Vital Soil, Volume 29: Function, Value and Properties (Developments in Soil Science)

Developments in Soil Science - Volume 29

Vital Soil Function, Value and Properties

Developments in Soil Science

Series Editors: A.E. Hartemink and A.B. McBratney

Cover illustration: "Vital Soil" as illustrated by Dr. Ruben Smit; www.outdoorvision.nl. The picture is taken five kilometres from Wageningen (the Netherlands) on 'the Grebbenberg'.

Developments in Soil Science - Volume 29

Vital Soil Function, Value and Properties

Edited by

Peter Doelman Doelman Advise Wageningen, The Netherlands

Herman J.P. Eijsackers Environmental Science Group Wageningen University and Research Centre, The Netherlands

2004

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CONTENTS List of contributors Foreword Preface Acknowledgements I Introduction 1. Leading concepts towards vital soil H. Eijsackers II General aspects of vital soils 2. The formation of soils N. van Breemen 3. Vegetation, organic matter and soil quality W.H.O. Ernst 4. Soil biota and activity H.A. Verhoef III Integrative approaches in soil biology 5. The key role of soil microbes W. Verstraete and B. Mertens 6. The use of soil invertebrates in ecological surveys of contaminated soils N.M. van Straalen 7. Balance and stability in vital soils P.C.deRuiter 8. Soil and higher organisms: from bottom-up relations to top-down monitoring N.W. van den Brink IV Natural attenuation and monitoring 9. Fate of contaminants in soil W.J.G.M. Peijnenburg 10. Ecological soil monitoring and quality assessment A.M. Breure V Synthesis and outlook 11. Synthesis for soil management P. Doelman Subject index

vii xi xiii xvii 1

21 41 99

127

159 197

215

245 281

307 333

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LIST OF CONTRIBUTORS Chapter 1: Leading concepts towards vital soil Author: Prof. Herman J.P. Eijsackers

e-mail: [email protected] Herman Eijsackers studied ecology, ethology and applied entomology at Leiden University. In 1970 he moved to the Research Institute for Nature Management for his PhD-studies on the "Sideeffect of 2,4,5-T on the soil fauna". His research work broadened to ecotoxicological research on the impacts of acidification, heavy metals and pesticides on soil biota. Next to his scientific work he became involved in research management. Since 2000 he has been scientific director of the Environmental Sciences Group of Wageningen University and Research Centre. He has also two extraordinary professorships: at the Vrije Universiteit Amsterdam and at the University of Stellenbosch SA.

Chapter 2: The formation of soils Author: Prof. Nico van Breemen

e-mail: [email protected] Nico van Breemen graduated from Wageningen University in 1968, where he obtained his PhD in 1978, and became Professor of Soil Formation and Ecopedology in 1986. His scientific interests involved formation of acid sulphate soils, problem soils for wetland rice, forest soils influenced by atmospheric deposition of sulphur and nitrogen, relationships between climate change and soils, and soil-plant feedbacks and their evolutionary implications.

Chapter 3: Vegetation, organic matter and soil quality Author: Prof. Wilfried H.O. Ernst

e-mail: [email protected] Wilfried H.O. Ernst obtained his PhD at Miinster University, on the subject "Ecology and phytosociology of plant communities on heavy metal-enriched soils in Central Europe and the alpine mountains". He has been professor in Botany for 30 years, guiding more than 50 PhD students. His research can be summarized by the title of his valedictory lecture of 2002: "Living at the boundaries of life".

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List of contributors

Chapter 4: Soil biota and activity Author: Prof. Herman A. Verhoef

e-mail: [email protected] Herman Verhoef obtained his PhD at the Faculty of Biology of Vrije Universiteit Amsterdam. The title of his dissertation (1978) was "An ecological study on water relations in Collembola". After having been involved in the ecophysiology of soil animals, he has turned his attention to community ecology, including food web ecology. He is professor in Soil Ecology at the Vrije Universiteit in Amsterdam, focussing on "soil-plant biocomplexity" and the relations between spatial heterogeneity and biodiversity.

Chapter 5: The key role of soil microbes Authors: Prof. Willy Verstraete, Birgit Mertens

e-mail: [email protected] Willy Verstraete obtained a degree as bio-engineer from Ghent University and a PhD from Cornell University. He is director of the Laboratory of Microbial Ecology & Technology at Ghent University. His team focuses on microbial processes based on complex microbial communities. Birgit Mertens holds a degree as bio-engineer from Leuven University. She is still working on her PhD at Ghent University on the bioremediation and monitoring of HCH polluted soils. She was actively involved in the writing, the conceptualization and editing of Chapter 5.

Chapter 6: The use of soil invertebrates in ecological surveys of contaminated soils Author: Prof. Nico M. van Straalen

e-mail: [email protected] Nico van Straalen obtained his PhD in 1983 on "Comparative Demography of Collembola" at Vrije Universiteit Amsterdam. He turned to mechanistic research on how Collembola are able to tolerate high exposures to heavy metals and the molecular genetics of metallothionein induction as a basis for evolution of tolerance. He has been a professor of Animal Ecology at the Vrije Universiteit since 1992. He is the leader of a research group on Evolution and Soil Ecology, conducting studies of stress responses in the soil community.

List of contributors

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Chapter 7: Balance and stability in vital soils Author: Prof. Peter C. de Ruiter e-mail: [email protected] Peter de Ruiter obtained his PhD in 1987 on "Predation by a Carabid Beetle" at the Utrecht University. He has been full professor at Utrecht University in the field of Environmental Sciences and his main interest is in the community structure and ecosystem functioning.

Chapter 8: Soil and higher organisms: from bottom-up relations to top-down monitoring Author: Dr. Nico W. van den Brink e-mail: [email protected] Nico van den Brink is senior scientist at Alterra, with a background in ecological risk assessment involving vertebrates and contaminants. His PhD "Probing for the invisible" at Groningen University focussed on the monitoring of global background levels of contaminants in Antarctic penguins and petrels. He has been working on the development of animal friendly methods in ecotoxicological research. He has led several projects on the assessment of risks of contaminants for wildlife in the Netherlands. His current research activities are focussed on the development of spatially explicit risk assessments.

Chapter 9: Fate of contaminants in soil Author: Dr. Willy J.G.M. Peijnenburg e-mail: [email protected] Willie Peijnenburg is a senior scientific staff member of the Laboratory for Ecological Risk Assessment of the National Institute of Public Health and Environment. His PhD study was on "Photochemically induced rearrangement reactions of organic compounds" at Eindhoven University. His main scientific interests include the implementation of bioavailability of heavy metals in risk assessment procedures and the development and application of QSARs for the estimation of physical-chemical properties and transformation rates of chemical substances in the environment.

x

List of contributors

Chapter 10: Ecological soil monitoring and qualify assessment Author: Dr. Anton M. Breure

e-mail: [email protected] Anton M. Breure is head of the Laboratory for Ecological Risk Assessment of the National Institute of Public Health and environment. He obtained his PhD at Amsterdam University on "Anaerobic waste water treatment". His main scientific interest is impact assessment of environmental and managerial stress on ecosystem quality.

Chapter 11: Synthesis for soil management Author: Dr.ir. Peter Doelman

e-mail: [email protected] Peter Doelman studied Plant Pathology at Wageningen University. He obtained his PhD in 1978 at Groningen University on the "Effects of lead on the soil microflora". After 20 years of scientific research on the effect of heavy metal on soil biota and on bioremediation of POPs at the Research Institute of Nature Management, he joined the practical world for "the great clean up". He prefers and enjoys a role between knowledge and practice in the biology of soil.

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FOREWORD In the world of environmental affairs people are largely aware of what is harmful to human beings and, as a result of extensive research, the risk of chemical pollution to humans is well documented. In most countries, too, environmental legislation focuses on reducing human risk to below an acceptable level. Knowledge of the impact of pollution on the wellbeing of our ecological system is, however, much less developed. The effect on bacteria and other micro-organisms, as well as on larger animals living underground, still requires a great deal of research. The editors of this book, Vital Soil, invited a dozen highly qualified and well-known Dutch and Belgian scientists to give their views, based on their field of expertise. They were challenged to think outside the scientific box and come up with new perspectives. As a result, policymakers can now refocus on newly acquired knowledge and experience. Life on earth appears to be very robust; it was established millions of years ago, and has taken all that time to evolve into what we see today. But careless human action can cause great destruction, with disastrous effects for many years to come. We must realise that we can use the earth's riches for a better quality of life, without destroying the ecological system that supports life. I hope that this book will help us to achieve that ambition. Pieter van Geel, State Secretary for the Environment

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PREFACE In 1979 the Dutch Ministry of Housing, Spatial Planning and Environmental Health started to cooperate with the Dutch Research Institute for Nature Management to carry out research on the impact of soil contamination. Results of this research have been used to strengthen the scientific rational for the Soil Protection Act. We have been intensively involved in that programme since then. During the years 1974-1983 joint international workshops were organised on such issues as the side effect of pesticides, and discussions were held on proposed methods to determine side effects and evaluate consequences. Terms such as flexibility, resilience and reversibility were used and we realized the consequences. Now, 25 years later, we present a book on soil life, in which scientific knowledge is presented in relation to its consequences for policy and practice. A book meant for researchers, students, soil managers and legislators. We aim for a book with a view, being aware that a certain degree of subjectivity (personal opinion) will always be involved. So we asked experienced scientists to co-author this book. As scientists with an impressive and long track record in all different aspects of soil science, they were invited to give their view and vision: what is the state of current knowledge, how can it be applied, what should be monitored, where should the development in research and in improved monitoring come?

Figure A. Soil scientists were invited to look into the soil and to give their view

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Preface

Nine soil scientists introduce their own field of expertise. They provide an account of the state of the art, the importance and the function of their field. Next they express their views on the possibilities and impossibilities of monitoring their aspects of the quality of soil. What are the vital issues involved in the functioning and quality of the soil. In answering this question they give special attention to soil contaminants. We structured the book in four major clusters. The introduction to the book provides a broad view of the problem of combining knowledge on the functioning of soil ecosystems with political and practical judgement and assessment (Eijsackers). Soil is the result of interactions of various processes and factors including climate, vegetation, geology, soil flora, soil fauna, land structure, water and time. These factors can be classified as the physical, chemical and biological aspects of the soil. These three groups of aspects have been combined into the four major clusters, the first one focusing on the formation of soil (Van Breemen), the formation of organic material (Ernst), and the recycling of organic matter (Verhoef). The second one is a specific soil biota cluster with the soil microflora (Verstraete and Mertens), the soil invertebrates (Van Straalen), a chapter on stability due to food webs by De Ruiter, and the chapter by Van den Brink dealing with the soil as a subway to food chains. Here soil microflora and soil fauna are exhibited as subgroups but also in connections, such as food webs and food chains. The third cluster provides knowledge on the nature and behaviour of contaminants in soil (Peijnenburg) and on the development of monitoring programmes (Breure). In the fourth cluster the various views are evaluated and combined in a synthesis for soil management (Doelman). As a result we present an up to date picture of the fascinating diversity in soil processes and structures, biotic as well as abiotic. In order to structure and focus this vast array of expertise on monitoring and assessing soil quality in practice, we provided the authors with a series of questions in advance: - What are the general principles, processes/roles/functions, and properties of vital soils? - How relevant are these characteristics for more or less homogenized soil systems like in cities, and heterogeneous areas where different soil forming processes can take place, like in rural and nature areas? How, and to what extent, are these characteristics threatened by contaminant groups such as heavy metals, POPs, and eutrophication? - Are there other threats and how can quality be established both in relation to soil protection and soil recovery? - What methodologies can we use to monitor these qualities?

Preface

xv

As a result, various options were presented, mostly complementary, sometimes contradictory. In the synthesis these shared and opposing views come together. With this knowledge and views we aim to bridge the gap between science and practise. So we want "Vital Soil" to be a book with a vision, one in which the state of the art and developments are presented. It is a book that recommends monitoring approaches, methodologies and characteristics in arranged tables and figures, but it is not a manual. Soils have a number of shared general features that can be monitored with a generic monitoring scheme, but in addition different soils have their own specific peculiarities that demand a proper and experienced interpretation of the monitoring results. Therefore we suggest that, if necessary, these generic monitoring activities can be supplemented by tailor-made monitoring measurements. Therefore, we should not only look more careful but also look more intensely. This book encourages us to look into the soil itself! Peter Doelman & Herman Eijsackers, Wageningen, April 26, 2004

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ACKNOWLEDGEMENTS "What is an individual without society?" "What determines the structure of a community?" This book is the result of many actions and inter-actions. It has emerged as a result of inspiration, labour, control, listening and stimulation. Its main sponsor was VROM, the Dutch Ministry of Spatial Planning, Housing and the Environment. Ruud Cino was the initiator, while Miech de Steenwinkel was stimulator and controller. Trudy Crommentuyn and Sandra Boekhold of the Ministry, together with Joop Vegter of the Technical Committee Soil Protection, formed the internal scientific steering committee. Their regular support was indispensable. Without the harmony of attitude and views displayed by the authors of this volume, there would not have been any basic chapters and consequently no book. We thank the external scientific committee, Lijbert Brussaard and Alex Zehnder, for ensuring scientific soundness. Elsevier's Critical Review Committee and Mara Vos Sarmiento and Joyce Happee played an important role in shaping the resulting text. Marilyn Minderhoud Jones supervised the correctness and readability of the English text and Yvonne de Ruigh took care of layout and text editing. Where would editors be without such a team?

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Chapter 1

LEADING CONCEPTS TOWARDS VITAL SOIL H. Eijsackers

Abstract This introductory chapter gives an analysis of the basic elements of a vital soil, of soil protection policies and monitoring options. The background to this approach is the increase in soil functions and an overexploitation that has resulted in conflicts as well as in concern for the consequences for human health, the health of soil, and soil sustainability. These functions and their complexity are described in some detail to show the complex nature of the soil scientist's task (Chapter 2-10), assess the state of the art and assess monitoring options. Soil ecology started from phenomenology, describing the phenomena observed in the field, with a strong emphasis on the dynamics of single species. Now it is gradually entering a phase of understanding and elaboration: research based on the formulation of hypothesis. Major research items are derived from the observation of fluctuations in numbers and activities in space and time. In soil vitality, stability and the restoration of a solid state of functional diversity is of critical importance. It is important to know how many years it may take before the restoration of certain processes, functions or desired ecosystems will be completed. Vegetation may require hundreds of years and similar time frames may also be necessary for soil microflora and soil invertebrates. However, in the case of larger animals or toppredators it may take longer. Of the threats to soil that have been recognized worldwide like erosion, contamination, decline of organic matter, sealing, compaction, salinization, flooding and landslides, and loss of biodiversity, the impact of soil contamination is discussed extensively. Furthermore, we also deal briefly with current political thinking on soil protection. The consideration of soil protection proceeds from the principle of multifunctionality in which the maintenance of each potential function must be secured for the benefit of future generations. This has meant a more tailor-made approach that is related to a realistic use of the potential of a specific soil at defined sites set in the perspective of high costs and the impacts of the remediation actions. In the European Soil Strategy currently being developed, the prime aim is to adjust existing policies that have a side effect on soil to a common strategy towards soil protection. Soil in all its diversity of structure and processes demands an integrated approach to both monitoring and research. Soil combines abiotic and biotic processes in a static, dynamic and structural way. Monitoring and studying biodiversity from a functional perspective is a major item.

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1.1 Vital soil for a vital world 1.1.1 Soil is the solid basis of life

Soil is the solid basis of the world and society. Vital soil is essential for a vital world. Soil provides the basis for houses, roads and industries, enables crop to be grown and food to be produced, as well as harbouring groundwater supplies, producing and recycling organic material, and storing and recycling organic waste. These different types of functions have all too easily become combined. By steadily putting higher demands on each function, problems have been created that threaten sustainability. Hence, choices have to be made concerning the combination of functions that fit best in specific circumstances. What is appropriate for sites that have been impaired by increased levels of nutrients and what should happen to sites that should have been preserved better because of their pristine character. For centuries we have effectively used the multiple benefits derived from primary production, to produce plants, food and fodder. More recently, secondary production has continued to develop with new technologies to recycling the elements and purify soil and groundwater. What happens when one of the soil's natural functions becomes heavily exploited? If a support function becomes maximized because the soil surface has been sealed with bricks or concrete, other functions will be hampered. Soil life under these sealed soils will be reduced because of reduced moisture content and so the capacity to clean rapidly disappears. However, local people are convinced that rainwater will be re-routed to other non-sealed areas, where incoming rainwater will increase and with all probability, surpass the natural infiltration capacity of the soil. The result may well be torrential surface flows that can lead to landslides and soil erosion. Bulldozing soils, deep-ploughing and topping up soils with sandy layers to improve the support-functions can also destroy soils natural structure and lead to all kinds of erosion. In agriculture, the nutrient supply has been maximized through the application of animal manure and chemical fertilizers to such an extent that groundwater locally has become overloaded with nutrients. In a number of European countries, for example, this has resulted in groundwater with nitrate levels that fail to meet the requirements set for health, and to surface water with increased nutrient levels. Monocultures and lack of crop rotation result in heavy pesticide use to ensure crops are protected. Incidental overdosing can result in elevated pesticide contents in ground and surface water. To make a potentially long story short, combining soil functions or piling them up in the same area with the result that they are over-exploiting, together

Leading concepts towards vital soil

3

with accidental spills and illegal dumpings has lead to a growing concern about human and soil health. As a result legislation has been put in place to deal with these new developments. As far as soil health is concerned law should be developed on criteria based on natural functioning. When soils are treated in a proper way and their characteristics used optimally, they can be developed sustainably. This does not imply that all functions have to proceed at the same speed everywhere, but it does mean that all functions remain possible at all places, the so called multi-functionality of soil. Bearing all these aspects in mind, measuring and monitoring soil quality must include the physical and chemical processes involved in the formation of the abiotic part of the soil, and the biotic processes, carried out by multiple life forms in their natural habitat. These include mutual interactions over soil issues that challenge soil scientists' knowledge, experience and understanding and suggest that soil properties should be seen as key items in soil sustainability. So practice and policy obtain monitoring tools for protection. The aim of this book is to contribute to soil policy and practice but in the context of the knowledge generated by the scientific world tentatively and schematically illustrated in Figure 1. This introductory chapter gives a historical analysis of the basic elements of a vital soil, and the elements of soil protection policy. A basic view is given as an introduction to the more detailed views of soil scientists. Among the threats to soil recognised in Europe such as erosion, contamination, decline of organic matter, sealing, compaction, salinization, flooding and landslides and the loss of

Figure 1. The view of professional soil scientists spinning off to policy and practice

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biodiversity, special attention is given to biodiversity, organic matter and soil contamination. Basic monitoring aspects will then be referred to and discussed. In this way I hope to pave the way for the specialist in the arts of vital soil. 2.2.2 Historical awareness of vital soil

Scientific awareness Soil science is the study and description of the processes in which geological bedrock material is transformed into a layered mixture of inorganic (mineral) and organic compounds. It describes the physical character of this mixture in relation to gas and water management. It also describes its nutritive value for agriculture, its supportive power for constructions (buildings, dikes and tunnels) and its filtering capacity for drinking water. We have also been using our soils as sink and, implicitly, as clean-up facilities. This reflects the present situation in which mankind has been using, managing and adapting soils for its own wealth and well-being. Consequently, a system has to be developed and applied that measures the qualities of our soils in terms of their present use value, bearing in mind their potential uses and our commitment to using our soils in a sustainable way. This system should also include the methods by which we measure these qualities and, implicitly the impacts on these qualities. This thinking and approach can be observed in the development of soil policy. Policy awareness Soil protection policy started in the 1970s, considerable later than water and air protection policy. These actions were triggered by the desire to protect the multiple functions of soil, and were speeded up by a number of soil contamination incidents in different parts of the world. These incidents differ in character and consequently resulted in different policy accents. Compare for instance the Superfund-problems in the United States with the brown fields in the United Kingdom, the industrial hotspots in Germany and the problem of diffuse over-fertilization and acidification of soils in the Netherlands. Major elements in soil protection have been: - impacts on human health, translated to uptake of contaminants in food crops and in drinking water; - impacts on dispersal of contaminants, especially in relation to groundwater, but also in relation to evaporation from soils under buildings; - impacts on soil ecosystem, measured by ecotoxicological effects on soil biota and bioaccumulation in terrestrial food chains. Of these three elements the impacts on the soil ecosystem have been given the

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least attention. In many countries extensive research programmes on ecotoxicological impacts have been carried out on the effects of hazardous concentrations (HC5 or HC50 methods). However, there has been considerable debate about the application of the statistically derived HC5-method, the use of reference levels, and ways of assessing the impacts of natural elements like heavy metals. Another fierce debate has been about the protection of the multiple functions of the soil. Should all potential functions be maintained so that an affected soil be allowed to recover its natural status, or should a present function be standardized and recovery levels set accordingly? Presently soil policies are moving towards a functional approach, translated into soil use standards for various types of land use. In the EU, policies on the maintenance of soil quality and soil remediation are merging into the new EU Soil Strategy (EU Commission, 2002). According to this strategy, remediation based decision support programmes like Clarinet and Caracas will lead to Risk Based Land Management. As a consequence it can be foreseen that the EU Water Directive will get a complementary EU Soil Directive. Table 1 shows historical landmarks in soil protection policy in European countries. 2.2.3 Reconstruction of soils

As a result of natural soil forming processes, soils have evolved into specific layered profiles, depending on the original mineral bedrock material, local environmental conditions and related soil life. Tillage and reclamation have changed this very considerably for instance when sometimes the ground is

Table 1. Examples of historical landmarks in soil protection policy in Europe Council of Europe Denmark

1972

European Soil Charter

1990 1992

Germany

1985 1989 1955 1988 1971

Act on Waste Deposits (revision older Act) Environmental Protection Act (update) Federal Soil Protection Act Contaminated Land inventory Land Act (general soil protection) Plan of Action of Contaminated Sites Draft Act on Soil Protection Soil Clean-up Guidelines Soil Protection Act Town and Country Planning Act

Norway The Netherlands

United Kingdom

1983 1987 1971 1974 1990

Control of Pollution Act Environmental Planning Act

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ploughed very deeply. Materials are consistently being added to the soil to maintain or improve its nutritive or supportive values, such as organic rich top soils, manure in various forms, town waste refuse and the large-scale addition of mine waste. So, next to a variety of natural soil types, there are various kinds of anthropogenic soils, which may even result in such typical soils as city soils, waste dump soils, and copper-loaded vineyard soils. These soils have been used as constructs rather than the natural system they are in essence. 1.2 Basic scientific assumptions and the underpinning role of soil scientists 2.2.2 Basic assumptions

When the protection of soils and the maintenance of their qualities in view of sustainable development is the basic aim of policy and management, soil structures and processes have to be sustainable. Research must therefore focus on providing the knowledge required to answer basic questions such as what aspects have to be protected and conserved; to what extent are disturbances allowed; what kind of timeframe has to be used when defining reversible impacts (when is it irreversible), what is effective resilience and recovery; has the observed structural diversity (the richness in species called biodiversity as well as the heterogeneity of various soil types) a role in robustness, resilience and recovery; and how will all these 'conserving' activities be interpreted in the perspective of natural adaptation and restoration. 1.2.2 Vitality as combination of internal structure and resistance to external impacts Using vitality of soils as a base, the first question is how to characterise vitality. The best description is: "vitality is the long-term ability to maintain a proper functioning of the soil system through a diversity of processes and organisms that carry out these processes". Maintenance includes robustness, flexibility and resilience to overcome adverse impacts and events. Essential is the ability to recover and regain the composition and functioning of the soil system. Therefore, vitality is a combination of different elements: Vitality = Robustness, Resilience, Recovery + (structural & functional) Richness, or in short: V = 4 x R. 1.2.3 Resilience, Robustness and Recovery Resilience, robustness and recovery can be characterized as the features of ecosystems that show how they react to disturbances and the extent to which they return to normal after a limited period. Disturbances in the form of fluctuations in numbers or activities of organisms have been described in

Leading concepts towards vital soil

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different ways. Orians (1974) distinguished five types of stability: constancy, amplitude, persistence, inertia and elasticity. These can be expressed for instance as different reaction patterns in time under undisturbed or disturbed conditions (Eijsackers; 1983) (Figure 2). Without an external impact, systems can either fluctuate marginally (Constancy) or continue to fluctuate to some extent. Moreover, systems can fluctuate with large or small amplitudes (Amplitude). For each type the most stable situation is indicated with "+s", the less stable with "s". With an external impact the system can hardly be changed (Inert, also called Robust) or the system can come back to its original equilibrium condition after a longer or shorter period (Elasticity, also called Resilience). The system can persist or die out. It can also reach a new equilibrium level after a disturbance (not in this picture). Typical robust systems function naturally under heavy or steady pressure, like riverbanks or coastal zones. Also at the species level there are clear differences in robustness of species. For example, the earthworm species Lumbricus rubellus occurs in all kinds of natural and disturbed systems and has a low sensitivity to contaminants (Eijsackers, 1996). For the assessment of disturbance, quantitative elements are relevant, such as how long does it take in time to come back from an extreme (= maximum or

Figure 2. Types of stability expressed as fluctuation pattern of a certain number or activity (N) over time in undisturbed or disturbed conditions (a vertical arrow), either in a low stable (s) or a high stable (+s) system (Eijsackers, 1983).

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minimum) situation to a more or less normal (= medium) situation, to what extent can the system be brought out of its original situation by an extreme event (height of the amplitude of fluctuation) yet still be able to return to the normal situation, what characteristics govern the recovery of an area. Many attempts to quantify those questions can be found in old (before 1987) and recent literature. An elegant exercise comes from Domsch et al. (1983), who used reproduction rates expressed in generation time to quantify the recovery of soil bacteria. Generation time expresses the general rate of metabolism and hence the rates of all kinds of restorative processes. To investigate the relation between amplitude and the duration of fluctuations, Domsch et al. (1983) studied fluctuations in soil bacteria numbers under a variety of situations including natural adverse conditions such as flooding. Results showed that fluctuations in general comprise a reduction of 90%. Hence, natural fluctuations up to one order of magnitude can potentially return to normal. Vegter (1988), Eijsackers et al. (1988) and Van Straalen and Van Rijn (1998) observed similar fluctuations in groups of soil-fauna. A simpler approach is to look at the seasonal fluctuations for soil organisms. Earthworms will take one season to recover from low numbers caused by summer dryness or winter cold (van Rhee, 1973). Jones and Hart (1998) concluded that earthworms are capable of recovering completely from a reduction percentage of over 90% caused by non-persistent chemicals (half-life < 50 days) within one year. With persistent pesticides (half-life > 50 days) only limited recovery was evident after one year. Another element in measuring natural recovery is the spatial re-invasion of a depleted area. Soil animals are mobile, and are able to move around in a directed way. Depending on their size and charge and on the moisture content of the soil, soil bacteria can be passively transported in groundwater (Bengtsson, 1997). Earthworms can crawl over the soil surface for distances of many meters. The lateral dispersal of earthworms in a year is approximately ten meters in soil that had never been inhabited by worms before (Hoogerkamp et al., 1983). Springtails and earthworms make a distinct choice when offered a gradient of increasingly polluted soil (Bengtsson et al., 1994; Eijsackers, 1981). Surface-active soil animals like beetles cover distances of several hundred meters per day. There are many more publications on dispersal rates, colonization and recolonization of soil animals in all kinds of soil systems, all showing the capability of soil biota to disperse and recolonize depleted areas. 2.2.4 Sustainability in relation to time

Fresco and Kroonenberg (1992) have indicated the ranges of various natural

Leading concepts towards vital soil

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development and recovery processes, for soil processes, for biodiversity and natural hazards (Figure 3). It shows that the sustainability of soil forming processes has to be assessed over periods that vary from one year to thousands of years. For most managers in practice and policy these are unimaginable periods of time. When we look at recovery of biodiversity by re-invasion or repopulation ranges become shorter: 1-10 years for microflora and invertebrates and 10-100 years for higher plants and animals. To be practical it could be decided to take a time period of one century. In fact the 'oldest' soil monitoring cases go back to Darwin's studies, "The formation of vegetable mould through the action of worms" (1888), and the long-term fertilization experiments started in Rothamsted in the UK in 1880 (Johnston et al., 1986).

Plant growth: 1. length of one growth cycle of annual crops 2. length of one growth cycle of perennial crops 3. length of growth cycle of production forest 4. average biomass turnover rates of tropical rainforest Soil processes: 1. time needed for complete erosion of topsoil 2. time needed for severe nutrient depletion by leaching in humid tropics 3. the same for severe nutrients depletion by leaching in the temperate zone 4. time needed for formation of fully developed topsoil Natural hazards: 1. frequency interval between gentle floods in alluvial areas 2. the same for large disastrous floods 3. frequency interval for andesitic volcanic ash falls 4. the same for destructive volcanic eruptions Biodiversity: I. time needed for restoration of microflora, invertebrates, macrofauna and macroflora after major disturbance

Figure 3. Time needed for sustainable development of various biotic and abiotic processes (modified after Fresco and Kroonenberg, 1992)

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1.2.5 Sustainability, in relation to acidification and organic matter content

A critical look at the management of soil resources leads to the conclusion that soils are 'consumed'. Due to the natural breakdown of bedrock there is a continuous supply of minerals and for a long time this has been seen as an everlasting supply. In recent decades, the limitations of that supply have become evident and can be observed in the exhaustion/depletion of three main steering characteristics: the acid buffering system, the maintenance of organic matter levels and the structure and texture (due to soil erosion and compaction) of soils. Since organic matter plays a central role in the structure and functioning of soil ecosystems, this issue will be treated in more detail, to explain its complexity and to underline the fact that knowledge is available. Many data from the research already carried out in the period 1960-80 are valuable for testing new hypotheses. The acid buffering system mainly consists of three systems: carbonate, silicate and aluminium buffer systems. In particular, in sandy areas with a high acid ammonia and nitrous oxide input, the quick reacting carbonate buffer is running out and as a result these soils fall back to silicate and aluminium systems that react more slowly. It will take longer to counteract acidifying inputs and negative acid-impacts will last longer. With respect to organic matter, cultivation enhances mineralization which means the chemical burning (oxidising) of the organic humus supplies formed over many millennia. This mineralization process began after the ploughing of grasslands, the irrigation of bogs and mires, and deep ploughing of layered sandy/organic soils. This raises the problem of how to restore organic matter content, when calculations show that mineralization rates exceed the rate of humification (Table 2). The humus production (gain) is 0.1% of the yearly primary production, whereas the percentage yearly humus loss is 1 - 2%. This is 30 and 60% for periods of a few to 60-80 years (Table 2). Long-term experiments in three agricultural systems with different amounts and types of organic matter input (Figure 4), showed that considerable extra external inputs of 13,500 kg organic manure per hectare (system III) were necessary to maintain the organic matter content of the soil. In Systems I and II there was a slow consistent decrease of 15% over 32 years. The total supply from different sources was 19,000 kg in System I, 34,000 kg in System II and 45,500 kg in System III. Continuous loss of organic matter without supply from other sources may lead to soil erosion. It is a problem in bare areas where plant cover does not prevent the upper soil from being washed or blown away by wind or water. This occurs in arid zones, but also to a lesser extent in temperate and flat areas.

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Table 2. Gain of soil organic matter due to plant production and humification, and loss due to soil cultivation (Van Veen and Kuikman, 1990; Eijsackers and Zehnder, 1990; Cole et al., 1989) Humus production in grassland 7.8 ton ha-1 v 1 Yearly net primary production 0.1% (resistant to mineralization) contributes to organic^ 0.028 ton org. C h a - ^ ^ = — gain matter Humus loss due to mineralization in a prairie soil 40 yea re aftpj- ctart nf ri||fivafinn Sandy soil 8-16 ton org. Cha-1 /* 0.2-0.4 ton org. C h a ^ y ^ - - ; . , Loamy /clay soil 10-26 ton org. C ha-1 \ » ,_ 0.22-0.65 ton org. C ha-Jjs^"^ Some other examples of org. C loss after cultivation virg in soils US Sandy prairie soil after 40 yr. cultivation 42-54% US Loamy-clay soil after 40 yr. cultivation 36-54% US grassland after 25 yr. cultivation 56% US virgin forest soil after 3 yr. cultivation 33% Canada grassland after 60 yr. cultivation 50-60% Canada grassland after 60-80 yr. cultivation (0-15 cm) 57% Canada grassland after 60-80 yr. cultivation (40-80 cm) 20%

Figure 4. Soil organic matter content in three crop rotation systems receiving different supplies of fertilization and manuring over the period 1952-1984 (modified after Kooistra, Lebbink and Brussaard, 1989). I. mineral fertilisers II. mineral fertilisers + green manure III. mineral fertilisers, ley and farmyard manure

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Next to the structure lost because of the actual disappearance of the top surface layer, there is a loss of soil texture due to mechanical forces like the trampling of wet clayey soils. From a perspective of sustainability these general examples show how difficult it is to maintain the main soil systems at their natural levels: the chemical system (pH-buffering), the bio-organic system (the organic matter content) and the physical system (loss of soil structure by erosion). It is difficult to realize improvement in conditions of deterioration. Therefore, an important question in relation to sustainable development is, how the temporal development of soil properties and processes can be defined in such a way that it is not only helpful in general policy and political discussions, but can also provide ways of bringing them under control in management and research. 1.2.6 Diversity in relation to functioning

One of the most intriguing and enigmatic questions in soil life studies has been the relationship between species diversity and soil process functioning (Anderson, 1975). Many studies have shown that soil systems continue to function even with a greatly impoverished soil organism diversity, when only a few species are active. Many scientists have queried whether a large proportion of species was in fact superfluous (redundant). But why then, has such great species diversity, as observed in soil, been developed during evolution? There are several explanations, such as the enormous horizontal and vertical spatial heterogeneity and huge variety in habitat scale and environmental conditions. This variety, combined with some general ecological feeding strategies saprophages (consuming dead organic matter), phytophages (consuming living plant and root material) and predators (consuming living animals) - may have lead to the pleomorphous composition of soil life as simplified in Figure 5. Although already known for decades, this information has serious consequences for monitoring that has to be carried out at considerably different scales. Another explanation for apparent redundancy is that physiological functioning may differ at 'micro food level' and under various environmental conditions. Therefore, maybe the majority of organisms in the soil will be inactive or resting. Over 90% of micro-organisms are normally resting and in soil fauna high numbers and high activities are interlinked with periods when numbers are much smaller and there is a low level of activity (Eijsackers et al., 1988). Functional diversity is the subject of the world-wide biodiversity debate. As Hopkin (1999) asked: "Are we using diversity as a probabilistic event because we cannot recognise the structure behind it? If so next to species

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Figure 5. Composition of a soil fauna selection, arranged according to feeding type (left phytophages, upper predators and right saprophages), size of the animals indicated, and habitat size expressed as various squares (modified after Van der Drift, 1963)

diversity we should study functional and life form diversity. And next to the study of fixed diversity 'patterns' we should study the total continuous process, because one of ecology's prime interest is to describe and predict the probability of things happening". 1.3 Major research topics; the challenge to integrate Monitoring the soil in all its diversity of structures and processes demands an integrated approach because policy and politics perceive the soil only at the highest level of integration - the generic responses of the soil system. Measurement is most precise and sensitive at the species level and further down to the gene-level. An integrated approach combines abiotic and biotic

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processes in a static, dynamic, structural and functional way to provide ways to combine molecular up to ecosystem level approaches. In this context it is necessary to identify some major research topics and their perspectives. These combine spatial and temporal heterogeneity, structural and functional biodiversity, and carbon cycling/energy transfer at the micro- to macro-level. 2.3.2 Combined spatial and temporal heterogeneity

Soil heterogeneity in combination with differences in numbers of soil animals is frequently studied either from a temporal or a spatial perspective. Combination is logical because these variations influence the perkiness and reproducibility of monitoring and assessment results. Vegter (1985) already showed a linearly related increase between temporal and spatial variability of populations and communities of forest-soil arthropods. Recently promising applications have been made in relation to geostatistics (Stein et al., 1992, Ettema et al., 2000; Ettema and Wardle, 2002). The consequence for monitoring is that natural variation in numbers over time has to be taken into account. Distinct seasonal fluctuation effects should be included in the monitoring scheme. Spatial variation is high even in homogenised arable soils, therefore a reasonable number of samples (normally 5-10) have to be taken per location. 1.3.2 Structural and functional biodiversity Classical biodiversity is based on the taxonomic characteristics of the number of species. Hopkin (1997) made a plea for studying biodiversity from a functional perspective: how do species function and optimize their survival and what consequences does this have for their life-form. By addressing the potential and actual specific functionality of organisms, succession will be better understood as well as the way ecosystems recover after disturbances. This may lead to a better understanding of the resilience of ecosystems in general. Approaches like the functional classification of soil fauna into 'leagues' as suggested by Faber (1991), and the distinction of twelve life history tactics of soil micro arthropods (Siepel, 1994) should be used to obtain a better understanding of structural and functional biodiversity, especially in relation to disturbances in the soil's environment. For monitoring purposes this scientific knowledge could be used to identify the functional species groups to be monitored. 1.3.3 Carbon cycling/energy transfer at the micro- to macro-level Litter decomposition is a classical issue in soil studies, especially the stepwise

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comminution and further mineralization of litter by soil fauna and microorganisms in an intimate interaction. This has been extensively studied for various litter types, soil communities and ecosystems in the International Biological Programme (IBP) both in the 1970s and subsequently. The conservation of carbon in the organic pools in the soil plays a prominent role within the climate debate and the soil threats debate. The formation of organic material has its origin in plant growth. The development of vegetations in relation to their specific carbon assimilation pathways (primer of humus formation) should also be monitored during long periods of time. These kinds of differences also comprise functional parameters like litter breakdown and carbon sequestration. 1.4 The impact of soil contamination Since soil contaminants such as heavy metals and persistent pesticides have for some time been considered chemical timebombs they get special attention. 1.4.1 Potential, immediate and derived impacts

When assessing the impacts of soil contamination one has to distinguish between potential, immediate and derived effects. Potential effects are the aim of the present preventive and normative approach, based on laboratory experiments measuring the effective dose (ED) on defined species. ED50 is an effect dose at which 50% of the exposed animals or 50% of the activity level of a process is hampered. These experiments are standardised tests. The exposure is maximum and there will be an immediate response when the tested compound is toxic. In nature animals provide their own living areas and due to heterogeneity there will be spots free of contaminants. From these safe spots (refugia) nonaffected individuals can re-invade a contaminated area after the contamination level has started to decrease. Decreases in organic contamination levels are mainly due to contaminant degradation or contaminants gradually binding to soil constituents, resulting in bound residues. Inorganic contaminants, such as heavy metals and nutrients, may also bind or be rendered insoluble, depending on the environmental conditions. Consequently, the bioavailability of contaminants under natural conditions may be lower than under experimental conditions. When there is no immediate effect derived impacts can still occur, because reproduction is hampered with a resulting decrease in population. Another possibility is that individual organisms become less mobile and unable to feed themselves sufficiently or alternatively they become more mobile and as a

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result become over-exposed to predators. To assess the negative impacts of soil contaminants in a generalised way the various soil contaminating substances are roughly divided into heavy metals, organic contaminants such as Persistent Organic Pollutants (POPs) and nutrients. Heavy metals comprise those metals that occur most commonly in the highest amounts: zinc and copper, essential elements for life-processes, and lead and cadmium that are non-essential elements. Similarly the nutrients referred to are mainly nitrogen, phosphorus, and sulphur. Most POPs are chemical compounds (such as insecticides) made by man that have been emitted into the environment. The response of organisms to natural substances is generally depicted as an optimum curve: small amounts have a positive effect that increases to a certain optimal level. When the amount is increased further, the response will become negative and toxic. Biodegradation of organics may take place under aerobic and various anaerobic conditions, depending on the structure of the compound. Degradation capacity is not unlimited. When a huge dose is deposited on the soil the toxic impact might be so massive that microbial life becomes eradicated. Only after re-succession from outside do area degradation processes return. Anaerobic conditions, that occur at high groundwater levels or in deeper soil layers, have a positive effect on binding and precipitation of anorganic compounds. Ploughing and harrowing causing aeration of the soil will have a positive effect on the availability of heavy metals. 2.4.2 Soil contamination in relation to other stressors

The mechanical cultivation of soil also has a negative impact on soil biota. Earthworms for instance are sensitive to soil cultivation and in normal agricultural practice cultivation does more harm than the use of pesticides (Edwards; 1989). Van de Bund (1979) carried out an extensive monitoring-study on 18 farms for several years in order to study the impact of agricultural practices on springtails and mites. The influence of soil type was small in relation to the crop type or crop rotation. Cultivation causes direct physical damage to and vertical re-distribution of the soil fauna. Van de Bund showed clearly that springtails and mites are sensitive to changes in soil organic matter and in particular soil moisture, which may be exemplary for the whole soil fauna. Also Krogh (1994) ranked agricultural systems according to their habitat quality for springtails, Collembola (expressed as range in numbers). He distinguished habitat quality on the basis of crop, crop age, and soil treatment.

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Chemical stress factors effectively played a more minor role than soil cultivation and the bareness of the soil surface. In ecosystems these stressfactors do occur naturally together with ecological stresses like food shortage, competing species and predators. 1.5 Monitoring Monitoring is the measurement of a number of pre-selected variables at a fixed pattern of locations at fixed times. Soil description has a long standing history as far as monitoring activities and soil description is concerned. Up to now the European Soil Bureau (part of the EU Joint Research Centre) collects, harmonises and distributes soil information from countries all over Europe. Soil ecosystem monitoring has been part of a number of research programmes in various countries (e.g. MATS and ISA in Sweden, Bodenuntersuchungsprogramm and SPP Umwelt in Switzerland, SPBO in the Netherlands). Programmes especially aimed at monitoring soil biota and functions, also in relation to above ground ecosystems, have been developed and started but so far not fully executed (e.g. Observatoire des Sols in France, MATS in Sweden and EMEP in US). The relatively small-scale Dutch soil biology monitoring programme, set up in addition to the soil quality network, has been running on a number of Dutch farms for several years. In rural (agricultural) areas monitoring is common and remediation monitoring is primarily a feature of urban life. For reasons of public interest and communication sound and simple explicable characteristics should be monitored there. Comparing data form contaminated areas with clean references is strongly recommended. For instance in the Netherlands abiotic reference sites were collected by Edelman in the early 1970s (Edelman and De Bruin, 1986) and used in the Dutch reference values. This could be combined with soil biological reference monitoring such as springtail sampling in forest soils (Vegter, 1985), or soil ecosystem studies of arable or forest soils (Brussaard et al., 1990; Berg et al., 2001). It would be worthwhile in my opinion to re-assess these kind of data on the composition of the soil and its soil life and to compare it to current levels and to relate them to contaminated sites. On a global scale a wealth of reference data is available from the International Biological Programme. In order to provide adequate support for soil policy and management a few parameters are needed rather than a whole set of parameters, specified according to soil type and pattern of use or contamination. However in order to pin-point the results, monitoring a specified set might be expected to reveal the fullest results. In principle, assessments should be made at the highest relevant

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ecological integration level: the ecosystem, which responds much slower than individual species or sub-individual biochemical biomarkers. Moreover, it reflects a greater soil surface than specific communities or species. In monitoring systems developed so far, a mixture is used of sub-individual biochemical biomarkers, relevant for biochemical processes, individual characteristics of a specific species, community characteristics like the biomass or performance of earthworms as a group, or the functional group composition of nematodes, and general ecosystem processes like primary production. These kinds of composite or integrative characteristics are not always effective in bringing the message across. Endangered, attractive or ecological key-species could be chosen as representative, because these send a direct message to the public. Unfortunately, soil life, despite its wealth of species in all forms and colours, offers no attractive pet species for the man in the street. Earthworms are one of the candidate groups, since they are well recognised and have a clear function in soil as diggers, litter comminutors and provide food for other species. However, we have to realize that key or indicator species can also be misused by over-generalising the relations between specific ecosystem conditions, species communities and types of contamination (Eijsackers and Lokke, 1996). Generally, a classic list of prerequisites is provided as far as the practical and logistic aspects of monitoring are concerned. The monitoring procedure must be reproducible, but also ecological relevant. It must of course be easy and cheap! Given the variability in soil type, and the heterogeneity in the soil, it is preferable to sample homogenous soil locations such as arable, grassland, park and garden soils. The sampling locations should be of sufficient size. Using relatively small locations (6x6 m) for extensive sampling of soil and soil life turned out to work negatively in the longer term (> 10 years) as shown by Edwards and Brown (1982) and Bembridge et al. (1998). 1.6 Way forward

It is clear that soils are complex systems. Hence, measuring, monitoring and assessing the quality of their structure and processes can only be achieved in an integrated way. The following chapters provide the ingredients for such an approach. They provide a vision on specific elements, their importance and how to measure and monitor soil quality. Most of the background data required are already available and can be used to build a tailored confection approach. Integration and assessment must be carried out at the scientific level, and subsequently interpreted and implemented at the policy level.

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References Anderson, J.M., 1975. The enigma of soil animal species diversity. In: J. Vanek (Ed.), Progress in Soil Zoology, Academia, Prague, pp 51-58. Bembridge, J., T.J. Kedwards and P.J. Edwards, 1998. 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, USA, 341-352. Bengtsson, G., 1997. Dispersal, heterogeneity and resistance: challenging soil quality assessment. In: N.M. van Straalen and H. Lokke, 1997. Ecol. risk assessm. contam. soil. Chapmann&Hall, London, 191-212. Bengtsson, G., S. Rundgren and M. Sjogren, 1994. Modelling dispersal distances in a soil gradient: the influence of metal resistance, competition, and experience. Oikos 71,13-23. Brussaard, L., J.A. van Veen, M.J. Kooistra and G. Lebbink, 1988. The Dutch Programme on Soil Ecology of Arable Farming Systems. I. Objectives, approach and preliminary results. Ecol. Bull. 39, 35-40. Berg, M.P., P.C. de Ruiter, W.A.M. Didden, M.P.M. Janssen, A.J. Schouten and H.A. Verhoef, 2001. Community food web, decomposition and nitrogen mineralization in a stratified Scots pine forest soil. Oikos 94,130-142. Bund, C.F. van de, 1979. De bodemfauna van bouwland in verband met neveneffecten van bestrijdingsmiddelen in de praktijk (Soil fauna of arable land in relation to side-effects of pesticides in practice). Landbouw en Plantenziekten, 7-8 Februari, Wageningen, the Netherlands, 34-50. Commission of the European Communities, 2002. Towards a thematic strategy for soil protection. Brussels, 35 pp. Domsch, K.H., G. Jagnow and T.H. Anderson, 1983. An ecological concept for the assessment of side-effects of agrochemicals on soil micro-organisms. Residue Rev. 86, 65-160. Drift, J.A. van der, 1963. The influence of biocides on the soil fauna. Neth. J. Plant Pathol. 69, 188-199. Edelman, Th. and M. de Bruin, 1986. Background values of 32 elements in Dutch top soils, determined with non-destructive neutron activation analysis. In: J.W. Assink en W.J. van den Brink (eds). Contaminated Soil. Martinus Nijhoff Publishers, Dordrecht. Edwards, C.A., 1989. The importance of integration in sustainable agricultural systems. Agr. Ecosyst. Environ. 27, 25-35. Edwards, P.J. and S.M. Brown, 1982. Use of grassland plots to study the effect of pesticides on earthworms. Pedobiologia 24, 145-150. Eijsackers, H., 1981. Effecten van koperhoudende varkensmest op regenwormen en op de kwaliteit van grasland (Effects of Cu-containing pig manure on earthworms and the quality of grassland). Landbouwkundig Tijdschrift 93, 307-314. Eijsackers, H., 1983. Onderzoek aan biotische kwaliteitskenmerken (Research on biotic quality characteristics). Nederlandse bodemkundige Vereniging, gebundelde voordrachten 2,1-19. Eijsackers, H., 1996. Natuurbeheer voor en door Milieubeheer. Openbare les, Vrije Universiteit, Amsterdam, 21 pp. Eijsackers, H., C.F. van de Bund, P. Doelman and W.-C. Ma, 1988. Aantallen en fluctuaties van het bodemleven (Numbers and activities of soil life). Report RIN 88/33, Arnhem.

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Eijsackers, H. and A. Zehnder, 1990. Litter decomposition: a Russian matriochka doll. Biogeochemistry 11,153-174. Eijsackers, H. and H. L0kke, 1996. Soil Ecotoxicological Risk Assessment. Ecosystem Health 2, 259-270. Ettema, Ch.H., S.L. Rathbun and D.C. Coleman, 2000. The importance of patch dynamics for soil Nematode species coexistence. In: J. Rusek (Ed.) Abstracts, XIII International Colloquium on Soil Zoology, Ceske Budejovice, 14-18 August, Icaris Ltd. Praha, 4. Ettema, C.H. and D.A. Wardle, 2002. Spatial soil ecology. Trends Ecol. Evolut. 17,177-183. Faber, J.H., 1991. Functional classification of soil fauna: a new approach. Oikos 62,110-117. Fresco, L.O. and S.B. Kroonenberg, 1992. Time and space scales in ecological sustainability., Land use policy 1992, 155-168. Hoogerkamp, M., H. Rogaar and H.J.P. Eijsackers, 1983. Effect of earthworms on grassland on recently reclaimed polder soils in the Netherlands. In: J.E. Satchell (Ed.) Earthworm ecology: from Darwin to Vermiculture. Chapmann&Hall, London, 85-106. Hopkin, S.P., 1997. Ecotoxicology, biodiversity and the species concept with special reference to springtails (Insecta: Collemabola). In: N.M. van Straalen and H. Lokke, 1997. Ecological risk assessment of contaminants in soil. Chapmann&Hall, London, 73-83. Hopkin, S.P., 1999. Life and death (mostly death). Contribution to Workshop 'Facing the 21st Century'. S. Rundgren (convenor), Lund, Sweden, 11-12 Dec. 1999. 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, USA, 247-267. Johnston, A.E., K.W.T. Goulding and P.R. Poulton, 1986. Soil acidification during more than 100 years under permanent grassland and woodland at Rothamsted. Soil Use Manag. 2, 3-10. Krogh, P.H., 1994. Micro arthropods as Bioindicators: a study of disturbed populations. PhD-thesis, National Environmental Research Institute, Silkeborg, Denmark, 96 pp. Orians, G.H., 1975. Diversity, stability and maturity in natural ecosystems. In: Dobben, W.H. van and R.H. Lowe-McConnell (Eds.), Unifying concepts in ecology. Junk Publ. The Hague, 139-150. Rhee, J.A. van and S. Nathans, 1973. Ecological aspects of earthworm populations in relation to weather conditions. Rev. Ecol. Biol. Sol. 10, 523-533. Scheffer, M., S. Carpenter, J.A. Foley, C. Folke and B. Walter, 2001. Catastrophic shifts in ecosystems. Nature 413: 591-596. Siepel, H., 1994. Structure and function of soil micro arthropod communities. PhD Thesis Wageningen. ISBN 90-9007450-3. Stein, A., R.M. Bekker, J.C.H. Blom and H. Rogaar, 1992. Spatial variability of earthworm populations in a permanent polder grassland. Biol. Fertil. Soil. 14, 260-266. Straalen, N.M. van and J.P. van Rijn, 1998. Ecotoxicological risk assessment of soil fauna recovery from pesticide application. Rev. Environ. Contam. Toxicol., 154: 83-141. Tunlid, A., 1999. Molecular biology: a linkage between microbial ecology, general ecology and organismal ecology. Oikos 85,177-189. Vegter, J., 1985. Coexistence of forest floor Collemabola. PhD thesis Vrije Universiteit, Amsterdam. Vegter, J.J., E.N.G. Joosse and G. Ernsting, 1988. Community structure distribution and population dynamics of Entomobryidae (Collembola). J. Anim. Ecol., 57, 971-981.

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Chapter 2

THE FORMATION OF SOILS N. van Breemen

Abstract

This chapter provides a bird's eye view of the complex set of physical, chemical and biological processes through which soils are formed. The road from igneous and metamorphic rock to parent material is left out; this chapter focuses on the road from parent material to soil layers. Soils form the thin outer skin of the earth's crust and are exploited by plant roots for anchorage, water and nutrients. Soils result from interactions and feedbacks between parent material, topography, climate, biota and time. Quick changes can take years or decades. Normal natural changes are slow, some taking millennia, some millions of years. A number of soil forming processes that influence soil chemistry, physics and morphology over relatively short time scales - years or decades - will be dealt with in detail, focussing on the processes of the physical ripening of soil. Chemical weathering and the role of ecological engineers are also referred to and given due attention. Soil is an essential resource for all terrestrial organisms including man. Improper and abusive management, careless land clearing and reclamation can lead either to the removal of vital top soils by man-induced erosion or to salinization, acidification or pollution. Housing, industry and transport also affect the soil. Man's inept handling of the soil has reduced soil formation properties in many areas. Whether or not the effects of homo sapiens on soils are a cause of concern depends on one's viewpoint and on the spatial and temporal scale being considered. At a global level, the soil's role in the biogeochemical cycling of N and C may be largely or wholly redundant. On a more local scale, soil management practices have probably always resulted in decreased biodiversity.

2.1 Introduction Darkle, darkle, little grain, 1 wonder how you entertain. A thousand creatures microscopic. Grains like you from pole to tropic Support land life upon this planet I marvel at you, crumb of granite! (F.D. Hole, 1989)

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Soils form a unique, irreplaceable and essential resource for all terrestrial organisms, including man. Soils do not simply form the thin outer skin of the earth's crust to be exploited by plant roots searching for anchorage, water and nutrients. Soils, as defined in Table 1, are complex natural bodies formed in solid rock or unconsolidated sediments under the influence of plants, microorganisms and soil animals, water and air. Their physical nature is determined by innumerable particles and interstitial spaces that form a continuous structure

Table 1. Definitions of terms commonly applied in soil science Term Soil

Decomposition Humus Mineralization Soil forming factors

Soil forming processes Soil horizon

Soil organic matter

Soil profile

Soil structure Soil texture Weathering

Definition "The collection of natural bodies in the earth's surface, containing living matter and (at least capable of) supporting plants out-of-doors. Its upper limit is air or shallow water. The lower limit normally coincides with the common rooting depth of native perennial plants" (after Soil Survey Staff, 1975) Transformation of plant litter or soil organic matter to nutrients, CO2, H2O and more highly decomposed organic matter Soil organic matter without a trace of plant cell structure Transformation of an element from a form bound in soil organic matter to an inorganic, plant-available form The five state factors (parent material, topography, climate, biota, and time) which determine the nature of the soil at any one particular location A process or set of linked physical and chemical processes (often caused or influenced by biota) that turn soil parent material into soil A more or less distinct soil layer parallel to the soil surface, with morphological, chemical and or physical properties that differ from those at shallower and deeper horizons, and that are caused by soil forming processes (rather than, for example, by geological processes) All dead remains of plants and micro-organisms, including humus and recognizable plant litter (but in practice indistinguishable from living micro-organisms) A vertical sequence of soil horizons, starting with an organic-rich horizon on top, to more or less unweathered parent material at some depth, usually between a few dm to tens of meters below the land surface The three-dimensional spatial arrangement of aggregations of soil particles and of larger pores (channels, cracks) in the soil matrix The size distribution of individual soil particles in a given soil sample Complete or partial dissolution of rock-forming ("primary") minerals, often resulting in the formation of new (secondary) soil minerals

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for the storage and transport of gas, water and solutes spanning nine orders of scale from the nanometre to the metre. Chemically, soils are made up of a number of crystalline and amorphous mineral phases and include organic substances that range from recently formed plant litter and their decomposed remnants and include humus. The physical chemistry of soils is determined by a very large, variably charged interface between the many solid phases and the soil solution, which continuously exchange ions supplied or withdrawn via flowing water and soil organisms. Soil biology is dominated by plant roots and decomposers. Plant roots and associated mycorrhizal fungi are active sinks of water and nutrients and sources of energy-rich organic substances, driven by solar energy. The decomposer food web responsible for cleaning up all waste is made up of numerous, largely unknown species of micro-organisms and myriads of soil animals, forming one of the most biologically diverse but poorly known sub-ecosystems on earth. Soils are normally far more suitable as a rooting medium for plants than the geological substrate they come from. In addition to their role as a medium for plant growth, soils play a dominant role in the biogeochemical cycling of water, carbon, nitrogen and other elements. In this way they help the oceans and the marine and terrestrial biosphere, and they influence chemical composition in the atmosphere and hydrosphere. Notwithstanding the generalities listed above, soils vary in appearance and properties from place to place, at scales that range from a few meters to hundreds of kilometres depending on the variation in factors involved in soil formation such as parent material, climate, biota, topography and age. Soil properties at any one place will change over time depending on these factors, sometimes quickly over years or decades, but often much more slowly over millennia or millions of years. Their apparent stability, plus the fact that we continuously tread on them and do not usually see their interior, are amongst the reasons why we tend to take them for granted. Worse, they are often referred to by names like "dirt" in North America that do not reflect their value or complexity. Improper and abusive agricultural management, careless land clearance and reclamation procedures, man-induced erosion, salinization and acidification, desertification, and air and water pollution, as well as the withdrawal of land for housing, industry and transportation, has meant that this vital resource is currently being destroyed more quickly than it can be formed. To appreciate the value of soils and their vulnerability to destruction we need to know how soils are formed and at what rates. This paper provides a bird's eye view of the complex set of physical, chemical and biological

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processes through which soils are formed. We will describe how soils evolve from their geological substrate: how soil minerals form from rock minerals, how nutrient pools change with time - increasing initially and decreasing in very old soils -, how individual mineral grains glue together creating soil structure with important consequences for the movement and availability of water and gases, and how these processes lead to different soil types depending on the factors involved in soil formation. 2.2 Soil forming factors Soil formation or soil genesis refers to changes of soil properties with time in one direction: the content of one component or mineral at a certain depth decreases or increases, sedimentary layering disappears, and so on. As stated earlier, most such changes are slow, and can be seen only after decades or millennia. However, there are exceptions such as the quick drop in pH when sulphides oxidise to sulphuric acid when sulphidic marshes are drained, and the formation of grey and brown mottles when a soil becomes very wet. Cyclic processes, caused by seasonal variations in weather, uptake of water and nutrients by plants, and the decomposition of plant litter by micro-organisms and soil fauna, are reflected by temporal variations in for example, soil pH, and the content of soluble and adsorbed nutrients. Because these changes are reversible on an annual or seasonal basis and do not constitute a unidirectional change in soil properties, they are not considered part of soil genesis. One of the central tenets in soil science is that all properties of a given soil are determined by five state factors, the soil forming factors (Dokuchaev, 1898 cited by Jenny, 1980): parent material, topography, climate, biota (including man), and time. Any particular combination of these factors will give rise to a certain soil forming process, a set of physical, chemical and biological processes that create a particular soil. The Dokuchaev/Jenny model, however, ignores biota's dependence on climate, topography, parent material and time, and also the strong interaction between soils and biota. These aspects are incorporated in the model proposed by Kemmers et al. (2000) and shown in Figure 1. 2.3 The stuff that soils are made of Geologists tend to view soil cover as a messy layer that hides the earth's crust and its wealth of information on the geological past. For soil scientists, on the other hand, the rock or sediments that form the geological substrate is merely the soil's parent material. Water, atmosphere, gravity and biota slowly but relentlessly chew away the rocks and minerals of the parent material, turning it

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Figure 1. The independent factors in rectangular boxes are major determinants of biota and soils, which mutually interact (modified after Kemmers et al., 2002). - The thickness of the arrows suggests the relative importance of the various influences. - The interactions between the dependent factors biota and soil give rise to feedback phenomena and spatial self-organization in the landscape (van Breemen and Meinders, 2004)

into soil material, which differs strongly from the parent material, chemically, mineralogically and physically. Weathering is the chemical, and mineralogical transformation of rock-forming minerals at the earth's surface to solutes and solid residues. The so-called primary minerals in igneous and metamorphic rocks were usually formed at high temperature and pressure at some depth within the earth's crust. They are unstable in the lower temperatures, lower pressure and the wet, oxic conditions prevailing near the earth's surface. Most primary minerals weather to iron and aluminium oxides, clay minerals and amorphous silicates, collectively called secondary minerals. Secondary minerals, together with primary minerals, plus any remaining primary minerals (often mainly highly weathering-resistant minerals such as quartz) form the inorganic elements in most soils. The organic fraction of the soil is admixed with organic remains from plants and micro-organisms, or "soil organic matter". Soil organic matter normally constitutes 1-5% of the mass of soil near the surface, with lower values at greater depth. Because organic matter usually has a high sorption capacity for cations and because it releases nutrients that are rare in rocks (N) or slowly available from minerals (P), its importance is much greater than suggested by its mass fraction.

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The mixture of primary and secondary minerals plus organic matter with its large surface area, normally with a negative charge, forms an important buffered store of all major and minor plant nutrients. Also in its interaction with water, soil material usually differs strongly from soil parent material because it has a much wider range of pore sizes. This is caused by soil structure rather than by the interstitial spaces between individual sand, silt and clay particles that are related to soil texture. These pores with widely varying diameters are caused by a range of processes including (1) the shrinkage and swelling associated with wetting and drying; (2) the spaces between and inside clumps of soil particles tightly glued together by organic remnants and polysaccharides exuded by plants and microbes, or held together by roots and fungal hypha; (3) between and inside densely packed faecal pellets produced by an array of soil fauna, and (4) relatively large channels caused by burrowing soil fauna and plant roots. With a wide range of pore sizes, soil can hold water and air at the same time: when capillary water is present in the smallest pores, gases can diffuse through the larger pores. In this way soil materials satisfy crucial, seemingly conflicting, requirements for plants: they simultaneously provide roots with water, gaseous oxygen, and nutrients. The differences between soil and parent material are most striking in the case of soils derived from solid rock, the so-called residual soils. They are less obvious in soils developed in sediments. Sediments normally consist of soil material eroded elsewhere and stripped of part of their organic matter and sorted again in specific grain sizes during transport and deposition. Freshly deposited sediments are usually made up of thin layers of well-sorted material, with a very uniform grain size. Such layers also contain (capillary) pores of uniform size. Depending on the moisture tension in the sediment, the capillary pores are either filled with water or with air, but not with both, as in (moist) soil material. The mixing of such layers, for example by burrowing fauna in soils or tidal sediments, will turn sediment into a body capable of supporting terrestrial plants: a soil. Soil material differs chemically and physically from its parent material and differs within depth, for example, within in the upper meters of the land surface. These differences are visible as horizontally layered so-called soil horizons. 2.4 Soil forming processes Differences between soil horizons, in fact, reflect different intensities of the various soil forming processes. The most common example is the formation of a dark-coloured, humus-rich, so-called A horizon, caused by a higher supply of

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plant litter at shallow depth. Each given set of soil forming factors result in a specific sequence of soil horizons, and is known as a soil profile. Conversely, the soil profile and the underlying soil properties reflect the soil forming factors, and form the basis of most so-called morpho-genetic soil classification systems. Examples of different soil profiles and the soil forming factors behind them are shown in Figure 2.

