeks in sewage sludge in the presence or the absence of earthworms; after week 6, the density of bacterivorous nematodes started to decline independently of the presence of E. andrei in the culture boxes. However, it was remarkable that the number of bacterivorous nematodes was always considerably lower in the presence of earthworms, and that these differences were statistically significant after 10 weeks (Figure 20.4A). Fungivorous nematodes did not appear in the sewage sludge during the 16 weeks of the experiment, but fungivorous nematodes appeared in cow manure after 6 weeks; their numbers increased strongly in the absence of earthworms and remained constantly low in the presence of E. andrei. After 16 weeks, the density of fungivorous nematodes was 150 ± 30 nematodes g−1dry weight in the treatment with no earthworms and only 10 ± 2 nematodes g−1 dry weight in the treatment with earthworms (Figure 20.4B).

APPLICATIONS OF VERMICOMPOSTING From the point of view of its commercial development and application, vermicomposting and any other biological treatment of organic wastes can be considered two-step processes. The first step is to convert the organic wastes into nontoxic products, eliminating or reducing human pathogen content and the concentrations of heavy metals and organic pollutants. A second step takes the process further by converting the new stable product into a valuable organic soil amendment with greatly increased microbial activities and by the humification of the organic material, which enhances the presence of plant growth promoters (Atiyeh et al. 2002b) (see Chapter 18 this volume) (Figure 20.5). So, a decision must be made, and the main criterion for this is the quality of the organic waste. If the residue is “bad,” then it may just be stabilized; if it is “good,” then there is the possibility of transforming it into a valuable organic soil amendment. A bad waste is, for example, sewage sludge because of its heterogeneous and highly variable composition and high concentration of human pathogens and organic and inorganic contaminants. In this case, the objective of the treatment should be to rapidly stabilize the material, and probably regular composting is a better solution than vermicomposting. A good waste, for example, is most animal wastes, wine residues, food wastes, or milk industry sludges. They are good because they have no pollutants, have a homogeneous composition, and have a good balance of nutrients; in this case, the objective could be to obtain a good organic soil amendment, and vermicomposting is probably the better choice (Figure 20.5).

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Organic residues

Stable product (nontoxic)

417

Valuable organic fertilizer

Pathogens

Fulvic and humic acids

Heavy metals

Specific microflora activity

Organic pollutants

Plant growth regulators

Residue quality Bad

Stabilization (composting)

Good

Organic fertilizer (vermicomposting)

FIGURE 20.5 Vermicomposting and any other biological treatment of organic wastes can be considered as two-step processes.

NEW PERSPECTIVES IN VERMICOMPOSTING RESEARCH Although an interest in vermicomposting research and technology has been increasing at a rapid rate and the body of knowledge now available is quite large, there are still many questions to be investigated. The topics of some of these questions are discussed here: the method of operation, the management of the process, the effect of the earthworms, the timing of the process or the resident time of the wastes to be converted into real biological soil amendments, the reason why vermicompost promotes plant growth, or the “mean life” or “expiration date” of these vermicomposts.

OPERATION

OF THE

PROCESS: HOW VERMICOMPOSTING WORKS

The classical approach in all research projects on vermicomposting and in the case study presented in this chapter consists of adding earthworms into organic residues and obtaining vermicompost (see Figure 20.6A). With a reductionist approach such as this, it is difÞcult to know how the vermicomposting system works and the role of the earthworms in the process. Figure 20.6B represents a simpliÞed model of the vermicomposting process; in ways similar to nature, earthworms obtain their energy from the organic matter, and their effects on the waste depend on the quality and quantity of the resource and on the earthworm species used. The vermicomposting process consists basically of two different subprocesses. The Þrst involves the earthworm gut-associated processes (GAPs), which include all the modiÞcations that the organic matter undergo during transit through the intestinal tract, including the transformation of nutrients, modiÞcations and increases in microbial diversity and activity, modiÞcations of microfaunal populations, homogenization, and the processes of digestion, assimilation, and excretion of the wastes. Once the earthworm GAPs end, the resultant casts are exposed to cast-associated processes (CAPs); here, the effects of the earthworms may be only indirect and include aeration of the substrate because of the burrowing activities. Moreover, earthworm casts are subject to an aging process and to the action of microorganisms and microinvertebrates present in the substrate; it is important to note that, during action of the vermicomposting systems, the casts are mixed with materials that were not eaten by the earthworms.

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A

? +

VERMICOMPOST

B RESOURCE QUALITY

Earthworm gut

ORGANIC MATTER

CASTS (t0 SPECIES

tx )

GAPs

CAPs

Nutrients addition Microbial activity Homogenization Digestion Assimilation Excretion

Aging Microbial activity Microfauna activity Mineralization Humification

FIGURE 20.6 Different strategies on vermicomposting research: (A) classical approach utilized in most studies; (B) split approach in two different processes, gut-associated processes (GAPs) and cast-associated processes (CAPs).

One example of these two different ways to study the vermicomposting process is given in Figure 20.7. The number of total coliform bacteria remains more or less stabilized in pig manure in the absence of earthworms and decreases sharply in the presence of earthworms (Figure 20.7A); the passage through the earthworm gut (GAP) does not affect human pathogen populations (Figure 20.7B), and the effect of the CAPs on human pathogen numbers after 2 months is quite different in the noningested waste, in the mixture of casts and waste, and in the casts, with these presenting the lowest numbers (Figure 20.7C). Studies of the vermicomposting process through such different compartments can give more detailed information of the biotic and abiotic changes in the organic wastes.