Figure 2. Typical soil profiles representing the different effects of soil formation depending on soil forming factors. The podzol shows a bleached horizon due to dissolution of iron and aluminium under the influence of acidic and complexing organic matter and its downward transport by percolating water into the darker horizon and the fibres at greater depth. Podzols normally form under coniferous vegetations in boreal climates (on many kinds of parent material), but also in more temperate humid climates on very nutrient-poor, quartz-rich parent materials (as the one shown here). The Plaggen soil shows a 70-cm thick dark surface horizon caused by centuries of applying manure-drenched sods to cropping fields on originally moderately lowfertility soils in temperate humid climates before the advent of fertilisers. In this way nutrients were extracted from a large grazing area and concentrated on much smaller areas of cropland. The chernozem is representative of fertile natural grassland soils in low rainfall areas in temperate climates. Under these conditions, nutrients from weathering and (natural) atmospheric inputs naturally accumulated in thick surface soils, creating the proverbial fertile black prairie soils that still form very productive grain land. The red tropical soil is typical of old stable land surfaces where intense weathering has proceeded for hundreds of thousands of years, leading to removal of all weatherable minerals and the accumulation of resistant iron and aluminium oxides

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So far, a number of soil processes have been mentioned that provide for anchorage, the simultaneous supply of water plus air, and make available the nutrients necessary for plant growth. These processes relate to weathering and the formation of secondary minerals, humus, and soil structure. These and a number of other soil forming processes that influence soil chemistry, physics and morphology on relatively short time scales - such as years or decades - will be dealt with in the most detail. Particular attention will be paid to the strong influence of biota on most of these processes, thus emphasising the links between biological soil qualities and soil formation. 2.4.2 Physical ripening

Roots need oxygen to grow and to function. Water-saturated, clayey sediments deposited under water are soft and have a high pore volume that is completely water-filled (in the order of 80%). Because of lack of oxygen, the roots of most plants cannot function in such sediments. Exceptions are plants that can supply oxygen to their roots through air tissue (aerenchyma), such as reed (Phragmites) and alder (Alnus) in temperate climates, and wetland rice (Oryza), or mangrove trees (Rhizophora and others) in the tropics. Water needs to be removed by drainage or evapotranspiration, and soil structure must be developed before terrestrial plants and most crop and forage plants can grow. The various physical and chemical changes taking place when water is removed are called soil ripening, a term coined in analogy to the 'ripening' of cheese (Pons and Zonneveld, 1965). Water-saturated clayey underwater sediments have a very low hydraulic conductivity, less than 1 mm per day under a potential gradient of 10 kPa/m (1 m water head per m). The material dries out very slowly by drainage alone. The reclamation of Dutch Zuyderzee polders and the soil formation that accompanied it has been greatly accelerated, for example, by growing Phragmites, which can extract water from such sediments. In tidal flats, burrowing crabs and the growth of tidal plants cause very high permeability even in very clayey sediments, and ebb-tide drainage contributes significantly to the ripening process (Smits et al., 1962). 2.4.2 Chemical weathering, acid buffering and formation of soil minerals

Bare rocks carry more vegetation than one might think. Lichens grow on rock surfaces or, in the case of endolithic lichens, in more sheltered conditions just below the surface. But bare rocks provide little or no anchorage, nutrients and water, and so are obviously poor substrates for plant growth. How do rocks weather to often fine-grained secondary minerals, creating soils containing

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enough nutrients, water, oxygen and room for anchorage necessary for a luxurious vegetation cover? A major factor determining the rate of weathering are the plants that also benefit from it. Plants and the soil fauna associated with them can be seen as ecological engineers, that change the physical and chemical nature of their substrate (Jones et al, 1994), with possible (evolutionary) feedbacks to those organisms (Odling Smee et al., 1996). Lichens exude polyphenolic acids that can stimulate rock dissolution, and rock-eating snails grazing on endolithic lichens can speed up the crumbing of calcareous rocks (Jones and Shachak, 1990). Vascular plants stimulate weathering by continuously pumping CCh from a low concentration in the atmosphere to a 10 or 100-fold higher concentration in the rooted soil. Pumping proceeds via root respiration and the microbial decay of dead roots and leaf and branch litter. The H+ produced by dissociation of CCh: CO2 + H2O - • H+ + HCO3

(1)

helps to dissolve minerals, releasing nutrients that were earlier locked up in unavailable forms (for example Ca2+, Mg2+, K+, and H2PO4) that can be taken up by plants. For example, weathering of the rock-forming mineral K-feldspar to the clay mineral kaolinite is stimulated by the production of H+, accompanied by the release of dissolved K and silica: Feldspar Kaolinite 2KAlSi3O8 + 9H2O + 2H+ ->• AbSi2Os(OH)4 + 2K++ 4H4SiO4

(2)

Organic acids exuded by rock ecto mycorrhizal fungi, are another way in which plants can indirectly stimulate rock weathering and the release of nutrients (Landeweert et al., 2001). Depending on the rock-forming minerals involved and other soil forming factors, other clay minerals, and oxides can be formed. Uptake of nutrients by plants and their adsorption by charged surfaces of clay, iron oxides and humus prevent the quick removal of these nutrients by percolating water. Eventually, however, practically all nutrients leach away to the oceans. As a result of weathering and leaching, the store of plant-available nutrients in soils tends to increase during the first centuries and millennia of soil formation, but eventually - over millions of years - nutrient contents becomes very low. Such highly weathered soils have lost all their primary minerals except the very resistant quartz sand, and often consist mainly of secondary minerals that are very stable in the conditions that prevail near the

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Earth's surface: oxides of Fe and Al, and kaolinite. Nutrients lost by leaching are largely replenished from the atmosphere: nutrients in rain water and the long-range transport of dust (Chadwick et al., 1999). Such low-productive, deeply weathered soils are typical of old land surfaces in the tropics, not only because weathering is more intense in tropical climates, but perhaps because such land surfaces have not been disturbed by the large-scale geological events associated with glaciation, volcanism and plate tectonics. It can be concluded that, in the long run, geological processes are essential to maintaining plant growth on earth, by supplying new injections of plant nutrients with fresh parent material. 2.5 Changes in soil pH in the course of soil formation Soil pH is one of the most important soil properties for plants for several reasons. The solubility - and therefore the plant-availability - of many elements (both nutrients and toxic substances) strongly depend on pH. For example, the solubility of many metal ions decreases with pH, so that toxicity of Al develops in very acid soils (pH < 3.5), while metal deficiencies (for example of Fe and Zn) may occur in alkaline (pH> 8.5) soils. By contrast, the availability of molybdenum increases with pH. Also, microbial and soil faunal activity is strongly influenced by pH. Bacteria and earthworms, for example, also generally favour near-neutral pH values and fungi are the dominant decomposers in acidic soils. Nitrifying bacteria are more active at high that at low pH, which explains why NO3- rather than NH4+ is the main form of plantavailable N in non-acid (pH >5) soils. Due to such effects and their interactions, decomposition and mineralization (see Table 1) are usually much faster in nearneutral than in strongly acidic (pH < 4) soils. As a result, near-neutral soils are more likely - but not necessarily - better supplied with plant nutrients than acidic soils. Near neutral to slightly alkaline (pH 6.5-8.5) soils are normally calcareous and, therefore, well supplied with Ca, while acid soils (pH = change in pH due to the addition of a mole of strong acid or base. Equilibrium pH values (small circles) are determined by the balance of alkalising effects of minerals and the acidifying effect of COi that is invariably present in higher-than-atmospheric concentration in soils. The equilibrium pH (and soil pH) tends to decrease high values in the presence of feldspars (such as anorthite (an), albite (alb) and microcline (m) to lower values typical due to formation of secondary minerals such as gibbsite (g), allophane (a), kalolinite (k) and iron oxide (f). Buffer curves for H+ and OH- in pure water (pH 7) are shown as dotted lines

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greater depth, due to the presence of organic matter that consists of plant litter in various stages of microbial decomposition. Because of its large surface area with common carboxylic and phenolic surface groups, soil organic matter interacts strongly with both mineral surfaces and solutes. Moreover, the mineralization of plant litter soil organic matter is the major proximate source of most plant nutrients. Therefore, even when its mass represents no more that a few percent of the total soil - as in most surface soils - its importance for soil processes and properties is huge. Most of the biomass that reaches the soil as litter and animal material is decomposed and mineralized to water, CCh and nutrients. Only a minor part persists for longer or shorter periods as soil organic matter, which is called humus if it lacks microscopic cell structure. Traditionally humic substances, which constitute 70 to 80% of organic matter in mineral soils, were characterised as dark-coloured partly aromatic, acidic, hydrophilic, molecularly flexible polyelectrolyte materials (Schnitzer, 1986). Recently, hydrophobic, aliphatic components and chemically recognisable remnants of plant and microbial cell material, for example, polysaccharides and proteins, have been recognised as important constituents of humus. As a result the traditional model is now being replaced by a model involving closely stacked, smaller units of such recognisable compounds (conformational structure) with a large apparent molecular weight (Piccolo, 2001) Different vertical sequences of plant litter and soil organic matter on and in the soil can be encountered in nature, and classified into different humus forms (see Muller, 1878; Klinka, 1981). These forms reflect the particular environmental conditions that affect plant growth and species composition, as well as the environmental conditions that affect decomposition and mineralization. Three main groups of humus forms can be distinguished: mull, moder and mor. When practically all soil organic matter is intimately mixed with and bound to minerals without recognisable plant remains, the humus form is called mull. In a natural mull, mixing is usually the result of the activity of burrowing soil fauna, such as earthworms, ants and termites. The soils in question are usually well supplied with plant nutrients and the soil pH is generally neutral to slightly acid. By definition, however, mixing via ploughing causes arable soils to have a mull humus form, regardless of other soil properties. Where conditions are unfavourable for strong mixing, as in acid, lowfertility soils under natural vegetation, the mineral surface is usually buried under a layer of pure, relatively less decomposed soil organic matter. If this socalled ectorganic horizon or forest floor is thicker than 2 cm, the humus form is

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either a moder or a mor. Within the ectorganic horizon, three sub-horizons can be identified. From top to bottom these are: L layer of almost undecomposed litter (Litter layer), F layer of recognisable, but fragmented organic remains (Fragmentation layer), H layer of humified organic matter, in which plant remains are not recognisable (Humus layer). A mor humus form has a thick F horizon with mainly fungal hypha on top of a relatively thick H horizon. A moder has a thick F horizon dominated by faunal excrements on top of a thin, or non-existent H horizon. Recently, Kemmers (2002) transformed Klinka's classification for Dutch soil landscapes and suggested a subdivision of the different humus forms in sub-classes indicative of various terrestrial and semi-terrestrial ecological conditions. Because the humus form may change rapidly, possibly within decades, it is very useful to characterise the soil ecological conditions in nature management schemes (Figure 4). Certain tree species (especially conifers) are normally associated with mor humus forms, and others (such as the deciduous species Tilia and Acer with mull. This has led to a debate about the question of which came first, the humus form or the tree (Binkley, 1995). These particular deciduous trees, however, appear to maintain high levels of Ca in the surface soil by the active uplift of dissolved Ca from deep soil layers, in contrast to the shallower rooted conifers with their mor humus form (Dijkstra and Smits 2003). Together with common garden experiments that show quick changes in soil pH, soil Ca and humus form under different tree species (Van Breemen and Meinders, 2004) these results strongly indicate that the plants and many other biological factors such as earthworms, determine the humus form. 2.5.2 Formation of sub-surface horizons soil classification

Depending on the soil forming factors, a range of sub-surface horizons caused by specific processes are recognised by pedologists, mainly on the basis of morphological criteria visible in the field. In addition to the ones discussed earlier, these horizons include different eluviation and illuviation horizons as well as horizons indicative of permanent or temporary anoxia. Eluviation horizons have been preferentially leached of specific compounds, such as fine clay, or iron and aluminium (hydr)oxides. Illuviation horizons, which normally underlie eluviation horizons, have been enriched by materials such as clay, humus, salts or iron and aluminium oxides, leached from

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Figure 4. The conversion from mull to mor. - After the disappearance of earthworms due to Cu- and Zn containing fungicides, within a few decades an orchard soil with a mull humus form (right) developed a 5cm thick mor-like ectorganic horizon (left). - At the same time the soil structure changed from porous granular and sub-angular blocky to compact, angular blocky and weak coarse prismatic down to 50 cm depth (Jongmans et ah, 2003). O, Ah, Bw and Cg indicate the different soil horizons present. - After the disappearance of earthworms similar effects have been observed in grassland soils, due to the long-term application of acidifying fertiliser (ammonium sulphate) at the Rothamsted Experimental Station, Harpenden, Herts, UK

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shallower depth. Hydromorphic soils can be recognised by so-called gley phenomena. These refer to grey and brown mottles caused by seasonal water-saturation of the soil horizon(s) in question, or to a uniformly neutral, greenish or bluish grey soil colour indicative of permanent water saturation and the associated absence of oxygen. The presence (or absence) of particular surface and subsurface horizons forms the basis of most, so-called genetic soil classification systems. The properties of the resulting soil groups at different hierarchical levels (Orders Suborders - Great groups - Subgroups) are sometimes highly relevant ecologically, but more often than not, soil classification provides little ecological information. For example, soil structure is considered only in certain soil groupings, and man-made pollution effects are not recognised in soil classification systems. In general, genetic soil properties that change quickly, that is within years or decades, tend to be more useful for an ecological soil characterization than soil properties that change very slowly. Humus forms and gley phenomena change relatively quickly and are, therefore, very useful for ecological soil characterization. 2.6 Man interferes with soils Man is and has been a powerful soil forming factor. For hundreds of thousands of years, hunter-gatherers burned vegetation to stimulate grazing. In the past millennia smelters involved in metal production consumed extensive forests, and farmers reclaimed forests and grasslands for agriculture. These interventions sometimes had a dramatic effect on soil formation. Since the industrial revolution and the strong increase in population that went with it, man's interference with soils has intensified. Most man-made soil effects involve properties that change relatively quickly, on a time scale of decades and centuries, rather than the millennia and millions of years required for mineral weathering. Therefore, except where man-made erosion has resulted in the complete removal of soil cover, texture and soil mineralogy are usually not affected by man's activities. Before the advent of fertilisers, the transformation of naturally vegetated grass or forestland to arable land implied the loss of organic matter and associated nutrients. Frequent physical soil disturbance due to tillage and harvesting helped accelerate the decomposition of soil organic matter and the mineralization of nutrients. Unless counter measures were taken, therefore, soils were in fact "mined": the nutrients accumulated in soils since the advent

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of soil formation were partly removed in crops and through increased drainage. Nutrients loss through drainage was facilitated by the removal of the permanent roots of natural vegetation. Declines in soil organic matter implied the loss of nutrient adsorption capacity, and a decrease in structural stability. As a result, crop yields fell and, with some exceptions, for example wetland rice cultivation, counter-measures were required to allow continuous cropping of the same land to proceed. The proverbial richness of prairie land reclaimed for grain crops in the decades before and after 1900 in North America was largely based on the mining of a rich resource. Today, farming on the Prairies requires the application of fertiliser. More traditional measures include the application of animal manure, collected from animals grazing on adjacent lands, or including fallow periods as in shifting cultivation. At present, in large parts of Africa, lack of fertilisers forces farmers to exhaust their soils with dire consequences for soil quality and food production (Lynam et al., 1998). In stark contrast, the abundant availability of cheap fertilisers has led to a strong increase in the pools of nutrients available in the agricultural soils of many rich countries. This in turn has caused increased emissions of such nutrients by volatilization into the air (of ammoniaN) and by drainage to ground and surface waters (mainly of nitrate-N and ortho-phosphate) and greatly contributed to the large scale eutrophication of the environment in large parts of the world (Vitousek et al., 1997). The application of biocides, and of acidifying fertilisers in the absence of liming, may affect soil biota to such an extent that the dynamics of soil organic matter and the structure of the soil are strongly affected (see Figure 4). Similar physical degradation is often caused by the use of heavy machinery for tillage and harvest operations, particularly on wet soil. Other effects associated with agricultural land use include soil salinization or alkalinization due to irrigation with inadequate drainage. Over large regions in the industrialised parts of the world, man's interference with the local hydrology has affected wetlands and has had a dramatic effect on soil properties. This has had a detrimental effect on biodiversity even when these soils were not turned into agricultural land. Soils under natural vegetations have also been affected by the atmospheric deposition of pollutants, a process that usually leads to increased soil acidification. 2.7 ... but so what? Man's interference with soil and its hydrology may increase or decrease the capacity of land to produce crops. The same activities generally decrease the capacity of land to sustain a certain level of biodiversity. Moreover, interfering

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with the soil often impacts on its capacity to regulate biogeochemical cycles. For example, land drainage has usually decreased denitrification - which prevents the world from slowly turning "nitric" - and land reclamation, drainage and deforestation have decreased the storage of CO2-C as soil organic matter which plays a role in buffering the CO2 concentration in the atmosphere. In addition, heavy fertiliser applications, land drainage and atmospheric deposition have caused increased emissions of substances from soil to ground and surface waters which usually causes detrimental ecological effects elsewhere. Not all of the concomitant changes in soil properties can be considered part of the soil formation processes, but most such changes are reflected in morphological, chemical and physical changes in the soil profile. Whether or not the effects of homo sapiens on soils are a cause of concern depend on one's point of view and on the spatial and temporal scale being considered. On a global scale, the soil's role in the biogeochemical cycling of N and C may be largely or wholly redundant, given those same functions in the marine realm. On a more local scale, soil management practices that have helped increase agricultural productivity on cropland or grassland have probably always resulted in decreased biodiversity in situ, and very often in the immediate and wider surroundings as well. However, wherever they get the chance, there are other species that are very good in doing away with others via their effects on soil. Sphagnum spp, for example, is one such culprit (Van Breemen, 1995). Nevertheless, today human being are causing a decline in biodiversity at all levels from local to global, an effect that is probably unprecedented for any single species. Habitat destruction, associated with deforestation and increasing agricultural land use, however, is but one cause of species loss resulting from human activity and related soil changes are mostly an associated phenomenon rather than a primary driver of decreasing biodiversity. In the Netherlands, at the local level however, blanket soil acidification and eutrophication from air pollution, plus the removal of smallscale spatial soil variability due to land consolidation and regional changes in hydrology have caused an enormous decline in biodiversity. Such effects, and the widespread, wanton destruction of increasingly rare pristine wilderness areas, have resulted in an outcry from activist groups in the media about man destroying life on Earth. From the perspective of geological time, however, man's tinkering with the environment is probably insignificant compared, for example, with the major meteorite impacts, that occur not infrequently geologically speaking and cause global disasters that beat even the bleakest anthropogenic doomsday scenarios. Moreover, such disasters never stopped life evolving and repeatedly recapturing the earth. The meteorite that

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did away with dinosaurs in fact paved the way for us mammals! Fortunately, we are not nearly so powerful and probably never will be. But of course, a more important and immediate concern is whether we will destroy ourselves via soil degradation. In 1960, a major Dutch writer W.F. Hermans (1960), who also happened to be an earth scientist, wrote as a postscript in his book on erosion "No philosopher will be able to disprove the statement that man will ever disappear from the surface of the earth". Hermans presumably had in mind the destructive effects of soil erosion. In May 1934, widespread and severe soil erosion caused by improper land use in the Great Plains led to a cloud of top soil that blanketed the Eastern United States with deposits settling as far as 2,400 km away. This triggered the establishment of the US Soil Conservation Service, which since 1935 effectively countered the problems until the service was abolished in 1991. Elsewhere, however, humanaccelerated soil erosion and physical and chemical soil degradation continue to occur world-wide, as the findings of the Global Assessment of Human-induced Soil Degradation (GLASOD) project have shown (Oldeman et al., 1991). So far increases in per hectare productivity have more than offset these negative effects. More important, production increase has, so far, kept up with population growth worldwide. As long as supply from affordable sources of energy continues, and we keep taking measures based on learning from past failures, I do not think our tinkering with soils is going to be our undoing. If I am right, of course, this does not mean we should continue to mistreat this element that is so basic to our existence, neither does it mean that we are in for a secure future! 2.8 Implementation in soil management Most "natural" soil forming processes leave their mark on soil properties over centuries, millennia or even longer time spans, rather than within the few years that form the time horizon of most people and policy makers. However, in many industrialised countries, and in the Netherlands in particular, soils have been widely influenced by man via tillage, drainage or some other form of reclamation for more than 2000 years. Many man-induced soils have been formed since then. Especially during the last thirty years or so, profiles have been formed in polluted areas that have completely "new" characteristics. Some soil scientists even define special diagnostic characteristics for highly maninfluenced soils, such as "urbic", "garbic" and "spolic" materials (Fanning and Fanning, 1988). Monitoring the morphological criteria of these man-influenced soils may be an aspect of soil management.

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References Binkley, D., 1995. The influence of tree specious on forest soils: processes and patterns. In: D.J. Mead and I.S. Cornforth (eds), Proceedings of the Trees and Soil workshop, Spec. Publ.; 10 Lincoln University Press, Canterbury New Zealand, pp 1-33. Chadwick, O.A., L.A. Derry, P.M. Vitousek, B.J. Huebert, and L.O. Hedin, 1999. Changing sources of nutrients during four million years of ecosystem development. Nature 397, 491-497. Dijkstra F.A. and M. Smits, 2002. Tree species effects on calcium cycling: the role of calcium uptake in deep soils. Ecosystems 5, 385-389. Fanning, D.S and M.C.B. Fanning, 1989. Soil morphology, genesis and classification. John Wiley & Sons Inc. 395 pp. Hermans, W.F., 1960. Erosie. Heijnis Zaandijk, 188 pp. Hole, F.D., 1989. See www.soils.wisc.edu/soils/poets/holel.htm. Jenny, H., 1980. The soil resource: origin and behaviour. Ecological studies 37, Springer Verlag, NY. Jones, C.G. and M. Shachak, 1990. Fertilization of the desert soil by rock-eating snails. Nature 346, 839-841. Jones, C.G., J.H. Lawton and M. Shachak, 1994. Organisms as ecosystem engineers. Oikos 69, 373-386. Jongmans, A.G., M.M. Pulleman, M. Balabane, F. van Oort and J.C.Y. Marinissen, 2003. Soil structure and characteristics of organic matter in two orchards differing in earthworm activity. Appl. Soil Ecol. 24, 219-232. Kemmers, R., R. de Waal, B. van Delft and P. Mekkink, 2002. Ecologische typering van bodems: actuele informatie over bodemkundige geschiktheid voor natuurontwikkeling. Landschap 19, 89-103. Klinka, K., R.N. Green, R.L. Towebridge and L.E. Lowe, 1981. Taxonomic classification of humus forms in ecosystems of British Columbia. First Approximation. Province of BC. Ministry of Forests. Canada. Landeweert, R., E. Hoffland, R.D. Finlay, Th.W. Kuyper and N. van Breemen, 2001. Linking plants to rocks: ectomycorrhizal fungi mobilize nutrients from minerals. Trends Ecol. Evolut. 16, 248-254. Lynamn, J.K., S.M. Nandwa and E.M.A. Smaling, 1998. Introduction to the special issue "Nutrient Balances as indicators of productivity and sustainability in Sub-Saharan African agriculture", Agric. Ecosyst. Environ. 71,1-4. Muller, P.E., 1887. Studien iiber die natiirlichen Humusformen und deren Eiwerkungen auf Vegetation und Boden. Julius Springer. Berlin, 324 pp. Odling-Smee, F.J.K., K.N. Laland and M.W. Feldman, 1996. Niche construction, Am. Naturalist 147, 641-648. Oldeman, L.R., R.T.A. Hakkeling and W.G. Sombroek, 1991. World Map of the Status of HumanInduced Soil Degradation: An explanatory Note (rev. ed.), UNEP and ISRIC, Wageningen. 34 pp. with maps. Piccolo, A., 2001. The supramolecular structure of humic substances. Soil Sci. 166, 810-832. Pons, L.J. and I.S. Zonneveld, 1965. Soil ripening and soil classification. ILRI publ. 13, Veenman, Wageningen, 128 pp. Schnitzer, M., 1986. Binding of humic substances by soil mineral colloids. In: P.M. Huang and M. Schnitzer, eds., Interactions of soil minerals with natural organics and microbes. SSSA

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Special publication 17, Madison, Wise. USA, pp 77-101. Smits, H., A.J. Zuur, D.A. van Schreven and W.A. Bosma, 1962. De fysische, chemische en microbiologische rijping der gronden in de IJsselmeerpolders. Van Zee tot Land, no. 32. Tjeenk Willink, Zwolle, 110 pp. Soil Survey Staff, 1975. Soil Taxonomy, Agric. Handbook 18 Washington DC, 503 pp. Van Breemen, N. and W.G. Wielemaker, 1976. Buffer intensities and equilibrium pH of minerals and soils: II Theoretical and actual pH of minerals and soils. Soil Sci. Soc. Amer. Proc. 38, 61-66. Van Breemen, N., 1995. How Sphagnum is bogging down other plants. Trends Ecol. Evolut. 10, 270-275. Van Breemen, N. and M. Meinders, 2004. Processes in self-organization of soil-vegetation patterns. Proc. Tenth Cary Conf., Inst. Ecosystem Studies, Millbrook, NY. Vitousek, P.M., J.D. Aber, R.W. Howarth, G.E. Likens, P.A. Matson, D.W. Schindler, W.H. Schlesinger and D.G. Tilman. 1997. Human alteration of the global nitrogen cycle: sources and consequences. Ecol. Applications 7, 737-750.

41

Chapter 3

VEGETATION, ORGANIC MATTER AND SOIL QUALITY W.H.O. Ernst

Abstract Soil life depends on water, inorganic chemical elements and organic matter. Due to photosynthesis there is the growth of vegetation, primary production. The role of vegetation and humus in relation to their function and the many characteristics of their composition have been described in relation to soil quality and soil health. Soil is an environmental component and is permanently changing due to the often cyclic processes of litter supply. This chapter focuses on vegetation and humus and their various interacting components. The overview looks at present day knowledge and classical scientific knowledge. Moreover, it emphasizes the possibility of applying neglected data to present day soil management. Differences in the chemistry and structure of soils and litter cause great diversity in vegetation types. The way plants cope with various concentration of nutrients in the soil is partly formalized in the indicator values for plant species (Ellenberg) that can be deduced from soil type, moisture regime and nutrient regime. In general higher nutrient levels as well as chemically extreme environments result in lower plant diversity. The adaptation of plants to changing environmental components is a permanent process, resulting after decennia or centuries in resistant ecotypes, as in the case of a surplus of heavy metals in specifically metal-resistant ecotypes of the grasses Agrostis and Festuca, or in endemic species such as the Zinc violet. Mycorrhiza can protect roots against a surplus of heavy metals. The role of organic matter, its quantitative production and the various humus forms such as Fulvic Acid (FA) and Humic Acid (HA) have been evaluated. Soil types have been arranged according to organic matter conversion types in Mull with very active fauna and microflora, Mor that is the opposite, and Moder which lies between both types. The functionality of soils and their vegetation are often evaluated on a very limited spatial scale. They should be considered in the spatial context of a landscape above and below the soil surface. Several very specific indicators can be applied for chemical changes in the soil. A very general indicator of soil quality is the change in plant species composition, evaluated by Ellenberg's indicator values and related to the effects of manure, heavy metals, acidification and desiccation. The chemical composition of macro-elements and metals of the above ground vegetation and its analysis can provide information on the behaviour and availability of metals in the

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W.H.O. Ernst

underground. This may reflect the position and capacity of the root and mycorrhiza system especially in buried (or covered) soil areas. The species-specific accumulation of chemical elements in leaf and above ground tissues and in litter has to be incorporated in the evaluation of soil quality and health, but is still missing in most ecological soil studies and in soil legislation as far as ecological risk assessment is concerned. There is hardly any vegetation and soil in any ecosystem that has escaped mechanical damages as a result of heavy machines, and modification to its chemistry by human activities. Physical impacts on the soil, resulting in soil compaction can be evaluated by root and mycorrhiza development that is restricted in compacted soil. In evaluating soil vitality, past, present and future land use has to be defined. Also impacts of physical, chemical and biological disturbance and the consequences for vegetation, soil organic matter, decomposition rate, and the impact on the soil profile have to be characterized. In general, plants and micro-organisms and to a lesser degree animals, have sufficient genetic potential to colonize each soil if water is available and changes occur slowly. The biodiversity of vegetation and fauna on soils with extreme saline and heavy metal-enriched environments may be very low. Such soil will not support the growth of edible plants, but the soil itself is still very vital.

3.1 Introduction Since the first recognition of the relations between humus forms and vegetation (Miiller, 1889) and between soil profiles and vegetation (Braun-Blanquet and Jenny, 1926), different approaches have been used to relate plant species and/or vegetation with aspects of soil quality. Oberdorfer (1946, 1994) presented the first flora that included information on the relation between plant species and their habitats, especially soil structure the form of and humus. Ellenberg (1974) has attempted to give all plant species in Central Europe indicator values for acidity, fertility and humidity, with a later addition of some indicator values for heavy metals and salinity (Ellenberg et al., 1991). Modified versions of this latter approach were applied for the evaluation of soil quality for forestry purposes in several countries (Noirfalise, 1984, for Belgium; Klinka et al., 1989, for British Columbia; Hawkes et al., 1997, for Great Britain), but not in non-wooded ecosystems. In virgin environments each soil is the result of interactions between various environmental components, i.e. (1) the parent material with its differences in structure and chemistry (lithosphere), (2) the water supply of the site (hydrosphere), (3) the climate (atmosphere), and (4) the biota (biosphere) with the relief as a modifier, all of which change with time (Figure 1). In soils modified by man the natural interactions of some components have been disturbed, as will be discussed later.

Vegetation, organic matter and soil quality

43

Decomposition (soil organisms)

Figure 1. The various interacting components of an ecosystem in a virgin environment

3.2 Plant productivity as the initiator of pedogenesis and roots as modifier of soil structure and chemistry 3.2.1 Netto primary production

The most prominent contribution of the biota to the soil formation processes (pedogenesis) is that of the vegetation which delivers the litter (above and below ground), the root exudates as principle organic ingredients, and the soil organisms decomposing and synthesizing humus. Above ground biomass production can be measured with a high reliability, whereas below ground biomass production is very unreliable (Table 1). One of the problems is the short life and rapid turnover of fine roots: 21.4% to 38.6% of the roots do not live more than two weeks, a further 38.3 to 45.8% of roots not longer than 2 months so that between 63.4 and 84.4% of roots disappear within two months (Ellenberg et al., 1986). By analysing the starch in soil samples and taking temperature into consideration, Marshall and Waring (1985) conclude that fine roots may account for as much as two-thirds of annual biomass production. Recent models include additional parameters such as mineral concentrations, carbon content, and soil moisture to estimate the below ground biomass production of roots. However uncertainties remain high (Vogt et al.,

W.H.O. Ernst

44

Table 1. Above ground and estimated below ground netto primary production (NPP) and above ground litter fall (in brackets in % of above ground NPP) for climatic forest types on different soil types (based on Vogt et at., 1996) Climatic forest type

Soil order

Above ground NPP (g m 2 yr1)

Above ground litter fall (g nv2 yr1)

Estimated below ground NPP (g m 2 yr 1 )

Boreal Cold temperate Warm temperate Subtropical Tropical

Inceptisol Inceptisol Ultisol Ultisol Ultisol

348 1180 1008 1660 1258

246 (70.7%) 436 (36.9%) 453 (44.9%) 1117(67.3%) 289 (23.0%)

n.c. 265 59 n.c. n.c.

Broadleaf evergreen Subtropical

Ultisol

1130

649 (57.4%)

1317

Tropical

Oxisol

761

844 (100%?)

379

Coniferous evergreen Boreal

Spodosol Spodosol Spodosol Ultisol

602 877

256 (42.5%)

Cold temperate Warm temperate Subtropical

n.c. 441

Broadleaf deciduous

2156 1660

290 (33.1%) 526 (24.4%) 839 (50.5%)

409 110

n.c. = not calculated

1996). Another difficulty in the estimation of the below ground biomass is the root exudation of organic acids, sugars and amino acids. Currently, it is not possible to obtain real quantitative data on the in situ exudation rates of organic compounds, because organic acids are immobilized or decomposed to a variable degree (Tyler and Falkengren-Grerup, 1998). Most available data are derived from laboratory studies. In addition, plants modify the degree of root exudation in response to exogenous and endogenous factors. This responsiveness and especially the practical problems of the study in situ have resulted in a discrepancy in the evaluation of the below ground biomass and root exudates as far as their ecological importance and role in soil forming processes are concerned (Ryan et al, 2001; Jones et al, 2003). The data in Table 1 are only reliable for annual netto primary productivity (NPP) of the above ground biomass and the annual litter fall, whereas the estimation of the annual below ground biomass production by excluding root exudates may be too low. Some of these values are, therefore, essentially lower than the net primary productivity given in older publications (Lieth, 1975; Murphy, 1975). The above ground litter fall removes between 23 and 71% of the annual netto primary production to the forest soil each year, if the value of the

Vegetation, organic matter and soil quality

45

broad-leaf evergreen tropical forest is excluded. In general, the contribution of the herb layer of forests to the annual netto productivity is less than 10%. Where trees cannot cope with prevailing environmental conditions, forests and woodlands are substituted for grasslands. The net primary productivity of most grasslands is more dependent on the water supply than on soil chemistry, except for sites with high concentrations of chemical elements (Table 2). In addition to the above mentioned problems in measuring netto primary production, the degree of herbivory above and especially below the soil surface creates an additional difficulty when it comes to establishing a realistic productivity data, especially in grasslands (Prins, 1987; Scholes and Walker, 1993). Only exclosures that hamper the access of herbivores and the application of pesticides, can help to obtain an indication of grassland primary productivity (van der Veen, 2000), but still the root exudation cannot be measured. The majority of plant species, including all tree species and most herbs and graminoids, bind atmospheric CCk by ribulose-l,5-biphosphate carboxylase (Rubisco) with a compound with three carbon atoms, the C3-pathway.

Table 2. Above ground and estimated below ground netto primary production (NPP) of grassland ecosystems Grassland ecosystem

NPP (g dry nlass nv2 yr1)1 Above Below

West African grass savannah Loudetia simplex Andropogon schirensis and Hyparrhenia species Sahelian grass savannah Schoenefeldia gracilis (annual) Andropogon gayanus (perennial) Diheteropogon hagerupii Salt marshes USA Spartina alterniflora Juncits roemerianus Carex stricta Phragmites australis Heavy metal grassland Pionier phase Optimal phase n.c. = not calculated

References Menaut and Cesar, 1986

830 1540

1320 2040

200 600

n.c. n.c.

800

n.c.

90-400 60-140 130 270

n.c. n.c. n.c. n.c.

29 600

n.c. n.c.

Penning de Vries and Djiteye, 1982

Reimold, 1977

Ernst, 1974

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W.H.O. Ernst

In warmer, more water-limited regions many tropical grasses and the crops maize, sugar cane and sorghum have modified the CO2 fixation by an initial incorporation of CO2 via phosphoenolpyruvate (PEP) carboxylase into a compound with four carbon atoms (C4-pathway) followed by the final CO2 fixation via the C3-pathway. Plant species with the C4-pathway use less water for the same amount of biomass production than C3-species and have lower leaf N-concentrations. The PEP carboxylase discriminate less against 13C than does Rubisco in C3 plants. This discrimination results in an enrichment of C4-plant biomass with 13C that also has consequences for the decomposition of plant litter (Ball and Bullimore, 1997). Decomposer micro-organisms prefer 13Cdepleted compounds for respiration and 13C-enriched molecules for biomass synthesis, so that the humus compounds of the soil will have another signature. Therefore, all data on primary netto production underestimate real productivity. At a low estimate between at least 250 and more than 2000 g litter m-2 y r i c a n b e u s e c [ for the decomposition and synthesis of new organic compounds such as fulvic and humic acids in the soil. The only exceptions are at those sites where physical constraints (substrate structure, such as rocks and ice shields, substrate mobility such as dunes, extremely low precipitation as in desert conditions, and too low or too high temperatures) and/or chemical constraints exceed the physiological capacity of plant species to adapt. 3.2.2 Roots affect soil structure and oxygenation

Plant roots can affect soil structure. Depending on the diameter - varying from very fine roots to very large roots - the length and the lateral and vertical distribution, they can modify the compactness of the soil and exploit different soil pores. Roots with a small diameter can penetrate rigid soil volumes whereas roots with larger diameter cannot do this. Root morphology is specific for plant species and ecotypes (Kutschera and Lichtenegger, 1982, 1991, 2002). Rooting depth is specific for ecosystems with a mean maximum rooting depth of 0.5 m in the tundra and 15 m in tropical grassland and savannas (Canadell et al., 1996). The greatest root depth was found in trees in the Kalahaari desert: 68 m for Boscia albitruncata, 60 m for Acacia erioloba (Jennings, 1974) and 40 m for Acacia karroo (Selaolo, 1998). Strong roots that have a deep penetration enable access to deep groundwater and make it possible for trees in miombo woodlands (Ernst and Walker, 1973) and Acacia-savanna's to flush their leaves months prior to the first precipitation. In addition, the release of deep, nutrientpoor groundwater by the roots in the top soil (hydraulic shift) modifies the hydrobalance in the grass and herb layer and creates islands of fertility around isolated trees in interaction with the top soil (Scholes and Archer, 1997;

Vegetation, organic matter and soil quality

47

Caldwell et al., 1998). In addition to water, the transport of inorganic and organic compounds along the roots to deeper soil horizons has been demonstrated for podzol profiles in birch forests and for oak trees near heavy metal smelters. In contrast, clumped roots, for example, those of the grass Avenellaflexuosa,keep back all the nutrients and water so that even after a long rainy event the soil below the roots is still dry. In general, roots of plant species building up grasslands and forests can protrude to all soil horizons (Kutschera and Lichtenegger, 1982; 1991; 2002), but the differentiation among the species of a particular type of vegetation can be high as was shown many years ago for chalk grasslands (Muller-Stoll, 1935). In addition to structural differences, there is a lot of morphological plasticity in reaction to abiotic soil conditions, especially to the heterogeneous distribution of nutrients in the soil. Normally, at nutrient shortage, the root system is strongly enlarged to enhance nutrient capture whereas when there is a surplus of a particular element in the soil, roots are stunted (Marschner, 1995). The demands of nutrients and resistance to a surplus of a chemical element is species specific, as exemplified by the stunted roots of the non-metal resistant herb Thlaspi arvense at enhanced soil metal concentrations and the proliferating roots of the metal-hyperaccumulator Thlaspi caerulescens under the same soil condition (Whiting et al., 2000). Roots of moderately flooding resistant species can briefly respond to submergence by developing roots near the soil surface if flooded during summer time (Schat, 1984) whereas flooding-sensitive species, such as most crops, are unable to respond. Physical impediments in the soil also modify root growth (Harley and Scott Russell, 1979). Plants in wetlands and aquatic ecosystems have evolved a specific tissue, the aerenchyma, which allows the transport of oxygen from the shoot (atmosphere) to the roots in water-saturated soils. Roots, however, cannot keep all the oxygen in their tissue and some leaks into the rhizosphere, the so-called 'radial oxygen loss' (Armstrong, 1967). The oxygen released in this way can modify the availability of nutrients and non-essential elements in the soil solution of the rhizosphere which results in the oxidation of the highly plant-available Fe(II) and Mn(II) into the slowly available Fe(III) and Mn(IV) and in the oxidation of the highly toxic reduced sulphur into sulphate (Sand-Jensen et al., 1982; Ando et al., 1983; Otte et al., 1989). This radial oxygen loss contributes to the development of a so-called iron plaque (pseudo-goethit) on the root surface that can be seen as a brown mantle around the roots that contrasts with the black colour of the reduced elements in the soil profile. The iron-plaque can absorb large amounts of cationic nutrients and non-essential heavy metals on their way from the soil solution into the root. Depending on the degree of saturation of

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W.H.O. Ernst

the iron-plaque, a barrier to element uptake can be formed that either diminishes metal toxicity, or enhances metal uptake, thus causing metal toxicity (Howeler, 1973; Taylor et al., 1984; Otte et al., 1989). Radial oxygen loss is species-specific. The dominance of a species in a particular vegetation creates great differences in the concentration of oxygen and/or sulphide in the soil solution of the rhizosphere and a high local heterogeneity in soils of wetland ecosystems (Table 3). In the evaluation of metal and sulphide toxicity of wetlands, a very careful and subtle sampling procedure is required because a sample taken just outside the rhizosphere will give an over-estimation of toxicity as long as the roots oxygenate its rhizosphere. In contrast to the international taxonomy of biota, soil taxonomy is still biased by national approaches, for example the national systems of the Netherlands (De Bakker and Schelling, 1989), Germany (Schachtschabel et al., 1992), the United States of America (USDA, 1975), and many other parts of the world (FitzPatrick, 1991). A proposal for an international system by FAO (1998) has received little attention in recent scientific publications. Unfortunately, all these classifications exclude the organic soil layer. 3.3 The organic soil layer Plant species deliver organic carbon to the soil by litter and root exudation. Together with nitrogen-fixing micro-organisms they enrich the soil with

Table 3. Concentration of oxygen and sulphides in the rhizosphere of some wetland vegetation types in Switzerland (data from Yerly, 1970) Vegetation type Jiincehim subnodiilosi

Jiinco-Filipenduletum

Cardamino-Scirpetum

Sphagno-Trichophorehim

Soil depth (cm)

Oxygen (mg O2 L-i)

0 10 35 50 30 65 120 0 10 30 140

6.7 1.1 0.1 0.05 0.9 0.55 0.40 0.35 0.0 0.0 0.0

Sulphide (mg S- L-i) 0 0 0 0 0 0 0 0 2.0 3.6 14.4

Vegetation, organic matter and soil quality

49

nitrogen, thus determining the C/N ratio of soil organic matter (SOM). Flaig (1975) has stated: "In the long-term organic matter does not accumulate in the soil under natural conditions"; but he has forgotten that this statement is only correct for soils that have not been exposed to waterlogged conditions. At high water tables in marshes and bogs the decomposition of organic matter is delayed considerably, so much so that organic matter accumulates and organic soils develop. Litter as the starting point of SOM has two components, the mineral elements and the organic substances. Both are accumulated and synthesized in a plant-, ecotype- and/or genotype-specific way. The high diversity of chemical compounds in the litter determines the decomposition process and finally the humus form in a species-specific way. Altitude and related regional factors (climate, mineral richness of the parent material, relief) are the main modifiers in the humification process. Recently, Wilson et al. (2001) have visualized the interaction of two soil factors, i.e. the soil nutrient regime and the soil moisture regime, by means of a so-called "soil quality grid", but they admitted that the structure and the chemistry of the parent material has to be incorporated into a three dimensional grid. This is presented in Figure 2. 3.3.1 Inorganic elements in the litter

Plant species have the ability to acquire inorganic elements from the soil in various degrees and to transfer them in an element- and plant-specific manner to the above ground parts of the plant. The highest mass flux is found for the macro nutrients nitrogen, potassium and calcium, followed by magnesium, phosphorus and sulphur, and the micro nutrients boron, chloride, copper, iron, manganese, molybdenum, nickel, and zinc. During the senescence of leaves, plants with a short supply of nutrients withdraw and retranslocate these elements to surviving plant organs. As a result of a preferential re-allocation of nitrogen and phosphorus, litter is mostly poor in these macro nutrients, resulting in high C/N and C/P ratios. Exceptions here are trees that are associated with nitrogen-fixing micro-organisms. The ash (Fraxinus excelsior), which is not associated with nitrogen-fixing organisms has a low C/N ratio in its litter because nitrogen is not retranslocated from the senescent leaves in the autumn. Where there is a high supply, chemical elements are accumulated in senescent leaves and finally in the litter. The pattern of accumulation of chemical elements in senescent leaves depends on the chemistry of the weathered parent material and on species specific processes. Acidic substrates are rich in aluminium, iron, and manganese. Consequently, plant species on acidic soils are resistant to a surplus of available aluminium (Ma et al., 2001)

50

W.H.O. Ernst

Soil moisture regime VD = very dry (1-2) MD = moderately dry (3-4) F = fresh (5-6) MO = moist (7-8) W = wet (9-10)

Soil nutrient regime VP = very poor (1-2) P = poor (3-4) M = medium (5-6) R = rich (7-8) VR = very rich (9)

Figure 2. An Ecological Site Classification quality cube based on the idea of Wilson et al. (2001). The scale of the soil moisture regime and the soil nutrient regime is based on modified Ellenberg indicator values (in brackets). As far as soil structure is concerned only the dominant types in the Netherlands have been considered

and manganese (Ernst and De Neeling, 1986). The low toxicity of high concentration of iron does not demand a specific adaptation to its surplus. Several shrubs belonging to the family Ericaceae as well as some trees are not only resistant to these inorganic elements, they also accumulate them especially in their leaves. On acidic sandy soils the leaves of Betula pendula and Sorbus aucwparia have been found to contain high manganese concentrations (Table 4). Plants that grow on limestone have to cope with a surplus of calcium (Ernst, 1978) and with a low supply of manganese and iron, the latter resulting in the excretion of phytosiderophores (Gries and Runge, 1992) and organic acids to enhance the mobility of these micro nutrients. Certain plant species can accumulate elements that are not enriched in the soil. The leaves and litter of Betula (Table 4), Populus and Salix (Ernst, 1984) are

Vegetation, organic matter and soil quality

51

Table 4. The concentration of the micro nutrients Cu, Fe, Mn, and Zn in the mature leaves of tree species on podzol soil in Schoonloo, the Netherlands (Van der Werff, 1981) and on limestone at Brochterbeck, Germany (Ernst, 1978) Tree species

Micro nutrient concentration (mmol kg"1) Cu Fe Mn Zn

Acidic sand (podzol) Betula pendula Carpinus betulus Corylus avellana Quercus robur Salix caprea Sambucus nigra Sorbus aucuparia

0.05 0.10 0.11 0.13 0.13 0.09 0.08

2.3 5.8 5.7 5.7

0.10 0.43 0.17 0.14

3.7

6.6 4.8 7.7

38.8 29.4 33.8 10.8 16.8 9.9 37.5

7.68 0.66 0.99 0.66 0.87 1.71 1.00

Limestone (rendzina) Acer pseudoplatanus Betula pendula Carpinus betulus Corylus avellana Fagus sylvatica Fraxinus excelsior Quercus robur Sambucus nigra

0.11 0.19 0.16 0.15

1.9 5.0 4.3 3.6 5.1 4.4 4.3

0.6 1.4 2.7 0.4 4.6 0.5 1.9 1.2

0.51 2.49 0.92 0.50 0.66 1.24 1.02 1.12

characterized by relatively high zinc concentration, thus contributing to an increase of this micro nutrient in the humus layer (Van der Werff, 1981). As such they can stimulate the growth of highly zinc-demanding grasses, like Molinia caerulea, in the understorey (Taylor et al., 2001). But the zinc concentration of this tree litter is low in comparison to that of plant species on zinc-enriched soils (Table 5). Plants on metalliferous soils, i.e. metallophytes, can accumulate physiologically essential and non-essential elements in such concentrations (Ernst, 1974) that they become toxic to non-adapted plants and animals. The Zn-hyperaccumulator Arabidopsis halleri can have more than 3% zinc in its dry leaves. There is also an ecosystem-specific accumulation of chemical elements in litter. The high demand and accumulation of silicon by many grasses and sedges results in a silicon-enriched litter in wetlands and (semi)-natural grasslands (Epstein, 1999). Differences in the chemistry and structure of the soil and in the chemistry of litter is the reason for the great diversity of vegetation types such as heavy metal and serpentine vegetation, salt vegetation, dunes on acidic and alkaline

52

W.H.O. Ernst

Table 5. The concentration of the micro nutrients Cu, Fe, Mn, and 7M in the mature leaves of herbs and grasses of a heavy metal grassland in Blankenrode, Germany Plant species

Micro nutrient concentration (mmol kg 1 ) Cu Fe Mn Zn

Metallophytes Arabidopsis halleri Minuartia verna Viola calaminaria

0.21 1.20 0.16

5.4 7.4 2.4

1.0 0.4 1.2

501.0 81.0 43.6

2.2

0.1 0.1 0.4

17.4

Special ecotypes herbs Campanula rotundifolia Plantago lanceolata Ranunclus acris Rumex acetosa Silene vulgaris Thymus serpyllum

0.06 0.15 0.09 0.08 0.14 0.08

0.9 1.8 2.3 1.4 3.2

2.0 0.4 0.2

10.9 29.8 23.4 31.1 16.1

Grasses Agrostis capillaris Festuca ovina Molinia caerulea

0.17 0.14 0.02

0.7 0.4 0.5

2.3 0.8 0.3

8.2 6.7 1.6

soils, chalk grassland, marshes, bogs, and various types of woodland and forest ecosystems (Ernst, 1974; Beadle, 1981; Ellenberg, 1988; Walter and Breckle, 1999; Pott, 1992; Stortelder et al, 1999). The ecophysiological potential of plant species to cope with enhanced concentration of one or more chemical elements in the soil is partly formalized in the indicator values for plant species in Central Europe (Ellenberg et al., 1991). Plants with a broad ecological response at the species level, but with a high specialization at the inter-population level are unfortunately ranked as "indifferent" by Ellenberg et al. (1991) and very seldom as "highly specialized", which would be necessary for an adequate application in the evaluation of soil quality (see Ernst, 2003). The species-specific accumulation of chemical elements in the litter has to be incorporated in the evaluation of soil quality and health, but it is missing in most ecological soil studies and in soil legislation in relation to ecological risk assessment. Even the most progressive Ecological Site Classification in forestry in England (Wilson et al., 2001) and Canada (Pojar et al., 1987) do not consider the mineral elements in litter to be important. This omission is in strong contrast to the interest in and emphasis of the hyperaccumulation of metals for phytoremediation purposes (Lasat, 2002).

Vegetation, organic matter and soil quality

53

3.3.2 The input of nitrogen

Another important biological contribution to soil chemistry is the maintenance and increase of its fertility by the fixation of atmospheric nitrogen (diazotrophy) by either free-living (Stewart, 1975) or symbiotic bacteria (Nutman, 1976). Parent rock materials and their weathering products do not contain nitrogen. All the nitrogen that is demanded by all biological processes is only present in the atmosphere as inert N2 and has to be transferred into a biologically active form (NO3, NH4, NH2). Therefore, soil fertility is dependent on the input of nitrogen by precipitation (atmospheric nitrogen oxygenation by thunderstorms) and by biological nitrogen fixation. The degree of soil fertility is determined by the balance between N-input and N-losses by leaching and denitrification. Nitrogen fixation by free-living micro-organisms is small, generally estimated below 4 kg N ha 1 yr 1 and below the nitrogen input of 9 kg N ha 1 yr 1 by precipitation in pristine environments. In rice fields cyanobacteria fix up to 70 kg N ha 1 yr1. The nitrogen input into ecosystems by symbiotic nitrogen fixation (Table 6) is essentially higher although the carbon demand of microorganisms for the N2-fixation is very high with 100 kg C for 1 kg N. There are two symbiotic types of micro-organisms: Alnus species in swamps and riverine woodlands and Hippophae rhamnoides in dune shrubs and on river banks, and Myn'ca-species in wetlands associated with Actinomycetes. Leguminous trees in Acacza-savannas and herbaceous legumes in natural

Table 6. Amount of annual nitrogen input by symbiotic Ni-fixation Plant species Trees Alnus glutinosa Alnus incana Alnus viridis Hippophae rhamnoides Herbs Lotus corniculatus Medicago sativa Medicago lupulina Trigonella foenicutn-graecum Trifolium repens Trifolium, annual species Trifolium, perennial species Ulex europaeus Tephrosia savanna S-Africa

Nitrogen fixation kg N ha 1 year 1 60 -130 37- 55 6- 15 27-180

Reference Akkermans, 1971

Stewart and Pearson, 1967 Marrs et al. (1983)

116 56-460 144 - 205 110 - 570 49 -135 45 - 335 45 - 675 27 6- 9

Scholes and Walker, 1993

54

W.H.O. Ernst

grassland and arable fields are associated with Rhizobiaceae: both associations can result in an annual nitrogen input that varies from a few tens to some hundreds of kg N. In legume monocultures, nitrogen fixation can increase up to 700kgNha 1 yr 1 . The degree of biological nitrogen fixation determines productivity, dependent on the density of the tree, shrub and herb population and the hydrology of the soil (Kaelke and Dawson, 2003). In ecosystems with only one or a few nitrogen-fixing plant species and species that make a low contribution to biomass, such as Vicia sativa subsp. nigra, the biological nitrogen input remains low so that oligotrophic and mesotrophic species are characteristic for such ecosystems. The nitrogen concentration of mature leaves is a good indicator of the effectivity of nitrogen fixation and nitrogen input into the soil (Table 7). Several factors can hamper symbiotic nitrogen fixation. Insufficient light penetrating to the understorey of mature broadleaved woodlands restricts the nitrogen fixation in legumes in forests and woodlands. On soils with low concentration of the micro nutrients nickel and molybdenum, which are essential for the nitrogen fixation process (Marschner, 1995), nitrogen fixation is very low in leguminous shrubs (Genista anglica, G. tinctoria, Ulex europaeus) in

heath lands (Ernst and Baumeister, 1978; Ernst, 1978). Adverse environmental conditions also cause a reduction in biological nitrogen fixation because Rhizobium species are very sensitive to salinity, heavy metals and soil acidity. In situations of soil acidification the most host-adequate Rhizobium strain may be replaced. This problem is already recognized in tropical agriculture (Graham and Vance, 2000), but not sufficiently studied in other parts of the world, especially not in (semi-)natural ecosystems.

Table 7. Major nutrient concentration in the leaves of plant species in a coastal dune grassland (Ernst, 2002) Nutrient concentration (mmol kg 1 dry mass)

Plant species N Agrostis capillaris Carexflacca Festuca arundinacea Festuca ovina Lysimachia vulgaris Potentilla reptans Vicia sativa subsp. nigra

P

K

Ca

860 1100

10.0 27.7

228 375

273 200

650 607

8.3 8.0 19.7

208 184

200 435 405 440 855

1170 940 1790

15.1 26.1

186 253 216

Vicia sativa subsp. nigra is the only species with symbiotic nitrogen fixation

Mg 75 70 62 84 55 112 148

Vegetation, organic matter and soil quality

55

As a consequence of the biological N2-fixation, nitrogen is not withdrawn from senescent leaves of plant species associated with nitrogen-fixing microorganisms. Therefore, their litter is very rich in nitrogen, having C/N ratios of below 10. This high nitrogen supply allows the natural development of eutrophicated ecosystems with a species-poor understorey of highly productive herbs such as Solatium dulcamara and Urtica dioica. Often the N-fixation during the season is so high that nitrate is leached into the groundwater and surpasses the human health value of 50 mg NOs L1 (Stuyfzand, 1993). The balance between the degree of biological nitrogen fixation and biological denitrification determines soil fertility and soil quality. The response of plant species to the available nitrogen (and phosphorus) pool has been formalized by Ellenberg's "N"-indicator value. This is a valuable parameter in the evaluation of soil quality despite the semi-quantitative estimate. The often cyclic processes of nitrogen fixation and litter supply to the upper soil layer are the main reasons why a soil is an environmental component permanently in a state of change. 3.3.3 Organic substances

The second component of litter, the organic substances, is produced in a species-specific way (Kogel-Knabner, 2002) and has very profound impacts on the degree of breakdown. The rate of decomposition follows the general sequence: sugar, proteins > proteids > pectines > cellulose > lignine > wax, resins > tannins (Figure 3). The organic chemistry of the litter shapes the conditions for two quite different processes: (1) Litter can be fully mineralized when the chemical compounds of the litter are easy for micro-organisms and solar radiation to handle. (2) Litter with recalcitrant compounds is the start of humification. For most soil scientists the humus layer of a soil does not belong to the soil; therefore its thickness is given as a value above soil surface (Baron et al., 1992). However, the thickness of the litter layer at a site already determines germination, penetration to photosynthetic active radiation, and frost exposure of a species in the vegetation. Plant species of pioneer ecosystems on wellaerated and warm sites produce only a little amount of litter, which is low in wax, resins and tannins. With the increase of vegetative succession, perennial plant species produce a litter with high concentrations of recalcitrant compounds. Amongst these compounds are monomeric water-soluble phenolics. In the climax vegetation of woodlands and forests, the lowest concentration of phenolics has been reported from the freshly fallen litter of Picea abies and the highest concentration from the litter of Carpinus betulus and Corylus avellana (Kuiters, 1990). During leaf senescence and in the litter,

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W.H.O. Ernst

Figure 3. T/ie decomposition curves of the various groups of organic constituents in leaf litter over a period of 15 years (Minderman, 1968) (with the permission of Blackwell Publishing Ltd.)

polyphenolics (tannins) as dominant secondary plant products can react with amino acids and proteins to insoluble complexes (Kuiters, 1990; Facelli and Pickett, 1991). They determine not only condensation and polymerization reactions mediated by soil organisms, but also affect the activity of bacteria, the development of mycelia, the germination of spores and seeds, and the extent to which soil animals consume them. So they affect pedogenesis. Tannin concentrations are very low and even absent in grasses with a rapid litter decomposition (Holcus lanatus), increase in herbs (18 - 139 mg tannin equivalents (TAE) g-1) and are highest in the leaves of shrubs and trees on nutrient-poor soils (up to 300 TAE g1) (Jackson et al., 1996). The type of tannin is related to species-specific characters. Small-molecule tannins were analysed in birch humus and large-molecule tannins in pine humus (Suominen et al., 2003). In the further decomposition process bacteria and fungi synthesize these compounds to different type of humic substances, i.e. fulvic, humic and hymelatomelanic acid and humine. The litter produced by conifers and eucalypts is not only high in phenolics and tannins, it is also high in oleoresins, a complex mixture of volatile (mono- and sesquiterpenes) and non-volatile (resin acids) terpenoids, for example, myrcene and pinene, which results in the accumulation of highly lignocellulosic material (Klason lignin) on the forest floor (Dijkstra, 1996; Kainulainen and Holopainen, 2002).

Vegetation, organic matter and soil quality

57

Fulvic and humic acids and humine The precursors of humus all have different residence times (Table 8). All humic substances can stay between two to five thousands year in virgin soil. Fulvic acids (FAs) have a low molecular mass from 800 to 9,000 Dalton. They contain more functional carboxyl and phenolic hydroxyl groups than humic acids, but both acids can bind metal cations and organic pollutants. In general, fulvic acids are less rich in nitrogen and sulphur than humic acids. FAs are easily soluble in water and contribute more to the leaching of elements to deeper soil layers than humic acids. The leaching of FA occurs preferentially along plant roots and contributes to the high patchiness of podzol soils. Humic acids (HAs) have a molecular weight that can increase from 5,000 to 100,000 Dalton and they have low solubility in water. HAs make strong complexes with Al, Ca, Fe, and Mg. The hymatomelanic acids have received very little attention. Humins have the highest degree of polymerization and stability, nitrogen concentration, and the lowest mobility in soils. The long residence time of fulvic and humic acids reflects the long organic "memory" of the soil. It makes it necessary to consider its impact on the development of vegetation and soils over a long period because changes in the composition of plant species of vegetation on a site will affect the chemical composition of the humic substances. The ratio of FA/HA may be a good indicator of soil development under different climates and vegetation types, as shown for climatic gradients from the tundra to the continental steppe in Russia (Kononova, 1975). In cool temperate climates fulvic acids are the dominant humic substance in the soil resulting in CFA/HA-ratios from 1.5 to 3.7. In warm climates humic acids dominate with CFA/HA-ratios from 0.4 to 0.9. In grassland on chernozem there was more fulvic acid present than in a broad-leaved forest on the same soil type. Podzolization of chernozems permits an increase in the concentration of fulvic acids.

Table 8. Components of humic substances, their residence time, the range of the concentration in carbon and nitrogen (Schachtschabel et al., 1992) and the related dominant humus type (Kuntze et al., 1988) Component Fulvic acid (FA) Humic acid (HA) Hymatomelanic acid Humine

Residence time (years) 1,800 - 4,300 1,900 - 5,400

C (%)

N (%)

?

43-52 50-60 58-62

2,900 - 3,500

>60

0.5 - 2.0 2.4 - 5.0 4.0 - 5.0 5.0 - 8.0

Humus type mor

moder -

mull

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W.H.O. Ernst

With increasing pedogenesis the FA/HA ratio can change. In the decomposing litter of Pinus strobus the FA/HA-ratio changed from 3.6 in relatively fresh litter to a stable ratio of 0.4 over the next 11 years, with a steady increase in the absolute concentration of FA and HA (Quails et al., 2003). FA and HA can stimulate plant growth at low concentrations, but are inhibitive at higher concentrations. Combinations of FA and HA as found in soil profiles have different beneficial effects on plant species in the forest understorey (Ernst et al., 1987) and on the root growth of grasses (Zhang et al., 2002), but can inhibit bacteria which are able to degrade DDT in forest soils (Fujimura et al., 1994). Despite these impacts the ecological and pedological interest in FA and HA is still very low, even in extensive ecosystem projects such as the Soiling project (Ellenberg et al., 1986) and the Nylsvley project (Scholes and Walker, 1993). It would be very interesting to investigate the impact of soil acidification on the CFA/HA-ratio. The main terrestrial humus forms The synthesis of humus is a gradual process resulting in a stratification of the upper organic horizon into various layers: (i) the litter horizon (Q = L = OL), (ii) the fragmentation horizon (Oe = F = OF), and (iii) the fermentation horizon (Oa = H = OH). The FAO-system (FAO, 1998) differentiates two types of organic horizons: the O-horizon which is not saturated with water for prolonged periods, and the H-horizon which is layered organic material saturated with water. The H-horizon in the FAO system should not be confused with the "H" of the fermentation layer of an O-horizon. There are three main humus forms in the terrestrial environment, i.e. mor, moder and mull. Under water-saturated conditions the humification process is delayed resulting in semi terrestrial humus (peat) and aquatic humus forms of dy and gyttja (mud, ooze). Recently, in France (Brethes et al., 1995) and in British Columbia (Klinka, 1997) the main terrestrial humus forms have been sub-divided using a different classification. Ponge (2003) has recently summarized some of the biological features of humus forms in terrestrial ecosystems (Table 9). His classification, however, is restricted to forest ecosystems, and even then it has to be modified. Mor Coniferous trees on mor have a pronounced ecto mycorrhizal symbiosis and the grasses in these forests have an arbuscular mycorrhiza. In the temperate climate zone mor is the humus form with the poorest ecological quality. It is produced by the decomposing activity of fungi with little contribution from bacteria and

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59

Table 9. Intensity of some biological features of the three main humus forms in forests and heath lands (modified from Ponge, 2003, and elaborated) and the correlation with soil nutrient classes (Wilson et ah, 2001) Biological feature Phenolic concentration Humification Nutrient; - availability to plants - soil class Dominant - mycorrhizal type

- faunal group - microbial group Soil nutrient class

Mor moderate to high very slow

Moder moderate slow

poor high

indirect by mycorrhiza medium

low

arbutoid, arbuscular, ericoid, ecto mycorrhiza none none very poor/poor

arbuscular, ecto mycorrhiza

arbuscular, ecto mycorrhiza

enchytraeids fungi poor/medium

lumbricids bacteria medium/rich

Mull low

rapid direct by root hairs

soil animals. Mor is dominant under heath lands and coniferous woodlands, mostly on silicate-rich parent material, and on Sphagnum bogs where the soil has a high water content. The mor humus is tannin-enriched and together with the dysmoder, a moder with a thick Oa layer, is described as raw humus (Kubiena, 1953). Moder Moder is due to a hampered litter decomposition with insufficient integration of holorganic faecal pellets from a rich mesofauna. Most microbial biomass is made up of fungal hyphae. There is more humine than in the mor. Moder humus is associated with oligotrophic grassland, with deciduous (oak and beech) and coniferous woodlands and forests with poor ground flora and nutrient-poor litter. Mull The best humus form is mull which is formed from litter in deciduous woodlands with a rich herb layer growing on soils with a high carbonate content and with high soil fauna activity. This environment is very suitable for earthworms. Decomposition of grassland litter on chalkstone will result in a mull. The dominant mycorrhiza of trees on mull is ecto mycorrhiza in beech-, hornbeam- and oak forests and only arbuscular mycorrhiza in herbs, grasses and the subdominant trees such as Acer-, Prunus- and Sorbus-species.

60

W.H.O. Ernst

The fine-tuning for mull will be explained in the French typology. Eumull is a mull with a thin horizon of whole leaves (no fragmented leaves are present) and with a very rich nutrient regime (Wilson et al., 2001). An amphimull has a full sequence from whole leaves (Oi), fragmented leaves (Oe) and well decomposed leaves enriched with faecal pellets (Oa); a mesomull has an Oi and a thin Oa layer, and a dysmull has an Oi and a thick Oe layer. This subdivision makes it necessary to re-associate plant species with these fine-tuned humus forms. From three coniferous species in the Inner Alps, Pinus cembra is associated with mor, Abies alba with oligomull (previously moder), and Pinus sylvestris with dysmull (Michalet et al., 2001). Oberdorfer (1946, 1994) was the first to associate plant species with the three main terrestrial humus forms and soil types (Table 10). This rough relationship is useful in a first evaluation of soil quality, if only species with a narrow humus preference are considered. Humus forms and mycorrhiza Most plant species have evolved different types of symbiosis with fungi (mycorrhiza), belonging to several phyla and follow different biological strategies. Endo mycorrhiza Arbuscular mycorrhizal (AM) fungi (Zygomycetes) are often associated with graminoids and many herbs and may be independent of the humus form (Allen, 1991). Some broad-leaved, mostly sub-dominant tree species like Prunus, Tilia, and Ulmus have an endo mycorrhyza with AM-fungi mainly related to mull and exceptionally to moder. The actual penetration of Acer species into broad-leaved woodlands and forests does not only change the litter quality in woodlands and forests in the Northern hemisphere (Kuiters and Sarink, 1986), but also enhances the dominance of Vasicular Arbuscular (VA) mycorrhizas. The AM fungi penetrate into the roots of the plants and form in the cells of the root cortex branched forms (arbuscules) where the transfer of water and nutrients from the fungal cell to the root cell takes place. Members of the orchid family have a specific symbiosis with Rhizoctonia (endo mycorrhiza) which is also necessary for the germination of orchid seeds. This orchid mycorrhiza is more frequent in mull than in moder, mor or peat thus explaining the increased biodiversity of orchids on mull. The mycorrhizal fungi of heather, blue- and billberry (Ericales) are ascomycetes (ericoid mycorrhiza) that have a preference for mor and moder. The fungi also penetrate into the cortex and can substitute the rhizodermis and transfer nitrogen and phosphorus directly into the roots of the plants The association with ericoid mycorrhiza can restrict the transfer of

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61

Table 10. Some examples of the indicator function of dominant plant species in an ecosystem with regard to the humus forms and soil structure (data based on Oberdorfer, 1994) Plant species

Humus form

Soil structure

Dwarf shrubs Empetrum nigrum Calluna vulgaris Vaccinium myrtillus Vaccinium vitis-idaea Loiseleuria procumbens

mor mor mor mor mor, moder

sand, organic sand, sandy loam, organic loam sand, sandy loam stones

moder moder » mull, mor

sandy silt loam loam

mull mull mull mull

clay, loam clay, loam loam stony and sandy loams

mor moder mull mull mull

decalcified loam decalcified loam loam, clay loam stony loams

mor mor moder mull, moder mull, moder mull peat

loam, clay sand, loam, peat loam, clay loam, clay sand, loam, clay sand, loam loam, clay, organic

Herbs Teucrium scorodonia Oxalis acetosella Aegopodium podagraria Allium ursinum Corydalis cava Mercurialis perennis Graminoids Nardus stricta Luzula luzuloides Avenella flexuosa Melica uniflora Sesleria albicans Trees Picea abies Pinus sylvestris Abies alba Fagus sylvatica Quercus robur Carpinus betulus Alnus glutinosa

non-essential heavy metals which enables Calluna vulgaris to grow on metalcontaminated soils without having a high metal resistance itself (Bradley et alv 1982). Ecto mycorrhiza In woodlands, many dominant tree species like the conifers Abies, Picea and Pinus, and the broad-leaved Carpinus, Corylus, Fagus and Quercus have an ecto mycorrhizal symbiosis with Basidiomycetes, independent of the humus form. The fungi do not penetrate into the root, but surround the roots with a so-called

62

W.H.O. Ernst

Hartigs' net and take over the function of root hairs. Due to the indirect contact with the root cells they slow down the transfer of nutrients and water from the fungus to the tree root. These ecto mycorrhizal fungi often restrict the transfer of heavy metals to a tree that is less metal-resistant than the ecto mycorrhizal fungi (Colpaert et alv 2000). But in harsh environments the ecto mycorrhizal partner of some trees changes from basidiomycetes to the ascomycete taxon Cenococcum (Ponge, 1990; 2003). Terpenes of coniferous litter hamper the development of many mycorrhizal and amycorrhizal fungi from coniferous forests (Hintikka, 1970). The degree of mycorrhization of forest trees is one aspect in the evaluation of forest health in addition to needle age classes in coniferous trees, crown development and productivity. A modification of the ecto mycorrhiza is the arbutoid mycorrhiza of shrubs like Arbutus and Arctostaphylos (Read, 2002). 3.3.4 Soil profiles

After the destruction of nearly all ecosystems and related soil processes during the last glacial period in the Northern hemisphere, post glacial vegetation development has been characterized by a sequence of dominant tree species over time. It modified soil processes finally resulting in new soil profiles. Some of these soils may be preserved (paleosols) because of burying by later sediments (Saijo and Tanaka, 2002). The importance of time lags in the decomposition of litter makes it a potential historic factor that can influence the present composition of the vegetation (Facelli and Pickett, 1991; Koerner et al., 1997). The parent material (lithosphere) shows a high variability in structure. There are differences between (i) igneous rocks such as granite and diorite, (ii) sedimentary rocks such as dolomite, limestone, sandstone, shale, and (iii) metamorphic rocks such as schist, slate and serpentine. The texture affects water capacity, and thus differentiates vegetation types by their water economy (Quezel, 1965; Beadle, 1981). It also determines the ability of roots to penetrate downward and to spread in horizontal layers. The development of vegetation and soil profiles is strongly governed by climate. Already without any human impact ecosystems have, over the past 13,000 years, created a high diversity of vegetation and soils. Present vegetation and soils represent only a status quo, biased by the past: they do not make it easy to extrapolate and forecast the climatic-driven development that can be expected in the future. The autogenic post-glacial temperature increase caused relatively rapid changes in the dominance of tree species in the forests and woodlands of the Northern hemisphere (Ellenberg, 1988). Prevalence was also modified by other processes such as a strong decrease in elm after the outbreak

Vegetation, organic matter and soil quality

63

of the elm disease around 5,000 years BC (Molley and O'Connell, 1991), and, over the past 5,000 years the movement of the coastline of the North Sea (Streif, 1990). During historic times the decrease in temperature during the Little Ice Age from the mid-14th to the mid-19th century following the warm medieval climatic period was obviously caused by decreased sun spot activity and increased volcanic activity (Fagan, 2001). This had considerable consequences for vegetation and soils. The colder climate allowed glaciers to extend across Europe (Kaennel Dobbertin and Braker, 2001) and lowered the timberline in the Alpine mountains. In North America Fagus grandifolia declined and was replaced by Pinus strobus and Quercus species (Campbell and McAndrews, 1993). All these unpredictable events in the past make it difficult to predict how vegetation and soils will develop in the future, even if the modellers of the ICCP (International Climate Change Partnership) believe they can make such a forecast without being able to explain the autogenic increase of the carbon dioxide levels in the past. Some post-glacial developments are conserved in the present-day European climate gradient. Under permafrost condition, a tundra soil Ah/C profile with a small organic horizon developed on well-watered regions. Today these soils have a humus content of ca. 1% (Kononova, 1975). This development of the soil profile will also be related to the parent material. In temperate climates, chemical weathering of rocks rich in aluminium silicate result in an acidic substrate that determines the pedogenesis of podzols. These nutrient-poor soils support an open mixed forest with birch (Betula pendula) and Scots pine (Pinus sylvestris), sometimes enriched with mountain ash (Sorbus aucuparia). The understorey of these woodlands is dominated by ericaceous species having an ericoid or arbutoid mycorrhiza with fungi or grasses with arbuscular mycorrhizal fungi with very small fruit bodies in the soil. These AM fungi can interconnect plant species of different taxonomic ranks and thus enhance nutrient exchange between plant species (Van der Heijden et al., 1998). These recent findings make it necessary to reconsider all competition models and theories because they have excluded the nutrient and water transfer at the root level between different plant species. Dwarf shrubs produce allelo-chemicals that constrain other plant species and fungi (Souto et al., 2000). The high concentrations of phenolics and tannins in the litter of these shrubs and Scots pine stimulated the synthesis of fulvic acids as its major compound and a high C/N ratio ranging from 24 to 84 (Nierop and Verstraten, 2003). The soil profile of these podzols is strongly layered (De Bakker and Edelman-Vlam, 1976; Van der Werff, 1981) with a mor humus in the O-horizon (up to 10% humus), poor amounts of organic matter in the leached E-horizon, enrichment of humus

64

W.H.O. Ernst

material in the Bh-horizon and finally low organic matter in the C-horizon (Table 11). These various soil horizons are also different in the solubility and complexing capacity of mineral nutrients (Van der Werff, 1981). Deeper horizons have a negative effect on the root development of woodland herbs and will hamper plant productivity after soil disturbance (Ernst and Nelissen, 1979). With increasing depth, the C/N ratio of these podzolic soils declines and the growing 14C age indicates a higher stabilization of the organic material in the lower soil horizons (Rumpel et al., 2002). The ecto mycorrhizal fungi of Pinus sylvestris and Betula pendula are distributed throughout the soil profile with a maximum in the mineral soil of the B and C horizon (Rosling et al., 2003). They can transfer chemical elements from deep soil horizons to the organic layer. In addition the decay of their fruit bodies enriches the upper humus with nutrients (Tyler, 1980) in such a way that it contributes to a patchy arrangement of herbs and grasses in the forest environment. In natural forests the population of tree species is not even in age. This results in the maintenance of a high biodiversity of mycorrhizal fungi related to all tree age classes (Fleming, 1985) in contrast to the low fungal biodiversity of even-aged stands in present day man-made forests. Therefore, the biodiversity of mycorrhizal fungi is a good indicator of tree population management in forests and of environmental quality. In temperate Europe and in soils with more loam and clay, oak (Quercus robur) and beech (Fagus sylvatica) - as ecto mycorrhizal trees - dominate the climax vegetation (Pott, 1992; Stortelder et al., 1999). Oak and beech litter is high in phenolics and tannins (Ellenberg et al., 1986; Kuiters, 1990) that hamper many decomposer fungi (Harrison, 1971; Hering, 1982), but nevertheless improve the C/N ratio varying between 10 and 23 (Kononova, 1975; Ellenberg et

Table 11. Composition ofhumic substances in the different soil horizons and pH of a podzol profile near Schoonloo, the Netherlands (Van der Werff 1981) Horizon O Ai

E B2h B22

C

pH 3.84 4.08 4.16 3.88 4.03 4.15

FA = fulvic acid HA =: humic acid

% of soil FA

HA

3.30±0.28 0.65±0.03 0.42±0.07

8.29±0.81 2.40±0.31 0.88±0.34

3.44 5.29 8.74

1.36±0.45 0.44±0.03 2.92±0.38

6.23±0.01 0.79±0.13 0.78±0.10

1.03 9.64

C/N ratio

%Csoil CFA

55.93

CHA

8.99 37.68 45.23 14.27 32.80 25.61

CHA/CFA

2.61 7.12 5.17 13.85 34.02 0.46

30.7 14.9 43.8 52.2 14.3 88.7

Vegetation, organic matter and soil quality

65

al., 1986; Nierop and Verstraten, 2003). Weathering of rocks rich in calcium carbonate produces alkaline or neutral material (pararendzina) supporting species-rich beech-woodlands on moderately moist to semi-dry soils, hornbeam (Carpinus betulus)-woodlcLnds on dry soils and lime (Tilia)-woodlands on very stony soils on steep hills (Pott, 1992) with a C/N ratio of less than 15 (De Bakker and Edelman-Vlam, 1976). Where acidification and decalcification proceeded unhindered, pararendzina progressed to brown calcareous soils with a stimulation of oaks. Most of these tree species and all herbs have leaf litter that can decompose more rapidly than oak litter (Kuiters and Denneman, 1987) resulting in the synthesis of mull. The dominant soils are brown forest soils with an improved humus content of between 4-6% (Kononova, 1975). In general, it is estimated that the development of soil profiles takes hundred(s) to thousand of years. 3.3.5 The catena and the vegetation

The functionality of soils and their vegetation are often evaluated on a very limited scale, but they have to be considered in the spatial context of a landscape as soon as geomorphological relief is present above and below the soil surface. A landscape can be very dynamic as in the case of the coastal dunes or mountains, or very flat like some marshlands. In a dynamic landscape such as coastal dunes there is a zonation of vegetation types over distances of only a few hundred meters. The age of the dunes varies from post-glacial inner dunes to medieval or grey dunes, and the present day or fore-dunes. A consequence of this dynamic is that the vegetation ranges from very dynamic pioneer vegetation to highly stable climax forests with very different soil profiles and humus forms with an interdependence of the vegetation zones characterized by different pathways for rain and groundwater and a gradient of sand mobility as visualised in Figure 4. The sequence of vegetation zones on coastal dunes can be easily recognized. The sequence starts with the drift line where, during high winter floods, mainly algal material and shells are deposited and later overblown by sand, creating a buried O-horizon resulting in a C/Ob/C horizon, which is never described by soil scientists. Indicator plants are summer-annual nitroresistant species like Salsola kali and Cakile maritima. On the higher part of the beach, primary and secondary dunes (yellow dunes) are formed by pioneer vegetation that includes a few grasses like Elytrigia atherica, Ammophila arenaria, and Leymus arenarius, which catch and fix the mobile sand. These grasses are supported by their association with mycorrhizal fungi. Part of the calcium-rich groundwater of the grey dunes is exfiltrating into the yellow dunes. There

66

Vegetation zones: 1 = driftline 2 = foredunes 3 = wind-exposed yellow/grey dune 4 = Hippophae - shrubs

W.H.O. Ernst

5 = winter-annual grassland 6 - perennial dune grassland 7 = dune slack 8 = Crataegus woodland

9 = Betula pubescens woodland

10 - oak-beech-maple woodland

Figure 4. The vegetation zonation in the coastal dunes of North Holland, the prevailing groundwater stream (arrows) and the associated soil profiles. The groundwater streams are based on Stuyfzand (1993)

is no evidence of pedological horizonation on these yellow dunes (Ranwell, 1972; Wilson, 1992). The absence of these species indicates a high disturbance of the upper beach by tourists and a danger to beach stability. At the higher tertiary or grey dune level exposure to the prevailing wind and salt spray influences soil development in an open dune grassland with Ammophila arenaria, Festuca rubra and Carex arenaria on a wind-exposed site with

small soil development of an Ah/C profile. On wind-protected sites the litter of buckthorn (Hippophae rhamnoides) living in association with nitrogen-fixing actinomycetes builds up a thick humus layer (mull) with a low C/N ratio of 7 10, resulting in an O/Ah/C profile. The high nitrogen concentration of the litter supports the growth of nitrophilous shrubs like Sambucus nigra and herbs like Urtica dioica in the under storey. The surplus of nitrate in the litter cannot be fully used by the vegetation so that leaching into the groundwater is common (Stuyfzand, 1993), surpassing the European groundwater limits for nitrate. For a long time, it has been suggested that the dying back of Ammophila arenaria and Hippophae rhamnoides is caused by decalcification of dune sand. Causal analyses, however, have shown that the lack of sand transport from the beach allows

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pathogenic nematodes to increase their populations to such an extent that the plant populations of A. arenaria and H. rhamnoides degenerate (Oremus and Otten, 1981; Van der Putten et al, 1988). Behind the grey dune, a dune grassland with winter-annuals can become established where sand, blown from the beach and the fore-dunes, is deposited resulting in a soil with no profile (C) or, in the case of high annual variability of sand transport, in a soil with overblown O or Ah horizons. On stabilized inner dunes vegetation with perennial plant species builds up a soil with an O/Ah/C profile. Soil properties change with exposure time to the leaching effect of rainfall and the grazing intensity of rabbits (Ranwell, 1972). The thickness of the organic soil layer can vary between 11 cm and 73 cm and the C/N ratio between 14 and 19 (Sival and Grootjans, 1996). The presence of Corynephorus canescens is an excellent indicator of the decalcification of coastal dune grassland. On sites with a high water table, i.e. dune slacks, a species-rich slack vegetation is primarily composed of grasses and sedges. The high water table diminishes the decomposition so that young slack soils after a period of 5 to 15 years have a thick layer of organic matter (Ernst et al., 1996; Sival, 1996; Berendse et al., 1998). The seasonal variation of the water table results in bands of iron hydroxide in the C-horizon followed by a gley in the permanently reduced G-horizon, thus the O/Ah/G/C-profile (De Bakker and Edelman-Vlam, 1976). In such dune slacks, water rich in iron and calcium from the higher dunes exfiltrates and contributes to the maintenance of a chemical gradient in the top soil and oligotrophic vegetation. At the other site of the slack the water, enriched by nutrients from the decomposing litter, infiltrates and supports a mesotrophic vegetation (Grootjans et al., 1996). The presence of the sedge Schoenus nigricans together with its nutlet predator Glyphipterix schoenicolella

indicate that the dynamic of the water table is adequate for the maintenance of these dune slacks (Ernst et al., 1995). Dune slacks protected from exposure to salt spray support a Betula pubescens woodland. With increasing distance from the coast, a Crataegus woodland and finally a climax forest with oak on the warm coastal sands or with beech and maple on more clayey sands develops. There is a relatively rapid turnover of organic material resulting in a weakly developed O/Ah/E/C profile, due to the leaching of calcium out of the E-horizon. Ground water from the inner dunes is exfiltrating into the neighbouring marshlands. This interdependence of vegetation zones has been recognized in recent dune management (Van Til and Mourik, 1999), although recognition of the soil processes involved is still insufficient.

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3.4 The impact of human activities: shaping anthrosols The destruction of soils and the human influence on soil formation is directly related to human population increase and the impact of human activity (Figure 5). This began during the Stone Age and has increased dramatically in recent centuries. The development has a strong regional component and can be separated into different phases of intensity. In the Mediterranean and the temperate zone of the Northern Hemisphere woodland, as the climax vegetation of non-waterlogged soils, was transformed into meadows, arable fields and settlements on plateau areas with three major types of effects: (1) disturbance of soil structure and soil profiles, enhancing erosion, and changing the relief by terracing and levelling, (2) a change of soil chemistry by deposition of pollutants, by irrigation and the application of fertilizer, and (3) a change in

Figure 5. The aspects of human impact on soil processes are given in the encircled fields. The thickness of the lines indicates the importance of the impact

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the biological components of the ecosystems creating new ecosystems. Consequently, in areas of early human concentrations landscapes were deforested not only for the purpose of agriculture, but also for ship- and housebuilding, fire-wood and for use in the metal and saline industries. As a result of climate change, huge areas of the Sahara turned from semi-arid to arid grassland and deserts. In the Middle Ages a first wave of desertification affected Europe. This can be seen in the overblown soils in sandy areas that have lead to the conservation of paleosoils. The grazing of woodlands by cattle affected the regeneration of forest trees and changed the understorey. Much of the grassland and heather that are now seen as interesting for nature conservation were shaped by the grazing of sheep, cattle and pigs and by sodcutting, creating patches of soil of varying fertility and disturbed soil profiles on a landscape scale. In sub-tropical Africa the strongest impact of human activity was soil erosion, the result of overgrazing (Abel and Blaikie, 1989; Sutherland and Bryan, 1990; De Groot et al., 1993) and bush encroachment on a regional scale (Van Vegten, 1983; Prins and Van der Jeugd, 1993). Human population growth has resulted in the abandonment of fallows and rotation culture. In contrast to the general belief that fallows served to re-establish or restore soil fertility and suppress weeds (Ruthenberg, 1983), it has been demonstrated that the revegetation of arable fields by indigenous grasses leads to a decrease in the population of root-parasitic nematodes during the fallow period, resulting in an increased crop yield for the first one or two years at the end of the fallow period (Rodriguez-Kabana and Canuilo, 1992). The switch to monocultures and to uniform crop cultivars has lead to an increase in the parasitism of cereals by root-hemiparasites, i.e. Striga- and Afecfra-species (Parker and Riches, 1993), and by root-holoparasites of tobacco, sunflower and legumes, i.e. Orobanche-species (Parker, 1994). In the tropics, the impact of man is more recent. Large-scale deforestation of tropical forests and enhanced shifting cultivation has extended over-exploitation and soil erosion (Myers, 1991; Kleinpenning, 1993). In general, human activities may affect one environmental component, but they may also modify many of them by physical, chemical and biological disturbances as happened in the coastal dunes of the Netherlands. This includes lowering of the groundwater table (i) as a result of extraction of dune water for drinking water from 1860 onwards, (ii) by polder reclamation, (iii) by planting coniferous trees, and (iv) by burying deep ditches for the filtration of river water. It also includes the eutrophication of the soil by the infiltration of river water and nitrogen and sulphur deposition. Furthermore, it relates to breaking

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down geomorphological processes by (i) the fixation of mobile dunes, (ii) terracing dune valleys for potato fields when farmers were resettled after an outbreak of potato disease in Brabant in the mid-nineteenth century, and (iii) sand excavation for bulb fields. Diminishing biological diversity as a result of cattle grazing should also be included. There is no vegetation and soil in any ecosystem that has escaped mechanical damages and modifications of its chemistry by human activities. Physical impacts on the soil, resulting in soil compaction can be evaluated by root development. This is restricted in compacted soil. As far as chemical changes in the soil are concerned there are several, sometimes very specific indicators that can be applied. Ellenberg's indicator values evaluate the direction of the vegetation change together with the determination of the loading of soils with nitrogen, phosphorus, sulphur, and protons. Saturation of soils with phosphorus on a regional scale will not only diminish biodiversity, but also restrict the mycorrhization of plants so that the absence of water supplied by mycorrhizal symbionts will result in water deficits for plants during droughts. 3.4.2 Disturbance of soil structure and soil profiles

The disturbance of ecosystems and their soils has been a fundamental part of establishing arable fields for food production. Soils were modified by removing large areas of natural forests on non-flooded soils and by establishing agricultural fields and pastures on plains on a large scale. In the early days of agriculture only the harvestable parts of the crop was removed, so that the decomposition of stubble with its high content in lignins could contribute to humification. With improved and mechanized ploughs it became possible to increase ploughing depths to such an extent that the deeper parts of the soil profiles were destroyed. The result was the mixing of the A, B and C soil horizons (Figure 6). In soil classification terms these man-made soils are known as anthrosols (FAO, 1998) and have their own taxonomy: agrisols for agricultural fields, hortisols or fimic anthrosols for garden soils, urbic anthrosols for soils in cities and industrial areas where the organic soil layer has been removed and replaced by sand, sewage or industrial sludges (Schachtschabel et al., 1992). Ploughing increases the rate at which organic matter decomposes and affects the chemical composition of humus, and nutrient mineralization (Szajdak et al., 2003). It also exposed soil organisms in the deeper soil layers to drought and frost. The high frequency with which soil was turned over resulted in the natural selection of agricultural weeds with short life cycles and with different degrees of seed dormancy. Ploughing has also been extended to woodlands and forests prior to tree planting. This led to

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Figure 6. An anthrosol: a thoroughly disturbed soil profile caused by ploughing a podzol (modified after De Bakker and Edelman-Vlam, 1976, p. 48)

the destruction of the soil profile and allowed woodland herbs to be arranged in rows, for example, Senecio sylvaticus in Scotch pine forests. On a disturbed podzol profile, most plant species show a strong positive preference in biomass production for the O-horizon compared to the low productivity on Ai to B3 horizons (Ernst and Nelissen, 1979). Sod or turf cutting began on podzolic soil in the early Middle Ages, although recently evidence has shown that such soils can be dated back to the Bronze Age (Schachtschabel et al., 1992). Sod-cutting has two aspects. Firstly the removal of the upper soil layer of heath land and grassland and its transport to stabled livestock where it was used to absorb faeces and urine resulting not only in heath lands being depleted of nutrients but also in hampering the

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succession to woodland. Secondly the transfer of these nutrient-enriched sods to arable fields was the first organic manuring of fields, and resulted in 'plaggensol' (FAO, 1998). When sods are derived from grassland on brown earth, they are rich in phosphorus and have a C/N ratio between 12 and 20. When sods are dug up from heath land on podzol, they contain only half of the phosphorus and have a C/N ratio between 15 and 30. With the availability of industrial fertilizer, sod-cutting for agricultural purposes is no longer practised. However, it is still carried out in the interest of nature conservancy either to keep heath land attractive for tourists or to establish "juvenile" soils for pioneer vegetation (Bakker, 1989). Sod-cutting continues to have a disturbing effect on the natural development of soil profiles, but this does not seem to worry those involved in nature management. Intense litter collection in woodlands in former times by farmers and today - often illegally - by citizens for their gardens, deprives the forest floor of its seed bank, and deprives soil of organic material, enhances acidification and leads to a reduction in forest fertility (Bray and Gorham, 1964; Ellenberg, 1988; Dzwonko and Gawronski, 2002). The use of heavy machines has caused changes in soil structure as soil became compacted. In arable fields water collects and stagnates and crops are damaged because they are deprived of oxygen. In forests, tree harvesting compacted the soil to such a degree that it has severely hampered the restoration of stands by oaks (Gaertig et al., 2002). Ploughing without immediate sowing can cause soil to be eroded by water (in hilly landscape) and fertile upper soil removed by wind. Terracing of hills is another human activity that promotes soil erosion and sedimentation in river flood plains resulting in old soil profiles becoming buried. Decreasing water levels caused desiccation during periods of low precipitation, changing the turnover of nutrients in wetlands to such an extent that ecosystems eventually became affected. The construction of settlements and infrastructure, i.e. roads and railways have removed the organic layer and the A-horizon, covered the remnant soil profile with sand or gravel or closed up soils (urbic anthrosols). This will be considered later in the section of buried soils. The strongest impact of human activity on soil chemistry and ecosystems has been the harvesting of biomass because it removes part of the nutrient storage and disturbs ecosystem balances. 3.4.2 Selection of plant resistance by changing soil chemistry via deposition of pollutants, irrigation and application fertilizers

Man has modified soil chemistry in many ways. These include the loss of

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control over the emission of chemicals on a local as well as a global scale, by the irrigation of crops and grassland, and through a lack of care in the quality control of fertilizers. Soil quality has been intentionally modified by the application of fertilizers and pesticides in agriculture, horticulture and forestry, and affected by salinization of roads by de-icing salts. All these activities have affected natural vegetation and soil quality, often for tens of centuries. While evaluations of this change in soil quality are often carried out for the upper soil layer, they very seldom extended to the complete soil profile. Heavy metals and plant response The first pollution of soils with heavy metals dates from the advent of the open smelting of ores. This activity contaminated soils, lakes and ice with copper and lead, but not with cadmium and zinc (Figure 7). The effects sometimes went beyond the immediate locality or region. In the northern hemisphere it created a non-natural background dating back to the Roman period (Hong et al, 1994, 1996, 1997). The bad management of metal-processing industries that were established far from metal ore outcrops resulted in heavily metal-loaded fallout. The metals became bound to the organic soil material in the upper soil layer destroying the natural vegetation and favouring those plant species - mostly grasses belonging to the taxa Agrostis and Festuca - which have a high genetic

Figure 7. Lead deposition in three uC-dated bog profiles. The first period of lead contamination was due to open smelting procedures by Roman miners: Featherbed Moss in Derbyshire (England) is in the vicinity of Roman lead mining sites (Lee and Thallis, 1973), Venner Moor in Westfalia, Germany, (Ernst et al., 1974) is at 160 km NE and Engbertsdijksveen in Overijssel, the Netherlands, (Van Geel et al, 1989) at 200 km NNE of the Roman lead mining sites near Breinigerberg, Germany

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potential for metal resistance in their gene pool. They build up new grasslands in former wooded areas on mostly podzol soils (Ernst, 1989; Wickland, 1989). Where there is a delayed leaching of metals to deeper soil layers, the metal exposure of the deep-rooting, up to 2 m, grass Molinia caerulea can be delayed for more than 100 years, whereas the shallow-rooting, up to 15 cm, grass Agrostis capillaris is immediately exposed, as observed near a Cd/Zn-smelter (Dueck et al., 1984). The ecological functions of these long-term metalcontaminated soils depend on the degree of contamination. At moderate metal loading the microbiological processes are scarcely hampered due to the selection of metal-resistant micro-organisms (De Rore et al., 1994). Only nitrogen-fixing Rhizobium-species are sensitive to a surplus of heavy metals due to a low potential of the evolution of metal resistance (Wu, 1989), as is also demonstrated by their low presence in naturally metal-enriched soils (Ernst, 1974). In general, the biodiversity of these anthropogenic metal-contaminated soils is lower than that of naturally metal-enriched soils, although nearly every soil can be colonized by one or more metal-resistant plants. The application of metal-based pesticides, metal-loaded sewage sludge and manure, as well as micro nutrient enriched fertilizers in agriculture (copper on peaty soils, zinc in maize field) and horticulture and the emission of leaded gasoline by cars are other sources of metal enrichment affecting the environment which continue to increase the natural background world-wide. They modify the survival of plants, microbes and animals and promote the selection of metal-resistant organisms. In the case of anthropogenically metal-enriched soils, the analysis of the metal concentration in plants and soils can give a first indication of the situation. Due to the evolution of metal-resistant populations of plants, microorganisms and animals, all functions of an ecosystem may operate in the same way as they do in naturally metal-enriched soils. However, biodiversity will be low. Only laboratory experiments can identify the degree of metal resistance. In peaty soils the binding of these metals to soil organic matter will affect soil processes for decennia or even longer (Verloo, 1980; Smith and Siccama, 1981; Dabkowska-Naskret, 2003). Moreover it will cause environmental problems as soon as the organic matter is oxidized. On sandy soils, with low humus content, this metal attenuation is restricted to the upper soil layer. Cadmium can be easily be translocated within the soil profile and leached to the groundwater whereas the binding of copper and lead to soil organic matter is relatively high (Dijkstra, 1996).

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Sulphur emission and soil acidification Another pronounced human impact on ecosystems on a continental scale started around 1800 with the emission of gaseous air pollutants (sulphur dioxide, nitrogen oxides and ammonia, and many other new chemicals). This affected the epiphytic lichens diversity on the trees, and disturbed natural ecosystems at regional and global level and enhanced the acidification of soils with a low buffer capacity. Liming and so-called 'vitality' fertilization of forests as a de-acidification measure on acidic soils stimulated tree growth, and disturbed the biogeochemical cycle of nutrients and encouraged the leaching of nutrients to groundwater, rivulets and rivers, thus enhancing eutrophication and acidification of other environments (Kreutzer, 1995; Frank and Stuanes, 2003). Plant species with a high resistance to a surplus of sulphur, especially members of the Brassicaceae, have expanded their area in Europe during the last two decades (Ernst, 1993). The restriction of sulphur dioxide emission in the 1980s has lead to the recovery of many SCh-sensitive lichen species in several Western European countries (Hawksworth and McManus, 1989; LetrouitGalinou et al., 1992; Van Dobben and De Bakker, 1996) and caused a decrease in SCte-resistant lichens (Bates et al., 2001). However, due to the down-regulation of the SO2 emission, rape (Brassica napus) and other crops now suffer from sulphur-deficiency if farmers do not fertilize their fields to compensate for the lack of the free-of-charge S-fertilization provided by atmospheric pollution (Schnug, 1993; 1997). Interest in the impact of sulphur on vegetation and the soil quality is just beginning to grow. The dominance of many cruciferous plant species on ruderal areas demonstrate the high sulphur attenuation in soils which are not cropped with highly S-demanding species. Man-made N- and P-eutrophication, soil acidification and contamination with pesticides The establishment of settlements provide a nucleus for the eutrophication of soils and the expansion of eutrophic vegetation. The first selection of nitrogenresistant plant species occurred in the vicinity of settlements early in the Bronze Age when the first agricultural weeds were introduced. After the application of heather sods had been stopped, the production of industrial fertilizers enhanced the eutrophication of fields and meadows and, as a result of runoff, also affected aquatic environments with detrimental effects for oligotrophic and mesotrophic ecosystems. The most extreme eutrophication is the saturation of soils with phosphorus (De Walle and Sevenster, 1998) which will last for centennia. The saturation of soils with phosphorus on a regional scale will not

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only diminish biodiversity, but also restrict the mycorrhization of plants so that the absence of the water supply by mycorrhizal symbionts will result in water deficits in the plants during droughts. From the 1970s onwards, high amendment rates of manure and sewage have changed the humification process by increasing the aliphatic character of organic substances and by decreasing the free radical concentration of humic acids in agricultural soils (Rovira et al., 2003). The emission of nitrogen compounds has eutrophicated all ecosystems and diminished the competitive ability of oligotrophic plant species, especially on soils derived from sandy material, on a continental scale. Eutrophication has lead to a decrease in biodiversity and the persistence of oligotrophic ecosystems because it injured sensitive plants species (Van der Eerden, 1992) and stimulated the growth of nitrophilic grasses in many semi-natural grasslands, bogs and heathers (Bobbink and Willems, 1987; Ernst, 1998a; Limpens et al., 2003). In the vicinity of ammonia-emitting bio-industries, nitrophilic lichens are dominant on oaks and the nitrophilic forest herbs Ceratocapnos claviculata cover the forest floors (Ernst, 1998b; Van Dobben et al., 1999). Its presence is an excellent indicator of this type of air pollution (Figure 8). Over-fertilized meadows can easily be recognized by the presence of Rumex crispus and R. obtusifolius, and eutrophicated ditches by a thick carpet of Lemna minor. Another aspect of eutrophication is that the rooting system is restricted to the top soil and there is a reduction in the degree of mycorrhization. In highly fertilized soils the nutrient demands of the plants can be saturated by roots growing near the soil surface; symbiosis with fungi is not necessary for the exploitation of nutrients in small soil pores. A short rooting system and a low degree of mycorrhization, however, have negative consequences as far as a sufficient water supply during droughts is concerned and this increases the possibility of crop losses. Enhanced levels of nitrogen in the litter can increase decomposition (Tuma, 2002) and will affect the humification process. A spin-off of the eutrophication of the environment with SCfeand NH3is the acidification of soils, at least over large areas of Central and Northern Europe. As soon as SO2 is dissolved in water, a strong acid is formed which can be further oxidized to sulphate which is taken up by the plants while the protons remain in the soil solution. In contrast, precipitation enriched with ammonia has a basic reaction. However, uptake of ammonium from the soil solution by plants is correlated with the excretion of protons to maintain cellular cation balance. This process has been known in agriculture for a long time and farmers describe ammonium-enriched fertilizers as acid fertilizer and nitrate-based fertilizer as alkaline. Many plant species which have a preference for

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Figure 8. The extension of the area of the annual herb Ceratocapnos (Corydalis) claviculata during three periods of monitoring (left) in comparison with the ammonia emission in the Netherlands in 1980 (right). The data base for the period prior to 1950 and between 1950 and 1980 is from Van der Meijden et al. (1989), that for 1988 from Van der Meijden (1999). The ammonia emission is modified after RIVM/CBS (2001) and given in four classes: 0-25 kg N ha1 yr1 (white), 25-50 kg N ha1 yr1 (dotted), 50 100 kg N ha-1 yr1 (hatched), >100 kg N haA yr1 (black)

ammonium above nitrate uptake are acidifying the soil solution. In the top soil of silicate soils, the pH sometimes decreases by more than 1.0 pH unit (Ulrich and Pankrath, 1983). In these soils the main acidity buffering agents have changed from a stage of base cations to one based on aluminium, increasing the availability and finally the toxicity of aluminium to plants. As a consequence, below a pH of ca. 4.5 plants can only modify Al3+ toxicity by exudating organic acids (Marschner, 1995) or by an enhanced uptake of H+ (Degenhardt et al., 1998). The result is that Al-sensitive species disappear (Banasova and Sucha, 1998) and that the humus will change to a mor or mor moder with high acidity. The design of organic pesticides and their world-wide application has brought new chemicals to the soils with different degrees of persistence. The application of pesticides has not only diminished the weeds in arable fields, but has also resulted in the selection of herbicide-resistant weeds (Holt et al., 1993), and perhaps also micro-organisms. The longevity of the seeds and the binding

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of the pesticides to humic and fulvic acids (Hesketh et al., 1996) will preserve this memory of human activity for many decennia in the organic soil layer. 3.4.3 The introduction of new plant species and their impacts on soils

With the introduction of plant species with a harvestable plant part below the soil surface such as potato, sugar beet and bulbs and with a low lignin concentration, the frequency of soil disturbance was enhanced, the humification process decreased and erosion by wind and water facilitated. The introduction of maize (Zea mays) as a C4-plant into C3-based European agriculture has changed the 12C/13C isotope ratio in fields (Accoe et al., 2002). Many C4-grasses such as Digitaria ischaemum, D. sanguinalis, Echinochloa crus-galli, Setaria

verticillata and S. viridis are now colonizing maize fields as weeds. Urban environments, especially those in the cool-temperate zone, have higher air temperatures and lower air humidity. This enables the invasion of sub-tropical plant species. Some of these plants, such as Eragrosfe-species, have a C4metabolism. Therefore, the isotope composition of the humus will not only change in agricultural fields, but also in urban ecosystems as well. Due to the longevity of humic and fulvic acids these changes may be recognizable for some hundreds of years. Forestry and the selection of tree species are strongly steered by short-term demands. In Europe, coal mining created a demand for pit-props and stimulated the substitution of broad-leaved tree species for Scots pine. This has turned back climate-driven natural forest development and decreased the biodiversity of mycorrhizal communities in former broad-leaved forests. In addition, the planting of Austrian, Corsican and Scots pine has severely contributed to the desiccation of soils due to the high and year-round transpiration of these conifers, especially in winter times. The chemistry of their litter enhances podzolization on soils with a low carbonate content within 75 years after planting (Wilson, 1992; Nierop and Verstraten, 2003) and reinforces the effects of harsh climate and poor substrate quality (Ponge, 1990). In hilly and mountain areas the planting of Norwegian spruce has destroyed the understorey and the ecto mycorrhizal fungi of the broad-leaved woodlands and stimulated podzolization. The introduction of Douglas fir from North America to Europe was not accompanied by the anticipated high productivity in all European countries. One reason for this was the absence of the specific ecto mycorrhizal symbiont, Endogone lactiflua (Fassi et al., 1969) - which was not introduced to Europe and New Zealand at all - and that of Rhizopogon parksii, the other symbiont (Klinka et al., 2001) which was only introduced into German, French and British forests (Jiilich, 1984). In addition, the

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decomposition of the litter of Douglas fir does not contribute to improving of the soil quality because it only gives a hemimor or mormoder (Klinka et al., 2001). All in all, the emphasis on coniferous species is nowadays in strong decline for a number of reasons: a low demand of Scots pine for coal mining, the high sensitivity of spruce to ozone (forest die-back), and the low productivity of Douglas fir. The negative economic results of coniferous forestry have resulted in a change in forest management with broad-leaved forests replacing conifers in Western and Central Europe. However, this change has many consequences for forest soils. The input of terpenes has stopped and thus their negative effects on soil fungi (Hintikka, 1970). The diversity of ecto mycorrhizal fungi associated with coniferous trees will disappear. The concentration of tannins will increase in broad-leaved litter. As a result, the chemistry of the litter and thus the humus type of soils will change, as shown for humus forms in some chronosequences of Scots pine forest after beech had been planted (Figure 9).

Figure 9. Occurrence and frequency of humus forms within the chronosequence of a forest with permanent Scots pine for 84 years (Pine 84) and permanent beech for 91 years (Beech 91) compared to a change of 76 Scots pine to 34 years of beech (Pine 76/Beech 34) and 114 years of Scots pine to 57 years of beech (Pine 114/Beech 57). Modified after Fischer et al. (2002), with the permission of the authors

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Therefore, analysis of humus forms is an essential part of the evaluation of soils and is increasingly being applied in forestry (Wilson et al., 2001). Conifers and the introduction of exotic broad-leaved species such as Eucalyptus have severely deregulated biological processes in the soil. The litter of eucalypts is rich in those monoterpenes which cannot be broken down by micro-organisms outside the native eucalypt zone in Australia. Cyclic burning of this litter has stimulated soil erosion in (sub)Mediterranean areas, blocked soil development, and finally caused substantial economic losses due to fire damage in France, Spain and Portugal during the dry summer of 2003. The allelopathic compounds in eucalypt litter also diminish the species biodiversity of the ground layer. Substitution of one indigenous tree species by another can also modify the under-storey vegetation and soil processes. When a forest previously occupied by hardwood was substituted by softwood species, the result was a change from mull to moder (Arpin et al., 1986). The planting of poplar and willows in wetlands and other sites instead of alder has broken down the input of nitrogen to wooded wetland ecosystems, and enhanced the accumulation of zinc-enriched litter in the humus layer with negative effects on re-introduced animals (Ernst, 1994). Fields and meadows are managed by the harvesting of crops and the mowing of grassland. Removing the standing crop changes the shoot/root ratio and diminishes the input of organic substances into soils. The result is a permanent interruption of natural soil formation. Therefore, all anthrosols have a disturbed soil profile (Beyer et al., 1996). Frequent mowing diminishes the root biomass by more than 90% (Tainton, 1984). City lawns have, therefore, a badly developed soil profile and the monthly mowing of grasses makes them very susceptible to drought. Changing land use from agriculture, horticulture and pasture to forests - a general trend in Central Europe - has had inevitable consequences. It will probably take at least a century before the original state of soil profiles and the associated soil biota can be restored. Afforestation of former grasslands, gardens and agricultural fields in the Vosges Mountains have shown that when compared with ancient forests the previous land use system had stimulated the establishment of nitrogen-demanding species in the under-storey. The highest memory of land use was found in the upper soil layer of those soils formerly used as gardens (hortisols). They have high concentration of phosphorus and nitrogen and a high pH, even after they had been abandoned for 90 years (Koerner et al., 1997).

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3.4.4 Buried soils

A special aspect is the burying of soil profiles by aeolian, fluviatile, biological, and anthropogenic processes. In contrast to biological and partly anthropogenic burial, there are two specific features of aeolian and fluviatile burial. In one part of the region soil will be removed (erosion) and in another part it will be deposited (burial). When this is a natural process, it may be responsible for maintaining the ecosystems; when it is man-made, it will destroy ecosystems. Aeolian burial Aeolian dynamics in coastal ecosystems is a natural process that is necessary for the maintenance of the diversity of vegetation types in coastal zones (see Section 2.3.5). From 7500 years BC onwards the strong rise of the water table of the ocean and the North Sea, 2.1 m per 100 years, has pushed the coastal zone to the continent and changed the position of rivers (Freund and Streif, 1999). This transgression has buried soils in coastal areas under sediments and sand, forming new dunes above a former marshy environment where new soils and vegetation have developed (Wilson, 1992; Granja and Soares de Carvalho, 1992). Climatic change during the Little Ice Age has resulted in new aeolian activities burying soils near the coast (Granja and Soares de Carvalho, 1992) and in other inland sandy areas. The fixation of dunes by planting helm on the foredunes interrupts this natural process and has resulted in a loss of dynamics in coastal ecosystems. The construction of coastal barriers for harbours has modified the natural aeolian processes and contributed to an erosion of the coastline by wind and water. The restoration of this coastline by sand supply from far outside the coastal line has modified the aeolian processes and changed sand chemistry in such a way that less calcium-rich material is blown on the coast which accelerates acidification. Aeolian burial in semi-arid regions is due to overgrazing and ploughing in the dry season. Intensive trampling detaches the sod from the soil and gusty winds are then able to remove the soil from the land surfaces (wind erosion) stimulating desertification. In the Sudano-Sahelian zone and in the Kalahari sandstorms can deplete up to 80 ton sand ha"1 yr 1 from the agricultural fields and overgrazed pastures. The second part of this process is the deposition of similar amounts of sands in other parts of the regions (aeolic burial). It is one of the great constraints for crop production (Michels et al, 1995). Fluviatile burial The fluviatile burial of soil profiles is related to flood events in valleys. Each flood either removes part of the former soil profile or superposes a new

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sediment layer on the river plain or on the river-bank levees. The positive effects of burial was welcomed in Egypt for thousands of years when fertile sediments were deposited on the river banks of the Nile and guaranteed a good crop. Most fluvial burial, however, has negative environmental impacts. When rivers pass mining areas their sediments can be enriched with heavy metals. As soon as such sediments, during floods, are deposited on the river banks, they can increase the metal level to such concentrations that crops are toxified and a heavy metal vegetation is established. This is the case on the banks of rivers Oker and Innerste in the Harz, for example, and in the Geul river in Belgium and the Netherlands (Ernst, 1974; Hindel et al., 1996). After mining activities and cleaning up tailings or constructing dams in the upper part of the rivers have been completed, the metal loads in sediment are strongly diminished and the contamination of top soil is strongly decreased. The metal contamination of the deeper soil layers, however, is nevertheless conserved (Rang et al., 1986; Swennen et al., 1994). As a consequence shallow-rooting plant species of the heavy metal vegetation such as the Zinc violet (Viola lutea subsp. calaminaria) and Zinc pennycress (Thlaspi caerulescens subsp. calaminaria) starts to disappear. Instead of welcoming the decrease of metal contamination, nature conservation agencies are disappointed because a rare plant species is disappearing. Deeprooting (up to 2 m) grasses (Molinia caerulea) and herbs (Silene vulgaris) can penetrate to such metal-enriched layers and transfer metals to the above ground plant parts. When tailings ponds, containing a mixture of arsenic, copper, lead and zinc, fail to retain toxic waste and, as in the case of the Azualcollar mine in Southern Spain (April 1998), flood killing everything in their path, part of the poisonous spillage will be bound to the top soil (Davis et al., 2000). The burial of clean soil by these types of spillage will enhance metal uptake into the surviving vegetation and enrich litter and humus with heavy metals. For several centuries after such a disaster the health of plants on agricultural fields and marshland will be affected. In the case of the Azualcollar mine the herbivores in the adjoining national park Coto de Donana, declared a World Heritage Site in 1994 by the United Nations, will also be affected. Biological burial Burial of litter and the organic layer by hampered decomposition in a watersaturated environment results in the development of histosols, in soil taxonomy not ranked under buried soil profiles. On sites with stagnant fresh water and a high productivity amongst mosses and ericaceous species, the decomposition of organic material is slow or even absent resulting in the development of organic

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soils. Therefore, the sequence of post-glacial vegetation and soil development is best conserved in environments with water-logged soils, i.e. fens and bogs. The start of the pedogenesis in water-logged soils can be an eutrophic fen with Phragmites australis. With climatic changes in the Boreals, for example, and decreasing water levels, such a fen was overgrown by an oak-birch woodland indicating a change in plant-available nutrients that finally resulted in a more mesotrophic ecosystem with increasing soil acidity. During the Atlanticum the annual precipitation increased resulting in a rise of water levels. It destroyed the swamp woodland and permitted the establishment of an oligotrophic Scheuchzeria palustris fen, later followed by a vegetation complex made up of Sphagnum peat mosses on wet microsites, Eriophorum species on hummocks and a Betula pubescens woodland on the driest part. The contribution of the various plant species has resulted in different humus forms at the same site. On wetter soils, sedges and alder (Alnus glutinosa) expanded. The alder started to enrich the soils with nitrogen due to the high nitrogen-fixing activity of symbiotic actinomycetes (Table 2) in its roots and the high nitrogen concentration in its litter. This nitrogen-enriched environment stimulated the expansion of eutrophic under-storey species and the eutrophication of ecosystems in its neighbourhood. But the changes in precipitation in the Atlanticum has turned many of the alder brook woodlands into marshes with the conservation of organic matter building up the various soil layers (De Bakker and Edelman-Vlam, 1976). Nowadays many deep-rooting Carex species (Kutschera and Lichtenegger, 1982) can exploit this conserved nutrient pool up to a depth of 2 m. Burial by human activities Anthropogenic activities have added a new type of buried soil profiles. All the soils buried by human activities still require taxonomic treatment (Schachtschabel et al., 1992; FAO, 1998). Soils for growing tulips and other bulbous plant species are prepared by the removal of part of the clayey top soil and covering the ground with sand from the dunes to improve the micro climate and facilitate the harvest of the bulbs. These bulbs are, however, attractive to nematodes and, therefore, require heavy applications of pesticide. Another option is to flood these soils for a period of several months to get rid of the pests, but this blocks all decomposition processes and changes the nutrient status of the soil by nutrient leaching. The huge quantities of sand which are layered on marshy areas destined for housing construction can be compared to an aeolian overblowing of the soil profile. However, the difference is that such soils come under buildings and

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have no free excess to oxygen. The sandy soils in the rest of the settlement area are usually partially covered with the top soil that was removed before house construction began. When these soils are sown for amenity purposes, without fertilization, they support a very unproductive grassland with low biodiversity, especially when the mowing frequency is biweekly. If no care is taken, sites with a highly disturbed soil without any profile, allow the development of a vegetation with ruderal species, often with a high percentage of invasive species such as Buddleja davidii, Heracleum mantegazzianum, Impatiens parviflora, Polycarpon tetraphyllum, Polygonum sachaliense, Reynoutria japonica, and Solidago-

species (Child et al., 2003). Unfortunately, the soil development under these species has not yet been studied. However, it can be expected that the dense above ground vegetation cover in stands of Heracleum mantegazzianum, Reynoutria japonica, and Solidago altissima will conserve soil moisture, thus

stimulate the decomposition of the relatively nutrient-rich litter and allow the development of a mull humus. The low nitrogen concentration in the litter of Buddleja davidii, however, will tend to a moder humus. A completely different aspect of anthropogenic soil burial is the covering of traffic infrastructure with gravel (railways, highways) or sealing them with asphalt, tar and/or pavements. This sealing redirects the hydrology to such an extent that during droughts the vegetation, especially trees, suffer from water deficiency, and when de-icing salts are applied in the winter, they will experience the effects of sodium and chloride toxicity during early summer. Tree growth will diminish as a result but the establishment of salt resistant plants like Cochlearia danica and Puccinellia distans will be stimulated. On a

regional scale heavy rains cannot penetrate into the soil, but run-off and cause floods and economic damage, not only in the cities, but also in rural districts. These sealed soils create a warm micro-climate and give exotic plant species an appropriate environment. This can be seen in the expansion of the South African perennial herb Senecio inaequidens (Ernst, 1998b) and the Southern African grass Eragrostis pilosa in European cities and along highways and railways as far to the East as Poland. 3.5 Conclusion There is no ecosystem and no soil that has not been affected by human activities, either directly or indirectly whether the change has been in structure and/or the chemistry of the soil horizons or the soil profile as a whole. Up to now, interest in considering the soil in its full function, i.e. the organic layer as part of the soil profile and the entire soil profile, has been very limited in ecology and soil science and at a minimum in soil protection policy. The

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objective of the Soil Protection Act in 1988 in the Netherlands was to protect the soil and the groundwater from pollution and damage. However, surprisingly, protection of the soil profile has been considered only in specific soil protection areas in provincial plans, although it is a very essential aspect of soil health because plant species and their symbionts have access to different soil horizons. Therefore I propose the following steps in the evaluation of soil vitality (Figure 10). Firstly, define the land use in the past, present and future. Land use over the past 2000 years can have modified the geomorpholgy, hydrology and pedology and as a result the vegetation in such a way that the original vegetation has been substituted by semi-natural vegetation such as heather, chalk grassland and garrigues as indicated by earth and stone walls or/and surface reliefs in many (chalk)grasslands or by anthropogenic vegetation.

Figure 10. Assessment of the (expected) changes of vegetation and soil properties in relation to land use and the kind of disturbance by human activities

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Keeping an area in its natural state is also a kind of land use. Secondly, identify the realized or expected impacts on the physical, chemical and biological disturbance and the consequences for vegetation, soil organic matter, decomposition velocity, and the impact on the soil profile. Physical disturbance will primarily affect the soil structure, the soil profile and the hydrology. Analysis of the rooting depth and the rooting structure will help to evaluate the impact of soil compaction and the height of the water table. The emphasis on the upper organic layer does not sufficiently recognize the importance of the soil profile for root development and mycorrhization. Chemical disturbance can affect the loading of humus with pesticides and heavy metals. It will modify the activity of free-living and symbiotic bacteria and fungi and thus litter decomposition, the synthesis of fulvic and humic acids and finally the humus form. The longevity of humic and fulvic acids results in a memory of several centuries at least. Therefore, soil health has to be seen in a long-term perspective. Biological disturbances are manifold and include the introduction of new species, changing vegetation management via different intensities of mowing, grazing and harvesting, and the introduction of new herbivores and new pests. These changes will modify the amount and chemical components of litter, the quality and quantity of root exudates and the symbiosis with fungi and bacteria, and thus finally all processes related to litter decomposition. In an evaluation of soil vitality it is necessary to describe land-use related vegetation (inclusive biological disturbances) together with the rooting depth, the type and degree of mycorrhization and - for nitrogen-fixing species - the activity of the root nodules (with N-concentration as a substitute). In the case of changes of land use in the future an analysis should be made of the consequences for symbioses. Litter chemistry should be analysed in relation to the dominant litter-producing species and, in the case of forestry, this should also be done from a historic perspective. The types of organic layer (humus forms) and ratio of fulvic and humic acids should also be described. Changing the ecosystem on a site by human activities has great consequences not only for the biodiversity of plants, their association with mycorrhizal fungi (Ernst, 1994) and herbivores, but also for the litter decomposing fungi and micro-organisms. The prediction of ecosystem development and thus the expectation of soil vitality requires the development of a framework of long-term land use and the options for the development and maintenance of soil profiles. As a first approach Ellenberg' indicator values and the quality of the soil profile give an initial indication of soil quality. Ecotests with standardized, not

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ecosystem-relevant organisms (Van den Munckhof et al., 1998) will not give real information on soil quality. In every case, one has to be aware that evolution and selection of appropriate ecotypes is a permanent process in all biota, although only marginally investigated in microbial ecology and nearly not investigated at all in animals. In general, plants and micro-organisms, and to a lesser degree animals, have sufficient genetic potential to colonize each soil if water is available, temperature and radiation are in the range of biological processes, mobility of the substrate is low and the concentration of plant-available chemicals does not surpass the physiological regulation mechanisms and the decomposition of organic material. However, the biodiversity of the vegetation and fauna on soils in extreme environmental conditions such as extremely saline and heavy metalenriched environments may be low. Such soils will not support the growth of those plant species that are important for human and animal consumption, but the soil is nevertheless vital. Therefore, it is imperative that any evaluation of soil vitality should first define the scope of the activities. 3.6 Implementation in soil management Water, inorganic chemical elements and organic matter are the main environmental conditions for soil ecology. The direction of the development of the primary production and the soil biota is determined by source and sink. Source can be the seed inoculum and sink the structure of the soil environment. For practical purposes three levels of soil type, moisture regime, nutrients should be considered as predicting the development of the vegetation. This would result in "simplified-Ellenberg indicators", being useful monitoring items for soil management within spatial planning. The chemical composition on macro element and metals of the above ground vegetation and its analysis can provide information on the behaviour and availability in the underground. Especially in buried (or covered) soil areas this may reflect the position and capacity of the root and mycorrhiza system. Organic matter should be given more attention and should always be measured as the total organic carbon content of the soil. As soon as organic matter content increases its quality, such as tannine type, it should be measured as well as the ratio fulvic acid/humic acid. In view of the long-term biological memory of soils future use and potential negative effects can be evaluated if the history of the soil is known. That 'memory' contains information that may need translation in order to become integrated in soil management.

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Brethes, A., J.J. Brun, B. Jabiol, J. Ponge and F. Toutain, 1995. Classification of forest humus forms: a French proposal. Annal. Sci. Foretieres 52,535-546. Caldwell, M.M., T.E. Dawson and J.H. Richards, 1998. Hydraulic lift: consequences of water efflux for the roots of plants. Oecologia 113,151-161. Campbell, I.D. and J.H. McAndrews, 1993. Forest disequilibrium by rapid Little Ice Age cooling. Nature 366, 336-338. Canadell, J., R.B. Jackson, J.R. Ehleringer, H.A. Mooney, O.E. Sala and E.D. Schulze, 1996. Maximum rooting depth of vegetation types at the global scale. Oecologia 108, 583-595. Child, L., J.H. Brock, G. Brundu, K. Prach, P. Pysek, P.M. Wade and M. Williamson (eds.), 2003. Plant Invasions: Ecological Threats and Management Solutions. Backhuys Publishers, Leiden. Colpaert, J.V., P. Vandenkoornhuyse, K. Adriaensen and J. Vangronsveld, 2000. Genetic variation and heavy metal tolerance in the ectomycorrhizal basidiomycete Suillus luteus. New Phytologist 147, 367-379. Dabkowska-Naskret, H., 2003. The role of organic matter in association with zinc in selected arable soils from Kujawy Region, Poland. Org. Geochem. 34, 645-649. Davis, R.A. jr., A.T. Welty, J. Borrego, J.A. Morales, J.G. Pendova and J.G. Ryan, 2000. Rio Tinto estuary (Spain): 5000 years of pollution. Environ. Geol. 39,1107-1116. De Bakker, H. and A.W. Edelman-Vlam, 1976. De Nederlandse bodem in kleur. Pudoc, Wageningen. De Bakker, H. and J. Schelling, 1989. Systeem van bodemclassificatie voor Nederland: De hogere niveaus. 2nd edit., Pudoc, Wageningen. Degenhardt, J., P.B. Larsen, S.H. Howell and L.V. Kochian, 1998. Aluminium resistance in the Arabidopsis mutant alr-104 is caused by an aluminium-induced increase in rhizosphere pH. Plant Physiol. 117,19-27. De Groot, P., A. Field-Juma and D.O. Hall, 1992. Taking Root. Revegetation in Semi-Arid Kenya. ACTS Press, Nairobi. De Rore, H., E. Top, F. Houwen, M. Mergeay and W. Verstraete, 1994. Evolution of heavy metal resistant transconjugants in a soil environment with a concomitant selective pressure. FEMS Microbiol. Ecol. 14, 263-274. De Walle, F.B. and J. Sevenster, 1998. Agriculture and the Environment. Minerals, Manure, and Measures. Kluwer Academic Publisher, Dordrecht. Dijkstra, E.F., 1996. Development of humus profiles under Scots pine on driftsands. A comparative study of heavy metal polluted and non-polluted sites in Noord-Brabant (the Netherlands). Dissertation University of Amsterdam, Amsterdam, the Netherlands. Dueck, Th.A., W.H.O. Ernst, J. Faber and F. Pasman, 1984. Heavy metal emission and genetic constitution of plant populations in the vicinity of two metal emission sources. Angewandte Botanik 58,47-59. Dzwonko, Z. and S. Gawronski, 2002. Effect of litter removal on species richness and acidification of an oak-pine woodland. Biol. Conservation 106,389-398. Ellenberg, H., 1974. Zeigerwerte der Gefasspflanzen Mitteleuropas. Scripta Geobotanica 9, 1-97. Goltze Verlag Gottingen. Ellenberg, H., 1988. Vegetation Ecology of Central Europe. 4th edit., Cambrigde University Press, Cambridge. Ellenberg, H., R. Mayer and J. Schauermann (eds.), 1986. Okosystemforschung. Ergebnisse des Sollingprojekts 1966-1986. E. Ulmer Verlag, Stuttgart.

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Ellenberg, H., H.E. Weber, R. Dull, V. Wirth, W. Werner and D. Paulissen, 1991. Indicator values of plants in Central Europe (in German). E. Goltze, Gottingen. Epstein, E., 1999. Silicon. Ann. Rev. Plant Physiol. Plant Mol. Biol. 50, 641-664. Ernst, W., 1974. Schwermetallvegetation der Erde. G. Fischer Verlag, Stuttgart. Ernst, W., 1978. Mesophyten. In: Baumeister, W., Ernst, W., Mineralstoffe und Pflanzenwachstum. 3rd edit. G. Fischer Verlag, Stuttgart, p. 248-249. Ernst, W. and B.H. Walker, 1973. Studies on the hydrature of trees in miombo woodland in South Central Africa. J. Ecol. 61, 667-673. Ernst, W.H.O., 1984. Indicatoren van een overmaat aan zware metalen in terrestrische ecosystemen. In: Best, E.P.H., Haeck, J. (red.), Ecologische indicatoren voor de kwaliteitsbeoordeling van lucht, water en ecosystemen. Pudoc, Wageningen, pp. 109-120. Ernst, W.H.O., 1989. Mine vegetation in Europe. In: Shaw, A.J. (ed.), Heavy Metal Tolerance in Plants: Evolutionary Aspects. CRC Press, Boca Raton, pp. 21-37. Ernst, W.H.O., 1993. Ecological aspects of sulphur in higher plants: The impact of SO2 and the evolution of the biosynthesis of organic sulphur compounds on populations and ecosystems. In: De Kok, L.J., Stulen, I., Rennenberg, H., Brunold, C, Rauser, W.E., (eds.). Sulphur Nutrition and Assimilation in Higher Plants. Regulatory, Agricultural and Environmental Aspects. SPB Academic Publishing, The Hague, pp. 295-313. Ernst, W.H.O., 1994. Oeco(toxico)logische risico's van veranderend landgebruik. In: GleichmanVerheijen. E.C., Van Koten-Hertogs (red.), Veranderend Landgebruik en Natuurontwikkeling. Raad voor het Milieu- en Natuuronderzoek RMNO publicatie 101, 35-44. Ernst, W.H.O., 1998a. Ecotypic variation and environmental adaptation to air pollution and global change. In: De Kok, L.J., Stulen, L, (eds.), Responses of Plant Metabolism to Air Pollution and Global Change. Backhuys Publishers, Leiden, pp. 217-232. Ernst, W.H.O., 1998b. Invasion, dispersal and ecology of the South African neophyte Senecio inaequidens in the Netherlands: from wool alien to railway and road alien. Acta Bot. Neerlandica 47,131-151. Ernst, W.H.O., 2002. Seedbank in potentially moist dune valleys in the coastal dunes of the Amsterdam Waterworks Company and the chance of restoration of the moist dune valley "Klazeweitje" (in Dutch). Report Vrije Universiteit Amsterdam and Amsterdam Waterworks. Ernst, W.H.O., 2003. The use of higher plants as bioindicators. In: Markert, B.A., Breure, A.M., Zechmeister, H.G. (eds.) Bioindicators & Biomonitors. Principles, Concepts and Application. Elsevier, Oxford, pp. 423-463. Ernst, W.H.O., W. Mathys, J. Salaske and P. Janiesch, 1974. Aspects of heavy metal contamination in Westfalia (in German). Abhandlungen aus dem Landesmuseum fur Naturkunde zu Munster in Westfalen36(2), 1-31. Ernst, W.H.O. and H.J.M. Nelissen, 1979. Growth and mineral nutrition of plant species from clearings on different horizons of an iron-humus podzol profile. Oecologia 41,175-182. Ernst, W.H.O. and A.J. De Neeling, 1986. Response of an acidic and a calcareous population of Chamaenerion angustifolium (L.) Scop, to iron, manganese, and aluminium. Flora 178,85-92. Ernst, W.H.O., M.H.S. Kraak and L. Stoots, 1987. Growth and mineral nutrition of Scrophularia nodosa with various combinations of fulvic and humic acids. J. Plant Physiol. 127,171-175. Ernst, W.H.O., R.D. Vis and F. Piccoli, 1995. Silicon in developing nuts of the sedge Schoemis nigricans. J. Plant Physiol. 146, 481-488. Ernst, W.H.O., Q.L. Slings and H.J.M. Nelissen, 1996. Pedogenesis in coastal wet dune slacks after sod-

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Chapter 4

SOIL BIOTA AND ACTIVITY H. Verhoef

Abstract The biological activity of the soil is highest in the upper layer, predominantly the rooting layer, where energy is present. This upper layer is extremely biotic and is, therefore, called a living system. This living system is surprisingly diverse. The basis for this diversity is provided by this spatial and temporal heterogeneity. Mineralization of organic matter or recycling of elements is the collective responsibility of soil fauna and soil microflora. Soil fauna fragments organic structures and influences primary decomposers such as the soil microflora. They also play a role in the synthesis of organic matter and any kind of formation of soil material. Quantification of the role of single species or size groups is extremely difficult since many feeding interactions are still not well understood. In principle, this dynamic system of mineralization of organic matter is sustainable, because of the equilibrium and stability of the soil system. The three key functions related to soil quality are (i) dynamics and mineralization of organic matter, (ii) soil structure formation and maintenance, and (iii) support and control of plant production and species diversity. Although these functions can be carried out by alternative systems, and via alternative pathways, as routes in food webs, it is the author's opinion that some functions are irreplaceable. Therefore, there is concern that repeated stress will lead to impoverishment of the soil community. Disturbance by contaminants, such as heavy metals, can best be measured by fauna, looking at species diversity. In the majority of soils, micro-organism diversity is probably never reduced to such a level that it affects the functioning of the decomposer system. It is likely, that only those processes carried out by a few microbial species, such as nitrification and nitrogen fixation, have the potential to show clear responses. Defining the cause of disturbance, such as changes in land use, including desiccation, acidification or fragmentation, determines the direction of monitoring. Using the Dutch Soil Quality Indicator System may be the safest way to monitor soil quality. Indicators of soil quality should be physical, chemical, biological and visible. This selection should be based on land use, the relationship between soil function and the indicator, spatial and temporal patterns of variation and the importance of this variation, the sensitivity of the measurement to changes in soil management, and comparability with routine sampling and monitoring programmes.

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4.1 Introduction When surveying the vitality of soils it should be emphasized that we are dealing with living systems. An assessment of soil quality can only be performed if we focus on the active part of the soil in which biota and processes can be monitored and, if necessary, manipulated. In the following description of the soil, the active part of the soil forms a core that is surrounded by an abiotic environment: The soil is a living system in which plant (roots), micro-organisms and soil fauna interact with each other and with the abiotic environment.

The most active part of the soil, the upper ten centimetres, influences what is under ground. At the same time, what is under ground has contact with the upper parts of the ground through changing water levels and hydrological processes on a landscape scale which lead to the transportation and sorption of elements. An important character of the soil system appears to be soil heterogeneity, both spatially and temporally. Temporal heterogeneity means more than predictable seasonal changes. It involves changes in time due to physical transport, chemical and biological transition processes, the dispersal of organisms and also to soil management. Soil life activity can be visualized as shown in Figure 1. The right side of the diagram shows the ecologically important activities. These have been divided into the main soil processes, such as the mineralization of organic matter, including C-, N-, P-, S-cycling, and the synthesis of organic matter by the soil microflora and the metabolizing, transforming and transportation of soil material by the soil fauna. These soil processes are also indicated as the natural functions of the soil. On the left side of the diagram the so-called structural components of the soil - the soil organisms - together with their relative sizes and densities, are presented. These components run from bacteria (microedaphon) to earthworms (macro-edaphon). Based on these compositional structures and the natural functions of the soil, the dynamic quality of the top soil, has been described as an integral value of these two components, as far as soil use and environmental conditions on site are concerned (Filip, 2002). It seems difficult, however, to reach agreement among soil scientists as to how soil quality is to be interpreted and assessed (Sposito and Zabel, 2003). In my view, soil quality should be linked to the sustained functioning of soil organisms and soil processes and, therefore, include topics such as equilibrium and stability. Equilibrium implies the balance between the synthesis and breakdown of organic matter, the hindering of enrichment of soil organic matter or its depletion, and the balance between increase and decrease of species to conserve diversity and ecological functions. Stability on the other

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Figure 1. Soil organisms, their approximate counts and ecologically important activities (modified after Filip, 2002)

hand implies the recovery of soil communities and processes after disturbances. It should be noted here that we are not dealing with fixed soil quality indices (Karlen et al., 2003), but with dynamic values that are sensitive to variation in climate and human impacts (Sposito and Zabel, 2003). Soil quality is best assessed by properties that are neither so permanent as to be insensitive to management, nor so easily changed that they give little indication of long-term alterations (Gardi et al., 2002). This implies that the assessment of soil quality tracks, in time, the reactive behaviour of a soil to temporal imbalance, for example in organic matter in- and output, immigration and emigration of species, and to disturbances. Disturbances, such as contaminants, threaten equilibrium and stability and thus soil quality. In the section on disturbance indicators, attention will be paid to the adverse effects on soil quality of contaminants, together with land use change, and the role of soil biota as indicators. Ways to quantify/monitor the functioning of the soil system are proposed in the following procedure: i. decide which soil processes are important in determining soil quality; ii. detect which organisms play an important role in these selected soil processes; iii. monitor these organisms (alone or in communities), taking into account their population dynamics, both in time and space.

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4.2 Soil quality related processes Choosing which biological soil processes are related to soil quality is a difficult process, as it may depend on the use of the soil and the complexity of the soil system. A biological soil process, such as decomposition might be essential for a system that depends solely on the internal nutrient supply, whereas in other systems, with nutrient input from external sources, decomposition will be less important. Nevertheless, there is a relatively clear general view on the biological soil processes that are related to soil quality. Without prioritizing them the following processes can be distinguished: i. the dynamics and mineralization of organic matter; ii. soil structure formation and/or maintenance; iii. the support and control of plant production and species diversity. The rates of these processes are related to soil quality and account, in different combinations and with different priorities for the development of natural soil systems, agro-ecosystems, forest and range management, landfills, urban compaction and recreation. In my opinion, there is no reason to exclude city parks (see Karlen et al., 2003), road banks and Ecological Main Structures from this list. The methodology for quantifying the rates of processes such as decomposition, mineralization, primary productivity, diversity dynamics and soil structure formation, has been well-described and standardized (Alef and Nannipieri, 1995). In these processes, biota play an important role. Some groups can be indicated as direct drivers of the process. Often this involves a combination of groups sometimes extending as far as food web level. In the next section important groups for specific processes are indicated and some are elaborated. They refer to a range of soil systems such as forest, intensively managed agricultural and meadows/pastures and mainly come from microcosm studies. In this manner the role of soil biota in organic matter dynamics and mineralization has been quantified under standardized conditions, with simple organic substrates such as leaf mixtures, needle mixtures, humus material, sometimes with the addition of sugar as an extra carbon source. 4.3 Soil process drivers 4.3.2 Dynamics and mineralization of organic matter

Bacteria and fungi are the primary decomposers of the plant residues and organic compounds released from plant roots: N-, C-, P- and S-mineralization.