TIMING

OF THE

VERMICOMPOSTING PROCESS

AND

LONGEVITY

OF

VERMICOMPOSTS

Directly related to the explanation above is the timing of the vermicomposting process, such as when the process is completed, and if there is an optimum level of biotic activity in the vermicompost. The Þrst phase, the passage of the organic matter through the earthworm gut (GAP), is rapid, and it is important to determine the magnitude and importance of the biochemical and physical changes during this gut transit. Once the fresh earthworm casts are deposited, then something like a “maturation” process of these casts starts. During this aging, vermicomposts reach an optimum in terms of biological properties that promote plant growth and suppress plant diseases. There are no data about when this optimum is achieved, how it can be determined, and how to determine if this optimum has some kind of longevity or expiration date (Figure 20.8).

EFFECTS

OF

VERMICOMPOSTS

ON

PLANT GROWTH

Earthworms have beneÞcial physical, biological, and chemical effects on soils, and these effects can increase plant growth and crop yields in both natural and agroecosystems (Edwards and Bohlen 1996; Edwards 1998) (see Chapter 18 this volume). These beneÞcial effects have been attributed to improvements in soil properties and structure (Kahsnitz 1992), to greater availability of mineral

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A TOTAL COLIFORMS Control

Earthworms 0

10

20

time(days)

B

C GAPs

CAPs

TOTAL COLIFORMS TOTAL COLIFORMS

WASTE

CAST

WASTE

WASTE + CAST Time (´)

CAST

FIGURE 20.7 A case example of different strategies on vermicomposting research: (A) classical approach; split approach in (B) gut-associated processes (GAPs) and (C) cast-associated processes (CAPs).

nutrients to plants (Gilot 1997), to enhancement of mycorrhizal infection, to control of plant parasitic nematode populations, to increased microbial populations, and to biologically active metabolites such as plant growth regulators (Tomati and Galli 1995; Doube et al. 1997) and humates (Atiyeh et al. 2002b). The effects of vermicomposts on the growth of a variety of crops, including cereals, legumes, vegetables, ornamental and ßowering plants, and trees, have been assessed in the greenhouse and to a lesser degree in Þeld crops (Chan and GrifÞths 1988; Edwards and Burrows 1988; Wilson and Carlile 1989; Mba 1996; Thankamani et al. 1996; BuckerÞeld and Webster 1998; BuckerÞeld et al. 1999; Nethra et al. 1999; Atiyeh et al. 1999, 2000b,c, 2002a,b) (see Chapter 18, this volume). These investigations have demonstrated consistently that vermicomposts have beneÞcial effects on plant growth independent of nutrient transformations and availability. Whether vermicomposts are used as soil additives or as components of horticultural soilless bedding plant container media, vermicomposts have consistently improved seed germination, enhanced seedling growth and development, and increased plant productivity and yields much more than would be possible from the mere conversion of mineral nutrients into more plant-available forms. The greatest plant growth responses and yields have occurred constantly when vermicomposts constituted a relatively small proportion (10 to 40%) of the total volume of the plant growth medium in which they are incorporated. Usually, greater proportions of vermicomposts substituted in growth media have not increased plant growth as much as smaller proportions (Atiyeh et al. 1999, 2002a,b). This could © 2004 by CRC Press LLC

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stable

Vermicompost quality

nonstable

fresh casts Time

FIGURE 20.8 Timing of the vermicomposting process. There are no data on when a vermicompost can be considered optimum, on how this optimum can be determined, and if this optimum has some kind of “expiration date.”

be because of adverse growth factors, such as high levels in 100% vermicomposts, particularly those from animal wastes. In spite of all this research on the effects of vermicomposts on plant growth, there are still few data in the literature validating possible mechanisms by which vermicomposts produce these growth enhancement effects.

VERMICOMPOSTS NEMATODES

AS

SUPPRESSORS

OF

PLANT DISEASES

AND

PLANT-PARASITIC

Although there are not many studies regarding vermicomposts as suppressors of plant diseases and plant-parasitic nematodes, it has been shown that the incidence of plant diseases can be limited by vermicomposts. Substrates supplemented with vermicompost were suppressive to root rot of tomato caused by Phytophthora nicotiane var. nicotianae, and dipping cabbage roots in a mixture of clay and vermicompost decreased infection by Plasmodiophora brassicae (Szczech et al. 1993); they also reduced infection of tomato plants by Fusarium oxysporum f. sp. lycopersici (Szczech 1999). Vermicompost at a concentration of 40 µg/ml caused a 50% reduction of zoosporangia formation of Phytophthora cryptogea, and amendment of soil extract with 1000 µg ml−1 of vermicompost completely inhibited the pathogen sporulation. Peat drenched with vermicompost extracts immediately after planting of gerbera, ivy, carnation, or cyclamen significantly suppressed the spread of diseases. The compound applied at a concentration of 25% caused a decrease of about 50% of propagule numbers of Fusarium oxysporum f. sp. dianthi in peat naturally infested with the pathogen (Orlikowski 1999). Vermicompost incorporation at a 20% rate reduced the incidence of diseased plants of gerbera (Gerbera jamesonii H. Bolus), the area under the disease progress curve, and the disease growth rate of the fungi Rhizoctonia solani, Phytophthora drechsleri, and Fusarium oxysporum (Rodríguez et al. 2000). Chaoui et al. (2002) demonstrated suppression of Pythium, Rhizoctonia, and Verticillum by vermicomposts. Arancon et al. (2002, 2003) demonstrated consistent suppression of plant-parasitic nematode populations by vermicomposts under pepper, tomatoes, strawberries, and grapes in the field.

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Clearly, vermicomposts cannot only suppress plant pathogens and plant parasitic nematodes, but they can also promote germination, growth, yields, and fruiting of many plants.

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