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However, a clear picture of who is doing what as far as these micro-organisms are concerned is not well understood and most bacteria have still to be discovered (Wardle, 2002). Soil fauna can influence (positive, neutral and negative) decomposition either indirectly by modifying biomass and the composition and activity of soil microbial communities, or directly by commuting organic matter (litter transformers), by consuming detritus and releasing inorganic nutrients. The effects of trophic groups of soil fauna on soil micro-organisms are summarized in Table 1 (Mikola et al., 2002). Microbial-feeding micro fauna, such as protists and nematodes, have a positive effect on C- and N-mineralization and on primary production. Enhanced C-mineralization is the result of an increased turnover rate, and the activity and respiration of grazed microbial populations (Anderson et al., 1981, Kuikman et al., 1990, Mikola and Setala, 1998b). An enhanced N-mineralization is mainly due to the direct animal excretion of excess N (Woods et al., 1982, Lussenhop, 1992). The nutrients released in turn enhance plant growth (Clarholm, 1985). The effects of meso- and macro fauna are also mostly positive, but are more heavily dependent on specific circumstances, such as the time of the year (Teuben, 1991), the nutrient content of microbial resources (Hanlon, 1981), the number of grazers (Hanlon and Anderson, 1979) and the stage of degradation of litter (van Wensem et al., 1993). Similarly, effects of enchytraeids may depend on their abundance and microbial growth conditions (Wolters, 1988). The percentage of total net N-mineralization attributed to soil fauna is between 25 and 37% in a variety of soils that range from polder, forest and grassland to arable soil. Protists, nematodes and enchytraeids are the important contributors, the process drivers. Included here are also the fragmentation and consumption of litter, detritus and microbial biomass and the alteration of microbial communities. Interpreting how soil fauna affects soil processes is complicated by the fact that faunal groups also prey on each other and thus can lead to a cascade effect throughout the food chain. A reduction in the abundances of microbial-feeding nematodes as a result of predation effects mineralization and primary productivity in either a positive (Allen-Morley and Coleman, 1989), neutral (Laakso and Setala, 1999a) or negative (Bouwman et al., 1994, Mikola and Setala, 1998a, Setala et al., 1999, Laakso and Setala, 1999a) way. The limited data available from other faunal groups also point to several possible outcomes. Predation on Collembola has been shown to increase fungal biomass (Verhoef et al., in press), to reduce litter mass loss (Lawrence and Wise, 2000) and to

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Table 1. Effects of trophic groups of soil fauna on soil microbes and ecosystem processes (modified after Mikola et al, 2002) Fauna group

Effects on microbes Microbial activity

Microbial biomass Micro fauna Protists Woods et al., 1982 Kuikman et al., 1990 Bonkowski et al., 2000 Nematodes Woods et al., 1982 Ingham et al., 1985 Mikola and Setala, 1998b Setala et al., 1999 Mesofauna Micro arthropods Hanlon and Anderson, 1979 Hedlund et al., 1991 Beare et al., 1992 Bardgett and Chan, 1999 Setala, 2000 Enchytraeids Williams and Griffiths, 1989 Sulkava et al., 1996 Cole et al., 2000

-

Effects on ecosystem processes C- and/or NPlant growth mineralization

+

0

+ +

-

+

-

+

+

+/-

-/o

+

+ +

0/+

+ +

+

+/+

-

+

+/0

+

0/+

0

+

0

0/+

+

Macro fauna Macro arthropods Gunnersson and Tunlid, 1986 Hassall et al., 1987 Teuben and Roelofsma, 1990 Van Wensem et al., 1993 Earthworms Haimi et al., 1992 Saetre, 1998 Wolter and Scheu, 1999 Alphei et al., 1996

+ +

+

+

+/-

+/-

-10

0/+ +

0/+ -

+

0

+

+ 0 Heemsbergen et al., in prep. -, 0 and + indicate that fauna has been found to reduce, have no effect or increase microbial biomass

and activity or ecosystem process rate.

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enhance C-mineralization (Hedlund and Ohrn, 2000), while predation on fungal-feeding and detritivorous mesofauna was found to have no effect on plant growth, despite significant reductions in prey populations (Laakso and Setala, 1999b). These examples illustrate how trophic groups and their interactions in food webs significantly influence soil functioning. Therefore, a food web approach is warranted, since studies and models enable us to quantify the contribution of the different trophic groups to various mineralization aspects. An example of a recent food web model analysis in the soil horizon of a Scots pine forest (Berg et al., 2001) will be elaborated. It estimates the contribution of the different soil biota to C- and N-mineralization rates. Fungi are the most important group for C-mineralization in the L- and F-layers of the soil. Together with bacteria they constitute approximately 75% of the total Cmineralization. Deeper in the organic H-layer, the bacteria account for 63%. The contribution of the soil fauna to C-mineralization ranges from 25% (L, F) to 15% (H). The enchytraeids account for approximately 30 to 50% of the contribution of the total faunal community. Exclusion and species identity An alternative way of evaluating the importance of specific groups for a specific soil process is the elimination of groups from the food web, the exclusion approach. In this way Faber and Verhoef (1991) demonstrated that specific collembola played a decisive role in N-mineralization, rather than the complete fauna. The importance of species identity over species number has frequently been demonstrated for different soil types (Mikola et al., 2002). For example in a microcosm study with different detritivorous species combinations, the presence of one particular earthworm species, Lumbricus rubellus, had more effect on C-mineralization than the number of species (Heemsbergen et al., in prep.). With a similar exclusion approach, but using a dynamic model, Hunt and Wall (2002) simulated the food web response to the removal of different trophic groups such as all bacterial-feeding nematodes. The model calculated changes in net primary production (NPP) by taking into account flows of C and N. Removal of individual trophic groups altered NPP by only a few %, because of the adaptation of the remaining biota. When all faunal groups had been removed NPP was reduced by 40%. This implies that food webs maintain their functions even when perturbed. However, there are indications that their structural stability can be reduced if community composition changes (Dunne et al., 2002), and that they are more susceptible (less stable) to subsequent disturbances (Griffiths et al., 2000), as visualised in Figure 2. Therefore it is

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Figure 2. Diagram summarizing the hypothesis that a change in food web composition may have no demonstrable impact on ecosystem function until the community is subjected to a second stress (modified after Hedlund et ah, in press)

concluded that, although soils seem rather robust in their reaction to disturbances, the composition of species has an important impact on mineralization processes. Another clear example of the importance of species identity comes from the study of the role of ammonia oxidizing bacteria in the N-transformations in the soil of acid coniferous forests (Laverman et al., 2001). Using molecular techniques to identify these bacteria, it was shown that in coniferous forest soils with high N-deposition, only one band was visible on the gel, probably originating from one - or at least only a few - bacterial species. This could mean that the process of nitrification in these soils depends on the presence and activity of a few bacterial species, making them true key species and perfect candidates for the monitoring of soil quality. A similar unique position has been identified for the saprophagous enchytraeid Cognettia sphagnetorum. This species has strongly beneficial effects on plant growth and nitrogen acquisition and shows a low functional replaceability (Laakso and Setala, 1999b). These are illustrations of the incorrectness of the redundancy hypothesis for soil systems as presented by Andren et al., 1995; Mikola and Setala, 1998. The main biotic controls of soil ecosystem functioning are most likely the key traits of the dominant species, the community composition of the component organisms, and the nature of biotic interactions among organisms (Wardle, 2002).

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4.3.2 Soil structure formation and/or maintenance

The maintenance and/or formation of soil structure is largely complementary to decomposition and mineralization of organic matter. Organic matter has important effects on soil structure both in colloidal form and as larger particles. Further, the energy released through decomposition is used by organisms for bioturbation (Figure 3). Ecosystem engineers (Jones et alv 1997) are organisms that directly or indirectly control the availability of resources to other organisms by causing physical state changes in biotic or abiotic materials. Soil ecosystem engineers are mainly found within the macro fauna group and include invertebrates with an average length of more than 2 mm, such as termites, earthworms and large arthropods. They dig and/or eat their way through the soil and create specific structures to accommodate their movements and living activities as they create, for example, burrows, galleries, nests and chambers, and produce casts and faecal pellets through their feeding activities. They provide a modified habitat for soil organisms that operate on smaller spatial scales. Earthworms stimulate microflora and soil fauna in the structures they create, by improving aeration and moisture conditions, by making resources more accessible by burying litter material, and by secreting urine and mucus. The mucus-urine mix produced by the earthworm Lumbricus terrestris attracts the collembolan Heteromurus nitidus in forest soils (Salmon, 2001). Earthworm structures can be associated with the presence of abundant nematodes, myriapodes and hymenoptera (Maraun et al., 1999; Decaens et al., 1999).

Figure 3. Relationships between soil structure and major soil processes (modified after Lavelle and Spain, 2001)

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Mounds created by ants provide constant and favourable microclimatic conditions, a supply of fresh organic material, and reduced predatory macro fauna (Laakso and Setala, 1998). Ant-earthworm mutualistic associations have been discovered there (Laakso and Setala, 1997). Ant nests can be considered to be potential "compost heaps", greatly enhancing the heterogeneity of the forest floor (Lenoir et al., 2001). 4.3.3 Support and control of plant production and species diversity

Soil organisms, intimately associated with plant roots, have the potential to induce important above ground effects. These include mutualists, such as mycorrhizal fungi and nutrient-transforming bacteria, and antagonists such as root pathogens and root herbivores. Soil fauna affects plant growth and tissue nutrient concentrations. Setala and Huhta (1991) reported soil fauna to increase leaf N-concentrations of Betula pendula seedlings by a factor of >3. This should lead to a greater quantity and quality in the litter returning to the soil, which may in turn benefit decomposer organisms. Other examples of soil communities affecting above ground communities include nitrogen fixing bacteria and nitrogen transforming bacteria, determining the relative abundance of nitrate, ammonium and dissolved organic nitrogen in the soil. As different plant species respond differently to nitrogen forms, nitrogen transforming bacteria influence the outcome of plant interactions and thereby plant composition. Root pathogenic and mycorrhizal fungi can play a key role in influencing plant community structure (Van der Heijden et al., 1998). Interactions involving root herbivores, such as phytophagous nematodes and elaterid larvae, have been shown to alter plant growth and plant succession (de Deijn et al., 2003). This opens up possibilities of overcoming bottlenecks in succession development and of speeding up community succession in nature development, for example, by influencing the outcome of competitive interactions between plants. The possible routes by which below ground organisms influence above ground organisms are presented in Figure 4. 4.4 Disturbance indicators Important soil process drivers are not necessarily organisms sensitive to specific disturbances. A key species for decomposition is not necessarily sensitive to desiccation or high copper levels, and is consequently not an appropriate disturbance indicator. Land use change, for example, may have a strong effect on soil community composition, without influencing soil process rates, because the key species for that specific soil process are insensitive to that land use

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Figure 4. Routes by which below ground organisms may influence above ground organisms (modified after Wardle, 2002)

change effect. However, in a series or combination of disturbances, the chance that the key species in an impoverished community will be the victim of another type of disturbance may increase. Disturbances caused by contamination and land use change will be discussed. 4.4.1 Contamination

In reviews on soil organisms as indicators of pollution, a transition from laboratory results to more realistic field data can be observed (Van Straalen and Van Gestel, 1992; Verhoef and Van Gestel, 1995). The soil fauna referred to are protozoa, nematodes, isopods, millipedes, oribatid mites, collembolans, enchytraeids, mainly where heavy metals are concerned, and earthworms in the case of heavy metals and pesticides. The role of bacteria as potential indicators of disturbances is questionable, taking into account that our knowledge about their role is limited to the approximately 1-5% of organisms that can be cultured. However, its application of molecular biological methods makes it possible to come to a closer understanding of the composition of microbial communities and their specific potential functions. The application of meta-genomics techniques, also in soil microbiology, can be used to identify the role of dominant, non-culturable organisms in the biodegradation of toxicants, the resistance to xenobiotica, or the suppression of soil-borne plant pathogens (Brussaard, 1997). Changes in soil processes due to stress-induced changes in composition of

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bacterial communities (Degens, 1998; Griffiths et al., 2000), should be considered with reservation. Changes in soil process rates were explicable in terms of fumigation effects, altering the microbial community composition, but not in terms of effects on microbial diversity. Griffiths et al. (2000) reported that the microbial community of fumigated soil was less resistant to a second stress, such as copper additions. Whether this was the result of reduced diversity or altered community composition remains unclear (Wardle, 2002). In the majority of soils, micro-organism diversity is probably never reduced to such a level that it could emerge as a plausible mechanism for affecting the functioning of the decomposer subsystem. It is more likely, that only those processes carried out by a few microbial species, like those involving specific nutrient transformations such as nitrification, and nitrogen fixation, have the potential to be responsive to shifts in diversity (Wardle, 1999). The examples in which a different community composition leads to different process rates, concern processes governed by species-poor communities such as denitrifier communities from two different sites, methane oxidisers from five different sites, and ammonium-oxidisers from nine different forest soils (Nugroho et al., in press). 4.4.2 Land use change

This includes the transition from natural to human-transformed sites and vice versa, eutrophication, acidification, desiccation, and fragmentation. Conversions between natural and human-transformed sites Natural and human-transformed systems often differ tremendously in terms of below ground properties. The conversion of natural forests to production systems has important effects on below ground organisms and processes (Wardle, 2002): - soil microbial biomass is usually much larger in forests than in cropping systems, and sometimes larger than in grassland systems; microbial feeding nematodes are often more abundant in arable and grassland systems than in forests, while forests contain much greater numbers of micro arthropods; soil macro fauna show a variety of responses, depending on the group of organisms present, but are frequently most numerous in grasslands. In general undisturbed systems are dominated by plants that have a longer life, which have traits associated with lower litter quality and higher secondary metabolite production (i.e. trees) and lead to domination by fungal-based food

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webs in which micro arthropods are important. By contrast, systems that are regularly disturbed and dominated by plant species with relatively high litter quality (i.e. many agricultural crop species) tend to promote the bacterial-based energy channel and bacterial-feeding micro fauna. The return of agricultural soil to a more natural state causes changes in soil acidity, in redox conditions and in the macro-chemistry of the soil (Salomons, 1993). Together with changes in soil structure (porosity, compaction, size distribution and stability of soil aggregates) this transition causes changes in the amount and composition of soil organic matter, consisting of microbial biomass and dissolved organic carbon (DOC). DOC is an important transporting system of heavy metals, nutrients (phosphate) and hydrophobic organic contaminants. In this way, land-use change and contamination are linked in nature development. In this reverse route from production to natural systems, the transition from bacterial-dominated to fungal-dominated degradation plays an important role, especially because fungi are relatively rare nowadays in Dutch agricultural soils. There are indications that nature development, due to low densities and activity in fungi (Siepel, 1991), results in the absence of certain groups of organisms that play an important role in soil structure formation. Here we refer to earthworms (Hoogerkamp et al, 1983), organisms that are related to plant composition and succession, such as mycorrhiza (van der Heijden et al., 1998) and phytophagous nematodes and elaterid larvae (de Deijn et al., 2003). It has been suggested in Section 4.3.3 that the introduction of these key groups may speed up nature development processes. Eutrophication The direct effect of increased atmospheric N-deposition on soil fauna is a decrease in the numbers and diversity of different soil organisms such as nematodes, collembolans, mites and locusts (Denneman and Torenbeek, 1987). In laboratory experiments it has been shown that the positive effect soil fauna, such as collembolans, has on nutrient (NH4, NO3, K, Ca, Mg) mobilization from pine litter disappears at N-deposition (Verhoef and Meintser, 1991). The decomposition of litter in the early stages of degradation (freshly fallen litter), is accelerated by N-deposition, whereas that of later degradation stages (F-layer) is hampered (Fog, 1988). This is probably caused by the effect of N-deposition on fungi, N-deposition having a known negative effect on ecto mycorrhizal and saprophytic fungi (Termorshuizen and Schaffers, 1991; Arnebrant et al., 1990; Berg and Verhoef, 1998). Negative effects on ligninase-producing fungi that influence decomposition during the later stages of degradation, in combination with the faster loss of carbon due to the increased initial breakdown of litter,

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will cause energy limitation in the later stages of decomposition. This may result in a decrease in long-term decomposition. N-deposition influenced not only growth, but also the N-content of fungi. Elevated fungal N-contents cause negative secondary changes (growth and reproduction) in the fungivorous Collembola (Hogervorst et al., 2003). These qualitative and quantitative reactions of fungi to elevated Nconcentrations have implications for the food web structure. In N-rich conditions, a switch has been found to the bacterial-based channel, which is reflected in the food web composition: The bacterial dominance has an effect on the abundance and composition of higher trophic levels, such as the nematodes (see, for example, Ettema and Bongers, 1993). Reduced levels of N-input show an increased diversity of soil micro arthropods, due to the decreasing dominance of certain (nitrophilous) species at equal species richness (Hogervorst et al., 1995). Acidification The effects of acidification on soil organisms are not always clear. Depending on the composition of acidophilic or alkalinophilic species, acid deposition will have a positive or a negative effect on the abundance of the soil community (Hagvar, 1984). In micro arthropods there are acidophilous, alkalophilous and acidindifferent species (van Straalen and Verhoef, 1997; Salmon and Ponge, 2001). In coniferous forests on acid sandy soils, many nematode species live at the edge of their tolerance limits, and ongoing acidification will have important effects on nematode composition (Bongers and Schouten, 1991). One important side effect of acidification is the leaching of cations out of the soil. This may cause a shortage in the (bio)availability of nutrients, as has been determined for the cryptostigmatid mite Platynothrus peltifer, probably because of Mn-shortage (Hogervorst et al., 1993). Heavy metals can also be mobilized as a result of acidification, leading to an increase in Pb and Cd intake in the earthworm Lumbricus rubellus. pH-decrease may lead to a higher availability of these metals in the soil. This will affect earthworms in particular as these animals absorb these metals from soil pore water. For soil arthropods this is less important as these animals take up toxicants mainly via the ingestion of organic matter (van Straalen and Bergema, 1991). Desiccation In many nature areas, groundwater levels have dropped causing drier soils. This has consequences for the water economy of the vegetation, and for the

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nutrient cycle in the soil, which is stimulated at relative desiccation, but restrained at strong desiccation. At lower groundwater levels the soil will be better aerated, causing an increase in the O2 availability of the reducers. In spring, soil temperatures increase faster. Both effects lead to an increase in decomposition rates and to the faster mineralization of organic matter. Often desiccation has a (indirect) side effect, called internal eutrophication. Drought, as a disturbance factor, has been introduced in microcosm experiments with forest soil and birch and/or pine seedlings (Liiri et al., 2001). In general, micro arthropod and enchytraeid populations decreased during drought, but recovered to their original levels after the drought had ended. Micro arthropods appeared to be more resistant to the immediate effects of drought than enchytraeids. The death of tree seedlings due to drought resulted in predictable decreases in microbial biomass and the numbers of nematodes. The recovery of the micro arthropod communities, nematodes and microorganisms, half a year after the drought, was independent of the species richness of micro arthropods in the microcosms, and did not support the insurance hypothesis that states that diverse systems are better able to resist disturbance and recover faster from it than species poor systems. Spiders, carabid beetles, earthworms and snails can be considered to be good indicators for changes in soil water conditions. Fragmentation

It has been suggested for above ground communities that higher trophic levels perceive and use resources over larger distances than lower trophic levels, and thus are more susceptible to habitat change and fragmentation (Holt, 1996). Applying this idea to soil communities (Hedlund et al., in press), enables the prediction of the possible susceptibility of different trophic levels to fragmentation. Organisms that use local resources on a micro scale, but have the ability of dispersing both in time and space, will have potentially efficient means of recolonizing disturbed areas, and may be more adaptable to change. As the size and range over which soil organisms use their resources vary in terms of magnitude, a discrimination has been made between organisms in the bacterial and in the fungal pathway. Bacteria act over pm-mm scales and their predators, the protozoa and nematodes, migrate up to centimetres, with bacterivorous nematodes moving slower than root feeding ones (Hunt et al., 2001). In contrast, the fungal pathway is characterized by larger scales because of strongly extending hyphal networks. Fungivorous micro arthropods can move on a cm-m scale, and their predators use even larger areas for foraging. Temporal dynamics on the spatial use and channel formation increase potential

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distribution of bacteria by several orders of magnitude, allowing an almost instantaneous response to resource influx and extending resource use up to the m scale. This also accounts for the recolonization capacities of bacterivorous nematodes (Ettema and Wardle, 2002). The fungal pathway cannot increase its potential distribution as much as the bacterial pathway. So, it can be concluded that fragmentation influences the soil community structure. 4.5 Monitoring biological soil quality 4.5.1 Soil quality indicators

Three recent examples of soil quality are presented to illustrate increasing complexity. The first Soil Quality Indicator system (Gardi et al., 2002) focuses on one soil group (Collembola or micro arthropods), based on one single sampling date, that compared grasslands (undisturbed) with arable fields (disturbed) in Northern Italy. The second Soil Quality Indicator system (Schouten et al., 2001) combines different combinations of soil groups related to relevant soil processes and is based on a series of sampling dates in successive years that compare agricultural areas in the Netherlands. The third Soil Quality Indicator system (Filip, 2002) focuses on micro-organisms related to relevant soil processes and is based on a series of sampling dates in two successive years which compare different soil types - clean and polluted - in five European countries. The relative sensitivity of the selected microbiological and process parameters are presented as sensitive indicators of soil quality on the bases of these data. The Biological Soil Quality index (BSQ) This soil quality index is based on the life-form approach and leads to the assembling of soil animals characterized by the same convergent morphological features (Parisi, 2001). Eco-morphological index (EMI) tables allow a score to be associated to each group, and the BSQ index can be calculated by adding the score of each group. According to the author, BSQ computation evaluates the adaptation of the soil animals to the edaphic environment, avoiding classification at species level. One BSQ is proposed based on micro arthropods (BSQ-ar), and a second based only on collembolan species (BSQ-c). In a recent application of this approach permanent grasslands, with low or no chemical inputs and high plant diversity, were compared with intensive agro-ecosystems, with high herbicide pressure and few plant species (Gardi et al., 2002). The soils were compared using standard soil quality indicators, such as organic carbon and aggregate stability (rather than process rates), and BSQ indices. The species-based BSQ-c values show a definite relationship with the

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increasing intensity of land use. It is claimed, however, that the BSQ-ar values, based on general taxonomic groups, also show the same trend, but are not very convincing as a tool for assessing biological soil quality. Identification to the species level and knowledge of their adaptive traits to specific soil characteristics seems indispensable for soil quality assessment. Furthermore, this knowledge should be coupled with soil process rates. The authors do not suggest that this method is appropriate for monitoring purposes. The Biological Indicator for Soil Quality (BiSQ) Since 1999, the so-called Biological Indicator for Soil Quality (BiSQ) has been used in the Netherlands for making an inventory of soil ecosystems, in combination with the Dutch Soil Quality Network (DSQN). This project ranges from concept development, through monitoring, to diagnostic instruments and prognostic response models. The indicator makes a cross-section through the soil ecosystem, analyzing biodiversity and functional groups. Up till now 175 locations in mainly agricultural areas have been sampled, with the emphasis being placed on differences in soil-type and land use. The aim is to come to aggregate complex field data in simple indicator values (quality indices). In Table 2, the soil processes - grouped in life support functions - are presented together with the selected soil organisms. This approach, of frequent sampling in time and space, also makes it possible to track the effects of the disturbances mentioned above. One may wonder if the soil quality indices based mainly on agricultural areas can be extrapolated to other types of soil, such as urban green, recreational areas, nature and Ecological Main Structures. Although it is preferable to monitor an increased number of soil types, general rules for the relationship between process rates and soil community structure can be deduced from the present data set, and can be applied to other soil types. There are, for example, the soil quality test kits - including mainly soil abiotic factors, with soil respiration as process parameter -, which proved to be good screening tools for both agricultural soil quality and conditions in Central Park in New York City (Karlen et al, 2003). Another example of a tool that has been developed for one type of soil use and can be made applicable to others, is the Detrital Food Web Model. This model, originally developed for short grass prairie ecosystems (Hunt et al., 1987), is being used to analyse the relationships between community structure and ecosystem processes, in agricultural soils (de Ruiter et al., 1993) and coniferous forests as well (Berg et al., 2001). This model has recently also been

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Table 2. Soil Quality Indicator System (modified after Schouten et ah, 2001) Life support functions Decomposition of organic matter

Nutrient recycling

Nutrient bioavailability (for plants) Soil structure formation Stability Soil Ecosystem

Indicators Worms + enchytraeids (1) Mites (2) >n Bacterial pathway (3) Organic matter transformatic Mushrooms (4) Microbial genetic diversity (5) Trophic interactions (6) Nitrogen mineralization =1+2+7+8+9+10 (in number and biomass) Subprocesses: - Microbial activity Micro-organisms (7) (bacteria and fungi) Protozoa (8) Nematodes (9) - Microbivory Springtails (10) Mites (2) Nematodes (+2 + 10) - Predation Mites (+9 +10) - Root herbivory N-, P- and H2O-uptake Mycorrhizal fungi (4)

Processes Fragmentation

Nitrification Bioturbation + aggregate formation Trophic interactions

Nitrifying bacteria (11) Worms + enchytraeids Community structure =1+2+7+8+9+10 (in number and biomass)

adapted to analyse the dynamic behaviour of populations and processes after a disturbance has occurred in the Ecosystem Stability Analysis project, that includes the effects of contamination on the functioning of soil organisms in terms of cycling of organic matter, energy and nutrients. An ecologically-based approach to assessing soil quality The goal of this international project is to find reliable indicators of soil quality for a broad range of soil use types (Filip, 2002). The different soils include crop fields, meadows, forests and urban soils in the Czech Republic, Hungary, Russia, Slovak Republic and Germany. The soil types are common in Europe and include calcic chernozem, dystric cambisol, luvic phaeozem, stagnic phaeozem and podzol. For each soil use type a clean site, as a control, and a

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polluted site were selected. Pollution consisted of organic pollutants, airborne pollutants, heavy metals, poly aromatics or mineral oil. It was the aim of the project to estimate whether specific soil characteristics are able to indicate anthropogenically-caused alterations in soil quality, when clean and polluted soils are compared. For a period of two years, 49 soil sites were sampled on a regular basis, 3 to 4 times a year. It was assumed that for the assessment of the total sustainability of the natural functions of the soil and different uses, key indicators should include biological and biochemical soil parameters (Figure 5). Micro-organisms were focused on because of their potential key role in ecologically important biogeochemical processes. In total more than twenty parameters were used to characterize microbial biomass, the composition of microbial communities and biochemical process-linked activities. Table 3 indicates the relative sensitivity of a selected group of parameters. Selectivity is measured by comparing the parameter values from the anthropogenically-affected soils with those of the control soils (undisturbed soils). It is concluded that N2-fixing bacteria, total microbial biomass, soil respiration, dehydrogenase activity and perhaps also the humification activity of soil micro-

organisms can serve as sensitive indicators of soil quality. This approach to investigating soils is to be preferred to toxicity tests as suggested by Rombke and Moltmann (1996) and many others. These tests may

Figure 5. Selected biological and abiotic parameters for the monitoring of soil quality (modified after Filip, 2002)

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Table 3. Relative sensitivity of the selected microbiological and biochemical parameters for the assessment of soil quality based on long-term soil analyses from 49 different anthropogenically-ajfected European soils (modified after Filip, 2002) Parameter Microbial biomass Composition of microbial communities Copiotrophic bacteria (colony forming units) Oligotrophic bacteria Actinomycetes Microscopic fungi Proteolytic spore forming bacteria Cellulose decomposer N2-fixing bacteria Pseudomonads Biochemical process-linked activities Respiration (CO2 release) Ammonification (NH4+ release) Nitrification/denitrification Dehydrogenase activity Humification activity * Sensitivity (relative to a control soil): - not detected + low ++++ maximum

Relative sensitivity*

+/++ ++ ++ ++ -/+ +/++ ++++ -/+ +++ +++ ++/+++ +++/++++ ++

be simple to use, and easily standardised, but they oversimplify both biological and abiotic structural complexity and the heterogeneity of the soil and tend to disregard soil ecological functions. 4.5.2 Standardization of soil quality indicators

Due to the wide range of uses to which soils are put and probably also due to the complexity of the soil system itself, it is difficult to identify general indicators for soil quality (Nortcliff, 2002). It has, therefore, been suggested by Sojka and Upchurch (1999) "that the search for a single, affordable, workable soil quality index is unattainable". Nortcliff (2002) has suggested grouping together attributes that might be used as indicators of soil quality: - physical attributes; - chemical attributes; - biological attributes; - visible attributes; soil organic matter as an attribute.

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The selection of the indicator attributes should then be based on: land use; relationship between soil function and the indicator; spatial and temporal patterns of variation and the importance of this variation; sensitivity of the measurement to changes in soil management; comparability with routine sampling and monitoring programmes. These criteria reflect my own personal selection from the list presented by Nortcliff (2002). The importance of spatial and temporal patterns of variation is closely linked to the chosen sampling strategy. Many soil properties, including microbial soil properties, are highly variable spatially over distances of a few square micrometers, even in relatively homogeneous landscapes. These problems, however, can be solved by adapting the sampling scheme (Laverman et al., 2002) or even turned to advantage by linking the soil community composition with hot spots of high process activity (Ettema et al., 1998; Verhoef et al., in press). Standardization of the evaluation of soil quality is necessary to allow comparison within and between countries. The International Standardization Organization (ISO) has developed a programme for the development of standard methods of soil analysis across a range of soil properties. It is my belief, however, that selection of soil indicators should be based on a thorough knowledge of the primary function of the specific soil, followed by a monitoring procedure of the related process rates and responsible organisms. 4.6 Implementation in soil management The recent observation that soil organisms also affect plant succession, leading to diversity, is a good reason to be very careful with practical suggestions for applications in soil management. It is my personal fear that a given set of recommendations might be used in an enthusiastic (result driven) way without too much reflection on the implications and meaning of the data. Therefore the soil quality indicator set, as published by Schouten et al. (2001) may be the safest one to handle. References Alef, Kv and P. Nannipieri, 1995. Methods in Applied Soil Microbiology and Biochemistry. Academic Press, Harcourt Brace & Company, Publishers, London: pp 576. Allen-Morley, C.R., and D.C. Coleman, 1989. Resilience of soil biota in various food webs to freezing perturbations. Ecology 70: 1127-1141.

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Alphei, J., M. Bonkowski and S. Scheu, 1996. Protozoa, Nematoda and Lumbricidae in the rhizosphere of Hordelymus europaeus (Poaceae): faunal interactions, response of micro-organisms and effects on plant growth. Oecologia 106:111-126. Anderson, R.V., D.C. Coleman and C.V. Cole, 1981. Effects of saprotrophic grazing on net mineralization. Terrestrial nitrogen cycles. Ecol. Bull. 33: 201-216. Andren, O., J. Bengtsson and M. Clarholm, 1995. Biodiversity and species redundancy among litter decomposers., p. 141-151., In: H. P. Collins, et al., eds. The significance and regulation of soil biodiversity. Kluwer, Dordrecht. Arnebrant, K., E. Baath and B. Soderstrom, 1990. Changes in microfungal community structure after fertilization of Scots pine forest soil with ammonium nitrate or urea. Soil Biol. Biochem. 22: 309-312. Bardgett, R.D. and K.F. Chan, 1999. Experimental evidence that soil fauna enhance nutrient mineralization and plant nutrient uptake in montane grassland ecosystems. Soil Biol. Biochem. 31:1007-1014. Beare, M.H., R.W. Parmelee, P.F. Hendrix, W. Cheng, D.C. Coleman and D.A.J. Crossley, 1992. Microbial and faunal interactions and effects on litter nitrogen and decomposition in agroecosystems. Ecol. Monographs 62: 569-591. Berg, M.P., and H.A. Verhoef, 1998. Ecological characteristics of a nitrogen-saturated coniferous forest in the Netherlands. Biol. Fertil. Soils 26: 258-267. Berg, M.P., P.C. de Ruiter, W.A.M. Didden, M.P.M. Janssen, A.J. Schouten and H.A. Verhoef, 2001. Community food web, decomposition and nitrogen mineralization in a stratified Scots pine forest soil. Oikos 94:130-142. Bongers, T., and A.J. Schouten, 1991. Nematodengemeenschappen als potentieel diagnostisch instrument voor chemische verontreinigingen., p. 175-186., In: G.P. Hekstra and F.J.M. van Linden, eds. Flora en fauna chemisch onder druk. Pudoc, Wageningen. Bonkowski, M., B.S. Griffiths and C. Scrimgeour, 2000. Substrate heterogeneity and micro fauna in soil organic 'hotspots' as determinants of nitrogen capture and growth of ryegrass. Appl. Soil Ecol. 14: 37-53. Bouwman, L.A., 1994. Short-term and long-term effects of bacterivorous nematodes and nematophagous fungi on carbon and nitrogen mineralization in microcosms. Biol. Fertil. Soils 17: 249-256. Brussaard, L., 1997. Biodiversity and ecosystem functioning in soil. Ambio 26: 563-570. Clarholm, M., 1985. Possible roles for roots, bacteria, protozoa and fungi in supplying nitrogen to plants., p. 355-365. In: A. H. Fitter, et al., eds. Ecological Interactions in Soil: Plants, Microbes and Animals. Blackwell, Oxford. Cole, L., R.D. Bardgett and P. Ineson, 2000. Enchytraeid worms (Oligochaeta) enhance mineralization of carbon in organic upland soils. Eur. J. Soil Sci. 51:185-192. de Deijn, G.B., C.E. Raaijmakers, H.R. Zoomer, M.P. Berg, P.C. de Ruiter, H.A. Verhoef, T.M. Bezemer and W.H. van der Putten, 2003. Soil invertebrate fauna enhances grassland succession and diversity. Nature 422: 711-713. de Ruiter, P.C, J.A. van Veen, J.C. Moore and L. Brussaard, 1993. Calculation of nitrogen mineralization in soil food webs. Plant Soil 157: 263-273. Decaens, T., P. Lavelle and J.J. Jimenez Jean, 1999. Soil surface macrofaunal communities associated with earthworm casts in grasslands of the eastern plains of Colombia. Appl. Soil Ecol. 13:

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87-100. Degens, B., 1998. Decreases in microbial functional diversity do not result in corresponding changes in decomposition under different moisture regimes. Soil Biol. Biochem. 30:1989-2000. Denneman, W.D. and R. Torenbeek, 1987. Nitraatimmissie en Nederlandse ecosystemen: een globale risico-analyse. RIN-rapport 87/23. Rijksinstituut voor Natuurbeheer, Arnhem. Dunne, J.A., R.J. Williams and N.D. Martinez, 2002. Network structure and biodiversity loss in food webs: robustness increases with connectance. Ecol. Letters 5: 558-567. Ettema, C.H., and T. Bongers, 1993. Characterization of nematode colonization and succession in disturbed soil using the Maturity Index. Biol. Fertil. Soils 16: 79-85. Ettema, C.H. and D.A. Wardle, 2002. Spatial soil ecology. Trends Ecol. Evol. 17:177-183. Faber, J.H. and H.A. Verhoef, 1991. Functional differences between closely related soil arthropods with respect to decomposition and nitrogen mobilization in a pine forest. Soil Biol. Biochem. 23: 15-23. Filip, Z., 2002. International approach to assessing soil quality by ecologically-related biological parameters. Agric, Ecosyst. Environ. 88:169-174. Fog, K., 1988. The effect of added nitrogen on the rate of decomposition of organic matter. Biol. Rev. 63: 433-462. Gardi, C, M. Tomaselli, V. Parisi, A. Petraglia and C. Santini, 2002. Soil quality indicators and biodiversity in northern Italian permanent grasslands. Eur. J. Soil Biol. 38:103-110. Griffiths, B.S., K. Ritz, R.D. Bardgett, R. Cook, S. Christensen, F. Ekelund, S.J. Sorensen, E. Baath, J. Bloem, P.C. de Ruiter, J. Dolfing and B. Nicolardot, 2000. Ecosystem response of pasture soil communities to fumigation-induced microbial diversity reductions: an examination of the biodiversity-ecosystem function relationship. Oikos 90: 279-294. Gunnarsson, T. and A. Tunlid, 1986. Recycling of fecal pellets in isopods: micro-organisms and nitrogen compounds as potential food for Oniscus aselhis L. Soil Biol. Biochem. 18: 595-600. Hagvar, S., 1984. Effects of liming and artificial acid rain on Collembola and Protura in a coniferous forest. Pedobiologia 27: 341-354. Haimi, J., V. Huhta and M. Boucelham, 1992. Growth increase of birch seedlings under the influence of earthworms. A laboratory study. Soil Biol. Biochem. 24:1525-1528. Hanlon, R.D.G., 1981. Influence of grazing by Collembola on the activity of senescent fungal colonies grown on media of different nutrient concentration. Oikos 36: 362-367. Hanlon, R.D.G. and J.M. Anderson, 1979. The effects of Collembola grazing on microbial activity in decomposing leaf litter. Oecologia 38: 93-99. Hassall, M., J.G. Turner and M.R.W. Rands, 1987. Effects of terrestrial isopods on the decomposition of woodland leaf litter. Oecologia 72: 597-604. Hedlund, K. and M.S. Ohrn, 2000. Tritrophic interactions in a soil community enhance decomposition rates. Oikos 88: 585-591. Hedlund, K., L. Boddy and CM. Preston, 1991. Mycelial responses of the soil fungus, Morteriella isabellina, to grazing by Onychiurus armatus (Collembola). Soil Biol. Biochem. 23: 361-366. Hedlund, K., B.S. Griffiths, S. Christensen, S. Scheu, H. Setala, T. Tscharntke and H.A. Verhoef, in press. Soil food web interactions in changing landscapes. Ecology Letters. Heemsbergen, D.A., M.P. Berg, J.R. van Hal, J.H. Faber and H.A. Verhoef, in prep. Complementarity in ecological functioning in soil detritivores as related to interspecific ecological distance.

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Hogervorst, R.F., H.A. Verhoef and N.M. van Straalen, 1993. Five-year trends in soil arthropod densities in pine forests with various levels of vitality. Biol. Fertil. Soils 15:189-195. Hogervorst, R.F., H.R. Zoomer and H.A. Verhoef, 1995. Effects of reduced nitrogen deposition on litter decomposition and soil fauna diversity. Ecosystems Research Report No. 20. ECSC-ECEAEC, Brussels. Hogervorst, R.F., M.A.J. Dijkhuis, M.A. van der Schaar, M.P. Berg and H.A. Verhoef, 2003. Indications for the tracking of elevated nitrogen levels through the fungal route in a soil food web. Environ. Pollut. 126: 257-266. Holt, R.D., 1996. Food webs in space: an island biogeographic perspective., p. 313-323., In: G.A. Polis and K.O. Winemuller, eds. Food webs - Integration of patterns and dynamics. Chapman and Hall. Hoogerkamp, M., H. Rogaar and H.J.P. Eijsackers, 1983. Effect of earthworms on grassland on recently reclaimed polder soils in the Netherlands., p. 85-105., In: J. E. Satchell, ed. Earthworm Ecology, from Darwin to Vermiculture. Chapman and Hall, London. Hunt, H.W. and D.H. Wall, 2002. Modelling the effects of soil biodiversity on ecosystem functioning. Global Change Biol. 8: 33-50. Hunt, H.W., D.H. Wall, N.M. DeCrappeo and J.S. Brenner, 2001. A model for nematode locomotion in soil. Nematology 3: 705-716. Hunt, H.W., D.C. Coleman, E.R. Ingham, R.E. Ingham, E.T. Elliott, J.C. Moore, S.L. Rose, C.P.P. Reid and C.R. Morley, 1987. The detrital food web in a short grass prairie. Biol. Fertil. Soils 3: 57-68. Ingham, R.E., J.A. Trofymow, E.R. Ingham and D.C. Coleman, 1985. Interactions of bacteria, fungi, and their nematode grazers: effects on nutrient cycling and plant growth. Ecol. Monographs 55: 119-140. Jones, C.G., J.H. Lawton and M. Shachak, 1997. Positive and negative effects of organisms as physical ecosystem engineers. Ecology 78:1946-1957. Karlen, D.L., C.A. Ditzler and S.S. Andrews, 2003. Soil Quality: why and how? Geoderma 114: 145-156. Kuikman, P.J., A.G. Jansen, J.A. van Veen and A.J.B. Zender, 1990. Protozoan predation and the turnover of soil organic carbon and nitrogen in the presence of plants. Biol. Fertil. Soils 10: 22-28. Laakso, J. and H. Setala', 1997. Nest mounds of red wood ants (Formica aquilonia): hot spots for litter dwelling earthworms. Oecologia 111: 565-569. Laakso, J. and H. Setala, 1998. Composition and trophic structure of detrital food web in ant nest mounds of Formica aquilonia and in the surrounding forest soil. Oikos 81: 266-278. Laakso, J. and H. Setala, 1999a. Population- and ecosystem-effects of predation on microbial-feeding nematodes. Oecologia 120: 279-286. Laakso, J. and H. Setala, 1999b. Sensitivity of primary production to changes in the architecture of belowground food webs. Oikos 87: 57-64. Lavelle, P. and A.V. Spain, 2001. Soil Ecology. Kluwer, pp 654. Laverman, A.M., A.G.C.L. Speksnijder, M. Braster, G.A. Kowalchuk, H.A. Verhoef and H.W. van Verseveld, 2001. Spatiotemporal stability of an ammonia-oxidizing community in a nitrogensaturated forest soil. Microbial Ecol. 42: 35-45. Lawrence, K.L. and D.H. Wise, 2000. Spider predation on forest floor Collembola and evidence for indirect effects on decomposition. Pedobiologia 44: 33-39.

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Lenoir, L., T. Persson and J. Bengtsson, 2001. Wood ant nests as potential hot spots for carbon and nitrogen mineralization. Biol. Fertil. Soils 34: 235-240. Liiri, M., H. Setala, J. Haimi, T. Pennanen and H. Fritze, 2002. Soil processes are not influenced by the functional complexity of soil decomposer food webs under disturbance. Soil Biol. Biochem. 34:1009-1020. Lussenhop, J., 1992. Mechanisms of Micro arthropod-Microbial Interactions in Soil, p. 1-33., In: M. Begon and A. H. Fitter, eds. Advances in Ecological Research, Vol. 23. Academic Press, London. Maraun, M., J. Alphei, M. Bonkowski, R. Buryn, S. Migge, M. Peter, M. Schaefer and S. Scheu, 1999. Middens of the earthworm Liimbricus terrestris (Lumbricidae): Microhabitats for micro- and mesofauna in forest soil. Pedobiologia 43: 276-287. Mikola, J. and H. Setala, 1998a. Relating species diversity to ecosystem functioning: mechanistic backgrounds and experimental approach with a decomposer foodweb. Oikos 83:180-194. Mikola, J. and H. Setala, 1998b. Productivity and trophic-level biomasses in a microbial-based soil food web. Oikos 82:158-168. Mikola, J., R.D. Bardgett and K. Hedlund, 2002. Biodiversity, ecosystem functioning and soil decomposer food webs., p. 169-180. In: M. Loreau, et al., eds. Biodiversity and Ecosystem Functioning; Synthesis and Perspectives. Oxford University Press. Nortcliff, S., 2002. Standardization of soil quality attributes. Agric, Ecosyst. Environ. 88:161-168. Nugroho, R.A., H.R. Zoomer, A.M. Laverman and H.A. Verhoef, in press. Environmental factors influence the ammonia-oxidising bacteria diversity and N-transformations. Soil Biol. Biochem. Parisi, V., 2001. La qualita biologica del suolo. Un metodo basato sui microartropodi. Acta Naturalia de 1' Ateno Parmense 37: 97-106. Rombke, J. and J.F. Moltmann, 1996. Applied Ecotoxicology CRC Press, Boca Raton. Saetre, P., 1998. Decomposition, microbial community structure, and earthworm effects along a birch-spruce soil gradient. Ecology 79: 834-846. Salmon, S., 2001. Earthworm excreta (mucus and urine) affect the distribution of springtails in forest soils. Biol. Fertil. Soils 34: 304-310. Salmon, S. and J.-F. Ponge, 2001. Earthworm excreta attract soil springtails: laboratory experiments on Heteromurus nitidus (Collembola: Entomobryidae). Soil Biol. Biochem. 33:1959-1969. Salomons, W., 1993. Lange-termijngedrag van stoffen in de bodem. Bodem 3:196-202. Schouten, A.J., M. Rutgers and A.M. Breure, 2001. BoBI op weg; tussentijdse evaluatie van het project Bodembiologische Indicator. 607604002/2001. Rijksinstituut voor Volksgezondheid en Milieu. Setala, H., 2000. Reciprocal interactions between Scots pine and soil food web structure in the presence and absence of ectomycorrhiza. Oecologia 125:109-118. Setala, H. and V. Huhta, 1991. Soil fauna increase Betula pendula growth: laboratory experiments with coniferous forest floor. Ecology 72: 665-671. Setala, H., P. Kulmala, J. Mikola and A.M. Markkola, 1999. Influence of ectomycorrhiza on the structure of detrital food webs in pine rhizosphere. Oikos 87: 113-122. Siepel, H., 1991. Recovering of natural processes in abandoned agricultural areas: decomposition of organic matter., p. 374-379. Proceedings 4th ECE/XIII. SIEEC, Godollo. (ed.) 1991. Sojka, R.E. and D.R. Upchurch, 1999. Reservations regarding the soil quality concept. Soil Sc. Soc. Am. J. 63:1039-1054.

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Sposito, G.Z. and A. Zabel, 2003. The assessment of soil quality. Geoderma 114:143-144. Sulkava, P., V. Huhta and J. Laakso, 1996. Impact of soil fauna structure on decomposition and Nmineralization in relation to temperature and moisture in forest soil. Pedobiologia 40: 505-513. Termorshuizen, A.J. and A.P. Schaffers, 1991. The decline of carpophores of ectomycorrhizal fungi in stands of Finns sylvestris L. in the Netherlands: possible causes. Nova Hedwigia 53: 267-289. Teuben, A., 1991. Nitrogen availability and interactions between soil arthropods and microorganisms during decomposition of coniferous litter: a field study. Biol. Fertil. Soils 10: 256-266. Teuben, A. and T.A.P.J. Roelofsma, 1990. Dynamic interactions between functional groups of soil arthropods and micro-organisms during decomposition of coniferous litter: a mesocosm study. Biol. Fertil. Soils 9:145-151. Van der Heijden, M.G.A., J.N. Klironomos, M. Ursic, P. Moutoglis, R. Streitwolf-Engel, T. Boiler, A. Wiemken and I.R. Sanders, 1998. Mycorrhizal fungal diversity determines plant biodiversity, ecosystem variability and productivity. Nature 396: 69-72. Van Straalen, N.M. and W.F. Bergema, 1991. Biologische beschikbaarheid en ecologisch risico van milieugevaarlijke stoffen., p. 27-37., In G.P. Hekstra and F.J.M. van Linden, eds. Flora en fauna chemisch onder druk. Pudoc, Wageningen. Van Straalen, N.M. and C.A.M. van Gestel, 1992. Soil. In: P. Calow, ed. Handbook of Ecotoxicology. Blackwell Scientific, Oxford. Van Straalen, N.M. and H.A. Verhoef, 1997. The development of a bioindicator system for soil acidity based on arthropod pH preferences. J. Appl. Ecol. 34: 217-232. Van Wensem, J., H.A. Verhoef and N.M. van Straalen, 1993. Litter degradation stage as a prime factor for isopod interaction with mineralization processes. Soil Biol. Biochem. 25:1175-1183. Verhoef, H.A. and S. Meintser, 1991. The role of soil arthropods in nutrient flow and the impact of atmospheric deposition., p. 497-506. In: G. K. Veeresh, ed. Advances in Management and Conservation of Soil Fauna. Vedam Books, New Delhi. Verhoef, H.A. and C.A.M. van Gestel, 1995. Methods to assess the effects of chemicals on soils., p. 223-257. In: R.A. Linthurst, et al., eds. Methods to assess the effects of chemicals on ecosystems. John Wiley & Sons, Chichester. Verhoef, H.A., M. Wortel and H.R. Zoomer, in press. Consequences of higher trophic level manipulation: cascading effects in a soil food web. Pedobiologia. Wardle, D.A., 1999. Biodiversity, ecosystems and interactions that transcend the interface. Trends Ecol. Evol. 14:125-127. Wardle, D.A., 2002. Communities and Ecosystems. Linking the aboveground and belowground components. Princeton University Press, Princeton. Wardle, D.A., H.A. Verhoef and M. Clarholm, 1998. Trophic relationships in the soil micro food web: predicting the responses to a changing global environment. Global Change Biol. 4: 713-727. Williams, B.L. and B.S. Griffiths, 1989. Enhanced nutrient mineralization and leaching from decomposing Sitka spruce litter by enchytraeid worms. Soil Biol. Biochem. 21: 183-188. 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., 1988. Effects of Mesenchytraeus glandulosus (Oligochaeta, Enchytraeidae) on decomposition processes. Pedobiologia 32: 387-398.

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Chapter 5

THE KEY ROLE OF SOIL MICROBES W. Verstraete and B. Mertens

Abstract

Soil is the "beholder of life" and determines the cycling of elements, the quality of groundwater, the presence of volatile compounds in the atmosphere as well as plant and animal health. Nearly every chemical transformation in soil the involves the active contribution of micro-organisms. Traditional physiological/biochemical as well as recently developed molecular monitoring methods for the most important soil ecosystem parameters are discussed, including their value as soil quality parameter. The transformation processes are related to environmental conditions. On a micro scale there are innumerable different habitats. This heterogeneity is the basis for a huge biodiversity of species and functions. The infallibility principle of nature, that states that for every natural organic component there is, in principle, a degradation pathway, should not be an excuse for leaving organic contaminants unmonitored. Too many uncertainties about their fate remain, therefore control activities such as measuring rates of degradation are important because of the occurrence of harmful intermediates. Soil quality must be founded on its capacity to support current forms of life and be a site of evolution for genomic new life. Out of the many soil functioning monitoring possibilities a selection of microbial functions and species diversity has been made. General functional aspects are soil respiration, nitrification, nitrogen fixation and bacterial DNA synthesis. General diversity aspects include ratio fungi/bacteria, mycorrhiza, suppressiveness to pathogens, and catabolic genes. Functions based on narrow diversity, such as nitrification and nitrogen fixation are most valuable in relation to monitoring adverse influences. The impact of heavy metals, persistent organic pollutants and eutrophication can be measured in many ways, but specially with respectively biosensors/biomarkers, catabolic genes and the ratio oligotrophs/copiotrophs. The most serious threat to soil is the fact that both politicians and the public are unaware of the importance of soil vitality. The setting up of environmental specimenbanks and 'soil nurseries' is recommended. The "Pareto-approach" is proposed for monitoring soil vitality. It is a new strategy to study effective associations: 80% of the tasks require approximately 20% of the overall diversity of the species to be involved. Therefore, the hypothesis is that 80% of a function such as respiration is carried out by 20% of the potential co-workers in a community. Monitoring actual functional diversity in relation to potential functional diversity is a good evaluation system.

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5.1 The importance of soil microbiology 5.2.2 Introduction

Almost every chemical transformation taking place in soil involves active contributions from soil micro-organisms. These transformations occur through numerous redox-reactions catalysed by an unknown number of enzymatic activities, triggered by soil microbes. In particular, they play an active role in soil fertility as a result of their involvement in the cycle of nutrients. Soil microorganisms are responsible for the decomposition of organic matter entering the soil (for example plant litter) and hence, also for the recycling of soil nutrients. Certain soil micro-organisms such as mycorrhizal fungi specifically increase the availability of mineral nutrients such as phosphorus to plants. Other soil microorganisms can increase the amount of nutrients present in the soil. For instance, nitrogen fixing bacteria can transform nitrogen gas present in the soil atmosphere into soluble nitrogenous compounds which plant roots can utilise for growth. In addition, soil microbes accelerate soil aggregate formation. Moreover, the soil microbiota provides a variety of services to society (Figure 1), such as the degradation of various undesirable organic chemicals in the

Figure 1. The importance of soil microbiology in biochemical soil processes towards society

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biosphere, such as xenobiotics, methane, and carbon monoxide. As such, the soil microbiota is of crucial importance in soil ecosystems, affecting the higher soil fauna and flora and human beings. In the past century, industrialization and the intense modernization of agriculture have had many recognizable benefits. Unfortunately, this civilization has also caused many environmental problems such as soil pollution and deterioration. In this context, the study of soil microbiology is of strategic importance. 5.1.2 Soil as the beholder of life

Only a decade ago, it was generally assumed that one gram of top soil contained a few hundred to maximum of a few thousand different bacteria, fungi, protozoa, etc. This assumption was based primarily on visual (microscopic) observations and on the cultivation of the organisms. The recent advance in molecular techniques has however shown that soil as such contains many organisms which cannot yet be grown and cultured, even with the best laboratory practices (Table 1). In fact, the soil matrix contains an abundance of genetic propagules, each of them with its own perfection and drive to exist and multiply. These organised bits of information, communicating with each other and with the environment around them, are called living entities or Operational Taxonomic Units (OTUs). Torsvik et al. (1990) reported approximately 4000 different genome equivalents per gram of soil, which suggests that there are perhaps 1000 or even more different species per gram soil. The most remarkable thing is that not only top soils are rich in micro-organisms, but also the soils deep under the surface. Hence, one certainly can state that soil, both at the surface and far below, is the beholder of life in all its complexity and genomic diversity.

Table 1. Culturability determined as a percentage of culturable bacteria in comparison with total cell counts (modified after Amann et al., 1995) Habitat

Seawater Fresh water Mesotrophic lake Tapwater Activated sludge Sediments Soils

Culturable (%) 0.001-0.1 0.25 0.1-1 0.1-3 1-15 0.25 0.3

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Obviously, the key to soil as a beholder of life is its heterogeneity. In the past great disrespect has often been shown to this key element. Man has tried to mix and homogenize soils as much as possible by various actions such as ploughing, mulching, planting mono-cultures, and irrigation. The fact that soils and sediments contain pores and spaces which are noncontinuous provides physical and chemical gradients. These gradients, for example in nutrient concentration, oxygen supply, and space availability, generate opportunities for all kinds of biological phenomena and thus form the basis of different organisms. They bring about diversity and generate complex biological communities. Modern molecular techniques very clearly substantiate the fact that physical and chemical heterogeneity is indeed essential to genomic richness. In Figure 2, heterogeneous versus more homogenous (surface versus vadose and saturated) soils are compared according to their microbial community diversity, represented by OTUs. History itself gives ample evidence of the importance of heterogeneity. For example, recent findings suggest that the desolateness of the Easter Islands, which once had a flourishing diversity of fauna and flora, was caused by extensive deforestation, i.e. de-heterogenization. Even the creation of life is based on heterogeneity. In fact, Nobel prize winner Christian De Duve states in his work "Vital dust" that wherever conditions are right, life will be created from cosmic dust, and wherever life is created, progress toward greater complexity is unavoidable (De Duve, 1995).

Figure 2. Operational Taxonomic Units (OTUs) abundance in surface, vadose, and saturated soils (modified after Zhou et ah, 2002)

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Although every functional gene or genome represents a marvellous entity evoking thousands to millions of years of evolution, it must be recognized that all genes are selfish and thus constitute a potential threat. Therefore, soils and soil ecosystems always deserve to be considered from a double perspective, i.e. a reserve of beneficial but also of potentially dangerous propagules. In this respect, another key life support function of the soil is represented by the ability of the soil microbial community to protect its surrounding ecosystem from these propagules. In fact, a healthy soil with a high diversity of soil microbial communities, has the capacity to counteract or suppress the activity of plantpathogenic or plant-deleterious (micro)organisms (Van Briiggen and Semenov, 2000). 5.1.3 Soil as the omnipotent degrader system

The soil has the remarkable and most valuable capacity to degrade a large variety of materials over time. In the 1970s, Martin Alexander challenged the myth that micro-organisms were infallible. If proper growth conditions were provided, it was generally assumed that they would find ways to degrade all the new xenobiotic chemicals released into the environment (Alexander, 1973). In those days, it sounded like heresy to pretend that bacteria would not in time succeed in cleaning up each single man-made chemical or product. The golden triangle in Figure 3 shows the omnipotence of the soil as a degrader system, given the appropriate environmental, microbial and contaminant conditions.

Favourable environmental conditions pH lectron-acceptors (Eh)

nutrients ionic strength porosity etc.

Figure 3. The golden triangle of biodegradation (modified after Doelman and Breedveld, 1999)

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Since the 1970s, a lot of information has been gathered on the potentials but also on the limitations of nature and although one still regularly discovers novel aspects of biodegradation and biodeterioration, it must be accepted that the degrader potential of the soil and its biotic community is limited. Moreover, degradation pathways can give rise to numerous unexpected and unfortunate conversions which render certain products more toxic. In addition, the highly applauded and widely practiced technique of natural attenuation needs some qualifications. Natural attenuation is now far too often an excuse for letting things stay as they are or in other words, doing nothing. Uncertainties regarding future extant dominant microbial metabolic reactions need to be addressed in order to decrease the uncertainty associated with natural attenuation strategies. It stands to reason that a site, in which natural attenuation based on microbiological processes is considered to occur, should be managed so that not only the benefits, but also the hazards of these processes are defined, hence the term MNA: Monitored or Managed Natural Attenuation. Moreover, the critical points of control and, in case of process failure, the points of action towards these microbiological processes, should be defined before evaluating natural attenuation. In addition to this regulatory response to temporal uncertainty, we should aim to reduce the amount of uncertainty. This can be achieved by site-monitoring protocols which explicitly characterize and confirm dominant extant metabolic activities and which can screen for potentially harmful intermediates (Smets et al., 2002). Changes of rates should be measured and interpreted in terms of the prevailing conditions. 5.1.4 Microbial soil quality redefined

The human species, very early in its evolution, has recognised the value of soil in terms of space, i.e. the horizontal dimension of soil. Actually, the soil was used to provide food, to bury all kinds of waste, and to build houses and dwellings. When no longer functional, the site was abandoned. However, the concept of soil quality, i.e. the full value of the substratum on which we live has, up to now, hardly received any attention. Notions of the intrinsic qualities of the soil such as biological stability, nutrient delivery, water-holding capacity, and the degradation of buried materials are difficult to find either in the literature or in the cultural heritage of peoples. At present, and maybe for the first time in the evolution of mankind, the issue of the intrinsic quality of the soil is being addressed. Yet, as evidenced in the December 2002 edition of the Dutch journal Bodem, the concept of soil quality is seen in a variety of ways. Remarkably, of the some 15 experts interviewed, none made a direct link between soil quality and life. In relation to

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the issue of soil being a cradle of evolving life, soil quality must be founded on its capacity to support current forms of life and be a site of evolution of genomic new life. The first aspect is already being explored by means of composite ecological indices (Schouten and Breure, 2001), but the second aspect is equally important. To stand for soil quality is to continuously strive for a better understanding and preservation of the mechanisms and means through which life exists and moreover evolves within the complex matrix of minerals, water and gases which we call 'soil'. In other words, soil quality is directly linked to present but also future life in all its diversity and newly unravelling capabilities. Indeed, in soil there is a constant evolution of new DNA within the individual organisms and new co-operations between the organisms themselves which generate new processes. 5.2 Monitoring soil health 5.2.1 Key functions in the soil

With regard to the indisputably important heterogeneity of the soil matrix, soil structure preservation goes without saying. In fact, the structural organization of the soil provides physical protection, favourable temperature, moisture and oxygen. In addition, heterogeneity assures the structural organization of different microbial groups. In this way, one microbial group can rapidly take advantage of metabolites that other groups have produced. Many microorganisms produce extra-cellular polysaccharides and other cellular debris, which function as a cementing agent that stabilizes soil aggregates. As such, soil microbes can play an important role in affecting water holding capacity, infiltration rate, crusting, erodibility, and susceptibility to compaction (Elliott et al., 1996). Heterogeneity is a necessary condition for preserving and beholding microbial diversity. This most important soil parameter is characterised by structural as well as functional diversity. Marshall et al. (1982) described the soil and its biotic component as "our most precious non-renewable resource". Linking the diversity of the soil biota to ecosystem functioning, therefore, remains a key ecological issue. Micro-organisms are key players in the nutrient cycling of the soil ecosystem (Pankhurst et al., 1997). The decomposition of organic residues results in the conversion to biomass or the mineralization of these residues to CO2, H2O, mineral nitrogen, phosphorus, and other nutrients. When eventually immobilised in biomass, the mineral components will also be released when this microbial biomass dies in the soil and becomes organic residue (Bloem et al., 1997).

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Another key characteristic concerns the hygiene of the soil. The presence or absence of potentially dangerous propagules or pathogens is of crucial importance to human, animal and plant health. The soil's life support function of suppressing these dangerous entities deserves full attention. The soil's ability to attenuate and transform or degrade potentially toxic compounds and elements is considered as a final key characteristic in this chapter. 5.2.2 Micro-organisms as indicators of soil health

The advantage of using micro-organisms as indicators for the monitoring of soil health lies in their ability to respond and rapidly adapt to changes in environmental conditions. This allows microbial analysis to be a more discriminating tool in soil health assessment than physical or chemical measurement. The way microbial populations and activities change, can therefore be used as an integrative indicator of soil health (Kennedy et al., 1995; Pankhurst et al., 1995). Based on the concept of environmental indicators, as described by Christensen et al. (1992), a microbial indicator in the present context can be defined: "A microbial indicator is a microbial parameter that represents properties of the environment or impacts on the environment, which can be interpreted beyond the information that the measured or observed parameter represents by itself" (Neiendam Nielsen and Winding, 2002). Because of the multi-functionality of soil, it is difficult to identify one single property as a general indicator of soil health (Paterson, 1998). Instead, Neiendam Nielsen and Winding (2002) proposed to use "end points of soil health", characterised by several soil ecosystem parameters. These soil ecosystem parameters can subsequently be characterised by several microbial indicators (Table 2). 5.2.3 Monitoring methods

A series of monitoring methods for the most important soil ecosystem parameters and their microbial indicators are addressed. The described methods can be divided into physiological and biochemical (traditional) methods and molecular methods. A schematic overview of these two approaches is presented in Figure 4. Traditional methods have as a shortcoming the fact that they depend on an analysis of phenotypic expression (for example respiration, catabolism) or on the ability to cultivate micro-organisms in the laboratory. As mentioned before, many micro-organisms are not culturable and phenotypic gene expression can be very low under test conditions (Torsvik et al., 1996). Molecular methods, i.e. nucleic acid technology, overcome these drawbacks. Indeed, a community's

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Table 2. List of microbial indicators for the characterization of soil ecosystem parameters for soil health monitoring (modified after Neiendam Nielsen and Winding, 2002) Soil ecosystem parameter Function C-cycling

N-cycling

General activities

Biodiversity

Root-activity General biomass

Biodiversity

Bioavailability of contaminants

Microbial indicator Soil respiration Metabolic quotient (qCCfe) Decomposition of organic matter Soil enzyme activity N-mineralization Nitrification Denitri fixation N-fixation Bacterial DNA synthesis RNA measurements Bacterial protein synthesis Community growth physiology Mycorrhiza Microbial biomass: direct methods Microbial biomass: indirect methods Microbial quotient Fungi Fungi-bacteria ratio Protozoa Structural diversity Functional diversity Marker lipids Suppressiveness to pathogens Biosensor bacteria Plasmid-containing bacteria Biomarker species Incidence and expression of catabolic genes

composition is most significantly represented by the complexity of information in nucleic acid molecules (Pickup, 1991; Stackerbrandt et alv 1993, Amann et al., 1995; Holben and Harris, 1995). However, it must be stressed that, because each technique has its specific advantages as well as disadvantages, it is important to combine several techniques, each specialized in a different aspect, in order to obtain a coherent overview. The principles, perspectives and limitations of structural and functional methods are discussed within the frame of biodiversity and nutrient cycling.

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Figure 4. Overview of traditional (cultivation and isolation) versus molecular methods (modified after Head et al., 1998)

Biodiversity Structural diversity

Shannon and Weaver (1949) were the first to quantitatively express the diversity of a community in the form of an index (Shannon-Weaver index). The species richness and the relative contribution each species makes to the total number of present organisms (i.e. species evenness) are considered within this index. Based on molecular techniques the phylogenetic diversity of soil microorganisms represents the genetic resources, i.e. actual and potential functions of the soil ecosystem (Neiendam Nielsen and Winding, 2002). For bacterial genetic diversity, 16S ribosomal RNA genes are most often studied, because these genes occur in all bacteria and moreover show variation in base composition among species, which provides phylogenetic information (Woese, 1989). Methods for examining the diversity of 16S rRNA sequences in microbial soil include denaturing/temperature gradient gel electrophoresis (DGGE/TGGE) (Muyzer et al., 1993/Heuer et al., 1997), terminal restriction fragment length polymorphism

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(T-RFLP) (Liu et al., 1997) and amplified rDNA restriction analysis (ARDRA) (Widmer et al., 1998). These genetic techniques assign so-called "fingerprints" to different soil communities, where each band in the fingerprint represents the 16S rRNA fragment of a certain microbial species (Figure 5). As such, these fingerprints give an image of the soil's microbial diversity. It has been shown that these methods are able to distinguish cultivated from non-cultivated soils (Bloem et al., 2002). Also herbicide applications in soil can be detected in this way (Seghers et al., 2003a). Fungal phylogenetic diversity on the other hand is traditionally studied by assessing the number and morphology of fungal fruiting bodies, spores and mycelium. However, the use of this method is often very difficult or even impossible when the formation of fruiting bodies fails to occur (Bridge et al., 2001). Molecular methods similar to bacterial diversity assessment, this time based on 18S rRNA, can overcome these problems. DGGE and TGGE are often used (Kowalchuk et al., 1997; Pennanen et al., 2001, Smit et al., 1999). Databases on fungal nucleic acid sequences are, however, still limited (Bridge et al., 2001; Smit et al., 1999).

Figure 5. DGGE fingerprints of bacterial communities in soil (Type I methanotrophic communities) and cluster analysis over two successive years (modified after Seghers et al, 2003a). Control soil (1K1, 1K2) and herbicide-treated (atrazine and metolachlor) soil (1A1, 1A2) sampled in September 2000. - Control soil (2K1, 2K2) and herbicide-treated (2A1, 2A2) sampled in March 2001. The DGGE bands indicated by the numbers 1 and 2 seem to be indicative for the long-term use these herbicides

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Ribosomal RNA fingerprinting techniques provide information on the genetic structure of microbial communities. However, a loss of function will not be detected in this way (Neiendam Nielsen and Winding, 2002). One way to overcome this problem is functional genome analysis, which looks at functional genes instead of rRNA genes. This might reveal differences between functionally related, but phylogenetically different micro-organisms, because the sequences of these functional genes might be more discriminative than genes encoding for rRNA (Palys et al., 1997). Functional genes have been used in fingerprinting studies, for example Wawer et al. (1995, 1997) used the (NiFe) hydrogenase gene to study the diversity and gene expression of Desulfovibrio species in environmental samples. Other examples are dinitrogenase reductase gene (nifH) to study the genetic diversity in Paenibacillus azotofixans strains in soil samples (Rosado et al., 1998) and the mer (mercury resistant) gene to analyze different sub-classes within microbial communities in soil and sediment (Bruce, 1997). Phospholipid Fatty Acid analysis (PLFA or FAME analysis) is based on the principle that microbes produce signature fatty acids that can be detected and quantified. These fatty acids are specific components of the cell membrane, exclusively present in viable cells. This makes the analysis a good indicator of recent microbial activity (Zelles, 1999). PLFAs are extracted from soil samples and subsequently analysed by gas chromatography (Zelles, 1999; Frostegard and Baath, 1993). Kennedy and Smith (1995) found that the diversity of soil microbial communities is directly related to the capacity of soils to suppress soil-borne plant diseases (i.e. suppressive soils). This means that the direct measurement of the microbial diversity represents a tool for assessing the soil's intrinsic potential to repulse plant disease (Van Elsas et al., 2002). Human pathogenic bacteria in soil are indicators of a putative danger for human health. Escherichia coli is the traditional indicator of faecal contamination and the related potential presence of other human pathogenic bacteria (Rhodes et al., 1988). However, preference goes out to the direct monitoring of human pathogenic bacteria (Morales et al., 1996) because some pathogenic bacteria may survive longer in the environment than E. coli. Pathogen enumeration is often done by cultivating methods (Marsh et al., 1998; Atlas, 1993). Molecular techniques are more accurate in estimating population sizes than cultivation-dependent methods. Examples of molecular techniques that can be used for the enumeration of specific microbial groups are the fluorescent in situ hybridization (FISH) (Szewzyk et al., 1993; Marsh et al., 1998) technique and quantitative Real-Time PCR (Lloyd-Jones et al., 1999).

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Functional diversity

Non-molecular methods are available to assess changes in functional attributes. An example is the technique of Community Level Physiological Profiles (CLPP), which is based on measuring oxidative catabolism of different substrates as potential sole C source (Garland and Mills, 1991). This measurement is done by using a multi-well plate (BIOLOG™), which tests 95 different substrates simultaneously and as such assesses the physiological diversity in the soil. One can also assess the enzymatic activity in soil, as the latter is mainly of microbial origin. Enzymatic activity of ecto-enzymes and free enzymes is used to determine the diversity of enzyme patterns in soil extracts. The enzyme activity is quantified by the incubation of the soil extract with commercial fluorogenic enzyme substrates (Hoppe, 1993) or colometric substrates (Wirth et al., 1992). The functional diversity of microbial populations in the soil may also be determined by diversity of mRNA within cells, because the mRNA concentration is correlated with the protein synthesis rate, and therefore also with cell activity. (Pfaffl et al., 2001). The technique of quantifying mRNA is still in its developmental stage. However, Lleo et al. (2001) have improved the sensitivity of the measurement. The nutrient cycle Because enzymes correlate directly to important soil activities, the direct measurement of these biocatalysts can be a useful tool in the assessment of metabolic reaction rates in the cycling of elements. Examples of enzymes as indicators of soil health are listed in Table 3 (Dick et al., 1996).

Table 3. Soil enzymes as indicators of soil health (modified after Dick et al., 1996) Soil enzyme

Enzyme reaction

Indicator of

Applicability

Dehydrogenase Beta-glucosidase Cellulase Phenol oxidase Urease Amidase Phosphatase Arylsulphatase

Electrol transport system Cellobiose hydrolysis Cellulose hydrolysis Lignin hydrolysis Urea hydrolysis N-mineralization Release of PCU3" Release of SCtf-

Microbial activity C-cy cling C-cy cling C-cy cling N-cycling N-cycling P-cy cling S-cy cling

Too general reaction Too general reaction Specific reactions Specific reactions Too general reaction Too general reaction Too general reaction Too general reaction

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The C-cycle

One way to monitor decomposition of organic matter is the measurement of soil respiration, a process in which aerobic micro-organisms oxidise organic matter to CCh, returning photosynthetically fixed carbon to the atmosphere. Measurement of soil respiration is one of the oldest, but still most frequently used techniques for the quantification of microbial activities in soil (Zibilske, 1994; Alef, 1995). Besides quantity, the decomposition rate of organic matter can also be assessed. A change in this rate will indicate a disturbance in microbial activity. Several techniques exist to measure this parameter (Verhoef, 1995; Harrison et al., 1988). The N-cycle

The nitrogen cycle includes the processes of N-mineralization, nitrification, denitrification and N2-fixation (Figure 6). Judging by sensitivity, nitrification is

Figure 6. Schematic representation of the N-cycle (modified after Zumpft, 1992). - Organic N (org-N) is mineralized to ammonium (NHf) by a variety of soil microorganisms (ammonification). - NH4+ is further immobilised by soil micro-organisms or oxidised to nitrite (NOr) and subsequently to nitrate (NO3) by aerobic nitrification. At this stage, leaching of NO3- to the groundwater may occur because of its high mobility. - Anaerobic denitrification to gaseous nitrogen (N2) via nitrous oxide (NO and N2O) may also take place, which can be carried out by a number of soil bacteria. - N2 can be transformed back into organic nitrogen through the process of nitrogen fixation, characteristic to certain microbial species

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believed to be a better parameter for monitoring of N-cycling than Nmineralization or denitrification, because fewer bacteria are able to complete the process of nitrification (Visser et alv 1992). Belser et al. (1980) described an ammonium oxidising assay as a measurement tool for this reaction. In addition, N-fixation can be measured by determining the frequency and diversity of the N-fixing species Rhizobium in soil. This can be established either by conventional counting of the number of nodules formed or through the isolation of the species by selective growth media (Laguerre et al., 1993; Bromfield et al., 1995; Tong et al., 1994). Recently developed molecular methods (DNA fingerprinting (Tan et al., 2001; Laguerre et al., 1994) and hybridization techniques (Laguerre et al., 1993)) can also map the diversity of this species. The P-cycle

In contrast to carbon and nitrogen, phosphor concentrations in soil are normally relatively low. The phosphate of nucleic acids and nucleotides is rapidly mineralized, but also quickly immobilised again by the micro-organisms, especially bacteria which have a high P requirement. In fact soil biomass is by far the largest pool of P apart from the inert inorganic and organic pools in the solid phase (Figure 7). However, not only do micro-organisms mineralise organic

P, some

species

of Aspergillus,

Arthrobacter, Pseudomonas

and

Achromobacter, abundant in the rhizosphere of plants, have the ability

Figure 7. The P cycle in a native prairie grassland in Western Canada. Figures in kg/ha*30 cm depth (modified after Halm et al., 1972)

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to secrete organic acids which attack insoluble Ca3(PO4)2 and subsequently release soluble phosphate (Bath and Nye, 1973). In addition, many plants have evolved cooperative life styles to enhance the absorption of P. One of the most important of these mechanisms is the mycorrhiza, a symbiotic association of fungus and plant root (Figure 8). The fungus is able to store P in its tissues, to be released to the host in times of P deficiency. In return, the fungus obtains carbon substrates and other growth requirements from the host (Allen et al., 1995). Moreover, the mycorrhiza directly contribute to soil aggregate stability as quantified by the production of glomalin, a glycoproteine (Figure 9). Stenberg (1999) and Van der Heijden et al. (1998) specifically proposed the group of arbuscular mycorrhizae in soil as an important indicator of plant and ecosystem health. Oehl et al. (2001) described the extraction and counting of spores of these mycorrhizae as a method to quantify them. Alternative methods, based on plant roots rather than on spores, were proposed by Kling et al. (1998) and Egli et al. (2001). Olsson (1999) described the use of arbuscular mycorrhizae specific PLFAs (Phospholipid Fatty Acid analysis). In the field of mycorrhizal research, molecular methods are also attracting considerable interest (DNA fingerprinting (Chelius et al., 1999; Jacquot et al., 2000; Redecker et al., 2001).

Figure 8. The symbiotic association between mycorrhiza and plant root, represented by the energy network diagram of hypothetical ecosystems. The additional contribution of the mycorrhiza is represented in the box (modified after Chilvers and Pryor, 1965)

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Figure 9. Glomalin and aggregate stability in soils from five different locations of the USA and Scotland (modified after Wright and Upadhyaya, 1998). Only the fraction of immunoreactive easily extractable glomalin is presented

5.2.4 Country side versus city green

Soil microbiology, whether it is looked at from a genomic point of view or from the perspective of the overall organisms, deals with sizes that are very small in comparison to higher soil or terrestrial fauna. The difference with the geophysical dimension is even more apparent. Because of the microscopic level of the indicators described, monitoring strategies for diversity will not differ with the ecological main structure (i.e. countryside or city green) being considered. Indeed, the microbial "village", is the level at which the microbial community functions and all tasks (niches) taken care of are situated at the soil micro-aggregate level. The latter comprises in the order of 100-500 micrometer. Actually recent progress in which the representativity of soil samples has been verified by molecular methods does corroborate the notion of micro-aggregate (Ellingsoe and Johnson, 2002; Ranjard et al., 2003). However, in case one has to deal with macro-systems such as river basins for example, the emphasis must be placed on the nutrient cycling function of the soil ecosystem. 5.3 Monitoring soil health in the presence of contaminants Polluted soils have become a public health problem and the demand for sitespecific risk assessment of soil pollution is increasing. In the traditional risk

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assessment of soil, chemical analyses are used in a model to calculate potential ecological risks. This method does not take biological availability into account although previous studies have confirmed that bioavailability is a major factor in the toxicity of environmental samples (Umbreit et al., 1988) and that the soil matrix is an important factor in determining a contaminant's bioavailability. Therefore, from an environmental point of view, the bioavailable fraction of a chemical compound may be a more relevant parameter in monitoring methods than the chemically extractable fraction. Because micro-organisms respond to the bioavailable fraction of a chemical compound in soil, they can be practical indicator tools. In this next section, the issue of monitoring soil health is oriented towards the most important groups of contaminants, i.e. heavy metals, Persistent Organic Pollutants (POPs) and nutrients (eutrophication). 5.3.2 Heavy metals

A very popular monitoring technique for the bioavailability of heavy metals is the method of biosensor monitoring. Biosensor bacteria are designed to respond to certain stress situations, such as heavy metals, through the use of reporter genes. The lux genes are reporter genes that code for bacterial luminescence. Using molecular techniques, these genes can be inserted into soil bacteria, resulting in new genetically modified micro-organisms (biosensors). The bioluminescence from the biosensors subjected to heavy metal pollution is proportional to metabolic activity and hence a simple, fast and sensitive way of detecting potentially toxic conditions in samples is provided by measuring light emission (Paton et al., 1997). Biosensor bacteria responding to mercury (Rasmussen et al., 2000) or chromate (Peitzsch et al., 1998) or zinc (Paton et al., 1997) are presently available. Heavy metal pollution can also be monitored by the use of a number of key species, or biomarkers, that are extremely sensitive or resistant to soil pollution. For example, it has been shown that protozoa bioassays can discriminate for soils with heavy metal contamination (Foissner, 1994; Campbell et al., 1997). Another group of micro-organisms frequently used as an indicator for heavy metal contamination is the group of cyanobacteria (Scherr et al., 2001). 5.3.2 POPs

Amongst the important classes of Persistent Organic Pollutants (POPs) are many families of chlorinated and brominated aromatics, and several pharmaceutical products (Jones and de Voogt, 1999). In recent decades, there have been several reports on the anaerobic and aerobic biodegradation of these

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persistent chemicals. Principles of degradation (Oldenhuis et al., 1992) and phenomena such as reductive chlorination (Bouwer and McCarty, 1983) and halorespiration (Fetzner, 1998) became general knowledge and have been implemented into bioremediation strategies. Monitoring strategies are manifold. Similar to the monitoring of heavy metal pollution, Burlage et al. (1990) and Layton et al. (1998) described biosensor techniques for the monitoring of naftalene degradation and the detection of poly chlorinated biphenyls (PCBs) respectively. Also the frequency of plasmidcontaining soil bacteria has been shown to be higher in polluted soils (Breen et al., 1992). Thus, measurement of numbers of plasmid-containing bacteria or numbers of plasmids in the soil can be used as a general indicator of environmental contaminants. Another possibility is to focus on catabolic genes. This approach is only possible when these genes are known to play a role in the degradation pathway of the contaminant. The principle is that the presence of a certain pollutant in the soil causes an increased expression of the respective catabolic genes. This higher gene expression is a consequence of either the growth of degrading bacteria or the spreading of the catabolic genes to other indigenous organisms. In addition, because catabolic genes can also be involved in the degradation of other naturally occurring or related compounds, the mere presence of these catabolic genes can be used as indicators of the soil's ability to transform or degrade xenobiotic compounds. Measurements of the incidence or expression of catabolic genes include the culturing of degradative micro-organisms (Sarand et al., 1999) or molecular methods for the detection of catabolic genes and the measurement of their expression (Wilson et al., 1999; Dennis et al., 2003). Certain key species (biomarkers) that are extremely sensitive or resistant to xenobiotic chemicals may be useful as indicators of soil POP pollution (Depledge et al., 1993). Amongst these the arbuscular mycorrhizae species are of specific interest (Cairney et al., 1999). It is clear that no single microbial parameter can be used universally in monitoring soil pollution, whether it be heavy metals or POPs. The key to successful monitoring lies within the combination of microbial activity and community measurements. The latter rely on the principle that pollution exerts a selection pressure, which induces a change in the community structure. Increased bacterial community tolerance, i.e. the PICT principle (Pollution Induced Community Tolerance), also indicates a change in community composition, with tolerant organisms becoming more abundant relative to sensitive organisms (Underwood, 1989; Rutgers, 1998). In this respect genetic

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fingerprinting can and has been used to show these genetic community structure alterations (Seghers et al., 2003b). 5.3.3 Nutrients

Eutrophication, the excess of nutrients, is considered to be one of the most important threats to soil quality. Eutrophication affects the capability of the soil to support biodiversity and affects the possibility of using groundwater for drinking water production. Eutrophication initially has an impact at the microbial level, as reflected in the quantity and magnitude of biochemical cycles that micro-organisms control (Reddy et al., 2002). Elevated nutrient levels can have a variety of effects on the structural composition and activities of the microbial communities involved in carbon cycling. Drake et al. (1996) reported the increased percentage of anaerobes, including methanogens, sulphate reducers and acetate producers. The ratio of oligotrophs, bacteria that require a low nutrient input, to copiotrophs, bacteria that require a high nutrient input, has been proposed to reflect the nutrient stress tolerance of the species present in the soil (van Bruggen and Semenov, 2000; Klappenbach et al., 2000; De Leij et al., 1993; Hattori, 1985). In Table 4 the most important indicators for the assessment potential of the different pollutants are given. 5.4 Novel considerations in relation to soil protection and recovery 5.4.1 Communication about soil

Overall, although the soil ecosystem has a number of essential functions of direct value to the 'consumer', the general public rarely express 'concern' in terms of the soil. The soil is a complex matrix and for most of the public 'terra incognita'. The ecologists can easily relate to air and water, and the criteria for their quality, because they observe in the water and in the air a multitude of living creatures. Yet, although so many plants and animals live on the soil, they are often not perceived to be directly related to the soil. Soil is a compartment of the environment which has been the depository of faecal matter, of corpses, of wastes (landfilling). Soil has been a material to be used, or to be more specific exploited by the farmer, the miner and the general user. One cannot expect successful legislation and concomitant implementation unless this reflects general public demand. Clearly communication about the soil should result in a better educated public and thus ensure proper legislation being developed and implemented. Apparently, the cry for recognition of biological monitoring aspects has been heard by European policy makers: an European monitoring and assessment

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Table 4. Soil ecosystem parameters to monitor the impact of different pollutants Soil ecosystem parameter Microbial indicator

Function

Biodiversity

Soil respiration Metabolic quotient (qCCk) Decomposition of org. mat. Soil enzyme activity N-cycling N-mineralization Nitrification Denitrification N-fixation General Bacterial DNA synthesis activities RNA measurements Bacterial protein synthesis Community growth phys. Root-activity Mycorrhiza

Assessment Assessment Assessment potential for potential for potential for heavy POPs nutrients metals

C-cycling

General biomass

Microbial biomass: direct Microbial biomass: indirect Microbial quotient Fungi Fungi-bacteria ratio Protozoa Biodiversity Structural diversity Functional diversity Marker lipids Suppressiveness to pathogens BioavailaBiosensor bacteria bility of Plasmid-containing bacteria contaminants Biomarker species Incidence and expression of catabolic genes

+ ++

+ + +

.

++ + -

. +

. ++ ++

++ +

++ +

++

.

.

+ +

. +

+ ++

+ +

++ + +

+ ++ + ++

. + .

+ + +

= not useful or no experience + = a useful parameter ++ = a very useful parameter

framework on soil has been proposed to provide policy makers with relevant information on soil from current national soil monitoring programmes (Huber et alv 2001). This framework has the specific objective of comparing biological properties with physical or chemical ones (Huber et al., 2001). The objective of

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the EU COST (www.isnp.it/cost/cost.htm) is even specially defined to investigate micro-organisms as indicators of soil health. 5.4.2 Microbiota as a -patrimony

Society has realized that certain parts of its heritage have so much value in terms of information, representation and remembrance, that they deserve to be preserved. Measures have already been taken to protect unique soils and soil ecosystems such as dunes, polders and podzols as top layers. But what about the sub-soils and the deeper horizons, that also store a wealth of information? Should one not consider the possibility of conserving certain zones against man-made alterations to the water table, against possible infiltration of contaminated groundwater, against input from genetically altered plants and other biological propagules? This would constitute a protection of a cultural patrimony of terrestrial model ecosystems both in a horizontal and a vertical dimension. For scientific reasons a totally different type of site deserves preservation. Indeed, due to the industrial revolutions of the past decades when wastes were often dumped directly into the soil, there are sites which can be typified as 'historically polluted soils'. Some of these, such as soils that have been polluted for almost half a century with high levels of various heavy metals, various mixtures of hydrocarbons, various degrees of nitrosated or chlorinated chemicals, represent sites of evolving adaptation of microbial communities from which we can learn a great deal in terms of new biochemistry and biology. The proposal to preserve certain 'historically polluted soils' should not be interpreted to mean that many such polluted sites should be kept untouched. The argument is that some of these sites which are in evolution and should in part be considered for preservation, because they can harbour potentialities for other sites already present or which may be encountered in the future. Finally, in line with the above considerations, it stands to reason that it would be most valuable for industries that produce chemicals that are released into the environment, to set up a 'soil nursery'. In these plots, treated with specified chemicals in a carefully documented way, new forms of biodegrative micro-organisms could evolve. This should be done for chemicals designed to enter the environment above a certain tonnage per year, and whose fate has not been totally established in terms of kinetics and mass balances. These sites could then be explored by means of advanced molecular tools for the appearance of new catabolic functions (Beller et al., 2002). It would be an act in line with the 'Charter of Responsible Care' and be in perfect agreement with the

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concept of the precautionary principle (i.e. scientific uncertainty + suspected harm). Such an initiative in which the producers of these chemicals help to contribute to a soil patrimony that allows the long-term documentation of the fate of chemicals in a series of monitored soil systems will be highly appreciated by the public. The proposed a soil patrimony may be reminiscent of the existing concept of environmental specimen banking. In the latter case, samples are taken and stored under liquid nitrogen for analysis and reference purposes in the future (Roling and Verseveld, 2002). However, the proposal to maintain microbiota as a patrimony is even more challenging. It requires the design and allocation of sites so that the natural physical, chemical and biological dynamics can be allowed to pursue their course and this for time frames relating to those of the rate-limiting processes involved, that is of the order of centuries. Continued monitoring and reporting are absolutely necessary in order to ensure that such processes proceed according to plan. 5.4.3 Major threats to the soil ecosystem

The law of Pareto stipulates that the flux of economic goods available will be distributed among the different consumers in a specific way such that on average some 20% of the consumers will take about 80% of the flux and thus concomitantly 80% of the consumers will have to do with 20% of the incoming means. This law, as shown in Figure 10 (Dejonghe et al., 2001) makes it clear that overall even the best and effective association of species can at most only properly deal with 80% of opportunities and tasks. A diversity of other species is, therefore, needed to complete other potentials and duties. An important advantage is that when a dominant species that has been successful so far gets into difficulties due to changes in the environment, parasitism or predation, a less important species from the large pool can replace it. There is an urgent need to examine the soil ecosystem in the context of the Pareto distribution. Methodologies should be developed to determine the success of individual species present in a soil sample so that Pareto's curve can be plotted. A good soil ecosystem should represent such a Pareto distribution of functional success versus species diversity. There should be no skewing either to the left, giving more emphasis to fewer species, or to the right, giving no chance to highly efficient performing species, in the dominance/species distribution of the soil's microbial community. Therefore, impacts on the soil which bring about shifts in the Pareto distribution of the soil microbial community should be considered as threats to the viability and health of the soil.

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Figure 10. Representation of different Pareto curves [energy flux=l I (fraction of the community)E] (Pareto, 1897). - According to Pareto, about 80% of all energy flux is governed by about 20% of the species in a community (modified after Dejonghe et ah, 2001)

At present there is awareness of the fact that strong impacts on the vitality of the soil biological system can be exerted by chemicals. The influences of the latter on the Pareto distribution still have to be documented. Moreover, there should be a quantification of the impacts of other threats such as soil homogenization by paving, ploughing, irrigation and growing mono-culture crops in terms of their impact on the normally less successful but equally important 80% of the other species present. 5.5 Conclusions towards implementation in soil management The soil is the beholder of microscopic life in all its genomic diversity and complexity. The latter issue truly deserves further exploration in terms of adequate measures to be taken to safeguard this patrimony. The analysis of microbial communities and the use of soil micro-organisms as biomarkers appear to be a sensitive method for detecting changes in soil health. They certainly ought to be included in monitoring programmes of soil protection policies. New developments in the field of molecular science offer possible ways of integrating soil quality management into mainstream economic life and at the same time encouraging public awareness and appreciation. The social dimension of soils, i.e. making the public aware of the soil quality and the services it provides, is of crucial importance and requires conventional soil scientists to interface with the public. The soil microbial community has been monitored extensively with respect

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to various impacts, particularly those exerted by chemicals. At present there is a need to consider the microbial community in the same way as any other economic community and learn to quantify the distribution of incoming goods and opportunities among the wide diversity of species present. Any deviation of the default Pareto 80/20 configuration should be examined with care. Hence not only chemicals but all anthropogenic actions impairing the diversity and effective distribution of tasks in the soil ecosystem warrant re-examination because they might have a particular effect on the large pool of non-dominant soil species of which the soil has - as no other system - up to now been the most effective beholder. Most probably many microbial interactions are still hidden. Therefore, politicians and regulators should be careful when judging soil quality and health based on a shortlist of narrowly focussed "test sets". Besides applying a general or a tailor-made monitoring set, we should also investigate the Pareto concept, especially in relation to respiration, dehydrogenase, nitrification, denitrification and nitrogen fixation. In addition, the expression of genes and mycorrhiza determination may become of special importance for the evaluation of soil management. Due to human exfoliation soil erosion is proceeding rapidly. Therefore there is a need for broad-based communication about the services of the soil to society in order to establish adequate soil management strategies. References Alef, K., 1995. Soil respiration. In: Methods in Applied Soil Microbiology and Biochemistry. Alef, K. and Nannipieri, P. (eds.). Academic Press, pp. 214-218. Alexander, M., 1973. Micro-organisms and chemical pollution. Bioscience 23 (9): 509-515. Allen, E.B., M.F. Allen, D.J. Helm, J.M. Trappe, R. Molina and E. Rincon, 1995. Patterns and regulation of mycorrhizal plant and fungal diversity. Plant Soil 170: 47-62. Amann, R.I., W. Ludwig and K.H. Schleifer, 1995. Phylogenetic identification and in situ detection of individual microbial cells without cultivation. Microbiol. Rev. 59: 143-169. Atlas, R., 1993. Handbook of microbiological media. Parks, L.C. (eds.). CRS Press, Inc., Boca Ratton, Florida. Bath, K.K. and P.H. Nye, 1973. Diffusion of phosphate to plant roots in soil. Plant Soil 38:161-175. Beller, H., S. Kane, T. Legler and P. Alvarez, 2002. A real-time polymerase chain reaction method for monitoring anaerobic hydrocarbon-degrading bacteria based on a catabolic gene. Environ. Sci. Technol. 36: 3977-3984. Belser, L.W. and E.L. Mays, 1980. Specific inhibition of nitrite oxidation by chlorate and its use in assessing nitrification in soils and sediments. Appl. Environ. Microbiol. 39: 505-510. Bloem, ]., P.C. de Ruiter and L.A. Bouwman, 1997. Food webs and nutrient cycling in agroecosystems. In: Modern Soil Microbiology, van Elsas, J.D., Trevors, J.T., and Wellington, E.M.H. (eds.). Marcel Dekker Inc., New York, pp. 245-278.

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Chapter 6

THE USE OF SOIL INVERTEBRATES IN ECOLOGICAL SURVEYS OF CONTAMINATED SOILS N. van Straalen

Abstract

This chapter on soil invertebrates, "the little rotters", provides an overview of field studies on the effects of soil pollutants on soil invertebrates published after 1990. It also assesses these studies in the light of the need to develop evaluation instruments (indicator systems) for soil quality that are both convincing and practical. Moreover, it proposes a scheme of invertebrate biodiversity assessment that can be used in ecological surveys of contaminated soils. Below ground diversity is essential for above ground ecosystem function. Soil fauna have a catalyzing role in the cycling of elements but also an important function in vegetation diversity and succession. Due to their abundant presence the soil fauna can be seen as the soil's ecological insurance. Species diversity makes a community more stable and secure against catastrophic events. An overview of vulnerable groups among the soil animal community clearly shows a difference in the sensitivities of the various species as far as contaminants are concerned. Springtails are more sensitive than enchytraeid and earthworms to polyaromatic hydrocarbons. Feeding habit and metabolic capacity play a role in this "sensitivity". The effects of POPs in the field have only been observed to a very limited extent, because of the lack of data. Most is known about the impact of heavy metals to which arthropods are less susceptible than earthworms. The composition of the nematode community is a nice tool for observing soil quality since it combines species diversity with functional aspects such as life cycle and feeding strategy. A list of many indices, such as susceptibility index, geomorphological index, Shannon-Wiener index and Maturity Index, are discussed with regard to the use of in situ soil management. Key characteristics in investigating disturbance and recovery are species richness and diversity (1), distribution of numbers over species (2), distribution of body size over species (3), classification of life cycle attributes (4), classification according to ecophysiological preferences (5) and the structures of food webs (6). During the last ten years research into practical soil qualification has not increased our understanding. As a result the boundary between soil regulation practise and soil scientific insight remains barren land. Soil evaluation can only be done when reference systems of certain classes of soil ecosystems are available. Among the great possible variety of bio-indicators the Maturity Index of nematodes, the ecological qualification of earthworms and the species richness of micro arthropods are the best candidates.

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6.1 Introduction The soil ecosystem harbours an extremely large biodiversity and it is increasingly recognized that this below ground diversity is essential for above ground ecosystem functions (Copley, 2000). Almost every class in the animal kingdom has at least some representative in the soil community. Figure 1 shows some common groups of arthropods, which is by far the most species-rich group of soil invertebrates. Other important animal groups are tardigrades, nematodes, flatworms, earthworms, slugs and snails. Biomass, productivity and

Figure 1. Impression of soil animal diversity, showing representatives from the major groups of arthropods. Adapted from Parisi (2001), with permission from the Society of Medicine and Surgery and Natural Science of Parma, and the author

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organic matter turnover of the animal communities in soil have been studied extensively in programmes such as the International Biological Programme since the 1970s, and a great deal of the knowledge accumulated over time is now well documented in text books (Swift et alv 1979; Curry, 1994). The issue of soil diversity and ecosystem function continues to receive attention, however, as testified by a recent meeting of the British Ecological Society on "Soil Biodiversity and Function" (Lancaster, March 2003) and the launch of a thematic programme "Biological Diversity and Ecosystem Function in Soil" by NERC in the UK (2003, see http://www.nmw.ac.uk/soilbio). The enormous biodiversity of "little rotters" has been called an enigma (Anderson, 1975), because of the apparent high trophic overlap between species. Soil biodiversity also offers a great opportunity for community ecology because it allows hypotheses of competition theory to be tested in a wide variety of settings (Lee, 1994). Societal interest in biodiversity has grown considerably since the Rio convention, held in 1992. Hagvar (1998) argued that the agreements made in the convention should also apply to soil conservation. One important argument in the discussion is that the maintenance of ecosystem functions and soil fertility critically depends on biodiversity of soil life. In addition, ethical reasons compel mankind to conserve the variety of odd creatures in soil. Recent progress in soil ecology has considerably deepened our understanding of the role of soil invertebrate fauna. Not only do soil fauna have an important catalyzing role with respect to microbial processes such as nitrogen mineralization (Verhoef and Brussaard, 1990), soil fauna also influence - by selective feeding on plant roots - above ground processes such as plant diversity and succession (Brussaard, 1998; De Deyn et al., 2003). Another recent insight is that soil fauna contribute to the "ecological insurance" of the soil ecosystem. The concept of ecological insurance was introduced by Naeem and Li (1997). They conducted experiments, using experimental microcosms, in which different assemblages of algae, bacteria and protozoans were cultured and the long-term stability of the community was assessed. It turned out that if the trophic groups of the community contained several species with a similar ecological role, fluctuations in the total biomass were smaller than in systems with only one species per functional group. This paper was one of the first experimental proofs to show that the apparent functional redundancy of species in a community is of crucial ecological importance: it prevents the collapse of the system when one of the species falls back. In other words, biodiversity provides insurance against fluctuating environmental conditions and prevents catastrophic consequences.

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Further experimental research, mostly in plant community ecology, has shown that species richness may not be the most crucial aspect of biodiversity as related to ecosystem reliability. The diversity of species attributes seems to be more important (Walker et al., 1999). In other words, to make a community more stable and to secure it against catastrophic events, it does not simply need more species, but more species with certain ecological roles. Loreau et al. (2001) reviewed the recent studies on the relation between biodiversity and ecosystem function, of which most relate to above ground plant diversity. This review has reconfirmed the idea that a minimum number of species is necessary to make an ecosystem function, and a greater number of species is necessary to confer stability. It is also increasingly realized that the influence of below ground biodiversity is not limited to the soil itself, but extends to plant growth and even to the herbivores feeding on these plants (Wurst and Jones, 2003). Despite the considerable progress in fundamental soil ecology illustrated above, the position of ecological arguments in practical soil quality evaluations has not improved very much over the last ten years (Van Straalen, 2002). The question is: why should this be so? One reason might be that the progress in soil ecology theory has not led to practical strategies for soil evaluation. Despite the improving knowledge on fundamental processes and interactions in soil, the regulator is still at a loss when it comes to evaluating the risks of soil pollution, the quality of soil's services, etc. In terms of the analysis provided by Turnhout (2003) one might say that the boundary zone between soil regulation practice and soil scientific insight is still barren land. Turnhout (2003) argued that ecological indicators are useful instruments to bridge the gap between science and policy. Following this argument, one way of improving the position of ecological arguments in soil quality evaluations is to develop indicator systems that can be used in soil quality assessments. The aim of this chapter is therefore to: 1. provide an overview of field studies on the effects of soil pollutants on soil invertebrates published after 1990; 2. assess these studies in the light of the need to develop evaluation instruments (indicator systems) for soil quality that are both convincing and practical; 3. propose a scheme of invertebrate biodiversity assessment that can be used in ecological surveys of contaminated soils. 6.2 Effects of soil pollution on invertebrate communities Communities may change under soil pollution for various reasons. One obvious effect is the differential sensitivity of species. The biodiversity of

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ecotoxicological responses is ultimately determined by toxicological processes such as uptake of the toxicant, sequestration, biotransformation and excretion. There is an extreme diversity among the soil-living animals in these processes and the presence of biochemical targets upon which toxicants may act (Van Straalen, 1994). Consequently, when a community is exposed to a toxicant, sensitive species will be suppressed and this in itself will cause community change. Differential sensitivity is not the only factor explaining community effects. Ecotoxicological work at the population level has demonstrated that the effects of toxicants also depend on the priorities that are set by organisms in their vital life-processes. For example, some organisms give priority to reproduction under all circumstances, while others give priority to survival at the cost of reproduction. These "strategies" can be characterized by the "sub-lethal sensitivity index" (SSI) proposed by Crommentuijn et al. (1995). Species have a high SSI if the effects on reproduction are apparent at exposure concentrations far below the LC50 (concentration causing 50% mortality among a group of test organisms). The distinction is comparable to the well known r-K dichotomy that was discussed in an ecotoxicological context by Eijsackers (1994). Obviously, in addition to direct effects, community changes are also brought about by indirect interactions, acting through the food-web or via physical changes in the habitat. For example, Pokarzhevskii et al. (2003a) argued that the effects of toxicants on soil animals might actually be due to the effects on microorganisms in the gut and in the environment. Micro-organisms constitute an essential link between detritus (which as such has low nutritional value to animals) and the animal's physiology. Likewise, the early work on pesticides in soil communities has already demonstrated how the numbers of some groups of soil animals may increase temporarily due to the suppression of their predators (Van de Bund, 1965). Stimulatory effects of herbicides on soil fauna have been described which are due to the increased accumulation of leaf litter on the soil surface (Edwards and Thompson, 1973). All these processes together will cause communities living in polluted soils to be different from those in clean soils. Conversely, community structure can be used to draw inferences about the quality of the soil. To investigate the potential for the development of such indicators, a short overview of the effects of soil pollution on soil invertebrates is provided in the sections below. The overviews focussed on the last ten years: no attempt is made to cover the literature before 1990. The review is limited to field studies because this paper aims to explore the use of soil invertebrates for in situ soil assessment.

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6.2.1 Polycyclic aromatic hydrocarbons and oil

Polycyclic aromatic hydrocarbons (PAH) constitute one of the most important components of soil pollution and are often the prime reason why a certain site poses a problem in land management. One of the reasons for this regulatory urgency is due to the fact that many PAH-contaminated sites are located in urban and residential areas and so decision-making comes under pressure as a result of economic and spatial planning considerations. Risk assessment is hampered by the fact that relatively little is known about the effects of PAHs on soil invertebrates (Van Brummelen et al., 1998). Most studies focus on metabolism and degradation, rather than on effects. Blakely et al. (2002) found weak correlations between PAHs and some groups of soil invertebrates (notably nematodes) in creosote-contaminated field soils. The abundance of many invertebrates was, however, better correlated with soil physical factors, especially bulk density, than with the concentrations of PAHs. It was suggested that PAHs influenced soil invertebrates not through direct toxicity but through an effect on microhabitat structure and food availability (bacteria and fungi). Eijsackers et al. (2001) evaluated the PAH contamination of peat sediment when this was left to age on the side of a ditch after dredging. The authors showed that the colonization of dredged sediment by earthworms contributed to the degradation of PAH. The ecological risk of PAH contamination in dredged sediment was considered negligible. That earthworms may enhance PAH degradation was also found in experiments reported by Ma et al. (1995) and Jager et al. (2000). Van Gestel et al. (2001) reported that chronic bioassays using earthworms could detect adverse effects of oil pollution in soils collected from a petroleum harbour. Pokarzhevskii et al. (2003b) employed a similar approach using enchytraeids and demonstrated low survival and reproduction rates amongst worms introduced into oil-polluted soil. Measurements of PAHs in field-collected animals have shown systematic differences between the major groups of animals. Earthworms seem to accumulate PAH more than arthropods such as insects and isopods (Pathirana et al., 1994; Van Brummelen et al., 1996; Achazi and Van Gestel, 2003). These differences may partly be related to the feeding habit of the species (Faber and Heijmans, 1996), but also reflect the capacity for metabolism, which is relatively high in isopods and springtails (Van Brummelen and Van Straalen 1996; Stroomberg et al., 2003; Howsam and Van Straalen, 2003), and relatively low in earthworms (Jager et al., 2000; Matscheko et al., 2002). Laboratory testing has demonstrated that on average, springtails are more sensitive to most PAHs than earthworms and enchytraeids (Sverdrup et al.,

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2002). Too little knowledge is available at the moment to draw any conclusions on differential sensitivity within the arthropods. So there is no good basis for predicting community change in PAH polluted soils. Effects in the field may be expected in individual arthropods and, when PAH contamination comes with an oil component, also in earthworms. However, community composition of soil invertebrates provides little clue as to such effects. Community inventories do not seem a promising tool for evaluating PAH contaminated soil. The use of more sensitive biochemical markers (metabolite concentrations, DNA damage) seems to be a better strategy. 6.2.2 Dioxins and poly chlorinated biphenyls

A review of dioxin toxicity to soil invertebrates and terrestrial wildlife was provided by Van Straalen et al. (1995). This review was mostly based on field studies at Seveso and a military base in Florida. The conclusion was that the extreme toxicity of some dioxins, such as 2,3,7,8-tetrachlorodibenzo(p)dioxin, also applies to soil invertebrates such as earthworms. However, despite this toxicity, effects were not observed at the field sites mentioned, which could be explained by the extremely strong binding affinity of dioxins to soil and consequently their very low bioavailability. Since 1995, no further studies on dioxin effects to soil invertebrates seem to have been made. Polychlorinated biphenyls (PCBs) are slightly more bioavailable than dioxins. The concentrations in invertebrates are often in the same order of magnitude as the total concentrations in the organic fraction in the soil (Hendriks et al., 1995; Lupetti et al., 1994), which is expected under the model of equilibrium partitioning (Van Gestel, 1997). However, in the superlipophilic range (PCBs with high chlorination) the residues in earthworms are lower than expected due to the fact that uptake from the food, which is less governed by equilibrium distribution than dermal uptake, begins to make a significant contribution to the total body burden (Belfroid et al., 1996). Although toxicity studies have indicated that PCBs can be toxic to earthworms (Fitzpatrick et al., 1992; Parmelee et al., 1997), effects are not expected to lead to major changes in the structure of the decomposer food-web in soils contaminated by chemical waste. Due to the accumulative properties of PCBs and their extreme persistence, environmental effects are much more relevant at higher levels in the food chain (mammals, birds). The evaluation of PCB and dioxin toxicity in soil itself should be based on biochemical effects seen in single species, rather than on indicator systems derived from community structure.

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6.2.3 Heavy metals

Heavy metals are the best investigated components of soil pollution. Metals are often used as model compounds for toxicity testing in the laboratory, because of their stability and ease of analysis. It is impossible to review all the work that has been done on the effects of heavy metals on soil invertebrates here. The account below is limited to some of the most significant field studies carried out in the last ten years. A review of the older literature is given by Bengtsson and Tranvik (1989) and Hopkin (1994). Copper is well known for its toxicity to earthworms and was already documented many years ago from field observations (Van Rhee, 1977). These observations have been confirmed in many laboratory studies (Reinecke and Reinecke, 1996; Khalil et al., 1996). The toxicity of copper to earthworms has been a major issue of concern in the use of copper-amended pig fodder, leading to copper-rich manure with restricted possibilities for disposal on land. Another well known issue is the use of copper oxychloride as a fungicide in vineyards. Due to liberal use of this material, many vineyard soils have become completely devoid of earthworms, with consequent soil deterioration (Abdul Rida and Bouche, 1995; Paoletti and Bressan, 1996). Belotti (1998) surveyed the abundance of three earthworm species in German vineyards. Disturbance was indicated at levels of total copper greater than 33 mg per kg of dry soil. Copper is also relatively toxic to snails (Gomot, 1997) and nematodes. Bongers et al., (2001) conducted experiments with nematode communities exposed to copper sulphate in a sandy soil and documented a great diversity of sensitivities among the nematode fauna. Predatory nematodes appeared to be the most sensitive to copper and the maturity index, a measure of community structure indicating the dominance of pioneering species versus persister species (see below), was strongly correlated with the copper treatment. This study and several others, for example, Korthals et al. (1996a) and Parmelee et al. (1997), convincingly demonstrated that the nematode community is highly responsive to soil factors, including pollution. Arthropods appear to be less susceptible to copper than earthworms. One of the few studies documenting the effects of copper on arthropods in the field is Bruus Pedersen et al. (1998). This study was conducted at an historically polluted site in Denmark (Hygum), where a hot spot of copper pollution had developed as a result of the use of copper sulphate for timber preservation in the past. The authors showed that the Shannon-Wiener diversity index of the arthropod community (Collembola, mites) correlated negatively with the total concentration of copper in soil (Figure 2). This study is similar to the work of

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Figure 2. Correlation between the Shannon-Wiener diversity index of the arthropod community (Collembola, mites) and the total concentration of copper in soil. Reproduced from Bruus Pedersen et al. (1999), with permission from Kluwer Academic Publishers and the author. - The study was conducted in a gradient laid out in a grass field contaminated by the historical use of copper sulphate for wood preservation

Hagvar and Abrahamsen (1990), who analysed changes in community structure along a natural gradient of lead pollution. The data from that study were reanalysed by Posthuma (1997) to show that the effect of lead on species richness of mites and Collembola occurs at around the same levels as extrapolated from laboratory toxicity data. In another gradient study, Zaitsev and Van Straalen (2001) analysed oribatid mite community structure in contaminated soils (copper, zinc and iron) near an industrial plant in Tula, Russian Federation. Overall species composition of oribatids was not clearly related to the pollution, however, there were some species that showed a clear response. The study demonstrates the necessity of identification at the species level, especially in oribatids, where the diversity of life-history strategies, feeding habits and susceptibility to disturbance is very large, even among closely related species (LeBrun and Van Straalen, 1995). Gillet and Ponge (2003) showed that changes in species composition in such gradient studies may be confounded by changes in vegetation cover: the most polluted site in their study attracted many heliophilic arthropods due to the lack of trees. They also showed that species might change their feeding habit if

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the accumulation of polluted litter forces them to feed on fungi in the deeper soil. Similarly, Pouyat et al. (1994) in a comparative study of natural and urban sites, found a negative correlation between fungivorous arthropods and metal concentrations which they ascribed to effects on the fungi, rather than on the arthropods themselves. Russel and Alberti (1998), however, noted that geogenic metal contamination favoured a dominance of surface-living Collembola, compared to the dominance of euedaphic species in a reference site. Cadmium is the most studied heavy metal, especially in laboratory studies. It can exert toxic effects at lower concentrations than copper, lead and zinc. However, in mixtures of metals in the environment, it is often zinc that is mainly responsible for the effect. Hopkin and Hames (1994) compared laboratory-derived effect thresholds with concentrations of various metals in the field near a smelting works and concluded that zinc, not cadmium, was responsible for the absence of the isopod Porcellio scaber in the vicinity of the works. Hopkin and Spurgeon (2001) proposed a simple procedure to decide which metal in a mixture is likely to be the main reason for the observed effects. They developed an index called "relative toxicity factor" and defined it as follows:

C T TF. =—J- -£where TFi is the relative toxicity factor for metal i, G is the measured concentration of metal i, G is the measured concentration for a reference metal (for example cadmium), T. is a toxicity parameter (for example EC50) for metal i and Tr is the same parameter for the reference metal. The metal with the largest TF value is the one most likely to be causing toxic effects in the field. The authors calculated this index for earthworms, snails, springtails and isopods, and concluded that at their study site, all four species indicated zinc as the major metal responsible for toxicity. Similarly, Lock et al. (2003) found a relationship between soil zinc concentrations and Collembola species diversity in a mine site contaminated by zinc, lead, cadmium and other metals. Spurgeon and Hopkin (1996, 1999) sampled earthworm communities at various distances from the Avonmouth metal smelter in the United Kingdom and found that species composition was affected at distances closer than 3 km from the works, where zinc concentrations in soil exceeded a value of 1290 mg/kg. There were significant species-dependent sensitivities, for example Aporrectodea caliginosa did not occur at zinc concentrations higher than 903 mg/kg, while Lumbricus castaneus survived up to 3627 mg/kg. The authors

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suggested that species with less well-developed calcium glands would be more susceptible to the toxic effects of zinc. Smit et al. (2002) monitored nematode communities in zinc-spiked natural soils left to age for several years in an outdoor experimental field. The nematode community comprised a great diversity of susceptibilities to zinc. Some species (for example Aphelenchoides sp.) increased in numbers even at zinc levels higher than 2000 mg/kg, while others (Eucephalobus striatus) could not maintain a viable population at zinc levels above 200 mg/kg. Effects on species richness were apparent above 320 mg/kg. However, a sensitive multivariate statistical analysis could demonstrate effects of zinc at concentrations as low as 56-100 mg/kg, so directly above the background level. It looks as though zinc has a natural differentiating effect on soil communities, comparable to soil factors such as pH and salinity. Posthuma et al. (2001) argued that the semifield study demonstrated that effect thresholds derived from laboratory toxicity testing can indeed be used in the forecasting of effects in the field, although it will not be possible to predict what will happen exactly or that it will always happen. 6.2.4 Acidic precipitation and eutrophication

The acidity of rainfall is mainly influenced by gaseous air pollutants such as SO2. Since the drastic emission reductions taken in the course of the 1970s and 1980s in many parts of the world, SO2 and acid rain have become less of a problem. However, other air pollutants (ozone, oxides of nitrogen) remain. In regions with intensive farming, ammonia concentrations in air are often extremely high and ammonia may contribute significantly to acidification of soils by stimulating nitrification. Many field inventories have identified soil pH as the most important determinant of community structure, both in arthropods (Ponge, 1993, 2000; Salmon and Ponge, 1999; Chagnon et al., 2000a, b; Chagnon et al., 2001) and earthworms (Sanchez et al., 1997). Kuperman (1996) compared communities along a deposition gradient and showed that several invertebrate groups, especially earthworms, snails and dipteran larvae, responded negatively to the decreased soil pH caused by acidic deposition. The importance of pH is also demonstrated by the occurrence of acidophilic species around the base of trees, a microhabitat with a lower pH than the area in between trees, due to the acidity of stemflow (Kopeszki, 1992; Scheu and Poser, 1999; Kaneko and Kofuji, 2000). A review of soil invertebrate responses to acid deposition was published by Kuperman and Edwards (1997). Although some invertebrate groups are on average more tolerant to acidity

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than others (for example enchytraeids versus earthworms), all groups comprise both acidophilic and alkalophilic species. Van Straalen and Verhoef (1997) established that the diversity of responses to pH amongst Collembola and Oribatida was extremely large, some species having a clear acidic preference, others preferring neutral to alkaline substrates. However, Salmon et al. (2002) showed that pH preferences might depend on the ionic identity of the buffers used, so the predictive power of laboratory tests for the field situation may be limited. Hogervorst et al. (1993), in a five-year study of soil arthropods in forests with differing degrees of vitality from acidic deposition, showed that certain species of oribatid mite were very responsive to the chemical changes induced by soil acidity, while other species did not respond at all. The species that was most strongly negatively correlated with forest vitality, Platynothrus peltifer, was also one of the most alkaline-preferring species in laboratory tests. In the case of ammonia, high precipitation not only leads to acidification of the soil, but also to eutrophication. Several studies have shown that, in addition to humidity and pH, ammonia is a main factor limiting nitrification (Laverman et al., 2000; Krave et al., 2002). So under high rates of ammonia deposition, nitrification is released from limitation and increases considerably in rate, which often leads to excessive nitrate leaching from the soil. This together with increased acidity from nitrification may cause depletion of cations such as manganese and calcium. Obviously, nitrogen enrichment has a major effect on soil microbial communities and organic matter decomposition. However, the effects on animals are less clear and are mostly indirect. Hogervorst et al. (2003) conducted experiments to show that high nitrogen levels in the litter favour certain species of fungi more than others and this would imply an altered community composition of the fungi-dependent pathway of the decomposer food-web. The studies cited above all demonstrate that soil pH is a major structuring factor of soil invertebrate communities. With sufficient knowledge on speciesspecific preferences, the composition of a soil invertebrate community in the field can be used to indicate the effects of acid precipitation. Despite the fact that soil acidification receives little priority in environmental policy at the moment, changes of nitrogen status and soil pH continue to have a major influence on soil communities and concomitant soil processes. 6.2.5 Pesticides

The scientific literature about pesticide effects on soil fauna is very large and has a long history. The older literature has been reviewed by Edwards and Thompson (1973) and a recent review was published by Van Straalen and Van

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Rijn (1998). Among the diversity of products to which soil invertebrates may be exposed, insecticides are the group with the largest effects, followed by fungicides. This is partly due to the presence of specific biochemical receptors (for example cholinesterase, ion channels) that are the targets of insecticide action and that are shared with the pest species. However, there are many examples showing that the precise exposure route is another, very critical factor determining the risk of insecticides to soil invertebrates. Invertebrates active on the soil surface, such a ground beetles, rove beetles, darkling beetles, linyphiid spiders, erigonid spiders and epigeic Collembola are particularly at risk. This point is illustrated in a paper by Jagers op Akkerhuis (1994) on the toxicity of deltamethrin to spiders (Oedothorax apicatus) in winter wheat. The author demonstrated a strong positive correlation between the activity of the spiders (modulated by climatic conditions) and the toxicity of deltamethrin, estimated from decrease of trapping success. The toxicity of the pesticide (principally a neurotoxicant) was greater under conditions where the mobility of the animals was high. This effect is actually due to a secondary action of deltamethrin on the water balance; the toxicant appears to induce an active excretion of water from the body. In principle, spiders can recover from the neurotoxic effect (immobilization), however, the loss of water soon becomes lethal, especially when air humidity is low (Jagers op Akkerhuis et al., 1995). Wiles and Jepson (1994) also recognized the idea of surface contact being a crucial factor for susceptibility to insecticides, when they developed a "susceptibility index", which combined an exposure function (based on walking track properties) with intrinsic sensitivity (derived from topical application). Sherrat and Jepson (1993) included the idea of mobility in a metapopulation model of ground beetles in a landscape of fields treated at intervals with insecticide. In a very elegant analysis, the authors showed that the long-term survival of the whole metapopulation is favoured by intermediate mobility: low mobility always leads to extinction because after a time every field has received a dose, and high mobility leads to extinction because this extends the mortality effect to beetles from other fields moving into a sprayed field. A recent study demonstrating the susceptibility of surface-living invertebrates to pesticides is Kennedy et al. (2001). Some of their results are reproduced in Figure 3. Treatment with dimethoate caused significant shortterm reductions in catches of ground beetles (Carabidae) and spiders (Linyphiidae), while pirimicarb had no significant effect on either group. The study also showed that small enclosed plots with pitfall traps placed centrally, should not be used as an alternative to large open plots, because more species

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Figure 3. Showing the effects of dimethoate and pirimicarb applications at recommended rates on pitfall catches of ground beetles (left) and linyphiid spiders (right) in winter wheat. Reproduced from Kennedy et al. (2001), with permission from the author. - The data are mean number of animals collected per trap (n) in large open plots (top) and small enclosed plots (bottom). SEM: standard deviation of the mean. Circles: control (sprayed with water), rectangles: pirimicarb treatment (140 g/ha), triangles: dimethoate treatment (336 g/ha). The treatment time is indicated by an arrow on the x-axis

axe caught in large open plots and catch variability is smaller. Many other studies, for example Asteraki et al. (1992), Purvis and Bannon (1992), Frampton and Cilgi (1992), Wiles and Jepson (1993) and Duffield and Aebischer (1994), have confirmed that carabid beetles and linyphiid spiders are sensitive nontarget groups for the effects of sprayed insecticides. Another group of soil invertebrates that can be considered relatively susceptible to insecticides are Collembola. Although old research (for example Van de Bund, 1965) had qualified Collembola as insensitive, because they increased in numbers after predatory arthropods (mesostigmatid and

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prostigmatid mites) had been suppressed by DDT, more recent research has shown that modern insecticides, specifically chlorpyrifos and dimethoate, can have drastic effects on Collembola communities, from which some species do not even recover within the same season (Joy and Chakravorty, 1991; Duffield and Aebischer, 1994; Frampton, 2000). Finally, earthworms may also suffer from insecticides. Some compounds are well known for their toxicity to earthworms, in particular carbofuran, endosulfan, parathion and carbaryl (Kula and Kokta, 1992; Reddy and Reddy, 1992; Reddy et al, 1997). Herbicides have also been tested for their effects on field communities of soil invertebrates, but direct toxicity, at recommended doses, of modern herbicides has not been documented clearly. Badejo et al. (1997) showed that the effects of soil cultivation on the catches of springtails and mites in pitfall traps were more important than those of a mixture of two herbicides, metobromuron and metolachlor. Several authors have noted the positive effects of herbicides on litter-inhabiting invertebrates stemming from an increase in dead vegetation cover on the ground. Cortet and Poinsot-Balaguer (1998) observed that atrazine had a positive, rather than a negative effect on Collembola colonizing maize litter bags. The apparent low toxicity of herbicides to arthropods does not imply that these compounds will have no effect in the field. An interesting phenomenon, already documented by Eijsackers (1978) in his work on 2,4,5-T, is that many herbicides seem to have a repellent action towards arthropods and decrease their consumption. This could imply that arthropods avoid herbicidetreated patches and that their positive effects on organic matter decomposition are not expressed. This idea is supported by recent studies. Brust et al. (1990) noted that atrazine and simazine were not toxic to carabid beetles, but did have a repellent action. Akinyemiju et al. (2000) explained the effects of hexazinone on soil micro arthropods from the repellent action of the herbicide and Ponge et al. (2002) noted clear avoidance behaviour of Collembola colonizing soils treated with isoproturon. A possible alarming effect of the herbicide atrazine is the feminization of male amphibians, which is due to a disruption of the metabolism of sex hormones (Hayes et al., 2002). However, such effects have not yet been observed in soil invertebrates. Fungicides are expected to disrupt the fungivore pathway of the soil foodweb by suppressing the basal resource. However, the effects of fungicides observed in the field are usually due to direct toxicity to animals. Several fungicides have been shown to be toxic to earthworms. Benomyl and carbendazim are the most obvious examples. In a classic study, Stringer and Lyons (1974) reported that repeated applications of benomyl in apple orchards led to an accumulation of leaf litter on the soil, and this effect appeared to be

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due to the suppression of earthworms. Several recent studies have confirmed the toxicity of benomyl, and its metabolite carbendazim to earthworms and enchytraeids (Holmstrup et al. 2000; Didden and Rombke, 2001). The effects seem to be specific to oligochaetes. Frampton and Wratten (2000) did not find an effect of carbendazim on Collembola, although these animals did suffer from two modern fungicides, propiconazole and triadimenol. Among the other fungicides, the organotin compounds deserve mention. Triphenyltin has been shown to be toxic to isopods in microcosms (Van Wensem et al., 1991), but recent field studies are missing. Summarizing the effects of pesticides on soil-living animals is very difficult. The account given above has demonstrated that the effects are contingent on the species and the way these are exposed. Species-specific metabolic pathways further add to the complicated relationship between pesticides and the invertebrate community. Therefore, community analysis is not expected to lead to an overall indicator system for pesticides. Rather, for each pesticide a targeted indicator species or species group should be selected. 6.2.6 Summary of pollutants effects

Table 1 provides an overview of the effects of various soil pollutants on invertebrate communities, as evidenced by field studies. The table also summarizes pollutants not discussed in the text above (references available from the author). It is clear that the picture is very complicated, although for some groups general patterns are emerging. These data should form the basis for bio-indicator systems, in which aspects of community composition are exploited to discover, estimate or evaluate the effects of soil pollutants. Such bio-indicator systems could then become part of ecological surveys conducted to evaluate contaminated soils. This question is discussed in the next section. 6.3 Evaluation of soil quality using invertebrates 6.3.1 Philosophy of bio-indication

As stated above, the basic idea of bio-indication is that the relationship between soil factors and soil communities can be reversed (Van Straalen, 1997). When soil factors influence community structure, the structure of a community must contain information on the soil factor. This concept of bio-indication is equivalent to what was called "calibration" by Ter Braak (1987). He saw calibration as the inverse of regression: regression equations are quantitative expressions between one or more environmental factors and community structure, while calibration (bio-indication) is a quantitative relationship

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Table 1. Overview of vulnerable groups among the soil animal community, as apparent from field studies Pollutant group Polycylic aromatic hydrocarbons, azaarenes and derivates Persistent organochlorines (PCBs, dioxins) Chlorinated ethylenes, phenoles and benzenes Oil, BTEX Alkyl benzene sulfonates and other detergents Veterinary drugs, antibiotics, hormones Copper

Zinc

Cadmium

Lead

Herbicides

Fungicides

Vulnerable animal groups Isopods, Collembola

Vertebrates (Rodentia and Insectivora) Earthworms

Earthworms Enchytraeids, nematodes, earthworms No data available

Earthworms, slugs, snails, oribatid mites Enchytraeids, nematodes, earthworms, isopods, soft-bodied springtails Oribatid mites, spiders, some springtails, vertebrates (shrews, mole) Oribatid mites, shrews, mole No group in particular

Earthworms, enchytraeids, isopods

Remarks Little knowledge available. Large inter-species differences in metabolism. Metabolizers expected to be more sensitive than accumulators. Low toxicity to invertebrates. Effects appear higher up in the food-chain Earthworms are important in transfer. Toxicity due to general narcotic effects, probably small inter-species differences.

Toxicity partly due to changes in soil structure. Field data scanty. Laboratory data suggest highest toxicity to pore water-dependent species. Interactions in decomposer-micro-organism interactions expected, but not documented. Copper toxicity to earthworms well documented. Toxicity of zinc does not follow the main taxonomic groups of soil invertebrates. Many groups contain sensitive as well as tolerant species. Cadmium seems to be most toxic to invertebrates that take up the metal with the food. Due to food-chain accumulation effects appear in predators and vertebrates. Differences between invertebrate species relatively small. Main hazard of lead is higher up in the food chain. Low toxicity of modern herbicides to animals. Effects are mostly secondary (avoidance of sprayed leaves, loss of food, increase of litter cover). Benzimidazoles, carbamates and organotins are known for their considerable side-effects on animals

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(Table 1 continued) Pollutant group Insecticides

Acidic precipitation

Radiation

Vulnerable animal groups Many arthropod groups, in particular beetles, spiders, mesostigmatid mites and springtails

Snails, dipteran larvae, earthworms, some oribatid mites, some Collembola, some isopods Earthworms, oribatid mites

Remarks

Animals with high surface activity are particularly vulnerable. Large differences between species due to species-specific exposures and metabolic capacities. Many secondary effects among detritivores due to suppression of predators. Large differences between species within each group. Earthworms generally avoid acid soils. Many Collembola and mites are acid tolerant, but some are very alkalophilic and suffer from acid precipitation. Species-specific vulnerability due to exposure, rather than inherent differences in sensitivity. Permanent soil dwellers and soil ingesters receive high doses.

Note: this table only describes the general trends and ignores the many species-specific sensitivities related to metabolism, microhabitat choice and life-cycle

between community structure and environmental factors. In some cases, it is even possible to estimate environmental factors quantitatively from community structure. This approach has been quite successful in reconstructing sediment pH from diatom community structure and palaeotemperature from insect remains (Ter Braak and Juggins, 1993). Bio-indicator systems are at their best when there is a clear, over-riding effect of an environmental factor and if this effect is specific and occurs at low intensity. Specificity and resolution are the two main properties on which bio-indicator systems should be evaluated (Van Straalen, 1998). That the structure of a community contains information about (partly hidden) ecological processes has been recognized for a long time by aquatic ecologists, who have developed extensive systems for bio-indication in which the information content of community structure is used to assess water quality (Metcalfe, 1989). The biotic index (De Pauw and Vanhooren, 1983) and the RIVPACS-system (Wright et al., 1989) are examples of bio-indicator approaches that have found their way to regulation. Such systems have not yet been developed for soil quality (see Pankhurst et al., 1997). What properties of community structure can be used to retrieve information

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about soil quality? The following is a listing of the various aspects proposed in the literature. Species richness and diversity. This is the most simple of all measures. The number of species can often be established easily, since it does not require a count of all the animals in a catch. For diversity indices such as the ShannonWiener index, species need to be counted and their proportion in the community should be established. The distribution of numbers over species. This is often pictured in a rank-

abundance graph, which is an expression of the balance between dominant and rare species (see, for example, Takeda, 1987 and Vegter et al., 1988). The distribution of body-size over species (Warwick, 1986). In this analysis, the

species are not only characterized by their numbers, but also by their body size. This analysis expresses the degree to which the community consists of small, dominant species versus large, rare species. Classification of species according to life-history attributes. Traditionally the r-

and K-selection approach has been most popular (Bouche, 1977; Satchell, 1980; Greenslade and Greenslade, 1987; Eijsackers, 1994; Andersen, 1995). For mesofauna, Siepel (1994, 1996) has proposed a classification system that is based on life-history tactics in three dimensions. The maturity index for nematodes, as developed by Bongers (1990) is another example of this approach. Classification of species according to ecophysiological preferences. In this system

the preferences of species in a community are used to draw conclusions about a specific environmental factor. The idea is that the environmental factor is indicated by the overlap of species preferences. The only soil factor for which this has been developed is pH (Van Straalen and Verhoef, 1997). The structure of the food-web. This can be expressed in various topological indices, such as the number of trophic levels and the "connectedness" of the food-web (Pimm et al., 1991). When complemented by energy-flow information, additional indices may be derived, such as the pattern of interaction strengths (De Ruiter et al., 1995). These parameters may indicate the degree of top-down versus bottom-up regulation in the community. 6.3.2 Proposed bio-indicator systems

Employing one or more of the above listed aspects of community structure, the number of proposed "bio-indicator systems" using soil invertebrates is extensive. Table 2 provides an overview of the various proposals. Below I highlight some of the most promising approaches.

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Table 2. Overview of soil invertebrate bio-indicator systems proposed in the literature Indicator system Nematode maturity index

Principle Nematodes classified on a "colonizer" "persister" scale

Predatory mite maturity index

Mesostigmatid mites classified according to an r-K score

Earthworm life-history strategies

REAL model for earthworms

Enchytraeid Reaktionzahl

Application Can be applied to all soils; measures general response to stress (metals, acidification, eutrophication) Mostly limited to forest soils; measures soil properties related to mull/mor humus

Reference Bongers (1990), Yeates and Bongers (1999)

Earthworms classified according to position in the soil profile and burrowing behaviour

Can be applied to all soils with sufficient number of species; measures aspects of humus type, pH and cultivation (ploughing)

Bouche (1977), Paoletti (1999a)

Integrated data base of various aspects related to the ecological and agronomical role of earthworms Scores related to responses to acidity and humidity assigned to enchytraeids

Very wide application

Bouche (1996)

Applicable to situations where effects on soil pH are manifested, for example cement factories Data base on speciesspecific responses not yet operational; at the moment only applied to heavy metal pollution

Graefe (1993), Beylich et al. (1995)

Composition of isopod fauna indicates effects of soil cultivation in agricultural landscapes

Paoletti and Hassell (1999)

SIVPACS

Pollution responses of earthworms, isopods and spiders, comparable to RIVPACS

Woodlice life-forms

Classification of woodlice according to body shape and movement pattern

Ruf (1998)

Spurgeon et al. (1996)

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(Table 2 continued) Indicator system Macro invertebrate biodiversity

Ant functional groups

Diptera feeding groups

Arthropod acidity index

Oribatid mite lifehistory strategies

Life-forms of Collembola

Dominance distribution of micro arthropods

Biological Index of Soil Quality (BSQ)

Principle Enumeration of species richness of earthworms, beetles, isopods, spiders, ants, millipedes, centipedes, etc. Classification of ants according to groups reflecting susceptibility to stress Classification of dipteran larvae in five feeding groups

Application Reference Applied in orchards j Paoletti and j and other agricultural Somaggio (1996), j ecosystems to indicate Paoletti (1999b) j land use and copper 1 pollution | | Wide application; Andersen (1995) | used in evaluation of | nature restoration and effects of mining Reflects type of Frouz (1999) organic materials in soil; applicable to organic soils Classification of Allows quantitative Van Straalen and arthropods (Collemestimation of soil pH Verhoef (1997), bola, oribatids, from invertebrate Van Straalen (1998) isopods) according to community structure pH preference Classification of mites | Indicates intensity of | Siepel (1994), Siepel according to | anthropogenic | (1996) reproductive and | influence and | dispersal strategies | successional stage of | | forests and grassland | | ecosystems | Classification of I Indicates profile I Van Straalen et al. Collembola according | build-up and (1985), Faber (1991) to morphological | ecological processes | types reflecting | stratified according to | position in the soil I the profile; mostly I profile applicable to forest soils Lognormal General impression of Hagvar (1994) distribution of disturbance; applied numbers over species to effects of heavy metals and acid rain in forest and grassland soils System of scores Provides indication of Parisi (2001), Gardi assigned to groups of biodiversity; wide et al. (2002) soil micro arthropods applicability

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The best-known system is the nematode maturity index, developed by Bongers (1990). This index is based on a classification of nematodes according to "colonizers" and "persisters". Colonizers are opportunistic species, characteristic for early successional stages of a soil, while persisters are species that arrive in a later successional stage. The division is more or less similar to the famous r/K dichotomy. Colonizers are species with an opportunistic feeding habit, a short life-cycle and a high reproductive effort. Persisters are species with low reproduction and low dispersal capacity. Each group of nematodes is assigned a score, varying from 1 to 5, and the index (M) is calculated as: n

i-i

where vi is the colonizer-persister score (1 - 5) of nematode i (i = l...s), and pi is the fraction of i in the community. In nematodes, the life-history strategy is more or less similar within a family. To assign a colonizer-persister ("c-p") score to a nematode, it is not necessary to identify it to species; instead, the family level already provides enough resolution. This is a particularly strong point of the index. The nematode maturity index has proven useful in a wide variety of applications. Freckman and Ettema (1993) showed that nematode communities of agro-ecosystems that differed in the degree of human intervention could very well be discerned using the maturity index. A similar conclusion was reached by Hodda and Wanless (1994) who investigated two contrasting grasslands. Korthals (1996a, b) showed that nematode community structure responded to copper pollution. The effect of copper could be summarized very well as a decrease of the maturity index. De Goede et al. (1993) developed a graphical interpretation of nematode community structure by plotting the three most abundant groups of nematodes (c-p 1, c-p 2, and c-p 3 to 5) in a triangle ("c-p triangle"). An example, showing community changes during succession of grasslands, is given in Figure 4. Reviews of the nematode maturity index are given in Yeates and Bongers (1999) and Bongers and Ferris (1999). The index proves to have a very wide applicability. Perhaps the only real objection to the index is that it responds to virtually every soil factor. It is, thus, a typical general index providing a clear indication of stress, but not indicating the cause. The second relatively well-developed system for indicating soil quality is the ecological classification of earthworms. Many aspects of the ecology and physiology of earthworms correlate with their size and burrowing activities.

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Figure 4. Community structure of nematodes can be visualised in a triangle showing the relative densities of colonizer-persister ("c-p") groups. Reproduced from De Goede et al (1993), with permission from the faculty of Agriculture and Applied Biological Sciences of Ghent University, and the author. - The proportions of c-p groups 1, 2 and 3 to 5 are plotted in this diagram to indicate the state of the nematode community. - De Goede (1993) investigated the effect of disturbances, indicated by the large Arrows 1 and 2. For example, in Arrow 1, the proportion of c-p 1 increases and c-p 3-5 decreases due to increased nutrient supply (more colonizers, less persisters). Line 3 indicates the trajectory of succession and return to the initial condition. The diagram was inspired by community changes observed in grassland soils.

The classification of Bouche (1977) takes this into account by dividing the community into (1) surface-active, pigmented and non-burrowing species (epigees), (2) species that construct branching, horizontal burrows and live in surface mineral and organic horizons (endogees), and (3) species that make deep burrows and come to the surface at night to draw down litter (aneciques). The epigees are best represented in soils where there is an accumulation of organic matter on the surface, such as woodland and grassland. These worms have a high population turnover, suffer from predation and produce a high numbers of cocoons. The endogees dominate the biomass in temperate grasslands: They produce an intermediate number of cocoons. The aneciques are also common grassland species: They produce the lowest number of cocoons of all and they prefer to feed on fresh leaves from the surface, but if these are not available, they can also live on the dead organic matter in the soil,

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like the endogees. The proportions of the three groups of earthworms provide a simple but logical system for bio-indication. Given the important ecological role of earthworms as organic matter processors, changes in dominance of earthworm ecological groups can be directly translated into changes in the ecological functioning of the soil. The use of earthworms as bio-indicators of soil fertility is, therefore, generally accepted, although some investigators have made critical notes (Doube and Schmidt 1997) arguing that the positive effects of earthworms on crop productivity remain to be demonstrated. The third bio-indicator system to be highlighted is the species richness of micro arthropods. Parisi (2001) developed an index called QBS (Qualita Biologica del Suolo). The index is derived by assigning scores to each micro arthropod group (mostly orders such as Collembola, Microcoryphia, Orthoptera, Coleoptera, Psocoptera, etc.) and adding the scores for those groups present in a soil. For some groups a subdivision of scores is developed, based on the life-form ("ecomorphological index", EMI). The QBS thus combines elements from Gisin's life-forms (Van Straalen et al., 1985) with the biodiversity-survey system proposed by Paoletti and Bressan (1996). Gardi et al. (2002) applied the system developed by Parisi (2001) to evaluate soil quality in permanent grasslands, contrasted with agricultural soil. An objection to the QBS is that it leaves out the most species-rich group of micro arthropods, oribatid mites. Table 2 lists a great many more systems, some of them similar to the ones discussed above. Some of the systems have only been applied once or twice and none of them has achieved the scale of application valid for the nematode maturity index. 6.4 Ecological surveys in the soil quality diamond Invertebrate surveys are not yet a standard item in soil quality evaluations. The present overview has shown that there is, nevertheless, a sufficient scientific background to change this situation. Already in the 1970s, indicator systems have been proposed and the necessary knowledge has increased considerably in the last ten years. In sediment evaluations, a "triad" approach has already been practiced for a long time (Chapman, 1990). The sediment quality triad is based on three elements: - chemical measurements; - ecological surveys; - bioassays.

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Similar approaches have been proposed for soil evaluation by Indeherberg et al. (1998) and Rutgers et al. (2000). The various systems reviewed in this paper all focus on some form of ecological survey. It is important to note, however, that ecological surveys should not be isolated from the other two elements of a soil evaluation: chemical measurement (including not only the total concentration of a contaminant, but also soil properties affecting bioavailability), and bioassays (laboratory-based tests using soils from the site). Only in combination with these three elements can ecological surveys play a role. This is pictured in Figure 5 as the "soil quality diamond", which is inspired by the soil quality triad, as proposed by Rutgers et al. (2000). The question arises, which type of ecological survey, among the various systems reviewed in this paper, would be best suited to fulfil this role? The view of this author is that it should be a combination of (1) the nematode maturity index, (2) earthworm ecological classification, (3) micro arthropod species richness. This collection of systems covers most effects of pollutants as listed in Table 1. It is also the best choice among the various indicators listed in Table 2, from the viewpoint of applicability and widespread acceptance.

Figure 5. The "soil quality diamond": soil assessment consists of three elements, chemical analysis, bioassays and ecological surveys. - The latter involves assessment of three indicator groups of soil invertebrates, nematodes (maturity index), earthworms (biomass of epigees, endogees and aneciques) and micro arthropods (number of species ofCollembola and Oribatida). - Comparison of these survey outcomes with a system of reference values will indicate whether the site poses any reason for concern

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The maturity index can be applied without further development, since experience is already extensive. The situation is comparable for earthworm surveys. For micro arthropods some further development is still necessary. The system of Parisi (2001), although attractive, cannot be applied without further modification, since it ignores the species richness of the groups surveyed and it does not exploit the large biodiversity of mites. Rather than giving each group an arbitrary score, it seems better to focus on species richness itself. Species counts of Collembola and oribatid mites in Tullgren-extracted samples could provide a simple measuring rod for biodiversity. Species richness also has the great advantage that only the number of species need to be established, not the actual abundance of each species. Further work needs to be done to establish a protocol allowing a reproducible estimation of species richness of micro arthropods for a series of ecosystem types. Ecological surveys can only be used meaningfully in soil evaluations if the results can be compared to a reference system (for example, how many species of micro arthropod are minimally expected in a normal soil?). At the moment, such a reference system is lacking. However, there is a lot of information in the open scientific literature, in the "grey" literature and in the heads of soil investigators that remains to be exploited and made explicit. Pilot projects in the Netherlands (Sinnige et al., 1992; Schouten et al., 1999, 2000; Bosveld et al., 2003) and in Germany (Ruf et al., 1999; Rombke et al., 1999; Ruf et al., 2003) have taken the first steps in this direction. A much greater effort is needed. It is expected that a reference system needs to be specified for certain classes of ecosystems, for example "pine forest", "grassland", "barren industrial site", etc., because vegetation cover and soil properties have a major influence on the species composition of a soil. Once a reference system is agreed upon, soil quality evaluations can be placed on a firm ecological basis and the position of ecological arguments in decisions about land management will finally improve. 6.5 Implementation in soil management Bio-indicator systems based on invertebrate soil animals can be used to indicate various aspects of soil quality. A review of the scientific literature shows that invertebrates are sensitive to almost all soil pollutants, although there are sometimes large differences between the groups and, in many cases, also within groups. A great variety of bio-indicator systems of soil invertebrates have been proposed in the literature. Among these, the maturity index of nematodes, ecological classifications of earthworms, and species richness of micro arthropods are the best candidates to be taken up in a general system of ecological survey. This collection of systems covers most of the pollutant effects

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documented in field studies. Ecological surveys are one of the three elements in the "soil quality diamond", which involves chemical analysis and bioassays as well as ecological surveys. More information, some of which already exists, is needed to develop a reference system. It is expected that an ecological survey equipped with a system-specific reference system will greatly improve the position of ecological arguments in land management. Acknowledgements Many thanks are due to Peter Doelman, Herman Eijsackers and Kees van Gestel for stimulating comments. References Abdul Rida, A.M.M. and M.B. Bouche, 1995. The eradication of an earthworm genus by heavy metals in southern France. Appl. Soil Ecol. 2, 45-52. Achazi, R.K. and C.A.M. van Gestel, 2003. Uptake and accumulation of PAHs by terrestrial invertebrates. In: PAHs: An Ecotoxicological Perspective (ed. P.E.T. Douben). John Wiley & Sons Ltd, Chichester, pp. 173-190. Akinyemiju, O.A., M.A. Badejo and A. Oyeniyi, 2000. The persistence of hexazinone and its influence on soil micro-arthropods in a humid tropical environment. In: Pollutants and Their Effects on Terrestrial and Aquatic Ecosystems (eds. M.A. Badejo and N.M. van Straalen). College Press Ltd./Enproct Consultants, Ibadan/Lagos, pp. 84-98. Andersen, A.N., 1995. A classification of Australian ant communities, based on functional groups which parallel plant life-forms in relation to stress and disturbance. J. Biogeogr., 22,15-29. Anderson, J.M., 1975. The enigma of soil animal species diversity. In: Progress in Soil Zoology (ed. J. Vanek). Junk B.V., The Hague, pp. 51-58. Asteraki, E.J., C.B. Hanks and R.O. Clements, 1992. The impact of two insecticides on predatory ground beetles (Carabidae) in newly-sown grass. Ann. Appl. Biol., 120, 25-39. Badejo, M.A., J.L. Olaifa and N.M. van Straalen, 1997. Effect of Galex on the Collembola fauna of cowpea plots. Pedobiologia, 41, 514-520. 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. Belotti, E., 1998. Assessment of a soil quality criterion by means of a field survey. Appl. Soil Ecol. 10, 51-63. Bengtsson, G. and L. Tranvik, 1989. Critical metal concentrations for forest soil invertebrates. A review of the limitations. Water, Air Soil Pollut. 47, 381-417. Beylich, A., H.-C. Friind and U. Graefe, 1995. Environmental monitoring of ecosystems and bioindication by means of decomposer communities. Newsletter on Enchytraeidae, 4, 25-34. Blakely, J.K., D.A. Neher and A.L. Spongberg, 2002. Soil invertebrate and microbial communities, and decomposition as indicators of polycyclic aromatic hydrocarbon contamination. Appl. Soil Ecol. 21, 71-88. Bongers, T., 1990. The maturity index: an ecological measure of environmental disturbance based on

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The use of soil invertebrates in ecological surveys of contaminated soils need for a soil invertebrate prediction and classification scheme (SIVPACS). In: Bio-indicator Systems for Soil Pollution (eds. N.M. Van Straalen and D.A. Krivolutsky). Kluwer Academic Publishers, Dordrecht, pp. 95-110. 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 smelting works. J. Appl. Ecol., 36,173-183. Stroomberg, G.J., F. Ariese, C.A.M. van Gestel, B. van Hattum, N.H. Velthorst and N.M. van Straalen, 2003. Pyrene biotransformation products as biomarkers of polycyclic aromatic hydrocarbon exposure in terrestrial Isopoda: concentration-response relationship, and field study in a contaminated forest. Environ. Toxicol. Chem., 22(1), 224-231. Stringer, A. and C.H. Lyons, 1974. The effect of benomyl and thiophanate-methyl on earthworm populations in apple orchards. Pesticide Sci., 5, 189-196. Sverdrup, L.E., P.H. Krogh, T. Nielsen and J. Stenersen, 2002. Relative sensitivity of three terrestrial invertebrate tests to polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem., 21(9), 1927-1933. Swift, M.J., O.W. Heal and J.M. Anderson, 1979. Decomposition in Terrestrial Ecosystems. Blackwell, Oxford. Takeda, H., 1987. Dynamics and maintenance of collembolan community structure in a forest soil system. Res. Population Ecol., 29, 291-346. Ter Braak, C.J.F., 1987. Calibration. In: Data Analysis in Community and Landscape Ecology (eds. R.H.G. Jongman, C.J.F. ter Braak and O.F.R. van Tongeren). Pudoc, Wageningen, pp. 78-90. Ter Braak, C.J.F. and S. Juggins, 1993. Weighted averaging partial least squares regression (WAPLS): an improved method for reconstructing environmental variables from species assemblages. Hydrobiologia, 269/270, 485-502. Turnhout, E., 2003. Ecological indicators in Dutch nature conservation. Science and policy intertwined in the classification and evaluation of nature. Ph.D. thesis, Vrije Universiteit, Amsterdam. Van Brummelen, T.C. and N.M. van Straalen, 1996. Uptake and elimination of benzo(a)pyrene in the terrestrial isopod Porcellio scaber. Archives Environ. Contam. Toxicol., 31, 277-285. Van Brummelen, T.C, R.A. Verweij, S.A. Wedzinga and C.A.M. van Gestel, 1996. Polycyclic aromatic hydrocarbons in earthworms and isopods from contaminated forest soils. Chemosphere, 32, 315-341. Van Brummelen, T.C, B. van Hattum, T. Crommentuijn and D.F. Kalf, 1998. Bioavailability and ecotoxicity of PAHs. In: The Handbook of Environmental Chemistry, Vol. 3, Part J: PAHs and Related Compounds (ed. A.H. Neilson). Springer-Verlag, Berlin, pp. 203-263. Van de Bund, C, 1965. Changes in the soil fauna caused by the application of insecticides. Bolletino di Zoologia agraria e di Bachicoltura, Serie II, 7, 185-212. Van Gestel, C.A.M., 1997. Scientific basis for extrapolating results from soil ecotoxicity tests to field conditions and the use of bioassays. In: Ecological Risk Assessment of Contaminants in Soil (eds. N.M. van Straalen and H. Lokke). Chapman & Hall, London, pp. 25-50. Van Gestel, C.A.M., J.J. van der Waarde, J.G.M. Derksen, E.E. van der Hoek, M.F.X.W. Veul, S. Bouwens, B. Rusch, R. Kronenburg and G.N.M. Stokman, 2001. The use of acute and chronic bioassays to determine the ecological risk and bioremediation efficiency of oil-polluted soils. Environ. Toxicol. Chem., 20, 1438-1449.

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The use of soil invertebrates in ecological surveys of contaminated soils Zaitsev, A.S. and N.M. van Straalen, 2001. Species diversity and metal accumulation in oribatid mites (Acari, Oribatida) of forests affected by a metallurgical plant. Pedobiologia, 45, 467-479.

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Chapter 7

BALANCE AND STABILITY IN VITAL SOILS P. de Ruiter

Abstract

Soil vitality is the ability of soil ecosystems to stay in balance internally in a changing world. The author develops a characterization concept based on the concept of ideal soil fertility, in relation to soil quality management. The soil community is inextricably related to soil processes such as the decomposition of organic matter and the mineralization of nutrients. This implies that balanced process rates depend on balanced population and balanced community dynamics. These balances are indicated by trophic biomass pyramids that generate stabilising patterns of trophic interaction strength. The levels at which balances are maintained will vary among the different kinds of land use. Stability is needed, since it implies that the organic matter and nutrient equilibria are maintained during environmental change or disturbance. Diagrams of below ground food webs serve as representative examples. The species are aggregated into functional groups, based on food choice and life-history groups. Material flow calculations and stability analyses need data on materials, energy, nutrients, growth and decay rates in soil organisms and biodiversity, soil community structure and ecological soil processes and community food webs. For the biotic part it implies determining species into groups that share the same prey and predators, quantifying the biomass of the functional groups and quantifying the feeding rates among the functional rates. In principle the knowledge needed to adequately measure and monitor soil vitality, restricted to nutrients, organic matter and trophic pyramids such as food web communities is available. The approach 'learning by doing' is recommended.

7.1 Introduction: A System Integration Approach to Soil Vitality 7.2.2 Soil quality and environmental change The increasing awareness of environmental problems has initiated worldwide research programmes to answer questions relating to how environmental stress-factors, that are caused by human activities, alter the structure of ecological communities and the functioning of ecosystems. Issues that inspire ecologists include climate change and the pollution of water, air and soils. The

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general objective of these research programmes is to contribute to environmental policy and the design of management options in order to improve environmental quality. The present chapter focuses on the issue of sustainable soil quality management. Sustainable preservation and management of soil quality is important for the various functions of the soil, such as its agricultural production function, its habitat function for soil organisms and natural vegetation, and its buffer function as a source and sink for materials, energy and nutrients. Soil quality management is also important for the quality of other ecosystem compartments. For example, a high soil nutrient status may form a threat to the quality of surface and groundwater through the leaching of nitrate and phosphate, and to the air through emission of ammonium and greenhouse gases. Aspects of soil quality embrace physical, chemical and biological properties. Physical aspects include soil structure, aeration, and hydrology. Chemical aspects include soil organic matter contents and composition, nutrient status, pH and contamination. Biological aspects include the presence, abundance and activity of soil organisms. Some physical and chemical processes are the direct result of the activity of soil organisms, such as soil structure formation, respiration, mineralization, natural attenuation and organic matter decomposition. This chapter will focus on why and how all these aspects should make up soil quality. I will argue that including soil ecology may form the step from soil fertility to soil vitality. My argument will mainly be a scientific one. Yet, an attempt will be made to look at vitality in the context of practical soil management. 7.2.2 From soil fertility to soil vitality

Traditionally, soil fertility includes all the above mentioned aspects of soil quality, physical, chemical and ecological. The central position in our thinking of soil fertility is the concept of ideal soil fertility. This ideal soil fertility is described in terms of balance, i.e. the balance between the rates with which nutrients become available and nutrient uptake by the plants and the balance in the build up and break down of soil organic matter and the various compounds within it (Janssen, 1999). This balanced fertility is considered to be ideal because soil ecosystem function, for example, under agricultural practice, will not lead to unwanted fluxes of compounds and nutrients to the environment. Obviously, the ideal fertility will never be reached. Therefore, it is also referred to as target fertility (Janssen, 1999). The concept of ideal or target fertility is

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suitable for agricultural soils as well as for soils in natural ecosystems. The levels at which the balances are maintained will, however, vary among the different kinds of land use. Hence, there is no explicit ecological component in the definition of ideal fertility. One can argue though that the soil community is inextricably related to the soil processes, especially the decomposition of organic matter and the mineralization of nutrients. This argument implies that balanced process rates indicate balanced population and community dynamics. There are, however, two arguments for explicitly including soil ecology in the concept of ideal fertility. The first is that despite balanced nutrient flows and constant soil organic matter contents, there might still be an out-of-balance situation regarding the microbial decomposition of (particular) components in the soil organic matter, which is obscured within the overall process of organic matter decomposition. Such a hidden out-of-balance may cause changes in the organic matter and nutrient dynamics in the long term. A second, more principle argument is the need for including ecosystem stability in the definition of soil quality. Stability implies that the organic matter and nutrient equilibria are maintained during environmental change or disturbance. As soil organic matter dynamics and components within the cycling of nutrients are biological processes, the stability of these processes is directly linked to the stability of the soil community. Given the fact that ecosystems undergo changes that can be natural as well as human induced, the stability of balance is important to the sustainability of soil quality. Note that stability does not necessarily imply constancy. A soil may undergo changes along with environmental change in terms of new balances at the appropriate level. In this chapter, including ecosystem stability in the definition of soil quality forms the step from fertility to vitality. Soil vitality is, therefore, defined as 'the ability of the soil ecosystems to stay in balance in a changing world'. 7.2.3 Soil ecology and the assessment of soil vitality

Soil ecology, as a sub-discipline of the soil sciences, has developed along the three 'functions' of the soil: the agricultural production function, the buffer function and the habitat function. Originally, soil ecology has strongly focused on agricultural soil fertility, for example by looking at biological processes that influence soil fertility. This focus in soil ecology is apparent in various large research programs directed to the development of ecologically sustainable agricultural management practices (Hendrix et al., 1986; Brussaard et al., 1988; Andren et al., 1990; Brussaard et al., 1990; De Ruiter et al., 1993a). In the eighties, the buffer function attracted the attention of soil ecologists. The issue

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of climate change initiated studies on the role of soil organisms in the carbon source and sink function of the soil (Wolters et al., 2000). The increasing number of cases of serious soil contamination initiated the field of soil-ecotoxicology. At first, soil ecotoxicology focused on the adverse effects of soil contamination on single individuals or populations, but more recently attention has shifted towards the assessment of contamination effects on the level of community structure and ecosystem functioning (Baird et al., 2001). In order to come up with cost-effective strategies to clean up contaminated soils, soil ecologists have also studied processes like natural degradation and the attenuation of organic contaminants. Most recently, soil ecologists have begun to look at the role of the soil biota in processes such as the development and preservation of natural ecosystems, especially in processes of land use change (De Deyn et al., 2003). Soil ecological research has generated an impressive amount of knowledge and understanding about the way soil ecosystems function, and how soil ecosystems respond to environmental change and disturbance. Still, it is largely unclear how ecological parameters indicate soil quality. For example, soil community structure and the composition of heavily contaminated soil do not seem to be principally different from non-contaminated soils. At the same time, soils that have the same type and land use for decades still show very different community structures. Sometimes soil community structure and composition merely reflect the long-term history of the soils. For example, soils in Dutch reclaimed areas still show strong signs of their marine history. These confusing observations hamper translating scientific knowledge to practical measures and management options. In fact, assessing soil quality by means of comparing ecological quality indicators is almost impossible because of the lack of control, i.e. a common understanding of what soil community is to be expected in 'normal', undisturbed, soils. In this chapter, an attempt is made to see whether a system integration approach can be helpful in analysing soil vitality. This system integration approach is characterized by combining various ecological aspects of soil quality into ecosystem properties, instead of comparing single ecological aspects across different soils. The ecosystem properties include information regarding community structure as well as ecological process rates. The central concept in the present system approach is the definition of soil vitality in terms of balance and stability. A balance between input and output of materials, energy and nutrients, a balance between growth and decay rates in populations of soil organisms and hence in soil biodiversity, and a balance between soil community structure and ecological soil processes. It is especially this latter balance that is shown to be important to ecosystem stability.

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The system integration approach is based on the analysis of the structure of the soil community food web. There are two arguments to studying food webs in the context of soil vitality. First, food web interactions determine the availability of food and the avoidance of predation and, therefore, also determine the stability of the soil community. Second, food web interactions form the basis of soil processes, such as the decomposition of organic matter and the mineralization of nutrients (Moore and Hunt, 1988, De Ruiter et al., 1998). In this way food webs link community stability to process stability. In this chapter, soil vitality is analysed by looking at balance and stability in real and model food webs. It will be shown that apart from balance and stability in the nutrient, organic matter and population dynamics, there is also the balance between food web structure and stability. This means that food web structure may serve as a stability indicator. This will be illustrated in the next section by means of a simple model food chain. Next, the principle of balance and stability will be shown in real soil ecosystems by analyzing stabilising patterns in community structure and process rates, together with the biological mechanism behind these patterns. The chapter will end with a discussion of how the concept of vitality in soils can be used to develop management practices that should restore, preserve and ensure sustainable soil ecosystems. 7.2 The principle: Balance and Stability in Model Food Chains Why should balance be linked to stability? The principle of balance and stability can best be illustrated by means of a simple model food chain (Moore et al., 1993; Moore and De Ruiter, 2000). Consider a food chain of length 3 in which the dynamics of the trophic groups are modelled by Lotka-Volterra differential equations. The food chain is modelled using Lotka-Volterra equations:

X^X^+^X,] j=l

Productivity is defined as biXf, where X;* denotes the equilibrium biomass of the basal primary producers. Productivity levels are modelled by choosing values for bt for the basal species. Parameters are chosen to be energetically feasible and near values known from literature; the values of productivity is expressed in gir^yr1. The rate with which the basic resource is made availability to the food chain is named 'productivity'. With this model we can analyze the interrelationship between an ecosystem process, i.e. productivity, and community structure, i.e.

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food chain length. When productivity is gradually increased from low to intermediate levels, the model predicts that the length of the food chain is governed by the level of productivity, in that there should be enough energy for the subsequent trophic levels (Figure la). It should be kept in mind that the energy transfer between two levels has an efficiency of approximately 10%. At low productivity levels there is only enough energy for one or two trophic levels (Figure la, open symbols). At high levels of productivity there is enough energy to support all three trophic levels. However, when productivity is increased to further higher levels, the dynamics of the populations in the food chain show large oscillations that destabilize the food chain (Moore and De Ruiter, 2000). Together, these two effects lead to a 'hump-shaped' curve describing the relationship between productivity and food chain stability (Figure la, closed symbols). If we look at the model results in terms of population size distributions, we see that at low to intermediate productivity levels, population sizes are organized in the form of a trophic pyramid with decreasing biomass over trophic levels (Figure lb). At high productivity levels, the pyramidal biomass structure disappears and eventually turns into an inverse biomass pyramid. These inverse pyramids are accompanied by large destabilizing oscillations (Moore and De Ruiter, 2000). When the food chain model allows that a new, higher trophic level group enters the food chain as soon as there is enough energy to support this higher trophic level, the model predicts that the population size distributions will take a pyramidal shape and destabilizing oscillations disappear. Hence, the food chain maintains a relatively high level of stability as long as its length is in balance with the level of productivity. 7.3 Real soil ecosystems: Balance and Stability in Complex Food Webs Does the principle of balance and stability also operate in real soil ecosystems? Soil communities look far from simple food chains. Soils harbour a large part of the world's biodiversity (Wolters, 1997; Griffiths et al, 2000), organized in complex food webs. Real soil food webs derive their principle energy sources from various kinds of soil organic matter and root derived materials. The first trophic level in soil food webs consists of organisms that feed on these primary sources. At this level we find the micro-organisms, such as bacteria and fungi, and herbivorous nematodes. The microbes are by far the most dominant groups of soil organisms, in terms of numbers as well as biomass (Andren et al., 1990; Bloem et al., 1994). At the higher trophic levels, soil food webs generally contain a large variety of faunal organisms, like protozoa (amoebae, flagellates, ciliates), nematodes (bacterivores, fungivores, omnivores, herbivores and predators),

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Figure 1. 77ze influence of productivity on the length and stability of a food chain (Moore et al., 1993; Moore and De Ruiter, 2000). a. The influence of productivity on the food chain feasibility (solid symbols) and food chain stability (open symbols). Feasibility requires that in equilibrium all populations can persist, i.e. Xi*>0 (Moore et al, 1993). The decreasing part of the hump-shaped curve of stability is caused by the emergence of large destabilizing oscillation (Moore and De Ruiter, 2000). The grey symbols indicate the effect of the introduction of an extra trophic level on food chain stability. b. The effect of productivity on the organization of population sizes over trophic level. At relatively low productivity, population sizes are organized in the form of trophic pyramids. At high levels of productivity the pyramidal structure disappears and changes into inverse pyramidal structures. The grey arrows indicated the effect of the introduction of an extra trophic level on population size distribution

micro arthropods such as mites (bacterivores, fungivores, predators) and collembolans (fungivores and predators), enchytraeids and earthworms (Figure 2). Just as real soil food webs are much more complex than simple food chains, real soil processes form a much more complex network of processing and

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Figure 2. Diagram of the below ground food web from a farming system of the Lovinkhoeve experimental farm (NL) (Moore and De Ruiter, 1991; De Ruiter et ah, 1993a). Diagram serves as a representative sample. Species are aggregated into functional groups, i.e. based on food choice and lifehistory parameters groups (Moore et ah, 1988). Detritus refers to all dead organic material. - Material flows to the detritus pool through the death rates and the excretion of waste products are not represented in the diagrams, but are taken into account in the material flow calculations and stability analyses. - The trapeziums with numbers illustrate the trophic pyramid in biomass (kg C haA, 0-25 cm depth layer); the arrows with numbers illustrate the trophic pyramid in feeding rates (kg C ha^yr1, 0-25 cm depth layer)

recycling pathways than the one single process of productivity. This implies that when we look for a balance between soil community structure and soil processes, we have to link two complex structures: that of the soil community food web and that of the soil biogeochemistry. The linking of two complex

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structures asks for a simplification. Such simplification can be obtained by adopting the functional approach in defining and describing food web structure and ecosystem processes. Through the functional approach it becomes possible to achieve an explicit correspondence between the two structures, because both community structure and ecosystem processes are then described in the same 'currencies', for example kg (C), ha, yr. A first step in the functional approach is that (taxonomic) species are aggregated into groups embracing all species sharing the same prey and the same predators (Moore et al., 1988). The second step is to quantify the biomass of the functional groups, for example in kg C ha 1 . The third step is to quantify the feeding rates among the functional groups, for example in kg C ha 4 yr-1. Quantification of feeding rates can be done by means of food web modelling, using available information regarding population sizes, turn-over rates and energy conversion parameters (O'Neill, 1969; Hunt et al., 1987; De Ruiter et al., 1993b). This procedure creates a link between community and process, as the feeding rates among the functional groups represent specific components in the cycling of matter, energy and nutrients. By doing so for a series of real soil food webs, it has been found that population sizes as well as feeding rates are organized in the form of trophic pyramids (De Ruiter et al., 1993a, De Ruiter et al., 1998). In the model food chain, the single process productivity was analyzed from the perspective of food chain length and stability and it was shown that biomass pyramids indicate a stable balance between productivity and food chain length (Figure 1). The question here is whether the observed trophic pyramids in the real food webs - in population sizes as well as feeding rates are also indicators of balance and stability. This question is approached by looking more closely at the strength of the interactions among the populations. Interaction strengths are the per capita effects (in the case of the present energy flow webs per biomass effects) of the functional groups upon one another (Figure 3) and determine the role of the trophic interactions in the dynamics and persistence of the populations constituting the food web, and hence of the food web as a whole. Interaction strengths can directly be derived from the biomass of the functional groups and the feeding rates among the groups (Figure 3). Interaction strengths (yr1) are derived from feeding rates, population sizes and energy conversion efficiencies. Negative effects are feeding rates divided by the population size of the consumer. Positive effects are the production rates (i.e. feeding rates times energy conversion efficiency) divided by the population size of the resource (De Ruiter et al., 1995). Note that the negative effects are much

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Figure 3. Patterns in the interaction strengths along trophic level in the food web of a farming system at the Lovinkhoeve experimental farm (NL)

(two orders of magnitude) larger than the positive effects. By doing so, it appeared that the interaction strengths values are organised in trophic level dependent patterns, but in a different manner to population sizes and feeding rates, i.e. the patterning in the interaction strengths is characterized by relatively strong top-down effects at the lower trophic levels and relatively strong bottom-up effects at the higher trophic levels (Figure 3). Interestingly, this patterning in the interaction strengths is found to have a strong enhancing effect on food web stability (De Ruiter et al., 1995; Moore et al., 1996; De Ruiter et al., 1998). This finding supports the idea that the stability of the food webs are also the result of a balance between process and community structure, as it is

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the trophic pyramidal structures in population sizes and feeding rates that translated into the stabilizing patterns of interaction strengths. This first indication that trophic pyramids indicate balance and stability in real food webs can be underpinned by looking at the mechanism behind the stabilising patterning of interaction strengths. 7.4 The mechanism: Balance, Stability and Trophic Biomass Pyramids What are the biological mechanisms behind the patterning in the interaction strengths? And can we understand the stabilizing effect of this patterning? The answers to these questions are to be found in mathematical theory regarding the stability of complex structures. Recently, this has been done by analyzing trophic interaction loops in complex food web structures (Neutel et al., 2002). A trophic interaction loop describes a pathway of interactions (note not feeding rates) from a species through the web back to the same species without visiting the species more than once. Hence a loop is a closed chain of trophic links (see Figure 4). Trophic interaction loops vary in length and weight. The loop length is the number of trophic groups visited, and the loop weight is the geometric mean of the interaction strengths in the loop. The maximum of all loop weights is an indicator of food web stability (Neutel et al., 2002). The observed patterns of the interaction strengths in soil food webs (Figure 3) ensures that the maximum loop weight is maintained at relatively low levels as long as the population sizes are organised in trophic pyramids (Neutel et al., 2002 and see Figure 4). Moreover, the stronger the pyramidal structure, the lower the maximum loop weight and the higher the level of food web stability (Neutel et al., 2002). Therefore, in real food webs, trophic biomass pyramids indicate a balance between resource availability and community structure, and ensure ecosystem stability by keeping the weight of the trophic interaction loops at relatively low levels. 7.5 Managing vitality in soils 7.5.2. Measuring soil vitality

In this chapter the ideal soil vitality is defined as an extension of the concept of ideal soil fertility. This extension is in essence the addition of a stability criterion to the balance criterion. Stability is approached as community stability, but it is assumed to relate to the stability of ecosystem processes that directly result from the activity of the soil community, for example soil organic matter decomposition and nutrient

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Figure 4. Patterns in interaction strengths resulting from biomass pyramids in an omnivorous trophic loop (Neutel et ah, 2002). a. Diagram of the relative size of populations and feeding rates. - Thickness of the arrows denotes the size of the feeding rates. b. Trophic interaction loops. - Thickness of the arrows denotes the strengths of the effects. - Grey arrows are negative effects, open arrows are positive effects. - The loop of three feeding rates forms the basis of two trophic interaction loops of length 3: loop (1) from omnivores - microbivores - microbes (and back to the omnivores), including two negative and one positive effect, and loop (2) omnivores - microbes - microbivores, including one negative and two positive effects. - The numbers between brackets denote of the number of the loop. - The biomass pyramid causes a small negative effect between the omnivores and the microbivores. This ensures that the long loop with two negative effects contains a relatively small effect, which keeps the weight of this loop low (Neutel et ah, 2002). - The positive effects are much weaker than the negative effect and, therefore, play only a minor role in maximum loop weight

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mineralization. Stability implies that the equilibria are not easily disturbed through environmental change or disturbance. Stability does not mean that these equilibria are constant, as environmental change may cause changes in the equilibria. In some cases environmental change should cause such changes in order to maintain stable equilibria. For example, when high input agricultural land is converted into low-input natural ecosystems, organic matter and nutrient equilibria should change towards lower equilibrium values. This integrated approach of assessing environmental quality through analysing the structure and stability of trophic networks is not new, and somewhat similar to what has been proposed by Ulanowicz (1996), who, in turn, refers to the well known concept of Odum (1969) 'The strategy of Ecosystem Development'. Ulanowicz defines ecosystem quality in terms of overall trophic efficiency, structure of pathways of energy and nutrient cycling, species richness and community complexity. Looking at soil vitality in terms of equilibria and stability of community and processes is therefore close to the approach by Ulanowicz. Moreover, both approaches are aimed to bridge the gap between scientific understanding of soil quality and the development of soil management practices. Application of soil vitality in soil management practices requires that we should be able to measure vitality. The ideal vitality will never be reached; therefore we should allow for departures from the ideal stable equilibria. This means that it will be necessary to measure departures from balance and stability, and to evaluate these departures in the light of the need for, and success of, soil management practice. These measurements should include nutrient flow rates, decomposition and build up of soil organic matter and the structure of the soil community food webs. All these soil attributes are highly dynamic and extremely heterogeneous in space. Sometimes there are serious technical difficulties, for example, the identification of the diversity within soil organic matter and how this relates to the substrate available for the soil microorganisms. An apparent principle problem is that for the evaluation of the structure of the food web, there seems to be no clear understanding of what the community composition should look like, in terms of the presence and abundance of species. When should soil communities be bacterial or fungal dominated? When should particular groups of organisms be present, such as earthworms, predatory collembolans and nematophagous mites? One of the reasons for these uncertainties is that soil community composition is, in many cases, the results of present management and the long-term history of soil management. For example, polder soils are still very distinct from soil communities in the

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'old' land, although soil type and soil management have been similar for many decades. In terms of ideal soil vitality, we, therefore, seem to lack an understanding of what should be considered as the target where management practiced should adopt for. Fortunately, however, this 'lack-of-control' problem can be solved, at least in part, by just looking at the pyramidal trophic shape of the soil food web as an indicator of stability. Together with observations on nutrient and organic matter dynamics this should make an adequate set of measurements. However, for an appropriate use of the trophic pyramidal structure, there are two things we should bear in mind. First, stability does not mean balance. In soils that are severely contaminated with heavy metals, we sometimes see trophic pyramidal structures - indicating stability - together with out-of-balance organic matter dynamics leading to thick organic layers that negatively affect soil fertility and plant growth. Second, stability or the absence of stability does not indicate stress, whatever that is. Sometimes it is found that environmental disturbance negatively influences stability (Griffiths et al., 2000). But it is also a well known phenomenon that environmental disturbance can lead to community structures that are strongly governed by stability constraints. Then we expect stable configurations, i.e. in the form of (strong) pyramids (Neutel, 2001). Hence, despite difficulties, we should be able to obtain rough, but adequate measures of balance and stability in soil ecosystems that encompass all three soil vitality aspects: nutrients, organic matter and community structure. These measures can serve as a basis for developing management practices to enhance and preserve soil vitality. 7.5.2 Managing soil vitality

At present there are two main lines of soil management in the Netherlands. The first is soil protection, primarily aimed at sanitising contaminated soil or reducing the risks of soil contamination to human health and natural ecosystems. Recently, this policy is newly defined in terms of 'functional' protection and sanitation, which can be characterized as focussing soil management on the (wanted) use and function of the soil, as constituted in the Dutch BEVER programme. Besides this soil protection policy, the Netherlands has a long tradition in managing soil fertility. This has developed into an environmental policy instrument defined in terms of maintaining nutrients balances (MINAS). This policy-instrument is linked to the concept of ideal soil fertility and prescribes to what extent farmers are allowed to deviate from the ideal soil fertility in terms of nutrient balance. The societal legitimation for this policy-instrument is that

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nutrient imbalances can form a threat to the quality of natural ecosystems. Moreover, it is possible to manage nutrients in a way that prevent too large a deviation from the equilibria. Hence, the other aspects of soil vitality, organic matter and ecology, have not yet been addressed in current policy instruments. This is very unfortunate. Balanced organic matter dynamics are extremely important, as many agricultural soils in the Netherlands suffer from organic matter depletion, while in natural ecosystems, high levels of organic matter hamper the development of species rich plant communities. Organic matter equilibria can also be managed. In agriculture, organic fertilizers can be used, while in natural ecosystems upper soil layers rich in organic matter can be removed. The relationship between organic matter dynamics and microbial activity is, however, still a poorly understood aspect of soil ecosystem functioning. Apart from their role in organic matter and nutrient dynamics, the soil ecology is also important for the physical structuring of the soils. Direct management of the soil ecological parameters, however, does not (yet) seem feasible. Besides, the legitimacy for setting strict rules and targets for organic matter and ecology might be considered weak. Not only because of the above mentioned uncertainties regarding measuring and interpreting soil organic matter and soil community composition, but also because organic matter and soil ecology are not considered a potential threat to the quality of other (aquatic, air) ecosystem compartments. Summarizing, my conclusions therefore are: - we are able to adequately measure and monitor soil vitality; this ability is important in managing soil vitality; - what we can measure does not necessarily have to be managed. Feasible soil vitality management should be restricted to nutrients and organic matter; ecological parameters (trophic pyramids) provide necessary information regarding the need for, and the success of, soil management; - soil vitality management in the form of restrictions, rules, or even fines, is not yet justified. Managing soil vitality has first to be further developed, especially with respect to soil organic matter management; development of nutrient and soil organic matter management requires strongly applied research and development programmes carried out for various types of land use: agriculture, natural ecosystem, gardens, as well as in the context of land use change; these programmes are to be carried out by consortia consisting of the land owners (farmers) and soil scientists; - in these programmes different kinds of management practices are to be

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tested with respect to balance and stability in nutrient and organic matter dynamics at appropriate levels give the (future) function of the soil. Special attention should be given to the diversity of compounds within the organic matter pool, and the way in which these compounds serve as substrate for (diverse) microbial communities. In such an approach of 'learning by doing' we should be able to come up with operational plans to restore and preserve sustainable vitality in soil ecosystems. And maybe much sooner than we think. 7.6 Implementation in soil management The practical implementation is simple and based on three activities. Firstly divide all soil fauna families into functional groups: prey and predator. Secondly, determine the total biomass of those groups and thirdly determine the feeding rates of those functional groups. Unfortunately, only minor parts of the ecosystem approach based on balance and stability are being applied in practise. For all groups as mentioned in Figure 3, the first two aspects (functional groups and biomass) are measurable on a routine base. Growth rate of microbes can also be measured, but for the other groups this might be more difficult, even in scientific research. Routine measurements for this third step have yet to be developed. Acknowledgements I would like to thank Dr. B.H. Janssen (Wageningen University Research Centre) for his influential way of thinking. References Andren, O., T. Lindberg, U. Bostrom, M. Clarholm, A.-C. Hansson, G. Johansson, J. Lagerlof, K. austian, J. Persson, R. Petterson, J. Schniirer, B. Sohlenius and M.I.O. A. Wivstad, T. Lindberg, K. Paustian and T. Rosswall (Editors), 1990. Ecology of Arable Land - Organisms, Carbon, and Nitrogen-cycling. Organic carbon and nitrogen flows. Ecol. Bull. 40: 85-125. Baird, D.J., T.C.M. Brock, P.C. de Ruiter, A.B.A. Boxall, J. Culp, M.P. Eldridge, U. Hommen, R.G. ak, K.A. Kidd and T. Dewitt, 2001. The food web approach in the environmental management of toxic substances. In: D.J. Baird and G.A. Burton, editors. Ecological Variability: Separating Anthropogenic from Natural Causes of Ecosystem Impairment. SETAC Press, Pensacola, USA, pp 83-122. Bloem, J., G. Lebbink, K.B. Zwart, L.A. Bouwman, S.L.G.E. Burgers, J.A. de Vos and P.C. de Ruiter, 1994. Dynamics of micro-organisms, microbivores and nitrogen mineralization in winter wheat fields under conventional and integrated management. Agric, Ecosyst. Environ. 51: 129-143. Brussaard, L., L.A. Bouwman, M. Geurs, J. Hassink and K.B. Zwart, 1990. Biomass, composition and

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temporal dynamics of soil organisms of a silt loam soil under conventional and integrated management. Neth. J. Agric. Sci. 38: 283-302. Brussaard, L., J.A. van Veen, M.J. Kooistra and G. Lebbink, 1988. The Dutch Programme on Soil Ecology of Arable Farming Systems. I. Objectives, approach and preliminary results. Ecol. Bull. 39: 35-40. de Deyn, G.B., C.E. Raaijmakers, H.R. Zoomer, M.P. Berg, P.C. de Ruiter, H.A. Verhoef, T.M. ezemer and W.H. van der Putten, 2003. Soil invertebrate fauna enhances grassland succession and diversity. Nature 422: 711-713. de Ruiter, P.C, J.C. Moore, J. Bloem, K.B. Zwart, L.A. Bouwman, J. Hassink, J.A. de Vos, J.C.Y. arinissen, W.A.M. Didden, G. Lebbink and L. Brussaard, 1993a. Simulation of nitrogen dynamics in the belowground food webs of two winter-wheat fields. J. Appl. Ecol. 30: 95-106. de Ruiter, P.C, J.A. van Veen, J.C. Moore, L. Brussaard and H.W. Hunt, 1993b. Calculation of nitrogen mineralization in soil food webs. Plant Soil 157: 263-273. de Ruiter, P.C, A.M. Neutel and J.C. Moore, 1995. Energetics, patterns of interaction strengths, and stability in real ecosystems. Science 269:1257-1260. de Ruiter, P.C, A.M. Neutel and J.C. Moore. 1998. Biodiversity in soil ecosystems: the role of energy flow and community stability. Appl. Soil Ecol. 10: 217-228. Griffiths, B.S., K. Ritz, R.D. Bardgett, R. Cook, S. Christensen, F. Ekelund, S.J. Sorensen, E. Baath, J. Bloem, P.C. de Ruiter, J. Dolfing and B. Nicolardot, 2000. Ecosystem response of pasture soil communities to fumigation-induced microbial diversity reductions: an examination of the biodiversity-ecosystem function relationship. Oikos 90: 279-294. Hendrix, P.F., R.W. Parmelee, D.A.J. Crossley, D.C. Coleman, E.P. Odum and P.M. Groffman, 1986. Detritus food webs in conventional and no-tillage agro ecosystems. Bioscience 36: 374-380. Hunt, H.W., D.C. Coleman, E.R. Ingham, R.E. Ingham, E.T. Elliott, J.C. Moore, S.L. Rose, C.P.P. Reid and C.R. Morley, 1987. The detrital food web in a shortgrass prairie. Biol. Fertil. Soils 3: 57-68. Janssen, B.H., 1999. Basics of budgets, buffers and balances of nutrients in relation to sustainability of agro ecosystems. In: E.M.A. Smaling, O. Oenema and L.O. Fresco, editors. Nutrient disequilibria in agro ecosystems: concepts and case studies. CABI Publishing, Wallingford, UK, pp 27-56. Moore, J.C. and P.C. de Ruiter, 1991. Temporal and spatial heterogeneity of trophic interactions within belowground food webs. Agric, Ecosyst. Environ. 34: 371-394. Moore, J.C. and P.C. de Ruiter, 2000. Invertebrates in detrital food webs along gradients of productivity, in D.C. Coleman and P.F. Hendrix, editors. Invertebrates as webmaster in ecosystems. Cabi, New York. Moore, J.C, P.C. de Ruiter and H.W. Hunt. 1993. Influence of productivity on the stability of real and model ecosystems. Science 261: 906-908. Moore, J.C, P.C. de Ruiter, H.W. Hunt, D.C. Coleman and D.W. Freckman, 1996. Microcosms and soil ecology: critical linkages between field studies and modelling food webs. Ecology 77: 694-705. Moore, J.C. and H.W. Hunt, 1988. Resource compartmentation and the stability of real ecosystems. Nature 333: 261-263. Moore, J.C, D.E. Walter and H.W. Hunt, 1988. Arthropod regulation of micro- and mesobiota in belowground food webs. Ann. Rev. Entomology 33: 419-439. Neutel, A.M., J.A.P. Heesterbeek and P.C. de Ruiter, 2002. Stability in real food webs: weak links in

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long loops. Science 296:1120-1123. O'Neill, R.V., 1969. Indirect estimation of energy fluxes in animal food webs. J. Theoretical Biol. 22: 284-290. Odum, E.P., 1969. The strategy of ecosystem development. Science 164: 262-270. Ulanowicz, R.E., 1996. Trophic flow networks as indicators of ecosystem stress. In: Food webs: Integration of patterns and dynamics (G.A. Polis & K.O. Winemiller, editors). Chapman & Hall, New York. Wolters, V., 1997. Functional implications of biodiversity in soil. Office for official publications of the European Community. Wolters, V., W.L. Silver, D.E. Bignell, D.C. Coleman, P. Lavelle, W.H. van der Putten, P.C. de Ruiter, J. Rusek, D.H. Wall, D.A. Wardle, L. Brussaard, J.M. Dangerfield, V.K. Brown, K.E. Giller, D.U. Hooper, O.E. Sala, J. Tiedje and J.A. Van Veen, 2000. Effects of global changes on above- and below-ground biodiversity: implications for ecosystem functioning. Bioscience 50:1-10.

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Chapter 8

SOIL AND HIGHER ORGANISMS: FROM BOTTOM-UP RELATIONS TO TOP-DOWN MONITORING N. van den Brink

Abstract

This chapter focuses on effects of soil quality changes, through changes in habitat structure and food web relations, on higher animals. Modern society has evolved in the company of higher animals, mainly mammals, birds and fish, but also amphibians and reptiles. Over the centuries the relationship between man and animals has been studied, at first focusing on domesticated and game animals. The last decades investigations were related to the effects of environmental changes on the occurrence and functioning of animals in more or less natural conditions. One of the identified mechanisms by which humans may affect animals was through changes in soil quality. Such changes may be direct or indirect. An example of direct effects is the accumulation of contaminants from soil to animals, through the food web, which may result in toxicological effects. Indirect effects in general involve changes in the occurrence of prey items, related to changes in soil quality. Eutrophication can result in a decrease in the size of insects, which may affect the feeding efficiency and breeding success of birds feeding on larger insects. Effects of heavy metals and chlorinated compounds in soil on higher animals strongly depend on their food web. When biomarkers indicate a possible risk effect it may be explained as an 'early warning'. When high concentrations of heavy metals are detected in kidney or liver and chlorinated compounds in fat-tissue it may be explained as a 'late warning'. The relation is complex and involves different routes of mechanisms. Scientific, ethical, ecological and practical criteria are discussed to emphasize the difficult but also the challenging task of this kind of research. An example of an ecological point is that spatial and temporal scales are extremely important. Animals may exploit their habitats on a larger scale than the scale at which soil characteristics vary. Also contamination levels in organisms may vary in time due to physiological changes. Practical criteria are the labour intensity, the timing of sampling and techniques, and, for example, storage of samples while being in the field. The different criteria types are discussed in depth in relation to the pros and cons of various sampling approaches. The incorporation of ecological factors in monitoring soil contamination is emphasized, like for instance habitat preference, life strategies and food preferences. Rules of thumb for monitoring are recommended. When monitoring animals, not only numbers and densities are important, but also parameters on the ecology of the animals,

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for example reproduction, growth and mortality. Monitoring of animals should be conducted as a scientific activity, with proper hypotheses, and should meet certain scientific criteria. Apply methods as elaborate as needed, but as simple as possible. Select proper risk limits and reference data and/or sites. Include relevant confounding factors in the monitoring design such that they can be separated from the prime factors of interest, for example, age of animals.

8.1 Introduction Modern society as we know it today, could not have evolved without the company of higher organisms such as the larger vertebrates - mammals, birds, amphibians, reptiles, and fish - which we will refer to here as animals. From the time that man began to hunt animals and later when he started to domesticate them, the effect of higher organisms on the structure of human society has been enormous. The domestication of animals enabled man to produce food products, to farm land, and employ animals for transportation. Wild animals have also structured man's environment, as the example of the habitat of the Rhine delta clearly shows (Vera, 1997). Another major impact on human society was the plague of the fourteenth and seventeenth century that killed up to 50% of the population and that was spread by flees on rats. The close relationship between animals and man has inspired many people to observe and study animals. Aristotle (384-322 B.C.) was one of the first to describe the natural history of organisms, with special emphasis on their shape. In the eighteenth century a philosopher (J.F. Martinet) described animals in relation to their habits, food choice and other characteristics (Martinet, 1778). These descriptions focussed on domesticated animals, and on animals that were plague or game animals. They were derived from a religious perspective (creationism). In the nineteenth century the distribution of animals became increasingly a focus of study, while in the first half of the twentieth century observations were published on the prey items of birds of prey (Hoekstra and Smeenk, 1992). In the second half of the twentieth century this lead to more ecologically oriented and less descriptive research into animals. Excellent studies have been published on the badger (Meles meles) (Kruuk, 1989), on starlings (Sturnus vulgaris) (Tinbergen, 1981), and, for instance, the knot (Calidris canutus) (Piersma, 1994). Whilst this overview is brief and incomplete, it does attempt to show how, in recent years, research has addressed questions that focus on the causality of animal behaviour and the way in which animals use their habitat. Over time the interest that man has displayed in animals has evolved from the question of

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'What are animals for?' (to serve man) to 'What has caused animals to be as they are and to behave as they do?' (a more evolutionary approach). As such animals have gained a more central place in observations, although it should be noted that this is closely related to the cultural background of western society. In other cultures, views on animals may be different. In recent years the development of observation has gone a step further. The effects that human being and their activities have had on animals have become an object of study. In the Netherlands this type of research has been stimulated by publications on the effects of pollutants on birds (Koeman, 1975), seals (Reijnders and De Ruiter-Dijkman, 1995), and birds of prey, for example. It has also been suggested that the extinction of animals such as the otter (Lutra lutra) in the Netherlands was related to exposure to PCB (Leonards, 1997). Other publications on the loss and fragmentation of habitat (Vos and Opdam, 1993), the effects of road-kills (Broekhuizen et al., 1994), and other human influences on animals have also increased awareness on this issue. One of the mechanisms through which humans can affect animals is through changes in soil quality. Soil quality is affected in different ways by human activities including contamination, acidification, eutrophication, desiccation and erosion. Higher organisms depend on the soil in many ways. Here, this relationship between soil quality and its human induced changes and the occurrence and functioning of animals will be addressed. The monitoring of animals as 'indicators' or 'ambassadors' of problems associated with changes in soil quality will be discussed by focussing on different types of relationships, occurrence and functioning of animals, differences in effects, and the relevance and criteria for monitoring the way animals can be used in monitoring (in future). In this chapter, I will focus on examples of terrestrial ecosystems, but much of what is discussed here also applies to aquatic or riparian ecosystems. In Figure 1 the different connections between soil and higher organisms are depicted very schematically, and will be addressed in Section 8.1.2 through Section 8.1.4. 8.1.1. Effect types

Before discussing Figure 1, the direct effects and indirect effects of soil quality on animals will be addressed. Figure 2 illustrates the difference. Direct effects have a mechanism that relates changes in soil quality to the effect this has on animals. Even changes in prey can affect higher organisms directly. Indirect effects comprise a cascade of direct effects. It is useful to differentiate since the underlying mechanisms may become visible. This will allow a better analysis of the effect routes, and a better definition of possible measures to deal with

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Figure 1. Relationships between soil quality and higher organisms

the problems. Furthermore, acknowledging two effect types enables us to classify the methodologies that are applied in monitoring animals and which will be discussed later. In Table 1 the different mechanisms are given, whereas a distinction is made between contamination of soils on the one hand, and soil characteristic as eutrophication, and acidification of soils on the other. This division may appear artificial, however it is effective because it recognises the different backgrounds of the impacts. In general contaminants display their effects through ecotoxicological mechanisms, while changes in eutrophication, acidification, and desiccation generally result in changes in habitat.

Figure 2. Types of effects that changes in soil quality may have on animals: direct effects that work directly from change to animal, and indirect effects which consist of a cascade of direct effects

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Table 1. Direct and indirect effects of changes in soil quality due to changes in soil characteristics or contamination Indirect effects Direct effects

Eutrophication, acidification, desiccation Contamination Habitat changes -^ food availability Food web changes -^ food availability Habitat changes -^ habitat use Ecotoxicological effects -^ mortality

8.1.2 Landscape structure

The landscape that organisms encounter is, in general, partly regulated by soil type. For instance, on peat soils no beech forest will develop, whereas on welldrained sandy soils it is unlikely that moorland will result. For meadow birds, a clear habitat preference has been reported. Large areas in the Netherlands are in agricultural use, and consist of grassland. Meadow birds have strong preference for grassland on peat soil, preferably with high water tables (Beintema et al., 1995). Grassland on drained sandy soils is less frequently inhabited by meadow birds, although this may vary between species. The preference for peat soils may be due to the fact that these soils are more easily penetrated while foraging and that prey items may be more abundant. So, although large areas are in agricultural use as grasslands, soil type is a major factor in determining the suitability of grassland for meadow birds. This example shows that on a larger landscape scale, soil characteristics may strongly affect the dispersion of animals. On a smaller spatial scale the changes in soil quality occur and that will be the topic here. 8.1.3 Habitat structure

Effects of changing soil quality on animals through changes in habitat structure (Figure 3) are mostly related to eutrophication, acidification, and desiccation rather than contamination. They result in alterations in habitat because of variations in food availability (indirect), or in changes in breeding sites, cover, and other characteristics (direct). Changes in habitat structure: indirect effects Changes in prey availability may be strongly related to changes in habitat, therefore predators, as higher organisms, may be affected indirectly. The decreasing numbers of the Red-backed shrike (Lanius collurio) in dune ecosystems has been attributed to decreasing diversity in insect species, which itself is related to changes in vegetation structure, and changes in nutrient conditions (Esselink et al., 2001). It appeared that the average prey size

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Figure 3. Relationships between soil quality and higher organisms through landscape structure

decreased due to eutrophication, also reported elsewhere (Siepel, 1990), which negatively affected the feeding efficiency of the shrikes. It has also been suggested that other bird species were also affected by changes in vegetation structure, due to eutrophication (Van Turnhout et al., 2003). Other more anecdotal examples include the decreasing numbers of reptiles in heather which is also related to changes in vegetation structure caused by eutrophication (Van Turnhout et al., 2003). So changes in soil quality may have direct effects on the occurrence of prey items, and as such indirectly affect predating organisms (Figure 2), clearly detected in decreasing numbers. To establish these causal relationships is very difficult. This issue will be addressed later. Changes in habitat structure: direct effects Changes in habitat may affect organisms directly (Figure 4), due to decreasing breeding areas or preferential feeding localities. In wetlands, terrestrialization and changes in nutrient balances have been related to changes in the occurrence and quality of reed (Phragmites australis). Increasing nutrient loads resulted in declining densities of reed, which affected the occurrence of Great Reed Warblers (Acropcephalus arundinaceus), breeding in reed (Graveland, 1996). So eutrophication can result directly in changes in bird densities through changes in habitat structure. In another study on passerine birds it was shown that calcium deficiency limited breeding success in forests (Graveland, 1995). The limited availability of calcium was caused by acidification of the soil. Snails, a major calcium source for passerines, were less abundant, and breeding birds could not meet their calcium demands, needed to produce normal egg-shells.

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Figure 4. Relationships between soil quality and higher organisms through habitat structure

Observations on higher animals should be well planned. Assessment of (local) number and densities is generally a first step, but this may not be sufficient. For instance, numbers of territorial species may not vary much, not even when population dynamics are affected by external factors (Klok, 1997). Hence, when monitoring animals in order to reveal changes in habitat related to soil quality, one should address aspects that relate to population dynamics such as reproduction and mortality, and feeding ecology, such as types of prey and size of prey, rather than ecology and the way the species functions. As stated earlier establishing relationships between soil quality and higher animals is very complicated. Feed-back mechanisms may dim the occurrence of effects, for instance, if higher organisms can use alternative food items in case of indirect effects. Time-lags in the occurrence of effects may prevent the detection of changes in habitat structure, due to soil quality, at an early stage. Research into such relationships demands a truly interdisciplinary approach which can capture all the different aspects of the relationships at different levels. Dedicated research will establish causal relations towards eutrophication (Esselink et al., 2001). It turned out that effects were a late warning for changes in soil quality rather then an 'early warning'. Other changes in the food web related to habitat changes may be evident more early. Changes in the occurrence of prey species for a given predator may occur before the predator itself is being affected. In multiple effect-routes, leading from soil quality to animals, higher animals may be more sensitive to changes in habitat structure

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than single prey item. Whenever signals in the occurrence and functioning of higher organisms have been established and linked to soil quality, this is generally followed up more quickly by publicity and changes in policy and management of the area. 8.2.4 Food web relations

In addition to changes in habitat structure, the occurrence of prey items may be altered by exposure to contaminants. Higher organisms accumulate contaminants through the food web, as a result from this accumulation there may be toxicological direct effects (Figure 5). Food web relations: indirect effects Organisms need food to meet their daily energy expenditure and nutrient requirements. Soil quality affects food abundance directly, via toxic effects. For instance, earthworms can be affected by copper (Ma, 1984), and through modelling exercises it has been shown that population densities may decrease (Klok, 1997). When this occurs, the food available for predators depending on earthworms decreases. In France, effects of cadmium, zinc and lead exposure to myriapod and isopod communities were established in field conditions (Grelle et al., 2000). These changes in the communities could not be directly related to toxicological effects on the organisms, but to changes in vegetation types related to metal exposure. Metal exposure resulted in a more metalliferous grassland and a different community of soil invertebrates. Such changes in the abundances of soil invertebrates may result in a decrease in food availability

Figure 5. Relationships between soil quality and higher organisms through accumulation in the food web

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and changes in feeding ecology of higher organisms as predators. The extent of the effects on higher organisms, depends on their feeding ecology. Omnivorous species may switch to other sources of food whereas specialist feeding species may be severely impacted. A decrease in the amount of prey generally results in a less favourable situation in which higher organisms cannot thrive very well. Food web relations: direct effects Vertebrates gain energy and nutrient through food, but together with their food uptake they also ingest other compounds. This includes contaminants, for which the uptake via ingestion is, in general, a major route of accumulation in higher organisms. The variety of contaminants that are accumulated via food uptake is vast, including persistent organic pollutants (POPs), heavy metals, and pesticides. The majority of these compounds is from an anthropogenic source, so the distribution is related to human activities. Nevertheless, even in remote places like the Arctic region or Antarctica contaminants like POPs have been detected (Borga et al, 2001, Van den Brink, 1997). On a smaller regional scale, environmental processes may result in the deposition of contaminants in other places than where they were released, such as in floodplains of rivers (Middelkoop, 2000). Differences in feeding ecology may result in completely different types of exposure. In the Biesbosch, a freshwater tidal area in the Netherlands, beavers {Castor fiber) are exposed to very high levels of cadmium, through uptake via the bark of salicideae, a food source that beavers strongly prefer (Nolet, 1994). Concentrations in this bark are high compared to other available plant species (3.5-13.1 ug/g dry weight), resulting in high levels of cadmium in the kidneys of beavers (467 [ag/g dry weight). This example shows that specialist feeding may result in unexpectedly high exposure to certain contaminants. Earthworms and other soil organisms are also capable of accumulating organic contaminants, and this may possibly result in elevated body concentrations (Ma et al., 1998). It has been shown that for the badger (Meles meles) in floodplains of the river Meuse levels of heavy metals of PCBs in corn were much lower than in earthworms (< 0.04 versus 2.2 ug/g) (Boudewijn et al., 2003). For individual badgers the choice of food regulates the level of exposure to contaminants to a large extent. For cormorants (Phalacrocorax carbo), it has been shown that a change in the average size of the fish they caught, decreased their exposure to POPs dramatically (Boudewijn et al., 1994). Hence, information on food choice and feeding ecology is of major importance when monitoring higher organisms in relation to contamination (Pascoe et al., 1996). Another important issue is where the animals feed. Soil quality in general

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varies in space, depending on such factors as deposition and erosion. (Kooistra et al., 2001). At a larger spatial scale, it has been reported that cadmium concentrations in badgers in the Netherlands show a significant variation. Highest concentrations have been detected in badgers found nearby rivers, while in badgers that live further away from rivers the cadmium concentrations were relatively low (Figure 6), indicating that rivers are a major source for cadmium. For copper such spatial variation could not be established, apparently this heavy metal is more evenly distributed (Van den Brink and Ma, 1998). Spatially explicit analysis of data concerning soil quality and higher organisms has only been touched on. 8.1.5 Why using higher organisms to monitor soil quality

Sensitivity Top predators may accumulate such high levels of toxins that effects of soil contamination become evident. Multiplication of small effects at lower levels in the food web, may result in a multi-stress situation for the higher organisms, which can consequently result in effects. These mechanisms show that although lower in the food web no changes may be detectable, effects on higher organisms may become evident. Hence, accumulation and the multiplication of effects may result in the fact that higher animals are most sensitive to certain changes in soil quality.

Figure 6. Cadmium and copper concentrations in kidneys from badgers in relation to the distance to the nearest river (Van den Brink and Ma, 1998)

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Communication Higher organisms are in general appealing to people, and as such, factors that negatively affect these organisms are likely to receive a lot of interest. This may be a strong argument for using higher organisms for monitoring purposes (Section 8.2). Spatial scale and policy relevance Soil quality is variable in small spaces. Higher organisms exploit their habitat in general at a larger spatial scale. This larger spatial scale is more important in policy terms especially as far as spatial planning is concerned. It appears that the monitoring of higher organisms is particularly suitable for addressing specific policy objectives. 8.2 How to monitor animals In order to detect effects and changes it is essential to monitor organisms. There are many different ways to monitor animals, depending on the type of information needed, and the species of organisms under study. Also, the question of why animals should be used to monitor changes in soil quality remains. The establishment of a causal relationship between soil quality and higher organisms is extremely difficult and as such it does not appear logical to use higher organisms. Several other factors make higher organisms excellent for this purpose, and some have been mentioned before. Monitoring activities should be well planned well designed and well focussed. In many cases, monitoring appears to be more focussed on the gathering of data, rather than in addressing the relevant issues. In a monitoring plan, practical issues like the variables to be monitored, the timing of sampling, the number of samples etc. need to be defined. However, the plan should also meet certain criteria in order to be effective in producing data that are valuable. Such criteria are, for instance, related to the questions on (i) how to design a monitoring scheme, including the formulation of hypotheses and statistical design (scientific criteria), (ii) how to perform the studies in an ethically sound way (ethical criteria), (iii) how to choose a species capable of monitoring (ecological criteria), and (iv) how to implement the monitoring scheme and communicate the result (practical criteria). These criteria will be addressed below. 8.2.1 Criteria for monitoring studies

Scientific criteria In some papers, monitoring is not considered to be a scientific activity (see

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Risebrough, 1992). This view is not shared by Van den Brink (1997) and Bignert et al. (1993), who suggest that a monitoring plan should be designed with more in mind than just the collection of data. Essential for an effective monitoring plan is the formulation of testable hypotheses, and an experimental design. The hypotheses should address the issues that gave rise to the need for monitoring. For instance, when one is interested in time-trends in pollutants levels in biota in industrialized country X, a hypothesis that could be postulated is: 'The concentrations of cadmium in the kidneys of badgers found dead in industrialized country X in 2003 are higher than the concentrations in kidneys from badgers found in 2004'. Explicit formulation of such testable hypothesis implies that (i) the monitoring activities are focussed (on badgers in industrialized country X), that (ii) the sampling-scheme is well designed (samples are needed in 2003 and 2004) and that (iii) methods are available to statistically analyse the results (in this case the use of ANOVA). If no specific hypothesis is defined, monitoring often results in the collection of samples that are more or less easily available. Consequently, there is little guarantee that good reference samples are collected, and most often the issue of the statistical analysis is only addressed after the samples have been collected. From a scientific point of view, monitoring animals is a challenging task. It is advisable to adhere to a more 'experimental approach'. This implies that factors that may affect the results of the monitoring programme, are included in the scheme. For instance, cadmium often shows an age-dependent accumulation pattern in animals, as has been shown in badgers for instance (Figure 7) (Van den Brink and Ma, 1998).

Figure 7. Cadmium concentrations in kidneys of badger from different age classes (Van den Brink and Ma, 1998)

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When monitoring cadmium or PCBs in animals from two different locations in order to detect differences in exposure, it is of great importance to know the age of the organisms being sampled. In the statistical analyses, age can then be included as a co-variable, which will prevent differences between sites being reported, that are in fact not related to the site but to the age of the animals at the different sites. So great care should be taken when defining the sampling scheme and analytical statistical methods. This should all be planned a priori. Ethical criteria

Working with animals is never value-free. When handling animals for scientific purposes, it should be realized that they may suffer more or less from the activities. In the Netherlands, each scientific activity that involves the use of vertebrate organisms, must be assessed by an 'Animal Ethics Committee' (Waelbers et al., 2003). This committee considers the scientific and social merits of the research and impact of the scientific actions that are proposed. It weighs these factors against the potential pain and inconvenience caused to the animals. The basis for this deliberation is the supposition that vertebrates are capable of suffering (Regan, 1983), or have an intrinsic value (Taylor, 1986). As such, whenever scientific activities result in any inconvenience, the researcher has the legal obligation to minimise this, and look for alternatives in either animal friendlier methods or in another design which does not involve the use of animals. In the end, the ethics committee evaluates whether the scientific merits counterbalances the inconvenience experienced by the animal. Under Dutch law, this ethical consideration also applies to wild-life (Waelbers et al., 2003). There are examples of how to decrease the impact on animals. In several studies animals have been captured in order to assess their distribution or population dynamics and habitat use. In such research it should be realized that captured wild animals experience high levels of stress. The methods applied should minimise this stress as much as possible. When using life-traps, for instance, food and bedding material should be supplied for the period that the animal is held in captivity, and this period should be kept to a minimum. For ethical, but also from scientific reasons it is not acceptable that the behaviour of animals is affected by the way they are treated and the methods used to study them. When using transmitters to track animals, the weight and shape of the transmitters should not restrict the animal's movement or behaviour. If it does, the way the animal functions may be altered. This also raises ethical issues about whether such research is justified or not. Furthermore, the data gained by research on animals whose behaviour deviates from normal practice may be of very limited use.

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Blood has been used in studies in order to assess organochlorine concentrations in birds (Van den Brink et al., 1998; Van den Brink and Bosveld, 2001). Feathers have been used to analyse heavy metals (Burger et alv 1994; Van den Brink et al., 2003). These methods may be invasive, but they have not affected successful breeding (Van den Brink and Pigott, 1996). Other excrements may also be used, such as the preen oil of birds, or otter (Lutra lutra) spraint (Gutleb and Kranz, 1998). From studies of the latter it appears that the spraint of otters can be used to analyse PCB levels and hormone patterns (Van den Brink et al, 2003b), while the DNA profile in the spraint can be used to identify the specimen that produced the spraint. These techniques are still being optimized, they show that ethically sound methods can be developed when needed. Ethical criteria for monitoring organisms are still under discussion. Most experience with assessing the distress organisms experience during scientific research has been gained with laboratory animals (Waelbers et al., 2003). However, wild ranging animals may react very differently to the stress of being handled. It is likely that being captured is more stressful to a wild organism than to a laboratory animal. This shows that ethical criteria are not static, and that each project requires specific consideration. Ecological criteria

In most cases animals are monitored in field situations in order to assess some sort of environmental change or stress. The choice of animal species and monitoring parameters should be made in relation to the ecological setting in which the animals range. In this context issues, such as spatial and temporal scales, ecological functioning, food web relations and the way the animal exploits its habitat, are important. Spatial scales link soil and animals. In general, contamination patterns vary at a relative small spatial scale. Deposition may vary in space, as has been shown in river floodplains (Kerkhofs et al., 1993; Kooistra et al., 2001). Animals exploit the habitat in general at a larger spatial scale than on which contaminants vary. Animals are connected with the smaller scales of contamination patterns through food chains. For little owls in the floodplains of the River Waal, a modelling approach was used to assess the spatial relations between contaminants and organisms (Kooistra et al., 2001). Little owls integrate contamination from a larger area, and show different contamination patterns than the underlying contamination patterns. Therefore, the choice of species also determines the spatial scale at which the result will be integrated. Contamination levels in animals may also vary in time, independent of the

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levels in the soil. Long-term changes in body concentrations should be kept in mind when monitoring organisms in relation to soil quality. Contamination patterns may also fluctuate over much shorter periods. For several species, for instance polar bears (Ursus ursus) and Antarctic birds such as Adelie penguins (Pygoscelis adeliae), significant changes in body concentrations of PCBs have been shown (Henriksen et al., 2001; Van den Brink et al., 1998), which could be related to physiological changes during the season (Figure 8). Temporal variation due to ecological factors should be uncoupled from the temporal contamination changes to be studied. Food web relations also play an important role in the choice of monitoring species. Large differences in the cadmium load between species of small mammals can be explained by the underlying food web (Ma, 1994). Insectivorous mammals, shrews (Sorex araneus) or moles (Talpa europea) showed higher body burdens of cadmium in their kidneys in comparison to two species of rodents (Microtus arvalis and Mycromys minitus) (Ma, 1994). The two rodent species mainly fed on plant material, while the insectivorous species fed on invertebrates like earthworms. It appeared that the plant-based food-web was less efficient in accumulating cadmium than the invertebrate-based food web. So, although the different species were of the same size and were captured at

Figure 8. Seasonal fluctuations of concentrations of several contaminants in blood of Adelie penguins, each period same 15 individuals (Van den Brink, 1997)

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the same location, cadmium concentrations differed up to two orders of magnitude (Ma, 1994). Therefore, when monitoring organisms in relation to soil quality it is very important to consider the efficiency of the different food webs in accumulating the specific compounds of interest. In Figure 9 the effect of food web transfer on contaminant concentrations in animals is illustrated for cadmium and PCBs. Species are ranked according to the concentrations generally found in the different species. This ranking is simplified and does not recognise variation between species due to, for instance, differences in sample type. Nevertheless, the figure illustrates clearly the effect of the food web on concentrations found in animals. In the case of cadmium, the concentrations are lowest in carnivorous and piscivorous animals and highest in insectivorous animals. For PCBs the highest concentrations are found in carnivorous and piscivorous animals. When selecting a species to monitor cadmium, an insectivorous animal is preferable given the fact that such species occur in the area of interest. As far as PCBs are concerned a carnivorous/piscivorous would be best.

Figure 9. Ranking of species according to the concentrations cadmium (left, jig/g dry weight, data S. Broekhuizen, unpublished) and PCBs (right, jig/g lipid weight, data van den Brink et al., 2000, Van den Brink and Ma, 1998, and Van den Brink, unpublished). Species in bold are collected at contaminated areas, arrows connecting normal and bold species indicate the effect of contamination in the soil

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The examples presented here on ecological criteria are mostly related to research that focuses on contamination. However, ecological criteria should also apply to research related to changes in habitat structure. Practical criteria

Research involving higher organisms is, in general, labour intensive. In order to conduct the research efficiently, practical aspects should be considered, such as timing of the sampling or observation of the organisms, the feasibility of applying sampling techniques under field conditions, storage of data and samples in the field. Techniques should be applied as elaborate as necessary, and as simple as possible. A well-designed and scrutinised sampling or observation scheme and methodology pays off in a scientific way when analysing the data, and is valuable when it comes to budgeting the work. Other practical criteria relate to communicating the results of monitoring research. Effects of changes in soil quality on animals are highly communicative! For instance, the environmental drawbacks of DDT only became known to the general public after its effects were detected in birds (Carson, 1963). Observations on malformations in frogs in the United States have increased the awareness about the environmental risks of endocrine disruption dramatically (Ouellet, 2000). The use of animals as test-organisms in monitoring projects, however, is less easy to communicate. In this sense, ethical criteria are of great importance. The means of communication and the final 'audience' that needs to be reached by the research, should be acknowledged in the design of monitoring programmes. When applying lethal methods in a monitoring project, the use of test animals may be justifiable as far as researchers using scientific arguments are concerned. The fact that researchers had to sacrifice organisms may lead the general public to ignore the results and conclusions of the research itself, as shown in the discussions surrounding the use of test-animals in medical research. Although the use of test animals has been presented as beneficial to humanity, their use in medical experiments is under pressure. The discussion has resulted in the development of animal friendlier methods. In general, it is unlikely that the results of monitoring programmes are very efficiently communicated. 8.2.2 Practical implementation of monitoring of higher organisms

Practical monitoring of indirect effects Methodologies to monitor indirect effects are generic, and focus on the feeding ecology of organisms and the occurrence of prey items. Generally, observations on the feeding ecology of the higher organisms will be the starting point of such

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programme. Such observation can be direct, for instance, at a bird's nesting site. There, it may be feasible to observe the carry over of prey items from parent bird to chick, and thus the diet, fed to the chick can be deduced. In other studies, pellets have been used to extrapolate diets, especially in cases on birds of prey. In the case of mammals, information on their diet may be extracted from remains in their faeces (Boudewijn et al., 2003). Such indirect observations should be applied with great care, because they tend to be biased towards prey items that are slowly or not easily digested. Observations on stomach contents from animals found dead may be helpful in verifying such bias, although in stomachs easily digested items may also be under-abundant. In the case of piscivorous animals feeding on fish, the use of otolithes, ear-bones, may be of special interest. Such ear-bones are in general resistant to digestive activities, they are species specific and give insight into the size of the fish that was eaten (Leopold et al., 2003). Results of observations on the feeding ecology of organisms are, in general, variable. This is partly related to the behaviour of predators. This variability means that repeated observations are required over time. Monitoring programmes based on such observations should, therefore, include repeated observations in time, and on more individuals. Observations of feeding habits are only of value when they can be compared to some sort of reference observations, showing the typical feeding habits of the animals under study. The choice of the reference site is extremely important, because many factors may cause differences in feeding habits. Other means of detecting indirect effects include studying the occurrence and distribution of food items. This involves the investigation of vegetation or prey items, and their spatial and temporal distribution, since they may reveal the underlying factors. The choice of parameters should be deduced from the habitat use and feeding habits of the animals of interest. Occurrence of smaller insect species may be indicative for eutrophication (Siepel, 1990), while other species may indicate the fact that the vegetation structure is changing. In addition, chemical analyses will have to be performed in order to reveal whether contamination has caused a decline in prey items. Practical monitoring of direct effects Direct effects occur through different mechanisms depending on the type of underlying mechanism. Eutrophication, acidification or desiccation in general result in changes in vegetation structure, which directly affect the occurrence and functioning of higher organisms as birds and snails, as presented in Section 8.1.3.

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Census techniques have been applied primarily in order to detect changes in soil quality, related to direct effects via habitat changes. Using such techniques, the numbers and densities of organisms can be assessed. Besides these numbers and densities, the functioning and habitat use of organisms should be addressed. For instance, Kleijn and co-workers (2001) did not detect any effects of altered agri-environment schemes (lower nutrient input, delayed mowing etc.) on the densities of meadow birds, and it was concluded that the implementation of management agreements did not increase the biodiversity of meadow birds (Kleijn et al., 2001). In contrast to this, the breeding success of black-tailed godwits {Limosa limosa) was reported to be higher in fields with management agreements (Schekkerman and Miiskens, 2000). It appears that the breeding success of meadow birds may be a better functional indicator for the success of management agreements, compared to bird densities. Although this examples is not directly related to soil quality, it clearly indicates that the assessment of numbers and densities may not be sufficient to detect the direct effects of habitat changes. In this respect the choice of monitoring parameters should be well considered, and deduced from the type of change in soil quality that needs to be addressed. In order to practically monitor contamination in higher organisms, chemical and toxicological related methods should be applied. The choice of endpoints and methodological approach, however, still needs to be filled in. One can approach contamination monitoring from an exposure point of view, based upon chemical observations, or from an effect-oriented point of view, based, for instance, on bio-markers. Bio-markers can be defined as a change in a biological response (ranging from molecular through cellular and physiological responses to behavioural changes) that can be related to exposure to environmental chemicals (Peakall, 1994). The use of bio-markers in monitoring mammals, birds and amphibians has been reviewed by Van der Oost et al. (in press). Both the exposure and effect-oriented approaches have their merits and limitations. The exposure approach directly relates changes in contaminant levels in higher organisms to a change in soil quality. Ma (2001) reported a dramatic increase (one to two orders of magnitude) in radioactive caesium in common shrews (S. araneus) in June 1986 and October 1987, when compared to animals trapped in 1984 and 1985 near Budel in the Netherlands. This increase was directly related to the fallout from the nuclear reactor accident in Chernobyl in the former USSR (April 26th, 1986). This example shows that, also in higher organisms, changes in the soil concentrations of certain contaminants may be detected without a time delay, although this is not always the case. An exposure-oriented approach is limited in detecting effects. The risks of effects can only be derived when

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relevant information is available on dose-effect relationships related to the compounds of interest. For certain groups of species and compounds such information exists, although in most studies the relationships have to be extrapolated from species related to the one being studied. In contrast to this is the effect-oriented approach. Numerous studies have been published in which bio-markers have been applied in order detect biological effects in organisms, mostly at molecular, cellular or tissue level. A classic bio-marker that has been applied extensively is the ethoxyresorufin-O-dealkylase (EROD) bio-marker (Bosveld, 1995). In this bio-marker the activity of a specific cytochrome P450 iso-enzyme is analysed, mostly in liver tissue. Induction of this iso-enzyme is directly related to exposure to dioxin-like compounds like for instance PCBs, PAHs and dioxins (for mechanism see Figure 10; Kennedy et al. 1993, Bosveld and Van den Berg, 1994). Such well established working mechanism between exposure to a certain group of compounds, and a biomarker response (as depicted in Figure 10), implies that this EROD biomarker may be applied in order to causally related possible exposure in field conditions to ecologically relevant effects.

Figure 10. Schematic overview of the mechanism for cytochrome P450 enzyme induction by dioxin like compounds, resulting in EROD activity. AhR: aryl hydrocarbon receptor; HSP90: heat shock protein; ARNT: Ah receptor nuclear translocator; P450s: cytochromes P450

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The application of bio-markers is, in general, considered to be of greatest merits as an early warning system for the detection of effects in organisms (Peakall, 1992). Bio-markers are generally related to a group of contaminants and not to a single compound. As such they are better in providing information on the effects in organisms than on the specific compounds that may be causing the effects. The choice of which approach needs to be applied in a certain case will depend on the type of information needed from the monitoring programme and the information on the contamination patterns. Without any information it may be advisable to screen organisms on contamination patterns as a first step (exposure oriented). If risks on effects are foreseen further research on possible effects may be the next step. Where information on contamination is available, or where simple monitoring of that contamination is the only objective, relevant bio-markers can be applied directly. Another aspect in monitoring is the type of samples that will be collected. This choice depends on the type of contamination. Organochlorine pollutants like PCBs and chlorinated pesticides have a high affinity to fat tissues, so when analysing these contaminants tissues containing fat should be collected. This may be fat deposits (Boudewijn et al., 2003). Other, non-destructively obtained sample types like blood (Van den Brink and Bosveld, 2001) or preenoil in case of birds (Van den Brink et al., 2003a) have also been used. Eggs and animals found dead have been collected in studies addressing organochlorine pollutants (Franson et al., 2000; Genot et al., 1995; Van den Brink and Ma, 1998). A major disadvantage in using dead animals, and non-fertile eggs is that they may not represent a non-biased cross-cut of the population of organisms. Older and weaker animals have a greater chance of being incorporated in the samples, and in case of the use of road-kills, specimens which wander more may be over abundant in the samples. When using infertile eggs, it may be that the eggs of young inexperienced birds, who may abandon their eggs more easily, are over represented in the sample. Young birds may have lower contamination loads, and so may their eggs. Such biases may not be easy to identify, however, they may have a considerable influence on the outcome of a monitoring programme. Heavy metals have high affinity for organs with a high blood flow (Ma, 1994), particularly the kidneys. Samples of kidneys have been used extensively in order to monitor for instance cadmium levels in wildlife (Ma, 1994). Nondestructive alternatives may be hair in mammals (Nolet, 1995), or feathers in birds (Denneman and Douben, 1993). The affinity of heavy metals for these tissues may differ and therefore their suitability as well. Eggs have also been used to monitor heavy metals (Franson et al., 2000). Bio-markers in general demand that animal tissues be used for further

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biochemical analyses. For the EROD bio-marker liver tissue is optimal, for the bio-marker related to acetylcholin-esterase brain tissue is required. In general, the application of bio-markers is destructive to the organism involved, although some non-destructive alternatives have been proposed (Van den Brink et al., 2000) and applied (Van den Brink and Bosveld, 2001; Fossi et al., 1997). 8.2.3 Rules of thumb and recommendations for monitoring -

-

When monitoring animals, not only number and densities are important, but also parameters on the ecology of the animals, for example reproduction, growth and mortality. Monitoring of animals should be conducted as a scientific activity, with proper hypotheses, and should meet certain scientific criteria When possible, monitoring of animals should be performed ethically sound, with animal friendly methods. Apply methods as elaborate as needed, but as simple as possible. Select proper risk limits and reference data and/or sites. Include relevant confounding factors in the monitoring design such that they can be separated from the prime factors of interest (for example, age of animals in case of contamination monitoring).

These rules of thumb are very general, and do not focus on a specific situation. A major factor not included in these rules is the choice of species, or animals, to be monitored. This choice is very case-specific, and it is therefore difficult to generate rules of thumb without specific examples. In order to initiate the process of selecting a species, a theoretical mechanism needs to be defined that relates the soil quality of interest to the species, such as food webs (Figure 9). In other case, for instance in parks within urban areas, animals may not feed on their normal prey items but may rely on offal or rubbish, and in such case may have no relation to soil quality at all. In such cases, animals may reflect environmental management more than the soil quality. The relationships between individual animals and the areas of interest is another aspect. Territorial species reflect local conditions much better than nonterritorial or migratory animals. For instance, the little owl is residential yearround, and territorial. This makes this species a perfect candidate for monitoring local conditions. It feeds on earthworms and other insects as well as small mammals and birds and is an excellent candidate for monitoring contaminants. However, habitat oriented monitoring should also be applied. The choice of species and parameters to be monitored and the related criteria are summarized in Table 2 and 3. Tables 2 and 3 are not exhaustive, other

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Table 2. Pros and cons in the choice of species to be monitored in relation to various criteria Species / criteria Small mammals (shrews, moles, mice)

Scientific High numbers can be collected, small spatial scale

Ethical

Larger mammals (otter, badger)

Small numbers, larger spatial scales, references available? age dependent accumulation

Non-destructive methods may be applied, roadkills?

Longer life span, individual behaviour possibly more determining exposure

Birds of prey

Numbers in general not high, intermediate or large spatial scale, age dependent on accumulation

Non-destructive methods may be applied, roadkills and eggs?

Longer life span, individual behaviour possibly more determining exposure

Need to be sacrificed

Ecological Territorial (shrews, moles), short life span, high metabolism (shrews)

Practical

Life traps relatively easy, not very sensitive in communication, small samples for analysis Difficult to trap or collect (roadkills), possibilities for communication, non-destructive samples small Difficult to collect (road-kills and eggs), possibilities for communication, non-destructive samples small

species and sample types may be applicable. It illustrates clearly that not any one single species or sample type is the 'golden' choice. Depending on the situation, different cases will be valued to a greater or lesser extent, and the implementation of certain species or methods may be limited. 8.2.4 Monitoring in urbanised areas

Up till now I have discussed studies and monitoring in more or less rural areas. However, results of monitoring of animals in urbanised areas may be more difficult to interpret, due to the presence of human activity in the habitat. As a result of such activities, animals may change their behaviour and feeding habits. They may for instance be attempted to eat garbage. The presence of dogs and cats may strongly affect their functioning and even determine their absence, while some bird species depend on nesting sites underneath roofs for breeding. For instance swifts an swallows, and other so called 'culture followers'. The occurrence and functioning of such species in urban areas is therefore more or less governed by human activities, and the relationship with

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Table 3. Pros and cons in the choice of sample type in relation to various criteria Sample type | Scientific / criteria I Use of non-fertile Eggs ; eggs may | introduce bias

Ethical

; Active collection j affects breeding | success, use of ! non-fertile eggs is | little problem Animals ;Use of animals 1 Active collection i found dead may i not preferred, use ! introduce bias I of animals found i dead is little j problem Feathers I Limited set ! Generally little : problem, 1 chemicals, 1 indicative for body 1 individual needs ! concentrations?, itobe captured I collected in field ino relation to \ individual Blood I Limited set of 1 Generally little | chemicals, | problem, : indicative for body : individual needs | concentrations? itobe captured, : Temporal : infections? | variation Excrements | Limited set of | Generally little i chemicals, i problem | indicative for body i concentrations? | Link with I individual? Biopsies : Limited set of : Generally little (skin, fat, ..) | chemicals, | problem, : indicative for body : individual may | concentrations? ineed to be : captured, | infections?

| Ecological

\ Practical

| Sequence of eggj laying may affect I chemical content, ; resources for egg iproduction? ; Behaviour related i chances to be killed (bias)

; Availability of i non-fertile eggs | may be ; unpredictable

! Moulting pattern : may affect 1 chemical content

; Small samples, 1 difficult to analyse

I Season variation

| Small samples, | difficult to analyse

iUse of excrements | for marking | territory

| Small samples, i difficult to analyse

: Chance on | infections

: Small samples, | difficult to analyse

1 Availability of Idead animals may ibe unpredictable

the soil quality may become overruled by other aspects. This implies that in case of monitoring of animals in more urbanised areas, the first aspect that needs to be verified is whether the animals behave naturally and feed on the

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normal food items, so that they are related to soil quality in a 'normal' way. 8.3 Future directions

In general monitoring of higher organisms is focussed on assessing the current state and detecting effects or risks of effects. It is not necessarily focussed on solving problems, or mitigating the risks of effects. Information gathered in monitoring projects would increase in value if monitoring strategies could be designed that they lead to further steps in policy, planning or management. Currently, such steps are mainly focussed on minimising changes in soil quality, spatial planning may also lead to the mitigation of effects. 8.3.1 Life-strategies More information is needed on the food web that the animals exploit. In order to able to anticipate or even predict changes or effects, causal relationships will need to be established between soil quality and the occurrence of prey species. One way to approach this is to centre research around the life-strategies of the prey species. It is likely that species with similar life-strategies are affected in similar ways by changes in soil quality. If species can be categorised according to their life strategies, this may alleviate the burden of the amount of research needed to establish the causal relations. 8.3.2 Spatial dynamics

Soil characteristics vary in space on a different scale than organisms, and higher organisms generally exploit their niche at a different spatial scale to invertebrates. In current methodologies little attention has been paid to implementing spatial variability into the design of monitoring plans. Spatially explicit monitoring programmes enable the detection of spatial variability in soil characteristics, but also in habitat used by organisms. Based on this, areas with high risks (bad soil quality and high preference by organism) can be distinguished from areas with low risk (good soil quality or where the environment is not used by organisms). Such an approach may help in different ways to the alleviate risks of changes in soil quality for higher organism. Firstly, the determination of areas with high risks makes it possible to allocate resources for policy or management more efficiently. No generic measures need to be applied, but a tailor-made approach can be implemented in those areas with high risks. Secondly, the determination of areas that organisms exploit intensively provides information on their preferences. Based on this kind of information it may be possible to increase the attractiveness of areas within the

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home range of the organisms that are less affected by the decrease in soil quality. In such an approach it is not the soil quality that is the major focus in mitigating risks on effects, but the ecology of the animals being studied. 8.3.3 Risk assessment

For a sound assessment of the risks of effects resulting from changes in soil quality proper risk limits are needed. Model-exercises have been conducted for the risks posed by contaminants in order to derive risk-limits (Jongbloed et al., 1996). Such a modelling approach is valuable, but in my opinion, more empirical data is needed for validation. Derivation of dose-response under (semi-) field conditions should supply such validating data. Moreover, feeding ecology and possible temporal variation in toxicological sensitivity should be addressed. 8.3.4 Reference sites

For the assessment of effects, it is essential to know what the 'normal' situation is, including natural variation. Hence, good reference sites are of prime importance in this kind of research. In an ideal situation several reference sites are available, in order to assess the ecological band-width of 'normal' undisturbed situations. Considerable effort needs to be put into addressing the functioning and structure of these reference sites, in order to distinguish anomalies. Acknowledgements I would like to acknowledge Wim Ma, Sim Broekhuizen and Peter Doelman for their valuable comments on earlier versions of this manuscript, and for sharing their knowledge and expertise. References Beintema, A., O. Moedt, and D. Ellinger, 1995. Ecologische atlas van de Nederlandse weidevogels. Schuyt en Co uitgevers en importeurs BV. Haarlem, the Netherlands. ISBN: 90-6097-391-7. Bignert, A., A. Gothberg, S. Jensen, K. Litzen, T. Odsjo, M. Olsson and L. Reutergardh, 1993. The need for adequate biological sampling in ecotoxicological investigations: a retrospective study of twenty years pollution monitoring. Sci.Total Environ. 128: 121 -139. Bosveld, A.T.C., 1995. Effects of polyhalogenated aromatic hydrocarbons on piscivorous avian wildlife. PhD-thesis University of Utrecht, Utrecht, the Netherlands. Bdrga, K., G.W. Gabrielsen, and J.U. Skaare, 2001. Biomagnification of organochlorines along a Barents Sea food chain. Environ. Pollut. 113:187-198. Bosveld, A.T.C. and M. van den Berg, 1994. Bio-markers and bioassays as alternative screening methods for the presence and effects of PCDDs, PCDFs and PCBs. Fresenius J. Anal. Chem. 348:

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106-110. Boudewijn, T., S. Dirksen and M. Ohm, 1994. Zichtbare effecten van onzichtbare stoffen (visible effects from invisible compounds, in Dutch). Bureau Waardenburg, Culemborg the Netherlands. ISBN: 90-3690-255. Boudewijn T.J., N.W. van den Brink, C. Klok and B. van Hattum, 2003. Verontreinigingen in Maasuiterwaarden: blootstellingen en belasting van dassen. Alterra rapport 815, in Dutch. Broekhuizen S, G.J.G.M. Miiskens and K. Sandifort, 1994. Invloed van sterfte door verkeer op de voortplanting bij dassen. IBN-report 055. ISSN 0928-6888. In Dutch.Burger, J., I.C.T. Nisbet and M. Gochfeld, 1994. Heavy metal and selenium levels in feathers of known-aged common terns (Sterna hirundo). Archives Environ. Contam. Toxicol. 26: 351-355. Carson, R., 1963. Silent spring. Hamish Hamilton Ltd, London. Denneman, W.D. and P.E.T. Douben, 1993. Trace metals in primary feathers of the barn owl (Tyto albe guttattus) in the Netherlands. Environ. Pollut. 82: 301-310. Esselink, H., M. Nijssen, G.J. van Duinen, J. Jansen, M. Geertsma, J. Kuper and A. Bravenboer, 2001. Verkennende studie naar gevolgen van vermesting verzuring en verdroging en effectgerichte maatregelen op fauna, vegetatie en abiotitek in duinen op Ameland en Terschelling. Report Stichting Bargerveen, Nijmegen the Netherlands (in Dutch). ISBN: 90-806432-2-x. Fossi, M.C., C. Savelli, L. Marsili, S. Casini, B. Jimenez, M. Junin, H. Castello and J.A. Lorenzani, 1997. Skin biopsy as a non-destructive tool for the toxicological assessment of endangered populations of pinnipeds: preliminary results on mixed function oxidase in Otaria flavescens. Chemosphere 35:1623-1635. Franson, J.C., T. Hollmen, R.H. Poppenga, M. Hario, M. Kilpi and M.R. Smith, 2000. Selected trace elements and organochlorines: Some findings in blood and eggs of nesting common eiders (Somateria mollissima) from Finland. Environ. Toxicol. Chem.19:1340-1347. Genot J.C., D. Lecci, J. Bonnet, G. Keck and A. Venant, 1995. Data on the chemical contamination in the Little Owl Athena nocuta (Scop.) and it's eggs, in France. Alauda 63:105-110. Graveland, J., 1995. The quest for calcium. PhD-thesis University of Groningen, the Netherlands. ISBN: 90-9008-131-3. Graveland, J., 1996. The decline of an aquatic songbird: The Great Reed Warbler Acrocephalus arundinaceus in the Netherlands. Limosa 69:85-96 (in Dutch with English abstract). Gutleb, A.C. and A. Kranz, 1998. Estimation of polychlorinated biphenyl (PCB) levels in livers of the otter (Lutra lutra) from concentrations in scats and fish. Water, Air, Soil Poll. 106: 481 -491. Grelle, C., M.C. Fabre, A. Lepretre and M. Descamps, 2000. Myriapod and isopod communities in soils contaminated by heavy metals in northern France. Eur. J. Soil Sci. 51: 425-433. Henriksen, E.O., O. Wiig, J.U. Skaare, G.W. Gabrielsen and A.E. Derocher, 2001. Monitoring PCBs in polar bears: lessons learned from Svalbard. J. Environ. Monit. 3: 493-498. Hoekstra, B. and C. Smeenk, 1992. Historisch overzicht van de zoogdierfaunastiek in Nederland. In: Atlas van de Nederlandse zoogdieren. Eds. Broekhuizen, S., B. Hoekstra, V. van Laar, C. Smeenk, C., J.B.M. Thissen. Stichting Uitgeverij KNNV, Utrecht, the Netherlands. ISBN: 90-5011-051-7. In Dutch, pp 11-16. Jonbloed, R.H, T.P. Traas, and R. Luttik, 1996. A probabilistic model for deriving soil quality criteria based on secondary poisoning of top predators. Ecotoxicol. Environ. Saf.: 34: 279-306. Kennedy S.W., A. Lorentzen and J.A. James, 1993. A rapid and sensitive cell culture bioassay for measuring ethoxyresorufin-O-deethylase (EROD) activity in cultured hepatocytes exposed to

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halogenated aromatic hydrocarbons extracted from wild bird eggs. Chemosphere 7: 367-373. Kerkhofs, M.J.J., W. Silva and W. Ma, 1993. Zware metalen en organische microverontreinigingen in bodem, regenwormen en dassen in het winterbed van de Maas bij Grave. EHM rapport 14, Rijkswaterstaat, Lelystad (in Dutch). Kleijn, D., F. Berendse, R. Smit and N. Gilissen, 2001. Agri-environment schemes do not effectively protect biodiversity in Dutch agricultural landscapes. Nature 413: 723-725. Klok, T.C., 1997. A quest for the role of habitat quality in nature conservation. PhD-thesis University of Amsterdam, the Netherlands. ISBN 90-76894-01-9. Koeman, J.H., 1975. The toxicological importance of chemical pollution for marine birds in the Netherlands. Die Vogelwarte 28:145-150. Kooistra, L., R. Leuven, P.H. Nienhuis, R. Wehrens and L.M.C. Buydens, 2001. A procedure for incorporating spatial variability in ecological risk assessment of Dutch River floodplains. Environ. Managem. 28, 359-373. Kruuk, H., 1989. The social badger. Oxford University Press, New York. ISBN: 0-19-858703-1. Leonards, P.M.G., 1997. PCBs in Mustelids, analysis, food chain transfer and critical levels. PhDthesis Vrije Universiteit Amsterdam, the Netherlands. ISBN: 9010219. Leopold, M.F., C.J.G. van Damme, C.J.M. Philippart and C.J.N., Winter, 2002. Otoliths of North Sea Fish - Fish Identification key by means of otoliths and other hard parts. Cd-Rom. ETI Information Services Ltd, Wokingham, United Kingdom. ISBN: Hyb: 90-75000-22-7. Ma, W., 1984. Sub lethal 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., 1994. Methodological principles of using small mammals for ecological hazard assessment of chemical soil pollution, with examples on cadmium and lead. In: Donker, M.H., H. Eijsackers and F. Heimbach (eds). Ecotoxicology of soil organisms. SETAC Special Publication Series, CRC press Inc. Boca Raton Fl. ISBN 0-87371-530-6. Ma, W.C., A. Vankleunen, J. Immerzeeland and P.G.J. Demaagd, 1998. Bioaccumulation of polycyclic aromatic hydrocarbons by earthworms: Assessment of equilibrium partitioning theory in in situ studies and water experiment. Environ. Toxicol. Chem. 17:1730-1737. Ma, W., 2001. Insectivora. In: Shore, R., Rattner, B.A. (eds). Ecotoxicology of wild mammals. Ecological & environmental Toxicology Series. John Wiley & Sons Ltd, West Sussex, UK. ISBN 0-472-97429-3. Martinet, J.F., 1778. Katechismus der Natuur. Johannes Allart, Amsterdam, the Netherlands (in Dutch). Middelkoop, H., 2000. Heavy-metal pollution of the Rhine and Meuse floodplains in the Netherlands. Neth. J. Geosci. 79: 411-427. Nolet, B.A., 1994. Return of the beaver to the Netherlands. PhD-thesis University Groningen, the Netherlands. Oulett, M., 2000. Amphibian deformities: current state of knowledge. In: Ecotoxicology of amphibians and reptiles. Editors: D.W. Sparling, G. Linder and C.A. Bishop. Society of Environ. Toxicol. Chem., Pensacola Fl. USA. Pascoe, G.A., R.J. Blancher and G. Linder, 1996. Food chain analysis of exposures and risks to wildlife at a metals-contaminated wetland. Archives Environ. Contam. Toxicol. 30: 306-318. Peakall, D.W., 1992. Animal bio-markers as pollution indicators. Chapman and Hall

Soil and higher organisms:frombottom-up relations to top-down monitoring Ecotoxicological Series, Chapman and Hall, London, UK. Peakall, D.W., 1994. Bio-markers: the way forward in environmental assessment. Toxicol. Ecotoxicol. News 1:55-60. Piersma. T., 1994. Close to the edge: energetic bottlenecks and the evolution of migratory pathways in knots. Thesis University Groningen, Groningen, the Netherlands. Regan, T., 1983. The case for animal rights. University of California, California. Reijnders P.J.H. and E.M. de Ruiter-Dijkman, 1995. Toxicological and epidemiological significance of pollutants in marine mammals. In: Whales, seals, fish and man. Editors: A.S. Blix, L. Wallo and 6 . Ulltang 1995. Risberough, R.W., 1992. Chemical change in Antarctica-significance? Marine Pollut. Bull. 25: 9-12. Schekkerman, H. and G. Muskens, 2000. Do black-tailed godwits (Limosa limosa) breeding in agricultural grasslands produce sufficient young for a stable population? Limosa 73: 121-134 (in Dutch, with English abstract). Siepel, H., 1990. The influence of management on food size in the menu of insectivorous animals. Proc. Exp. Appl. Entomol. 1: 69-74. Taylor, P., 1986. Respect for nature: a theory of environmental ethics. Princeton University Press, Princeton, New Jersey, USA. Tinbergen, J.M., 1981. Foraging decision in starling (Sturnus vulgaris). PhD-thesis University of Groningen, Groningen, the Netherlands. Van den Brink, N.W. and K. Pigott, 1996. Effects of sampling blood and preenland oil on the breeding success of Antarctic birds. J. Field Ornithol. 67: 623-629. Van den Brink, N.W., 1997. Probing for the invisible. PhD-thesis University Groningen, the Netherlands. Van den Brink, N.W., J.A. van Franeker and E.M. de Ruiter-Dijkman, 1998. Fluctuating concentrations of organochlorine pollutants during a breeding season in two Antarctic seabirds: Adelie Penguin and Southern Fulmar. Environ. Toxicol. Chem. 17: 702-709. Van den Brink, N.W. and W.-C. Ma, 1998. Spatial and temporal trends in levels of trace metals and PCBs in the European Badger Meles meles (L., 1758) in the Netherlands: implications for reproduction. Sci. Total Environ. 222: 107-118. Van den Brink N.W., E.M. de Ruiter-Dijkema, S. Broekhuizen, P.J.H. Reijnders and A.T.C. Bosveld, 2000. Polychlorinated biphenyl pattern analysis: potential non-destructive bio-marker in vertebrates for exposure to cytochrome P450-inducing organochlorines. Environ. Toxicol. Chem. 19:575-581. Van den Brink N.W. and A.T.C. Bosveld, 2001. Spatial and seasonal trends of PCB concentrations and effects in common terns (Sterna hirundo) in the Netherlands. Marine Pollut. Bull. 42: 280-285. Van den Brink, N.W., N.M. Groen, J. de Jonge and A.T.C. Bosveld, 2003a. Ecotoxicological suitability of floodplain habitats in the Netherlands for the little owl (Athene noctua vidalli). Environ. Pollut. 122: 27-134. Van den Brink, N.W., J. de Jonge and H.A.H. Jansman, 2003b. The use of spraints for non-invasive monitoring of pollution levels and reproductive status in individual free ranging otters (Lutra lutra). Abstract book SETAC conference 2003, Hamburg, Society of Environmental Contamination and Toxicology. Van den Brink, N.W., M.B.E Lee-de Groot, P.A.F. de Bie and A.T.C. Bosveld, in press. Enzyme

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markers in frogs (Rana spec) for monitoring of risks of aquatic pollution. Aquatic Health Managem. Soc. Van den Brink, N.W. and J.M. Baveco, 2004. Spatially Explicit Risk Analysis: a new solution to contamination problems in the Metropolitan Delta. In: Planning metropolitan Landscapes. Editors: G. Tress, B. Tress, B. Harms, P. Smeets and A. van der Valk. Delta series 4, Wageningen the Netherlands. ISBN 90-807637-3X. Van der Oost, R., C. Porte Visa and N.W. van den Brink, in press. Bio-markers in environmental assessment. In: Den Besten, P.J. and M. Munawar, editors. Biological Testing in Marine & Freshwaters: Trends, Relevance and Linkages to Ecosystem Health. Society of aquatic and environmental health. Van Turnhout, C, S. Stuijfzand, M. Nijssen and H. Esselink, 2003. Gevolgen van verzuring, vermesting en verdroging en invloed van herstelbeheer op duinfauna, basisdocument. Report EC-LNV nr. 2003/153, Ede, the Netherlands (in Dutch). ISBN: 90-80643-2-46. Vera, F.W.M., 1997. Metaforen voor de wildernis: eik, hazelaar, rund en paard. PhD-thesis Agricultural University Wageningen, Wageningen, the Netherlands. ISBN: 90-5485-746-3 (In Dutch, with English abstract). Vos, C.C. and P. Opdam, (eds) 1993. Landscape ecology of a stressed environment. IALE studies in landscape ecology 1. Chapman and Hall London UK. ISBN: 0-412-44820-3. Waelbers, K., T.J. de Cockbuning and J. Vorstenbosch, 2003. Proefdieren zonder laboratorium, opvattingen over dierproeven buiten het laboratorium en de rol van Dierexperimentencommissies. University Utrecht, Utrecht the Netherlands (In Dutch). ISBN: 90-72920-22-8.

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Chapter 9

FATE OF CONTAMINANTS IN SOIL W. Peijnenburg

Abstract This chapter discusses the present status of our knowledge on the nature and behaviour of contaminants in the soil. The fate of contaminants is described according to the processes involved in natural attenuation: biodegradation, diffusion, dilution, sorption, volatilization and chemical and biochemical stabilization. The role of the uptake by vegetation and animals in attenuation is also considered. The contaminants are arranged in the structural groups heavy metals, nutrients, and polyaramites such as the chlorinated ones. Since the leading theory on the uptake of chemicals by soil organisms is the equilibrium-partitioning (EP) theory, its theoretical aspects and practical applications and limitations are discussed in detail. This provides us with the means to apply this concept to the relevant bioavailability and bioaccessibility fractions of the total contamination content in soil. The defined liquid, particulate and biotic phase play a leading role in the assessment of this EP. Three different tables contain overviews of commonly used extraction techniques for the three different contaminant groups. They contain extraction methods, endpoints of consideration and the soil properties of interest as organic matter content, pH, redox potential, clay content and dissolved organic carbon. Many controlling factors can be quantified and modelled on a routine basis. Unfortunately the various biotic species seem to differ in their uptake procedures. Therefore, there is no single expression for the bioavailable or bioaccessible fractions of soil contamination. It is concluded that monitoring the effective presence of contaminants should be based on specific extraction techniques, closely related to biological availability.

9.1 Introduction Many varieties of organic and inorganic compounds are present in soils. Due to natural and anthropogenic emission, point sources and veils of various chemicals are present in soil on mondial, continental, regional and local scales. We will discuss the fate of these substances in the soil in the context of the complexity of soil and the large variety of substances introduced into it. These substances or components are defined as contaminants when they are

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present in elevated concentrations. For reasons of simplicity and clarity, they have been divided into three structural groups: nutrients, (persistent) organic chemicals (including pesticides) and heavy metals. After an introduction to the soil environment, contaminants and their properties, the interaction between the contaminants and the soil matrix and its consequences will be described according to the definition of natural attenuation. The aspects of bioavailability and bioaccessibility, which may be seen as the key elements governing the persistence and the potential and actual exposure of biota to contaminants will be discussed. A monitoring strategy for bioavailable and bioaccessible contaminants present in soils is described in the final sections. This includes a brief summary of the methods available for measuring chemicals in the field. 9.2 Soil Environment Soil represents one of the most complex matrices in the environment because of its heterogeneity. It is an interplay of particles of different size, organic matter of different quality and quantity, pore water, (pore) air, and biota. It differs radically with regard to size as well as ecophysiological characteristics and it may lead to conditions that vary from aerobic to anaerobic. This interplay has been depicted schematically in Figure 1.

Figure 1. Microscopic associations between organisms, organic matter and mineral particles (modified after Paul and Clark, 1989, and Cuypers, 2001)

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Following interactions at micro- and macro-scales, soils act as sinks for a blend of organic and inorganic compounds of natural and anthropogenic origin. As such the soil provides large quantities of essential and non-essential elements, and compounds of a widely varying nature as sources of nutrition to biota. Contaminants are introduced into the soil by natural as well as anthropogenic sources. The contaminant groups discussed are persistent organics as pesticides (POPs) and hydrocarbons, heavy metals and nutrients. Since the current international focus is on POPs, they will receive the most attention here. 9.2.2 Polyaromatic hydrocarbons (PAHs)

Volcanic activities, hydrothermal wells and forest fires in combination with the burning of fossil fuels are examples of important natural sources of PAHs due to incomplete combustion of organic material. In soil, background concentrations in pristine samples can range from 50 to 1100 ug.kg-1 (Fritz and Engst, 1975). Concentrations of PAHs in contaminated soils, due to wood preservation, creosote production, gas works etc, may range upward to some thousands mg.kg4 (Wilson and Jones, 1993).The concentrations in the subsoils differ from the top soils. Here we focus on top soils. 9.2.2 Persistent organic pollutants (POPs)

Another group of organic contaminants which are present in many soils are polychlorinated biphenyls (PCBs). PCBs were first described in 1881 but have been produced technically to be used as additives in plastics, in electrical insulators, as basis compounds for pesticides, etc. Because of their varied use they have been spread to all environmental compartments and have been introduced into soils through various routes including application of pesticides, via refuse or via deposition of air particles. Although PCB production has been reduced in most countries since the 1970s and production was officially stopped in 1993, the total PCB content in background soil today amounts to 21,000 tons. A group of chemicals classified as being among the most persistent and most toxic compounds ever produced by mankind, are the chlorinated dibenzo-pdioxins and dibenzofurans. These chemicals are produced by biomass burning, but they are also the by-products of the large-scale production of chlorinated aromatic compounds (Alcock et al., 1998). Kjeller et al. (1991) measured the concentrations of chlorinated dibenzo-p-dioxins and dibenzofurans in soil samples from the United Kingdom taken from 1850 to the present day (Figure 2). Concentrations of 30-60 ng.kg4 before 1945 are related to the large-scale combustion of fuels (coal) and the smelting of metals for the production of iron

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Figure 2. The concentrations of chlorinated dibenzo-p-dioxins and dibenzofurans in soil samples taken from the UK since 1850 (modified after Kjeller et ah, 1991)

for an agriculture quantities. Examples of this include the group of phthalate esters, a group of chemicals that is widely used in enormous quantities as plasticizers. Most phthalate esters are difficult to quantify at levels exceeding the detection limits of the currently available analytical techniques. Also, nonhydrophobic chemicals like halogenated solvents, such as chlorinated ethenes and ethanes, usually do not bind strongly to the soil constituents and instead are either evaporated following emission to soil, or tend to reside in the groundwater. A specific class of chemicals that, other than those mentioned above, are intentionally emitted to the soil compartments and more or less by definition exert adverse effects, are the pesticides. Apart from their unique mode of entry (deliberately spraying) into the soil, pesticides behave like most persistent organic compounds and this class of compounds will not be dealt with exclusively in this chapter. Apart from classifying chemicals as naturally occurring, xenobiotic, essential or non-essential substances, it is also possible to classify them further on the basis of their inherent chemical properties and toxicity characteristics. POPs is a diverse category of bioaccumulative and toxic organic compounds of natural or anthropogenic origin that resist photolytic and chemical degradation,

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Figure 3. Structures of some of the priority persistent organic pollutants, as identified by the United Nations Economic Commission for Europe but condensed according to QSARs (Quantitative Structure Activity Relationship), and a number of PAHs typically found in soils, and some humus-like structures

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and which require specific conditions for microbial degradation. They are characterised by low water solubility and high lipid solubility, resulting in bioaccumulation in the fatty tissues of living organisms. POPs may be transported in the environment at low concentrations through the movement of fresh and marine waters. Recent evidence has shown that POPs are also transported long distances in the atmosphere, resulting in their widespread distribution around the earth even to regions where they have never been produced or used, like the Arctic. Concern about the possible long-term adverse effects on human health and the environment has initiated international negotiations on an 'International Legally Binding Instrument for Implementing International Action on Certain Persistent Organic Pollutants'. In these negotiations, twelve persistent organic pollutants have been identified as priority POPs (Table 1), requiring immediate legislative action. A selection of some typical substances is given in Figure 3. It also contains some typical substructures of humus-like soil components that make up the major part of the organic material present in soils. Most of the priority POPs have half-lives in the soil that vary from several months to several years. The half-lives given in Table 1 should be regarded as tentative because they vary with environmental conditions such as redox, temperature, and pH. Delicate interactions between chemical structure, soil properties, and mode of entry into the environment determine whether or not a specific chemical is persistent and hence of potential hazard to the soil compartment. It is the interaction with the constituents that make up the soil matrix in combination with soil-specific factors that determines whether a contaminant will be incorporated into the cell material of plants and animals, or is subjected to a number of natural attenuation processes. A schematic overview of the advective and diffusive processes governing the environmental fate of chemicals is given in Figure 4, as a more detailed example of the multitude of specific interactions typically occurring in top soil. The processes of sorption and precipitation grosso modo determine the fate of contaminants in the environment. 9.2.3 Metals and nutrients

In addition to the organic compounds mentioned above, metals and metal compounds as well as nutrients may affect the functioning of soil ecosystems. Ranges of metal concentrations typically encountered in soils, are summarised in Table 2. The table clearly shows the wide range of trace element levels that are typically found. In general the lower-end values given here represent natural background levels.

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Table 1. Properties, the half-lives in soil, and a general description of the characteristic environmental behaviour of the priority persistent organic pollutants identified by the United Nations Economic Commission for Europe (modified after Mackay et al., 1999) POPs Aldrin Chlordane Dichlorodiphenyltrichloroethane (DDT) Hexachlorobenzene Mirex Polychlorinated biphenyls (PCBs) Polychlorinated dioxins

Properties Kow

Kaw

~ 10"

~ 10"

~ 10'°

2 9

~ 1O -

~ 10'-2

~ 103-2

~ 10"

~ 10-i-3

9

5

Koa

~ 10 7 » ~ 10 8 « ~ 10"

Half-lives (h)

Environmental behaviour in soil

Ageing/ humification

10,000 - 30,000 10,000 - 30,000 10,000 - 30,000

Persistent Persistent Persistent

+ + +

~ 10"

~ 10»~10"

~ 106-8 ~ 10™ ~ 10 8 ' 8

>30,000 >30,000 >10,000

Persistent Persistent Persistent

+ + +

~ 10"»

~ io-2»

~ W"

3,000 -10,000

+

Polychlorinated furans ~ 1O-1

~ 10 3 - 2

~ 10"

10,000 - 30,000

Phenanthrene Fluoranthene Benzo(«)pyrene

~ 10«

~ io-«

~ 10"

~ 10-5-5 ~10"

-

2.5 - 4,400 2.5 - 4,400 1,368 -13,000

2,3,7,8,-congeners especially toxic Persistent 2,3,7,8,-congeners especially toxic Persistent Mineralization Persistent Persistent

~ 1O-

~ 10"

+

+ + +

Kow = octanol-water partition coefficient Kaw = air-water partition coefficient Koa = octanol-air partition coefficient

Metals and nutrients are introduced into the environment from natural and anthropogenic sources and due to (lack of) specific interactions with the soil constituents. Both enhanced and diminished levels of metals and nutrients may affect soil ecosystems. It is even possible that both deficient and excess levels of metals and nutrients are present in a single soil. The complexity of balancing the levels of contaminants in the environment is well illustrated by the fact that the introduction of (artificial) fertilisers to optimise crop yield is often associated with the release of heavy metals like Cd that are present at low levels in natural phosphate ores. Interpretation of the consequences of the long-term presence of contaminants for the functioning of ecosystems and human beings must take into account the site-specific potential of natural attenuation. Differences in availability for interaction of the contaminants with plants and animals are also important. Thus, whereas nutrients in a given soil may lead to adverse

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Figure 4. Schematic overview of the processes governing the environmental fate of chemicals as well as an indication of specific interactions taking place in the top soil Table 2. Concentration ranges of trace elements in soils (mg.kg-1) (Swaine, 1955) Element Silver Arsenic Boron Barium Cadmium Cobalt Chromium Caesium Copper Manganese Molybdenum Nickel Lead Selenium Tin Strontium Vanadium Zinc

Extreme range

Usual range

O.1-9.0 0.1-1,000.0

4RNHz + 3COz + HzO ROH + NH NH3 + H+ NO 3 + 2H* + HzO RNHz + 2COz + 2H2O 2N2 + SCOz + 7HzO 4NOz + COz + 3HzO RNHz + HzO 4RNHz + 7COz + 5HzO 4NO 3 + 4H+

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of the factors that affect the long-term performance of natural attenuation processes, and the monitoring of contaminants to ensure continued effectiveness. In the Netherlands this approach tends to become common practice. The contaminants that typically pose a threat to soil ecosystems are either extremely toxic, or there is a limit to their availability for NA processes and uptake by plants and animals. A schematic overview of the NA processes that affect the persistence of soil contaminants is given in Table 4. For POPs and PAHs, and heavy metals this will be discussed in terms of sorption, leaching, biodegradation and plant uptake aspects. The chapter of Verstraete and Mertens is appropriate for eutrophication and nutrient cycles. 9.3.1 Sorption in NA

An overview of the structures of the POPs, PAHs and the humus-like components (Figure 3), indicates that similar reactive sites are visible. Free electron-pairs along the oxygen atoms and at the edges of the chlorine atoms provide many opportunities of co-valent binding, and hence of tight interactions between these chemical structures and the constituents of the soil matrix. Subsequently, the contaminants become tightly bound to the soil matrix and cannot be easily extracted by means of conventional techniques, and are

Table 4. Overview of the natural attenuation processes that affect the persistence of typical soil contaminants Natural Attenuation processes in soil

Heavy metals POPs

PAHs

Nutrients

Sorption Diffusion Dilution Evaporation Microbial transformation Chemical/biochemical stabilization Uptake plants Uptake soil fauna

++ + + +1) +

+++ + + + ++++ +++

+++ + + - ++++ +++

++ ++ ++ + ++++ +

++ ++

++

++

++++ ++

++++ : +++ : ++ : + : : ') :

very general and strong general and strong regularly seldom/hardly not an electron acceptor (Mn, Fe, etc.)

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not available for interactions with soil biota. The phenomenon of incorporation of initially potential toxic compounds into soils is referred to as humification, bound-residue formation, or ageing. The complexation of organic contaminants with humic material can also be called humification and is part of ageing. The binding to humus decreases the amount of material and reduces the toxicity of the parent material. Jean-Marie Bollag (1992) emphasizes that this process should be exploited. He describes for instance the incorporation of 2,4dichlorophenoxyacetic acid in syringid acid, a humic acid (Figure 3). It is an oxidative coupling process. The products formed during the cross-coupling of pollutants and humic acids are highly heterogeneous and complex and, therefore, are extremely difficult to identify. In laboratory studies some mechanisms are elucidated by using model systems: fungi producing laccase incubating with chlorinated phenols in the presence of specific phenolic humic constituents. The process of humification has been neglected or remains unmentioned in the natural attenuation process, and therefore receives some special attention here. The higher the natural organic matter content of a soil, the more feasible the phenomenon of NA by sorption to and inclusion in organic matter. The effect of ageing is schematically depicted in Figure 5 according to the idea of Fuhr (1984) who took into account that chemicals may be partly transformed and ultimately bound, leading to a bound residue. From Figure 5 it can be deduced that the (non-bound) fraction of the chemical concentration that is available for interaction with the pore water and the biota present in the soil, diminishes quickly over time. This has severe consequences for extrapolating results form short-term tests, such as those usually done in laboratory settings, to realistic field conditions. For instance, toxicity testing is usually done after relatively

Figure 5. Schematic representation of the impact of time on the fate of chemicals in soil (modified after Fuhr, 1984)

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short equilibrium periods that typically vary from a few minutes to a few weeks. Equilibrium Partitioning As explained by Jager (2003a), the leading theory on the uptake of chemicals by soil organisms is the Equilibrium Partitioning (EP) theory, formulated and broadly adopted around 1990. Basically, this approach states that organisms do not take up chemicals from soils directly, but only from the freely-dissolved phase in the pore water. Thus it may be deduced from Figure 5 that the results of toxicity tests, for instance, will typically overestimate the truly bioavailable fraction and hence adverse effects in the real environment. A chemical will tend to distribute itself between the soil, water and organism phases until it is in thermodynamic equilibrium. This implies that the chemical residues in organisms can be predicted when we know the sorption coefficient of the chemical (partitioning between solids and water) and the bioconcentration factor (partitioning between water and organism). This is schematically depicted in Figure 6. EP theory takes the effect of ageing into account in risk assessment. EP has become an integral part of soil chemical risk assessment as far as predicting toxicity (from aquatic data) as well as body residues (from total concentrations) in soil-dwelling organisms are concerned. Currently, the term EP is often used in a broader sense, relaxing the precondition of equilibrium, and denoting the fact that (time-varying) concentrations in organisms can be predicted from the

Figure 6. Schematic overview of the processes underlying the equilibrium partitioning concept (modified after Jager, 2003a)

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(time-varying) concentrations in pore water. Despite its popularity, the limitations of EP have also been observed. The most striking deviations are discussed below. Sequestration or ageing, as mentioned above, is the process by which chemicals tend to become less available in time for uptake by organisms and by soft chemical extraction techniques. The most likely mechanism for this behaviour is that the chemical is moving deeper into the organic matrix as contact time increases, or because of possible chemical oxidation with humic substances as Bollag suggested. Sequestration has been presented as a deviation from EP, but in fact it is a strong support. Granted, the use of equations where sorption is estimated from hydrophobicity will fail to predict the effects of sequestration, but EP (in the broad sense) appears to be quite robust as long as good estimates or measurements of pore-water concentrations are available for the specific situation being studied. Another deviation from EP that is extensively discussed is feeding. Chemicals are not only taken up by organisms from (pore) water through the skin, but also from the gut. It is a generally held view that the existence of multiple routes of entry into an organism leads to deviations from EP predictions, especially for very hydrophobic chemicals that have a low solubility in water. It was predicted that feeding becomes an important uptake route for earthworms when log Kow exceeds 5. For sediment organisms there is evidence that feeding is important for very hydrophobic chemicals, and may lead to deviations from EP of up to a factor of 5. However, there are few studies that succeed in experimentally separating both uptake routes, and often conclusions on uptake routes are drawn without confirming that equilibrium was established, and without knowing the actual pore-water concentrations either. Furthermore, it is unlikely that chemicals are transferred directly from a solid phase to an organism without the intervention of a solution phase. Biotransformation may also lead to deviations from EP, but this process is not well studied. Biotransformation usually results in more water-soluble metabolites that are easily excreted. When the exchange with the pore water is slow, even low levels of transformation may affect toxicokinetics. Despite its limitations, EP is still the reference theory for discussing the accumulation of organic chemicals in soil organisms. This implies that body residues observed in biota should first be related to pore water concentrations before alternative theories can be explored. Intermedia transport is important in systems that contain more than one phase. Chemicals will migrate from one phase to another if the phases are not in thermodynamic equilibrium (do not have the same fugacity). Octanol is often

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considered as a surrogate for various condensed lipophilic materials present in natural phases such as non-living natural organic matter present in soils, sediments, or aerosols and certain lipid-like constituents of plants, animals, and micro-organisms. The fraction of such organic phases on a global scale is quite low. Organic phases, however, have a high affinity and a high storage capacity especially for organic contaminants. Organic phases, therefore, are the major sinks for hydrophobic contaminants. It has been observed experimentally that the ratio of concentrations in two phases is constant if the concentrations of the chemical in both phases are sufficiently low (thermodynamic equilibrium). In this case, at equilibrium conditions, the reversible distribution between phases can be described by a constant, which is known as the distribution coefficient (Kat>):

Kah = ^

(1)

^b

For solids - (pore) water systems, the equilibrium constant is known as the partition coefficient (KP) or distribution constant (Kd). Partition coefficients are available for many chemicals from laboratory and field measurements. As organic carbon present in water (Dissolved Organic Carbon), sediment or soil is the main sink for hydrophobic organic contaminants, the partition coefficients for these compounds are often adjusted (normalised) with respect to the organic carbon content of these compartments:

Kp=Kocxfoc=^-

(2)

Koc is the organic carbon normalised partition coefficient (L/kg), foe is fraction of organic carbon, and Cs and Cw are the chemical concentrations in the solid phase and the (pore) water phase respectively. Koc for neutral organic chemicals is often estimated from the octanol-water partition coefficient (Kow). It may be deduced from Equation 2 that partition coefficients of hydrophobic organic compounds in general are dependent upon both the chemical of interest (compound specific properties affect the value of Koc), and the properties of the medium in which it resides. Apart from the fraction of organic carbon present in the sorption phase, environmental factors also affect partitioning. These factors include temperature, particle size distribution, surface area of the sorbent, pH, ionic strength, presence of suspended material or colloidal

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material, as well as dissolved organic matter concentration. In addition, clay minerals may also act as additional sorption phases for organic compounds. Nevertheless, organic carbon-normalised partition coefficients for a specific chemical are fairly constant among soils, provided that the additional environmental factors impacting partitioning are kept reasonably constant. Highly hydrophobic contaminants like POPs and high-membered ring PAHs tend to sorb strongly to organic phases present in the solid and the dissolved organic carbon (DOC) in the aquatic phases (pore water). Uptake of POPs and other hydrophobic contaminants therefore usually takes place directly from the solid soil phase such as ingestion of solid material and consumption of food. Cationic pollutants like ionic substances and metals show deviant behaviour. Apart from organic carbon, additional constituents of the solid soil phase may act significantly as sorbing phases. Again, environmental conditions like pH and redox state strongly affect the partitioning of these compounds and as a result KP or Kd values vary by orders of magnitude among different soils. The effect of pH and redox conditions (Eh) is schematically illustrated in Figure 7 (see Salomons, 1995). Metal levels in the pore water will be especially high in aerobic soils at low pH (sandy soils), whereas it can easily be deduced from Figure 7 that metals will be dominantly sorbed or precipitated in anaerobic soils

Figure 7. Schematics of the effect of pH and Eh on metal speciation (modified after Salomons, 1995)

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at high pH. Hence, the latter soils in general pose no threat to organisms exposed via the pore water. The master variables, pH and Eh, affect the processes of adsorption, precipitation, complexation and oxido-reduction of the elements (heavy metals) and their solubility. Depending on soil type it can be concluded that acidification increases the solubility and therefore the mobility of the metals Cd, Pb, Zn, amongst others. Since Cr has a more extreme speciation values (Cr3+ versus Cr7+ in CrOc) it may behave differently. Desiccation, leading to more oxidizing conditions, may enhance mobility. Changing the groundwater level results in other environmental conditions and therefore other mobilities. Speciation of heavy metals in the aqueous phase, including sorption to DOC, strongly affects the availability of the metals for uptake and subsequently exerts adverse effects. Speciation is usually associated with the aqueous phase and involves the interaction of a contaminant with the various organic and inorganic constituents of the pore water. In the case of metals, these interactions not only include the formation of short-living complexes with a number of anions, like chloride, sulphate or carbonate, and complexation or binding to DOC, but also binding to the outer membranes of organisms and plants. An example of the various interactions in pore water is given in Figure 8. In the same way that soil organic matter is important for sorption of hydrophobic organics, the sorption to DOC is an important attenuation process for these

Figure 8. Schematic overview of the interactions of a metal ion in solution

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chemicals. They effectively reduce the amount of chemical present in the pore water that is available for sorption to living biota. 9.3.2. Leaching

Leaching is of relevance for all chemical substances that may dissolve in pore water or groundwater. Leaching is the advective movement of a pollutant towards the groundwater. Leaching is directly related to the total concentration of the contaminant in the pore water, including the fraction sorbed to the DOC. The hydrophobicity of an organic contaminant in combination with the amounts of organic carbon in the solid phase and the pore water determine leachability in a given soil. Leaching of metals is also directly proportional to the total metal levels in the pore water and is strongly affected by the soil and pore water properties that determine metal partitioning in the soil. Especially pH is an important parameter in this respect. As such, the leachability of a chemical is directly related to a delicate balance that results from the sorption and speciation processes discussed above. In general, higher concentrations of particulate organic matter and higher pH-values decrease the extent of leaching of any contaminant, except Cr and As. The Henry constant (H) of a chemical determines the volatilization. This is the air-water partition coefficient (Kaw). It can be expressed as a dimensionless ratio of the concentration of the chemical in the air and in the (pore) water. For any organic compound, it is the magnitude of the Henry's law constant in relationship to the value of the Koc that determines the extent of volatilization of the chemical. A multi-media fate model is an appropriate means for describing the extent of volatilization of a given contaminant in a given soil. 9.3.3 Volatilization in NA

Volatilization is relevant only for some POPs and PAHs. Volatilization may contribute directly to the attenuation of contaminant levels in soil. Many substances reach the soil as a result of wet and dry deposition from the air. Whether a volatile chemical will partition from the soil into the air depends on its interaction with various soil constituents. Thus, volatile hydrophobic chemicals like a number of PCBs that sorb strongly, for instance, to organic carbon will only volatilise in limited amounts. 9.3.4 Bio degradation and metabolism in natural attenuation

Biodegradation is mainly important in the case of organic contaminants. Biodegradation, or the transformation of chemical substances through the action of living organisms, is one of the major processes that determine the fate

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of organic chemicals in terrestrial environments. If the biodegradation of a chemical is slow, then the chemical may bioaccumulate and cause primary and secondary poisoning in the food web, or it may reduce the quality of drinking water and affect the various functions of surface waters. Biodegradation is, in general, equivalent to conversion into simple and relatively harmless molecules and ions, such as carbon dioxide, methane, water and other inorganic compounds. This phenomenon has been called ultimate biodegradation or mineralization and may be regarded as a true sink in aerobic soil compartments. The environmental conditions of the soil are the key factors that determine whether transformation and mineralization occur. Even in arable soil systems, soil aggregates host a variety of different redox-conditions and, therefore, it is plausible that the potential for degradation is there for most, if not all, organics. The degradation itself will be slow as Table 5 shows. As can be seen in Table 5 natural degradation rates of plant material are slow (k is the relative fraction degraded per time unit). Those rates of natural organics are presented in order to be aware of the extreme slowness of normal natural biodegradation and also as a warning and contrast for degradation rates measured under laboratory conditions. Degradation rates of mono aromatics such as benzene under field conditions are of similar value (Hutchins, 1991; Beller et al., 1992). Natural organic matter consists of carbon, hydrogen, nitrogen, oxygen, phosphor and sulphur, the same basic material as microbes. PAHs consist of carbon, hydrogen and oxygen and POPs of carbon, hydrogen, oxygen and chloride. So, for the degradation of PAHs and POPs the microflora needs at least extra nitrogen, phosphor and sulphur, and it has to adapt its degradation system. Degradation rates of hydrophobic PAHs and POPs are extremely slow, as the response of the microbial community towards natural as well as Table 5. Some typical examples of degradation rate constants (k) of organic matter of defined vegetations (modified after Werff 1992) Type of predominant vegetation Phragmites australis Phragmites karka Typha domingensis Typha glauca Typha Latifolia Typha angustata Typha ekphantina Scirpus fluviatilis

k 0,0035 0,0045 0,0078 0,0014 0,0043 0,006 0,0038 0,0018

Type of predominant vegetation Scirpus americanus Scirpus mucronatus Juncus squarrosus Juncus roemerianus Paspalum repens Carex rostrata Carex riparia Zizania aquatica

k 0,0021-0,0025 0,0044 0,0013 0,0016-0,0017 0,00717 0,0046 0,0029 0,077

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xenobiotic organic compounds does not depend on the total concentration of the contaminants but mainly on the water soluble concentration. This soluble fraction is relatively small for mineral oil, PAH, PCB and many chlorinated organic compounds such as hexachlorocyclohexane (HCH) and HCB, but not for tetrachloroethene (PCE) and trichloroethene (TCE). The sorbed or bounded fractions of xenobiotics are not available for microbial interference as microbial activity takes place in the soluble phase. The soluble concentration may be too low to support the measurable growth of the microbial population. However, even without growth micro-organisms may completely degrade pollutants. A brief review of literature data on persistent soil contaminants as PAHs and some POPs shows that PAHs are mainly biodegraded under aerobic conditions, but the smaller hydrocarbons also with NO3 or Fe2C>3 as electron acceptor (Cerniglia, 1992; Warith et al., 1992). HCH can be degraded under many different redox conditions (Bachman et al., 1988). PCBs are biodegraded under sequential anaerobic and aerobic conditions (Quensen et al., 1988, Abramowicz, 1990). Not much has been reported on complete degradation of POPs as drins and HCH, although one may question whether this has been investigated at all. Scientific research in the field of microbiology carried out during the last two decades has lead to great discoveries. Twenty years ago the anaerobic microbial degradation of halogenated compounds was generally considered as "not occurring". In the 1980s, reductive dechlorination (Bouwer and McCarty, 1983) was discovered as a new phenomenon. Then around 1992, partly as a result of the PhD theses of Oldenhuis, Van der Meer and Holliger, the facts and principles of the microbial degradation of chlorinated compounds became more clear. The anaerobic microbial degradation of chlorinated xenobiotics also became general knowledge, and applicable to bio-remediation. It became clear that bacteria could use PCE and other chlorinated compounds as electron acceptors. The principle of "halorespiration" was added to the phenomenon of recycling of elements. Suitable electron-donors and suitable environmental conditions became objects to study to apply dechlorination in insitu bioremediation processes. It took 12 years before the principle of biodegradation of HCH was partly unravelled. In addition, the role of the various prevailing electron acceptors, such as O2, NO3, Fe3+, SO42 and CO2 and various electron donors became clear. For example, the fact that in the presence of nitrate complete dehalogenation is impossible. Around 1995, the partial biodegradation of dioxins (Adriaens et al., 1995; Halden and Dwyer, 1997) was described and in 2003, the complete reductive dehalogenation of dioxins by the bacteria Dehalococcoides was reported (Bunge et al., 2003). Many parameters affect the biodegradation of chemicals in the environment

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as well as in biodegradation test systems used to simulate the environment. The bioavailability of the chemical to the enzyme has been identified as a major factor in determining biodegradation in nature, and this aspect will be dealt with in more detail in Section 9.5. The variables that potentially affect rates of biodegradation can be ranked into three categories: the soil related, the microflora related and the contaminant related factor, and are depicted in Figure 9. For detailed information on this triangle the chapter by Verstraete and Mertens is very appropriate. Figure 9 is presented here as an important reminder. Since the impact in soil of most of the variables given in Figure 9 is poorly understood, it is currently impossible to accurately predict how rates of biodegradation might vary from location to location for a given set of environmental conditions. Moreover, the use of second-order kinetics to evaluate rates of biodegradation has been validated only for a limited range of organic compounds. Transformation of chemicals in the environment can also occur by abiotic processes. The most important abiotic transformation processes can be divided into four separate categories, such as hydrolysis, alteration of the chemical structure by direct reaction with water; oxidation, a transformation process in which electrons are transferred from the chemical to a species accepting the electrons; reduction, the reverse of oxidation, electron transfer takes place from a reductant to the chemical to be reduced; and photochemical degradation, transformation due to interaction with sunlight. Usually, this last process is of limited relevance for the soil compartment. Transformation and mineralization processes can alter physico-chemical and toxicological properties and reduce exposure concentrations of environmental

Figure 9. Variables known to affect biodegradation rate constants (k) of organic contaminants

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chemicals. Where biotransformation is performed by higher organisms, the formation of polar transformation products (metabolites) can also provide an important method of detoxification. The rate of degradation of a specific chemical depends on its availability for reaction, its intrinsic reactivity, the availability of the reactant and the reactivity of the reactant. Biodegradation and metabolism of heavy metals by microbes does not occur in a strict sense in natural attenuation since these processes imply elimination. The microbial transformation of metallic minerals, however, has been known for hundreds of years. Bacteria are capable of changing the valences of metals and metalloids. The valences of these metals as well as their major microbial transformation reactions are given in Table 6. Knowledge obtained over the last 25 years may still follow this relatively old table. Fe and Mn do play a role as electron acceptor in the degradation of aromatics (Sumners and Silver, 1978), thus inducing natural attenuation of these chemicals. 9.3.5 Uptake as part ofNA

Uptake by plants and animals is, in general, the consequence of a large number of competing processes, both in the aqueous and in the solid phase, as well as at the interface between the biota and the pore water. In the definition of NA by USEPA (1993), the interaction with the living other part of the soil, such as roots and soil fauna, has not been mentioned. Those living parts may nevertheless play a role in the elimination of contaminants, both by means of degradation and by accumulation. Hyperaccumulating plants may be especially important in this respect as they are capable of trapping large amounts of soil contaminants. Vegetation plays a major role for nutrients and nutrient cycling in the soil, since most plants have optimized their modes of uptake and elimination of nutrients in order to cope with conditions of both deficiency and excess nutrients. Vegetation may also play a role for the attenuation of heavy metals, albeit that this is usually on a smaller scale than nitrogen, for example. Nutrients and

Table 6. Micro-organisms mediated heavy metal oxidation-reduction reactions (modified after Summers and Silver, 1978) Transformation

Metal

Reduction Oxidation Methylation Demethylation

As5% Cr6+, Fe3+, Hg% Hg2+, Mn4% Se4% Te4* As3*, Cr3+, Fe°, Fe2+, Mn2+, Sb 3t As5*, Cd2+, Hg2+, Se4+, Sn2+, Te4+ Hg2+

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W.J.G.M. Peijnenburg

heavy metals, such as cations or anions, are taken up by all plants via the pore water. Roots produce or excrete organics as root-exudates. These compounds may strongly affect the availability of contaminants. Figure 10 shows the role of roots. Uptake by roots of organics such as POPs and PAHs is, in general, unnatural. Exceptions occur as the result of damaged roots and organics that behave as cations or anions. Although there are exceptions, it may be stated that uptake of POPs and PAHs via plant roots and their subsequent transport to the above ground parts of the plants is of limited importance for leafy crops. For these contaminants and these plants it is atmospheric deposition on the leaves and subsequent transport through the cuticle that determines accumulation. The uptake of vegetation may be stimulated as well as hindered, or protected, by its mycorrhiza. The role of root exudates is manifold, such as increased availability of contaminants and a higher microbial activity, called the rhizosphere effect. The chapter by Ernst on vegetation and organic matter provides more information. 9.4 Bioavailability and bioaccessibility Due to the tendency that available fractions of contaminants in soil are considered to be far more relevant than the total fraction, extensive attention will be paid to the determination of their bioavailable and -accessible fractions. Persistence affects the potential for NA and environmental exposure because persistent chemicals exhibit higher concentrations per unit emission. A key role in the processes governing the persistence and toxicity of contaminants is

Figure 10. Schematic overview of the interactions at the root-soil-pore water interface determining uptake of nutrients and metals in particular

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played by the phenomena of bioavailability and bioaccessibility. Bioavailability may be defined as the fraction of the total amount of a chemical present in soil that, within a given time span, is either available or can be made available for uptake by (micro)organisms or plants, from either the organism's direct surrounding or the plant, or by ingestion of food. This definition of bioavailability implicitly suggests that the 'bioaccessible fraction' is defined as the fraction of the total amount of a chemical present in ingested soil particles that, at maximum, can be released during digestion. Different factors affect bioaccessibility. For example, the bioaccessibility of metals and ionizable contaminants is expected to be highly dependent on the pH values in the different compartments of the gastrointestinal tract. As a consequence, the bioaccessible fraction of ionizable contaminants and metals present in ingested soil is, in general, larger for mammals than for soil-dwelling invertebrates like, for instance, earthworms that regulate gut pH at around neutral. Uptake of chemicals involves the passage of compounds across a biological membrane, mediated by a carrier or a single solute. Compounds may enter tissues through passive diffusion, facilitated diffusion and by active transport mechanisms. Passive diffusion is the major uptake process for many organic chemicals as well as some metals and organometals. The driving force for uptake is a fugacity difference between water and the organism, as explained on the basis of EP theory. The fate and toxic potential of metals is, however, governed by more refined interactions at the micro level. Ultimately it is the free metal ion that is supposed to be capable of traversing biological membranes. As a consequence, metal availability and toxicity are functions of water chemistry, since speciation determines the free metal ion activity. It is the free metal ion concentration that often provides a better indication of availability and toxicity (Sunda and Guillard, 1976; DiToro et al., 2001) than total or dissolved concentrations. Studies with aquatic organisms (Campbell, 1995; Hare and Tessier, 1996), invertebrates (Kiewiet and Ma, 1991) and plants in nutrient solutions (Lexmond and Van der Vorm, 1981) revealed that metal uptake is also influenced by protons (H+) competition and divalent macro-ions such as Ca2+ as illustrated earlier in Figure 8. Morel (1983) and Pagenkopf (1983) introduced models in which the interactions of chemical species with organisms are used to predict trace metal uptake and toxicity. Most commonly this approach is called the free ion activity model (FIAM). The FIAM is gaining popularity in studies of soilplant relationships (Parker et al., 1995), even though some exceptions are known to exist (Campbell, 1995). Recent research (Gorsuch, 2002), has advanced

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the current level of understanding of metal availability to aquatic life via aqueous exposures, and the understanding of how to relate exposure levels to effects. The relationships between chemical forms of metals and the bioavailability of the metals are not sufficiently understood for terrestrial plants and soil-dwelling organisms, to permit the prediction of metal toxicity on the basis of the chemistry of the pore water. The examples given conceal a typical, fundamental difference between exposure in aqueous systems and in soils. Terrestrial plants may well be exposed in aqueous nutrient solution. Various data have been reported which either make clear the validity of FIAM, or from which additional chemical species that are actually taken up may be deduced. These species include complexes of metals, chelating agents, natural organic ligands, such as humic acids, and inorganic ligands such as chloride. There are very few studies with soil-grown plants, and often it is difficult to separate the actual ionic species that contribute to enhanced uptake. In addition, the soil itself exerts a diffusion limitation to metal transport because of the strong nature of metal binding to the solid surface and the tortuous nature of soil pores. In soils, the relationship between free metal activity and metal uptake is, therefore, not as close a relationship as had previously been assumed. It can be concluded that membrane or dermal uptake of contaminants can be described by the EP theory (Shea, 1988). However, regulated uptake of nutrients and essential elements is different and also the uptake of contaminants via the digestive tract of soil biota representatives. Also organisms responding to other environmental conditions may have different effects on uptake. This is shown for earthworms in Jager's (2003b) thesis. For example, plants will physically access more metal in the soil when root growth rates are high, so that temperature and nutrition may affect plant metal uptake directly through their effects on plant growth (McLaughlin, 2001). 9.5 Monitoring techniques for bioavailability and bioaccessibility Monitoring implies repeated measurements in time at fixed places. 9.5.1 General

The equilibrium-partitioning concept provides a proper means of operationalizing the definitions of bioavailability and bioaccessibility. The liquid phase, the particulate phase and the biotic phase play leading roles here. Equilibration processes are assumed to take place between all phases. The 'biotic phase' consists of a variety of species, each with a characteristic set of exposure routes and specific characteristics. They include relatively long

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equilibration periods as compared to equilibration periods for most physicochemical processes, concentration-dependent occurrence of toxic responses, and the role of some equilibria in determining toxicant uptake. The challenge of monitoring bioavailable and bioaccessible fractions thus boils down to linking chemistry in the solid phase and in solution, to uptake by biotic species. It determines whether or not toxic effects will actually occur. This affects the choice of extractant used to mimic bioavailability. Since there is not a single bioavailable or bioaccessible fraction, the terms potentially and actually available fractions have been introduced to enable the inclusion of the aspect of time. Any monitoring strategy to be developed should aim at providing estimates of the fractions that are potentially and actually available or accessible. 9.5.2 Assessment methods

Methods developed for measuring (mimicking) bioavailable and bioaccessible fractions can be grouped into three categories. Methods for assessing contaminants in (pore) waters - in the case of metals this includes an assessment of speciation and the activity of the free metal ion; single (solvent based) and sequential extraction methods, and methods with rigorous digestion procedures to determine total concentrations, including selective oxidation of organic matter and thermal desorption. The methods for assessing contaminants in (pore) water may provide an estimate of the fraction actually available and accessible, whereas extractions and digestions may provide estimates of potentially available and accessible fractions. Nutrients The search for chemical methods to determine the concentration of individual plant-available nutrients in agricultural soils began several decades ago. As a result of research focussing on estimating the quantities of nutrients that should be added to soil to achieve maximum crop yield, chemical methods were derived that provide reasonably predictive data of the bioavailability of inorganic ions necessary for plan development. Soil water can be analysed directly to determine the fraction of phosphate actually available. Certain extractants can then be employed to determine the phosphorus fraction potentially available. It is possible to extract specific mineral fractions of phosphate, and total phosphorus concentrations are determined after complete digestion with strong acids.

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POPs

Various extraction methods have been developed to estimate the (bio)available fraction of contaminants. These include Tenax-beads to extract the contaminant pool resulting in a multi-phase desorption profile of which the first, fast kinetic phase is considered the bioavailable fraction. Solvent extractions resulting in less vigorous extraction than total extraction with organic solvents and solventwater mixtures have been applied and related to uptake of PAHs and atrazine in earthworms. Results of these studies show an ageing effect for both solvent extraction and the uptake of these chemicals by the organisms. Besides solvent extractions, extractions with passive samplers such as C18-disks and semipermeable membrane devices (SPMDs) have also been used. Solid phase micro extraction (SPME) is another passive sampler that has been used in water, sediment and soil samples for some time now and SPME-derived pore water concentrations have been correlated to body burdens of contaminants in biota (Van der Wai, 2003). Metals The analytical determination and chemical equilibrium models exist for determining metal speciation in solution. Models are based upon experimentally derived parameters, and models cannot be any better than the assumptions and data on which they are based. Computational procedures, based on thermodynamic principles, allow the equilibrium speciation of a system to be calculated once total component concentrations are known. Various techniques can be used to measure free metal activity in solution. The most direct method for determining free metal species is possibly through ionselective electrodes (ISEs). Voltammetric methods such as differential pulse polarography (DPP) and differential pulse anodic stripping voltammetry (DPASV) are sensitive electrochemical methods and can be used to determine the concentrations of 'labile' metal. Exchange resins are also used to measure the free metal activity or the relative lability of metals (Beveridge et al., 1989). Donnan dialysis membranes also can be used to determine the speciation of solutions. Minnich and McBride (1987) have for instance used this technique to determine the free Cu2+-activity in Cu-salts and sewage sludge-amended soils. Recently, Temminghoff et al. (2000) developed the Wageningen-Donnan Membrane Technique for directly measuring metal activities. A separation technique via Donnan equilibrium across a negatively charged ion-exchange membrane was adapted for this purpose. Diffusive gradients in thin films (DGT) have been used to assess bioavailable forms of metals. The technique utilises a gel layer to remove ions from solution. This establishes a gradient. A

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resin at the base of the gel arrests the diffusion of ions. The DGT technique is supposed to allow free diffusion of labile species of metals (Zhang et al., 1998; Davison et al v 2000). Finally, information on metal speciation may be gained by separation and/or exchange methods. For instance, Gregson and Alloway (1984) used gel permeation chromatography to separate species of Pb based on molecular weight. Extractions represent an intermediate position between total digestions and (pore) water samples. Research concerning metal behaviour in soils has focused mainly on the use of various chemical extractions to describe forms of metal present. Generally speaking, six types of extractions may be distinguished (Table 7). In addition to single extractants, sequential extraction is often performed to differentiate the metal fractions present in the soil. A widely used version of a sequential extraction procedure was first formulated by Tessier et al. (1979). An example of a sequential extraction scheme is shown in Table 8. The most common approaches to total soil metal estimates aim at

Table 7. Overview of commonly used extraction techniques for extracting heavy metals from soil Extraction type (Weak) salt extractions

Reductive extractants Weak acid extractions Strong complexation methods Dilute strong acids Combined extractants

Examples of reagents used CaCk, Ca(NO3>2, NBitAc, NaNCfe, Mg-salts, BaCk, in concentrations from as low as 0.001 M and up to 1 M salt solutions sodium-ascorbate, hydroxylamine-HCl, sodium dithionite acetic acid, citric acid DTPA-TEA, EDTA, NTA HNO3, HC1, 'double acid' (HC1 + H2SO4) Ammonium oxalate-oxalic acid, Mehlich III (dilute acid, salt and EDTA)

Table 8. A general Tessier scheme of seauential extraction (Tessier et al., 1979) Extractant

Fraction obtained

1 M MgCb or 1 M NaOAc 1 M NaOAc+HOAc Reducing extractants like Na2S2O4 (hydrosulfite) + Na oxalate, or Na-hydroxylamine.HCl Oxidizing agents HNO3+ H2O2, later NH4OAC to prevent readsorption Digestion HF+HCIO4

exchangeable fraction bound to carbonates bound to Fe-Mn oxides, or easily reducible bound to organic matter residual fraction

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determining total recoverable metal. Many total metal digestions involve hydrofluoric acid additions to the digest to break down Si crystal lattices and extract metal bound or trapped in these crystal structures. Total recoverable metal digestions are typically designed to stop short of destroying crystal structures. Soil digestions also usually employ elevated temperatures. In recent years microwave digestion methods have gained popularity because of increased safety and shorter digestion times. Three USEPA microwave digestion procedures exist and include HNO3, HNO3-HCI and HNO3-HCI-HF respectively (USEPA, 1997). USEPA states that the first two procedures are designed to extract metals that can, potentially, become environmentally available. 9.5.3 Linking chemistry to biology

Element uptake is controlled by chemical availability in solution as well as the capacity of the soil or labile fractions present in solution to supply that element. Most plants and biota will accumulate many times the amount of metal available in solution at any given moment. In effect the soil solution is emptied and replenished. Uptake of contaminants is, therefore, not only dependent on the availability of the chemical in solution (intensity factor) and the uptake mechanisms, but also on the capacity of the soil solid phases to supply the fluxes of that particular element (capacity factor). Understanding bioavailability for species exposed via the pore water requires the consideration of both aspects: the intensity of exposure through the EP-concept or the FIAM, and the capacity of the soil to maintain this level of free contaminant in solution. The consequence for daily practice for risk assessment, the approach can on a regular basis only be applied for the aqueous compartment of ecosystems. At present, several programs directed at incorporating further refinements into the BLMs for copper, zinc and silver are in progress in order to extend the application of these BLMs to other types of organisms, thus making the BLMs more suitable for use in the development of water quality criteria. Application of the BLM framework to soil biota is in the early stage of development studies. An elegant recent approach for measuring effective soil solution concentrations and metal supplied from the solid phase, is the technique of diffusive gradients in thin films (DGT). This technique of surrogate chemical measurement, first reported by Davison and Zhang (1994), has promise as a quantitative measure of the effective bioavailable metal for plants, as measured DGT fluxes of Cu relate well to Cu uptake by Lepidium heterophyllum (Zhang et al., 2001). A second general approach towards relating available and bioavailable fractions to uptake is by means of multifactorial analyses. In this approach,

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accumulation characteristics like uptake rate constants, steady-state concentrations, bioconcentration factors, and biota to soil accumulation factors are statistically associated with water or soil characteristics such as the total concentration, rapidly or slowly desorbing contaminant fractions, pH, organic matter content, dissolved organic carbon content, and the cation exchange capacity (CEC). Observations are typically analysed by multiple regression analyses, to weigh the impact of these characteristics against each other. It should be noted that this general approach does not necessarily provide any insight into mechanistic influences of any of the variables mentioned on metal uptake by plants or organisms. Nevertheless, such approaches allow eliminating uptake routes that do not contribute significantly. Thirdly, instead of total soil concentrations, operationally defined extraction techniques have been used with the aim of assessing the pool of solid-phase contaminant that buffers the solution concentration, thus mimicking the concentrations of labile contaminants that are available for uptake by biota or plants, and accounting for the supply rate term capacity factor. Various successful examples have been published. Often there are large discrepancies among extraction techniques with regard to simulating bioavailable and bioaccessible fractions. Whereas one specific extraction technique may provide a good indication of the bioavailable or bioaccessible fraction for a specific species or plant, there may not be any correlation at all for other species or plants using the same extraction technique. In addition, when applied to soils with a broad spectrum of contaminant concentrations or soil characteristics, they often fail to adequately predict metal availability or accessibility. Therefore it is clear there is no single expression for bioavailable or bioaccessible fractions. 9.6 Towards a monitoring strategy The findings summarised above illustrate that the consideration of an EP approach as the basis for monitoring is an overly simple approximation of reality. A large number of simulation approaches for bioavailability and bioaccessibility are currently available and validation of the chemical techniques by biological uptake and toxicity experiments yield variable results. Whereas the paradigm of the free ion in solution in the case of metals or the non-complexed and truly dissolved POPs and nutrients as being the available species, was confirmed for some organisms or plants, other organisms or plants do not adhere at all to this paradigm. Instead, experiments with the latter biota clearly warrant the use of multivariate approaches and the plethora of extraction techniques developed so far. Although there is a large variance in species-specific, compound-specific and compartment-specific bioavailable and

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bioaccessible fractions, determining the most available or accessible pools (while not at all perfect) will provide a better prediction of risk than total concentrations or contaminants extracted by strong extractants. However, one should be aware that especially indices of actual availability or accessibility can change over time. This may be due to long-term changes in pH, redox conditions or organic matter content. Indices of potential availability or accessibility will respond less, and hence any monitoring strategy should take potentially available fractions into account. Total concentrations or compounds extracted by strong extractants provide a worst-case estimate of possible long-term changes and should therefore be taken as a starting point for monitoring, irrespective of the goal of monitoring. Before designing a monitoring strategy, it is of utmost importance to have clarity about the purpose of the monitoring programme. Issues to be considered include the composition of the ecosystem of interest in terms of environmental properties (like OM, DOC, pH, redox condition, etc.), the end point of determination (ecosystems versus specific populations, species or even individual organisms or plants), the relevant uptake and exposure routes as well as the chemicals of interest. In common practise such information will often be lacking, and instead of focussing on actually available or accessible fractions, the focus will therefore be the potentially available or accessible pools. A general monitoring strategy is depicted in Figure 11, and further elaborated in Table 9. Depending on the pre-set end point of determination and the critical total levels related to this end point, the first step of the monitoring strategy provides for a worst case approach. This can be used to deduce whether a worst-case estimate of the potentially available and accessible pool of contaminants is likely to be indicative of adverse effects. The limit indicated in Figure 11 against which the measured total content of contaminant of the potentially available contaminant concentration is to be judged, can be many. Examples include any environmental quality objective that has been set for the chemical of interest, any adverse effect level considered to be relevant for the specific site of contamination, and any effect level that is related to the objective of protecting individual species (like for instance endangered soil organisms or plants), populations, distinct parts of ecosystems, or human beings. In the case of the effect level against which contaminant concentration data are to be compared, it is preferable to use chronic effect data, rather than acute effect data. If the pre-set end point of determination is exceeded, then it must be determined in the next step whether the composition of the matrix of interest in terms of properties affecting potential availability or accessibility - will

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Table 9. Monitoring table Total content, including bound residues Components ; Methods available POPs and PAHs

Harsh extraction, like prolonged soxhlet[extraction | |

[End point of [consideration

Soil Iproperjties of interest

bM |Any quality jobjective or adverse I leffect level based on i jtotal concentrations. i lln addition, total hatural background

i

Examples of suited methods

;End point of ^consideration

-Tenax-extraction lAny adverse -Less vigorous solvent leffect level extractions based on -Extractions with C18- iextractable disks and SPMDs [concentrations -SPME

levels are to be taken! jinto account Iwhenever relevant

rleavy metals [Digestion, by lAny quality jmeans of jobjective or adverse [strong HNO3, [effect level based on |HNO3-HC1, Itotal concentrations. |HNO3-HCI- lln addition, total |HF hatural background

IpH, OM, -Weak salt extractions |Any adverse jclay -Reductive extractants leffect level jcontent, -Weak acid extractionsbased on Eh -Strong complexation iextractable

levels are to be taken! linto account Mutrients

Extractable contaminant content

Harsh jAny quality extraction or iobjective or adverse digestion leffect level based on itotal concentrations. ITotal natural backiground levels are to be taken into iaccount.

methods |concentrations, -Dilute strong acids lusing extrac-Combined extractants|tants of similar Extraction Strength Usually not relevant

|-

pocicontent |of pore [water, Eh, OM

Eh = redox potential OM = organic matter

sufficiently reduce the potential adverse impact of the total pool. If it is judged that the potential adverse impact is not sufficiently reduced by the composition of the soil matrix, then a large battery of experimental techniques for assessing the various contaminant pools may be applied. Depending on the end point under consideration and the compound-, matrix-

Figure 11. General outline of a monitoring strategy supporting risk assessment

and species-specific factors identified above, the proper set of measurement techniques is to be selected in order to make sure that site-specific impacts are adequately taken into account. 9.6 Concluding remarks Assessment of the potential and actual adverse effects of contaminants on soil ecosystems requires that the characteristics of both the soil environment and the contaminants are taken into account. A simplified strategy that takes these elements into account is to combine expressions of total amounts of pollutants present with expressions of bioavailable fractions. For the three classes of contaminants considered combinations of commonly available methods of analysis are recommended in the integrative assessment of soil quality and soil functioning. For POPs (including PAHs) the soxhlet extraction to determine potential risks, combined with tenax extraction to assess actual adverse effects. For heavy metals it is recommended to combine HNCb-extraction with a weak salt extraction (0.01 M CaCk). The pH of the soil, OM-content, clay content, and natural background levels need to be determined in order to make sure that expected adverse effects are not merely related to the composition of the soil, or to the varying ecological preferences of the species present in the undisturbed soil ecosystem). In the agricultural context many methods have been developed to assess the nutrient status. HN03-extraction will be sufficient to assess the potential impact of either increased or decreased levels of nutrients. Suitable reference soils should be included in the monitoring programme.

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Monitoring results of potential and actually available fractions that are mutually supportive, imply that a decision on the next steps in terms of more detailed risk assessment can immediately be taken. More detailed research, including expert judgement, is required where the expressions of potentially and actually available and accessible contaminants give mixed results. 9.7 Implementation in soil management The final question to be addressed is what are the consequences for practice. As far as the nature and behaviour of contaminants are concerned it boils down to the ultimate fate of the soil. Similar processes are also involved in areas called "City Green" (small green areas within the city) and "Ecological Main Structure" (large green locations outside cities). For practical purposes the factors that control contaminant behaviour have to be monitored. The fate of nutrients, heavy metals and persistent organic contaminants in soil have been described and explained in relation to recycling possibilities and soil characteristics such as organic matter, mineral fraction and environmental conditions such as pH and redox-conditions. These controlling factors can be quantified on a routine base, but they can also be modelled. Examples here include a model for extractability, bio availability for earthworms or bio-availability for moles, provided the uptake procedure of respectively earthworms and moles is included. In the foregoing it is shown that the knowledge, data and key characteristics on compound behaviour are available for application in soil management, provided the specific equilibrium partitioning pitfalls have been considered. The implementation itself is the responsibility of the one asking the questions: What is it you want? And of course there should be validation and verification of the model outputs, as unexpected differences can be better analysed in the field than from behind a computer screen. References Abramowicz, D.A., 1990. Aerobic and anaerobic biodegradation of PCB: a review. Crit. Rev. Biotechnol. 10:241-251. Adriaens, P., Q. Fu and D. Grbic-Galic, 1995. Bioavailability and transformation of highly chlorinated dibenzo-p-dioxins and dibenzofurans in anaerobic soils and sediments. Environ. Sci. Technology 29: 2252-2260. Alcock, R.E., M.S. McLachlan, A.E. Johnston and K.C. Jones, 1998. Evidence for the presence of PCDD/Fs in the environment prior to 1900 and further studies on their temporal trends. Environ. Sci. Technol., 32,1580-1587. Bachmann, A., P. Walet, W. de Bruin, J.L.M. Huntjens, W. Roelofsen and A.J.B. Zehnder, 1988.

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Biodegradation of alpha and beta hexachlorocyclohexane in a soil slurry under different redox conditions. Appl. Environ. Microbiol. 54: 143-149. Beller. H.R., D. Grbic-Galic and M. Reinhard, 1992. Microbial degradation of toluene under sulfatereducing conditions and the influence of iron on the process. Appl. Environ. Microbiol. 58: 786-793. Beveridge, A., P. Waller and W.F. Pickering, 1989. Evaluation of 'labile' metal in sediments using ion exchange resins. Talanta. 36, 535-542. Bollag, J.M., 1992. Decontaminating soil with enzymes. Environ. Sci. Technol. 26, no 10,: 1876-1881. Bunge, M. I. Kahn, J.D. Wallis, E.K. Miller and A.D. Wagner, 2003. Reductive dehalogenation of chlorinated dioxins by anaerobic bacterium. Nature 401, 23 January. Campbell, P.G.C., 1995. Interactions between trace metals and aquatic organisms: A critique of the free ion activity model. In "Metal speciation and bioavailability in aquatic systems" (A. Tessier, and D.R. Turner, Eds.), pp. 45-102. IUPAC Series on analytical and Physical Chemistry, John Wiley, New York, NY, USA. Cerniglia, C, 1992; Degradation of PAH's in soil, state of art of U.S.-research work. Dechema; 6-9 dec. 1992; Karlsruhe. Davison, W., and H. Zhang, 1994. In situ speciation measurements of trace components in natural waters using thin-film gels. Nature, London, 367,1995, p.546-548. Davison, W., P.S. Hooda, H. Zhang and A.C. Edwards, 2000. DG measured fluxes as surrogates for uptake of metals by plants. Adv. Environ. Res. 3, 550-555. De Vries, W. and A. Breeuwsma, 1987. The relation between soil acidification and element cycling. Water Air Soil Pollut. 35: 293-310, DiToro, D.M., CD. Kavvadas, R. Mathew, P.R. Paquin and R.P. Winfield, 2001. The persistence and availability of metals in aquatic environments. International Council on Metals and the Environment, Ottawa, Canada. Fritz, W. and R. Engst, 1975. Neuere Ergebnisse zur umweltbedingten Kontaminanten von Lebensmittel mit Krebserregenden Kohlenwasserstoffen. Techniek und Umweltschutz 10: 148-172. Fuhr, F., 1984. Praxisnahe Tracerversuche zum Verbleib von Pflanzenschutzwirkstoffen in Agragokosystemen. Rheinisch-Westfalische Akademie der Wissenschafften, Vortrage N326: 5-35. Gorsuch, J.W. (ed.), 2002. Comparative Biochemistry and Physiology. Part C: Toxicology and Pharmacology. Special issue: The Biotic Ligand Model for Metals-Current research, Future directions, Regulatory implications. Amsterdam: Elsevier Science. Gregson, S.K. and B.J. Alloway, 1984. Gel permeation chromatography studies on the speciation of lead in solutions of heavily polluted soils. J. Soil Sci. 35, 55-61. Halden, R.U. and D.F. Dwyer, 1997. Biodegradation of dioxin related compounds: a review. Bioremediation J. 1: 11-25 Hare, L., and A. Tessier, 1996. Predicting animal cadmium concentrations in lakes. Nature. 380, 430-432. Holliger, C, 1992. Reductive dehalogenation by anaerobic bacteria. PhD Thesis LUW. Hutchins. S.R., 1991. Biodegradation of mono aromatic hydrocarbons by aquifer micro-organisms using oxygen, nitrate or nitrons oxide as terminal electron acceptor. Appl. Environ. Microbiol. 57: 2403-2407.

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Jager, T., 2003a. Worming your way into bioavailability. Modelling the uptake of organic chemicals in earthworms. Ph D-Thesis, University of Utrecht, Utrecht, the Netherlands. Jager, T., 2003b. Worming your way into bioavailability. Modelling the uptake of organic chemicals in earthworms. Ph D-Thesis, University of Utrecht, Utrecht, the Netherlands. Kiewiet, A.T., and W.C. Ma, 1991. Effect of pH and calcium on lead and cadmium uptake by earthworms in water. Ecotoxicol. Environ. Saf. 21, 32-37. Kjeller, L.O., K.C. Jones, A.E. Johnston and C. Rappe, 1991. Increase in the polychlorinated dibenzop-dioxin and -furan content of soils and vegetation since the 1840s. Environ. Sci. Technol., 25, 1619-1627. Lexmond, T.M. and P.D.J. van der Vorm, 1981. The effect of pH on copper toxicity to hydroponically grown maize. Neth J Agr Sci. 29, 209-230. Mackay D., W.Y. Shiu and K.C. Ma, 1999. Physical-chemical properties and environmental fate degradation handbook. (CD-ROM). Boca Raton, FL, USA: Chapman & Hall CRCnetBASE, CRC. McLaughlin, M.J., 2001. Bioavailability of metals to terrestrial plants. In "Bioavailability of metals in terrestrial ecosystems; Importance of partitioning for bioavailability to invertebrates, microbes and plants" (H.E. Allen, Ed.). SETAC Press, Pensacola, Fl, USA. Meer, J.R. van der, 1992. Molecular mechanisms of adaptation of soil bacteria to chlorinated benzenes. PhD Thesis Wageningen University. Minnich, M.M. and M. McBride, 1987. Copper activity in soil solution: I. Measurement by ionselective electrode and Donnan dialysis. Soil Sci. Soc. Amer. J. 51, 568-572. Morel, F.M.M., 1983. Principles of Aquatic Chemistry, 1st ed., John Wiley, New York, USA. Oldenhuis, R., 1992. Microbial degradation of chlorinated compounds. PhD Thesis Groningen University. Pagenkopf, G.K., 1983. Gill surface interaction model for trace metal toxicity to fishes: role of complexation, pH and water hardness. Environ. Sci. Technol. 17, 342-347. Parker, D.R., R.L. Chaney and W.A. Norvell, 1995. Chemical equilibrium models: Application to plant nutrition research. In: "Chemical equilibrium and reaction models" (R.H. Loeppert, A.P. chwab and S. Goldberg, Eds.). SSSA Special Publication Number 42, Madison, WI. Soil Science Society of America Inc., American Society of Agronomy Inc. Paul, E.A. and F.E. Clark, 1989. Soil microbiology and biochemistry. Academic Press, San Diego, Calif. Quensen. J.F., J.M. Tiedje and S.A. Boyd, 1988; Reductive dechlorination of polychlorinated biphenyls by anaerobic micro-organisms from sediments. Science 242: 752-754. Salomons, W., 1995. Long-term strategies for handling contaminated sites and large-scale areas. In: Biogeodynamics of pollutants in soils and sediments; editors: Salomons& Stigliani. SpringerVerlag Berlin Heidelberg New York, pp 1- 30. Shea, D., 1988. Developing national sediment quality criteria. Environ. Sci. Technol. 22,1256-1261. Summers A.O. and S. Silver, 1978. Microbial transformation of metals. Ann. Rev. Microbiol. 32: 637-672. Sunda, W.G., and R.R.L. Guillard, 1976. Relationship between cupric ion activity and toxicity of copper to phytoplankton. J. Marine Res. 34, 511-529. Swaine, D.J., 1955. The trace element content of soils. Commonwealth Bur. Soil Sci. Tech. Commun. 48,157 pp. Temminghoff, E.J.M., A.C.C. Plette, R. van Eck and W.H. van Riemsdijk, 2000. Determination of the

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chemical speciation of trace metals in aqueous systems by the Wageningen Donnan Membrane Technique. Ann. Chim. Acta. 417, 149-157. Tessier, A., P.G.C. Campbell and M. Bisson, 1979. Sequential extraction procedure for the speciation of particulate trace metals. Anal. Chem. 51, 844-851. USEPA, 1993. definition of natural attenuation. USEPA, 1997. Microwave assisted acid dissolution of sediments, sludges, soils and oils. 2nd ed. In: "SW846 Test methods for evaluating solid waste, physical/chemical methods". USEPA Publication No. 955-001-00000-1. Washington, DC, U.S. Government Printing Office. Van der Wai, L., 2003. Bioavailability of organic contaminants in soil. Solid phase micro-extraction predicts uptake in Oligochaetes. PhD-Thesis, University of Utrecht, Utrecht, the Netherlands. Warith, M.A., R. Ferehner and L. Fernands, 1992; Bioremediation of organic contaminated soil. Hazardous Waste Haz. Mat. 9: 137-177. Werff, van der, P.A., 1992. Applied soil ecology in alternative agriculture. Reader University Wageningen, the Netherlands. Wessel, W.W., 1997. Metal nutrient dynamics in organic surface layer of an acidifying forest soil. PhD thesis University of Amsterdam; ISBN: 90-6787-050-1. Wilson, S.C. and K.C. Jones, 1993. Bioremediation of soil contaminated with polynuclear aromatic hydrocarbons (PAHs): a review. Environ. Pollut. 81: 229-249 Zhang, H., W. Davison, B. Knight and S. McGrath, 1998. In situ measurements of solution concentrations and fluxes of trace metals in soils using DGT. Environ. Sci. Technol. 32, 704-710.

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Chapter 10

ECOLOGICAL SOIL MONITORING AND QUALITY ASSESSMENT A. Breure

Abstract

This chapter describes the huge variety of attempts made during the past 50 years to develop a soil quality monitoring system. The monitoring of agricultural performance began around 1950 and lead to advice on optimizing crop production and information about the presence of pest organisms in the soil. Since the 1990s RIVM has managed the Dutch Soil Quality Network (DSQN), with the aim of obtaining policy information on soil status trends for about 70% of the surface area of the Netherlands. Nowadays, within EU, OECD, FAO and several European countries there are various initiatives to come to an ecological characterization of the soil. The sustainability of ecosystems should always be described in terms of their interaction with the economy and society. The ecological economic model distinguishes four types of capital: natural, human, manufactured and social. For general monitoring the most well known and, therefore applicable biological characteristics are microbial biomass, nematode diversity and earthworm numbers and diversity. The diversity and abundance of higher soil organisms provide more insight into the stability and resilience of the system in this authors view. A tailor-made monitoring system for ecological systems follows the demands and the desires of the users. Research has to be done to obtain more quantitative insights. If the ecological quality of the soil has to be derived from the "fitness for use", there are major knowledge gaps. It is possible to derive indicators for status and trends, based on ecological insights proxies, models, and monitoring data. A major bottleneck is the collection of monitoring data. There is an urgent need for data.

10.1 Importance of ecological functions of the soil Soil is a scarce resource in the Netherlands and in other urbanized, ruralized and industrialized countries. Environmental stewardship should therefore address soil quality management. Nowadays, environmental policy is increasingly focussed on the sustainable use of the soil, both in the Netherlands and in international organizations, such as in the European Union (EU), and the Organization of Economic Co-operation and Development (OECD). That means

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that the current use of soil should meet the needs of the present without compromising the ability of future generations to meet their own needs, or compromising the use of soil elsewhere (WCED, 1987). Soil is a key component of the earth's environment. It forms the interface between the hydrosphere, atmosphere and the organisms inhabiting it, and it regulates natural material and energy cycles. Soil is sensitive to climatic impacts and to human and historical activities. Therefore, its structure and characteristics are the product of an age-old process, making it a non-renewable resource. Soil has different functions and in many of them soil organisms play an important role. A number of processes occur in the soil that may be described as ecosystem services or life support functions. These processes include the transfer and cycling of energy and matter among trophic groups and with the abiotic environment. They include the degradation of organic matter, nutrient cycles and degradation of pollutants, which is very important for groundwater quality. The soil contains a huge amount of organisms with a high biodiversity that live in close interaction with each other and with the above ground organisms. The sub-soil processes are performed by consortia of organisms, including bacteria, fungi, protozoans, nematodes, oligochaets, and arthropods. These organisms are also largely responsible for the aggregate formation and the structure of the soil, governing water holding capacity and facilitating plant rooting. Moreover soil plays an important role in hydrology, it carries the buildings and infrastructure and it is the substratum for agriculture and nature. Soil organisms interact in a soil food web, where each trophic layer is food for the next trophic layer. In general a soil food web is based on the degradation of roots and dead organic material. The stability of the performance of an ecological function is dependent on the stability of the soil food web. There are several definitions of stability, but more interaction stabilizes the process. It may also lead to more fluctuation in species abundances. When predators have increasing choice in their food, the chance that they will prey on a specific species until it is extinct decreases. (Polis, 1998; Hanski, 1997). This is important in terms of the biodiversity of soil organisms. In the chapters by Verstraete and Mertens (5), Van Straalen (6) and De Ruiter (7) much has been written about all these processes in the soil and functions of the soil, and this will not be repeated here. 10.2 Recognition of the importance of soil ecological processes in policy and legislation The importance of the biotic components of soil is recognised in soil protection

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policy and can be found back in legislation. Article 1 of the Dutch Soil Protection Act contains a number of definitions. The definition of soil is "the solid part of the earth with the aqueous and gaseous compounds and organisms therein". Article 1 also confirms the importance of protecting the soil by the preventing, reducing or undoing of changes in the properties of the soil that might result in a decrease or threat to the functional properties of the soil for man, plant or animal. In other words, soil and soil-biodiversity should be protected because of their functional properties. In the United Nations Convention on Biological Diversity (UNCBD) the first and most important reason for protecting biodiversity is its intrinsic value. This rational is similar to that underlying legislation in Germany where the purpose of the Federal Soil Protection Law is stated in its paragraph one and is defined as the "protection or restoration of the functions of the soil on a permanent sustainable basis". These actions shall include the prevention of harmful soil changes, rehabilitation of the soil, of contaminated sites and of waters contaminated by such sites; and taking precautions against negative soil impacts. Where impacts are made on the soil, disruptions of its natural functions and of its function as an archive of natural and cultural history should be avoided as far as possible. Harmful soil changes are impacts on soil functions that may result in hazards, considerable disadvantage or nuisances for individuals or the general public. In both soil-protection Acts, the functions profitable for man - are mentioned first. The German Soil Protection Act also sums up the most important ecological functions: - a basis for life and a habitat for people, animals, plants and soil organisms; - a part of natural systems, especially through its water and nutrient cycles; - a medium for decomposition, balance and restoration as a result of its filtering, buffering and substance-converting properties, and especially groundwater protection; a function as an archive of natural and cultural history; - a storage place for raw materials; - a surface for living and recreation; - a place for agriculture and forestry; a place for other economic use, traffic, waste disposal. This definition can also be found in European objectives for soil policy. A general European policy on soil is to be launched because of the nature of soil as a non-renewable resource and its functions that require it be used in a sustainable way. Ecological functions and soil biodiversity are important because of their role in structure formation, the stability of structure and

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functions, fertility, buffering and in providing the potential for using the soil as a carbon sink - important in the United Nations Framework Convention on Climate Change (UNFCCC). 10.3 What is ecological quality of the soil? Soil quality can be defined as the capacity of soil to function within ecosystem boundaries such that productivity and environmental quality are maintained and plant health and animal health are promoted (Doran and Parkin, 1994; Stenberg, 1999). Within ecosystem boundaries means that there are differences between soils, partly due to soil use. The definition is focussed on soil function and benefits for mankind. There is no argument for biodiversity conservation. There are no absolute quality criteria and the quality of each soil must be determined in relation to natural variability in soil type, soil use and climate. The ecological quality of soil may be defined as the extent to which the composition and functioning of the soil ecosystem meets the quality criteria (fitness for the proposed use). It is hard to answer the question of which criteria have to be met before soil quality can be characterised as good or bad. In fact this question cannot be answered objectively, and social choices have to be made. Policymakers have to perform this process of choice. The outcome should be based on ecological insights and social motives. The sustainable use of a soil implies a minimal shift of negative effects to the future and the surroundings. Here it is important to consider the physical, chemical and biological characteristics of the soil in an integral way, taking into account technical, social and economic factors. 10.4 Determination of ecological soil quality 10.4.1 Why European soil policy aims at a sustainable use of the soil (resolution European Parliament 19 November 2003). It is thought that the development of the ecological quality of the soil might be a good measure for the extent to which soil use can be said to be sustainable. This requires that it is possible to measure such quality, including its chemical, physical and biological aspects. Many ecosystem functions are related to the organisms present in the soil. The relation between the biodiversity in the soil, environmental pressure and the functioning of the soil ecosystem is very complex and is not always the same. A considerable amount is known about the qualitative relationships between ecosystem structure and functions. However, much research is necessary to clarify the quantitative relationships between the ecological quality

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of the soil, including soil biodiversity and the functioning of soil ecosystems, in relation to optimal and sustainable soil management. 10.4.2 How

The increasing need to quantify ecosystem quality for the assessment of sustainability demands a holistic approach, taking chemical, physical and biological characteristics of the soil into account. Such approaches recognise the complexity of ecological interactions and the importance of ecosystem processes as a reflection of underlying functions, including soil characteristics. 10.4.3 The need for indicators for ecological quality

Ecological quality is composed of many different variables and a complete description of the soil ecosystem will not be possible. Therefore, it is necessary to identify proxies, which are representative of ecosystem quality. An indicator ought to be derived from a statistical and/or mechanistic interpretation of data from these three characteristic domains that relate to soil use. Ecological indicators may be used to identify the possible reasons for change. There is a need to assess when ecosystem changes attributed to anthropogenic disturbances have gone beyond some threshold of acceptability. 10.4.4 Criteria for indicators

In general indicators should meet some pragmatic criteria. They have to answer a relevant question. That means that before an indicator may be developed the question of the user of the information must be clear. Indicators for ecological quality of soil should give insight into the structure and function of the soil ecosystem, in relation to soil use. The questions to be answered for political purposes relate to the sustainability of soil use, the resilience and functionality of the soil, changeability of land use and the effectiveness of policy and management. These subjects need a definition before they may be quantified. In the development of indicators for soil quality this means that information must be produced on: - state of the soil ecosystem, what is present; trends. Is the quality, which has to be defined separately, changing; - valuation. If quality criteria exist, status may be quantified in terms of good and bad, the indicator gives insights as to whether changes in the quality are positive (wanted) or negative (unwanted); - indication of future developments. An indicator must provide information, which opens up possibilities of influencing the wanted or unwanted value. That means that there must be a known relationship between the outcome of

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the indicator and human acting; - GIS. It is very useful to incorporate indicator information in a GIS system; - the indicator value must be unequivocal. It must be clear what the indicator value shows; appealing for politics and society. This is mainly a communication aspect. Policy makers and the public should regard the outcome as highly relevant; the indicator must be based on data that can be obtained in a cost-effective way. 10.5 Providing data for the indicators: monitoring activities Monitoring activities are necessary to provide the data necessary for tracking changes in the environment and ecosystems over time. Indicators need data for monitoring. It is possible to provide data for different indicators using the same monitoring data. Therefore, there must be an alignment between (collection of) monitoring data and the user of the information produced by the indicators. 20.5.2 The use of monitoring data for indicators

Considerable extra information is required before useful information can be derived from descriptive monitoring data. Roots (1996) describes a way of moving from observations to useful information using the staircase of knowing as illustrated in Figure 1. If observations are quantitative, recorded against some agreed standard that is independent of how we feel at the time and independent of what will happen to the result, we call those observations measurements. We say that measurements must be as objective as possible. However, when we combine these measurements and give them more meaning, we have to add a purpose, a selected methodology and some judgement of quality. In doing so, the result becomes increasingly subjective. Observations and measurements, when verified to agreed standards become data. Data, properly selected, tested and related to subject areas can become information. Indicators are developed to integrate and interpret monitoring data by use of ecological insights, statistical techniques, and models. That means that indicator outcomes, such as quality criteria, are subjective and partly the result of choices made by man. Scientists collecting the data and users of the information have to walk up and down the staircase continuously to communicate managerial questions and the interpretation of data and observations.

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Figure 1. The staircase of knowing (modified after Roots, 1996)

10.6 History of soil monitoring in the Netherlands The Netherlands is an agricultural country and as such has a long history in soil characterization. This resulted in the first geological soil maps produced in the middle of the twentieth century and finally in the Soil Map of the Netherlands at 1:50,000, which included ground water levels. This is kept up to date at Alterra1, in Wageningen. In addition there is agricultural monitoring. Since 1950 institutes have been responsible for giving fertiliser advice and information on the presence of pest organisms in the soil. Since the 1990s RIVM2 manages the Dutch Soil Quality Network (DSQN). The main aim of DSQN is to obtain policy information on trends in soil status. It deals with a selection of sample sites representative of 70% of the surface area of the Netherlands as far as soil type and land use are concerned. A complete field sampling 'round' is carried out continuously over a period of five years. The network contains 200 locations with 10 soil type/land use combinations (20 replications). Every year 40 locations are sampled. Most 1 2

Expertise centre for sustainable soil management National Institute of Public Health and the Environment

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sample locations are on farms that are between 5 and 100 ha in size and consist of combinations of soil type (sand, peat, clay) and land use (grassland, pasture, arable soil and horticulture). Also samples from organic farms and nature areas are collected. The database contains data on clay, lutum, grain size, pH, organic carbon content, heavy metal content, and different types of agriculture and nature. 10.6.1 Development of a biological indicator for soil quality in the Netherlands Since 1997, RIVM, Alterra, and Wageningen University have been working on the development of a biological indicator for soil quality (BISQ). This indicator system aims at providing an integral view of the ecological state of the soil relative to a desired or optimal quality with respect to a series of specific life support functions (LSF) of soil. Schouten et al. (1997) surveyed the important functions and processes in the soil and the organisms involved, and selected as a starting point (i) fragmentation and degradation of organic material, (ii) recycling of nutrients, (iii) soil structure evolution (bioturbation and aggregate formation), (iv) availability of nutrients for plants, and (v) stability of ecosystems (food web). Indicative variables are potential rates of various processes, food web structure, and diversity within an abundance of functional groups of organisms, resulting in 12 distinct variables. The data obtained can be used in trend analyses, such as positive or negative trends or desirable and undesirable soil qualities. However, so far there are no ecological soil quality criteria. The biotic data derived are combined with the database of BISQ on abundance and diversity of soil organisms (nematodes, enchytreaids, earthworms, arthropods, bacteria) and on activities of soil processes (C- and Nmineralization rates, activities and growth rate of bacteria). In addition, this database is filled with data from literature and other monitoring programmes (Breure et al., 2002). 10.6.2 Requirements for ecological monitoring

Ecological monitoring requires an infrastructure that allows biotic and a-biotic parameters to be measured simultaneously. Decisions have to be made about the temporal and spatial frequency of the samples being taken. Ideally, the setting up of an indicator system and a monitoring system should be done simultaneously. Data must be obtained at reasonable cost and at the desired levels of accuracy and frequency. Van de Brink (Chapter 8) gives a detailed overview of the complexity of ecological monitoring of indicator animals.

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10.7 How to derive ecological quality criteria for the soil? The political question to the developers of the indicator system was related to a possible quantification of sustainable land use, the resilience and functionality of the soil, changeability of land use and the effectiveness of policy and management. What do these questions mean for the derivation of soil quality criteria? 20.7.2 Ecological quality and sustainability

As far as sustainability is concerned we have to define what kind of sustainability we mean. Hueting and Reijnders (1998) define sustainability as the use of the vital functions of biophysical surroundings in such a way that they remain infinitely available. Sustainability of ecosystems must always be described in interaction with economy and society (Lorenz, 1999) as illustrated in Figure 2. Costanza (2000) describes the ecological economic model that distinguishes four types of capital: - natural capital: including ecological systems, mineral deposits, and other aspects of the natural world; - human capital: the physical labour of humans and the know-how stored in their brains; manufactured capital: machines and infrastructure of the human economy; social (cultural) capital: web of interpersonal connections, rules and norms that allow human interactions to occur. In this model sustainable development is specified by a minimum condition. The sum of the four capitals should be non-declining at all times. Assumptions

Figure 2. The view that sustainability of an ecosystem must be in interaction with economy and society

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regarding substitutability among the different types of capitals lead to a spectrum of possible sustainable states. At one end of this spectrum lies weak sustainability, where there are no constraints to the substitutability of capital, in particular for natural capital. At the other end of the spectrum lies strong sustainability where substitutability of natural capital is either not possible or at least severely constrained. The anthropocentric perspective focuses on the maintenance of life support services, essential for human survival, (biochemical cycling, space) in order to keep clean soil, clean air and uncontaminated food. These assets are called critical natural capital and cannot be substituted. Those supportive to the strong sustainability perspective advocate the precautionary principle with regard to substitution. If strong sustainability is adopted politically the weight of, for example, biodiversity loss is different from the situation in which weak sustainability is adopted. In the latter case only the preservation of ecological function is important. 10.7.2 Ecological quality and resilience

Resilience of soil indicates the extent of recovery after disturbances caused by stress. From many sides indications emerge that show that species rich ecosystems are more stable that species poor systems (Lawton and Brown, 1994; Van Straalen, 2002). Therefore, a possible proxy for resilience of the soil is the diversity within functional groups of organisms (organisms with the same ecological function). The sensitivity of organisms for different stresses is not equal for all organisms. That means that the loss of certain species within a diverse functional group might result in the dominance of other species - a changed biodiversity - but the function will continue It can be deduced that a decrease in diversity within a functional group will also lead to a decreased stability of the function. 10.7.3 Ecological quality and functionality of the soil

Functionality of the soil is particularly important with respect to ecosystem functions. Functionality of the soil is a requirement for its optimal use. It can be determined by measuring soil processes, or by modelling such processes, based on monitoring data from relevant organisms and abiotic conditions. 10.7.4 Ecological quality and the changeability of the soil

Changeability of land use is important, especially when land is scarce. This property is hard to quantify from an ecological point of view. Based on insights into the relationship between the abiotic and biotic characteristics of the soil and soil use, a proxy may be found in modelling activities.

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10.7.5 Ecological quality and effectiveness of policy and management

Effectiveness of policy and management requires insights into the effect of abiotic conditions and management measures on the presence and performance of organisms and ecosystems. Such relationships can be derived statistically if there is a sufficiently large database with monitoring data. The first results of a prognostic modelling of soil organisms and ecosystem functions based on monitoring data recently became available (Mulder et al., 2003; Schouten et al., 2004). It enables us to predict the result of some management actions on the changes in soil biodiversity for a limited number of combinations of soil type and soil use. 10.7.6 Dilemmas in deriving ecological quality criteria

Undisturbed ecosystems are rare in the Netherlands. As a consequence, there is implicit acceptance of the fact that man is influencing soil ecosystems. The criteria to be met to characterise soil quality as good or bad is already partly subjective, and social choices have to be made to determine the type of ecosystem wanted. Policymakers have to perform this process of choice. The outcome should be based on a combination of ecological insights and social motives. The sustainable use of a soil implies a minimal shift of negative effects to the future and the surroundings. Here, it is important to consider all soil characteristics in an integral way, taking into account technical, social and economic factors. When deriving references for ecological soil quality it might only be possible to state in general terms whether a certain quality is higher or lower that a certain fixed or chosen count down point. Optimal quality cannot be given. It may result in qualitative soil quality criteria, based on local conditions and managerial requirements. The approach to derive policy goals for soil ecosystems might have analogies with activities that have been performed in ecology, nature and water quality management. 10.8 Experience with the derivation of quality criteria for nature A reference book with 130 target nature types (Bal et al., 1995) is used in the Netherlands for nature policy goals. As far as vegetation composition is concerned, the approach rests on plant sociology, a description of ecosystembased knowledge of plant species occurring together (Schaminee et al., 1995). The fauna and the abiotic characteristics that belong in these ecosystems are also described. Effects of environmental pressure can be assessed using different models, such as MOVE (De Heer et al., 2000) in the case of vegetation, or DEMNAT (Witte et al., 1992) for changes in water management. These

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correlative models use statistical habitat - response relationships - and are based on extensive databases of monitoring data. 10.8.1 Derivation of quality criteria in aquatic ecosystems and its consequences for soil

Internationally there are different systems to assess the ecological quality of aquatic ecosystems such as RIVPACS. In the Netherlands, the EKOO system is widely used by regional managers. They use the system to determine trends and to examine the effectivity of management. They are extremely interested in assessment and forecasting the effect of management measures. This demands tailor-made developments. A number of characteristics may be distinguished in ecological aquatic assessment systems that follow from the demands and the desires of the users. The most important concern the diagnostic and/or prognostic description of the system, the scale of the system, the species community in the system, the kind of system such as marine or fresh water, the literature knowledge about specific monitoring, and the controlling variables. Assessment systems must have different characteristics for different users. International and national governments are mainly interested in a description of state and trends, if possible with references. Regional and local governments would like to be able to assess the effectivity of management measures as well. All assessment systems for aquatic ecosystems are based on extensive monitoring programmes. The reference descriptions are partly deduced from literature data (Bal et al., 1995), an ecological databank, expert judgement and the use of prognostic models. At a later stage, specific research may be performed to determine the control variables still lacking. In the Netherlands about 100 reference descriptions are in use for aquatic ecosystems (STOWA et al., 2001). A similar assessment system to that developed for aquatic ecosystems should be built for the soil, although there will be important differences between assessment systems for soil and assessment systems for water and nature. The quality of nature is based mainly on the intrinsic value of nature. In aquatic ecosystems, intrinsic value also plays an important role, although the function of such systems is also taken into account, for example safety, drinking water supply, or recreation. For the soil the most important protection base is its utility function. What is the soil used for? And what must the soil do now and in the future? Efforts to develop an ecological characterization system for the soil should also allow for different users. A local terrain manager, for example, does not need the same type of information as a province. The following organizations can be identified as users of indicator values for

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terrestrial ecosystems. They can be distinguished by scale and aim. Ministry of VROM3: integral information on the sustainability of land use - Ministry of LNV4: agrobiodiversity - EU: management of ecosystem sustainability - OECD: production support biodiversity FAO: biodiversity/soil fertility - UN: use of ecosystems/fair sharing Local managers: local information, influence of management There are good ecological and social reasons for arguing that a sustainably managed soil can perform the desired utility functions with a minimum input of chemicals and energy now and in the future. Such a soil may be classified as good or healthy. Different approaches are possible to derive criteria for good quality soil. For agrarian ecosystems good might approximate the situation that existed pre 1950, analogous to the nature policy in the Netherlands. Another possibility is to qualify the soil quality on biological farms as good or desirable, or take the situation in a nature area with the same type of soil as the reference point. Other options are theoretical references, deduced from a desired functioning of the soil in relation to carbon and nitrogen mineralization, the stability of the food web, and/or the diversity of soil organisms within functional groups as a measure for the resilience of the soil ecosystem. 10.9 Dutch approaches to derive integrated ecological soil quality The various approaches, such as SSD, TRIAD, and BISQ, will be described briefly, and how they are used to derive soil quality criteria. 10.9.1 Species sensitivity distribution: SSD

In the Netherlands data on human health risks and ecotoxicological risks are determined and compared statistically and the resulting, most stringent value is used. Ecotoxicological risks are based on Species Sensitivity Distributions (SSDs) if a sufficient amount of toxicity data exists (Posthuma et al, 2002). The derivation of standards is based on the proposition that if a maximum of 5% of the species is exposed to a concentration higher than the No Observed Effect Concentration (NOEC), it means that the organisms in the ecosystem are sufficiently protected.

3

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10.9.2 Location Specific Risk Assessment: TRIAD

SSD without spatial site-specific differentiation gave rise to considerable managerial problems. Therefore, an addendum to the Soil Protection Act was enacted (VROM, 2002a,b,c). It states: "If the costs of the remediation, compared to its effects, do not justify that remediation is being performed in a way that all functional properties of the soil for man, plant or animal are maintained or recovered, it is acceptable to take measures to isolate and manage contamination, and to control the effects of the isolation and management measures. If the pollutant does not disperse or threaten to disperse, the soil can be used as planned." A methodology to assess whether this adaptation of the Soil Protection Act can be used for sound ecological management of sites where complete restoration is possible is being developed. The approach in the Netherlands for site-specific ecological risk assessment of soil contamination is based on the estimation of effects, such as HC50 values, from the presence of contaminants in soil. To assess the seriousness of soil pollution at a specific location, a decision support system is being developed based on the TRIAD approach. Chapman (1986) originally described this approach for the assessment of sediment quality. The TRIAD is composed of three elements. An assessment of risks from the presence of contaminants in the soil and in biota (substances directed approach), an assessment of risks from the results of bioassays with samples from the site, and biological field observations. (Rutgers et al., 2000). In this system a local reference is chosen to derive the target quality. The need for sanitation is deduced from the intended soil use and the observed quality of the soil in relation to a local target quality. In pollution gradients it is often possible to use a local, clean reference. In many cases such a reference is not available and other references have to be applied.

Figure 3. TRIAD is composed of chemistry and toxicity data of the contaminant and ecology data from the contaminated site

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10.9.3 Provisional quality criteria used in BISQ

A simple clustering of observations at similar locations is the first step leading to descriptions of soil ecosystems and is the way to derive the Biological Indicator for Soil Quality (BISQ). This leads to an overview of the ecological capital of the soil. However, it is not possible to give a quality assessment and it does not provide policy or management targets. Another approach is the building of a national database that can be used to derive correlations between the presence of different species, the soil type and the soil use. This makes it possible to derive references when a desired functioning of the soil is described. This approach also makes it possible to differentiate reference descriptions according to the soil type, and eventually the soil use. In this way an insight can also be gained into the dynamics of soil processes and the resilience of the soil ecosystem. Originally, the aim was to quantify ecological quality by assessing the stability of the soil food web, based on extensive monitoring. However, it was shown that monitoring efforts for such assessments were too heavy. Among other things, it meant that every location had to be sampled four times a year (Schouten et al., 2001). Therefore, it was decided to deduce stability from the diversity and abundance of species within a functional group, reasoning that, in general, there is a high redundancy of organisms, resulting in a high stability. However, when redundancy decreases because of environmental stress, only a few resistant species will remain. In such situations the ecological basis for processes may become very narrow. When the resistant species also disappear, or are inhibited, as a result of future and yet unknown human activities, a process stops and the function is permanently affected. Therefore, the indicator system has been based on life support functions of the soil by hypothesising "The threat to vital soil processes can be expressed by comparing the number of species in functional groups of a certain area with its reference." Due to the redundancy of species, a process is assumed to continue to exist with fewer species, when species disappear, in which case the risk of instability and uncontrolled fluctuations will increase. At present data for the development of the indicator has been collected mainly from the agricultural areas, from organic farms and nature areas. To provisionally assess the quality of the soil ecosystem, the average indicator value of the organic farms has been scaled as 100% and the indicator values of the other locations have been scaled against the organic farms. Different quality standards for different types of farms and for different types of soil were then scaled. In this way a quality goal for grasslands on sand, grasslands on clay and so on were obtained. It is clear that for each combination of soil type and land

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use a quality goal can be derived in this way. However, as yet there is not enough data for this to be accomplished. 10.9.4 Derivation of "general" quality criteria

The indicator value results from stress factors, soil type, soil use and vegetation. Consequently references have to be deduced from a high number of observations. The choice of a desired reference is a political issue. For a specific combination of soil use and soil type the reference might be the average of the 20 corresponding locations in the Dutch Soil Quality Network, or the average of 20 biological farms. Soils with a very high or a very low indicator value should be investigated further. Monitoring changes in time might diminish the importance of subjective reference values. Spatial vast- and long-term monitoring activities may not seem ideal, but it is probably the most realistic way of obtaining objective information about differences between ecosystems, temporary variation within ecosystems and the human influence on ecosystems. Based on the insight that the ecological quality of the soil is dependant on a combination of soil type and soil use, a system has been proposed to derive differentiated quality criteria based on monitoring data and ecological and statistical models (Breure et al., 2003). Firstly, different ecosystem services have been identified that have to be provided by the soil and subsequently ecological parameters important for that function are developed. This combination is given in Table 1. The different ecosystem services in different types of soil use have also been identified (see Table 2). From Table 2 it is clear, that not all ecosystem services have to be delivered in all types of land use. This makes it possible to discriminate between quality criteria for different types of land use. Given the way ideas and concepts are developing, there is a clear need for a "Handbook of Soil Reference Types". A quality system, with the degree of detail found in systems for aquatic ecosystems and nature seems a Utopia at the moment. However, it can be expected that the first rough version of ecosystem descriptions will be available within a few years. The first impulses have been given. 10.9.5 Proposed general outline of the indicator

To quantify ecological quality it might be useful to define targets and develop a distance to target indicator. In order to value the indicator outcome as good or bad, it is necessary to choose targets and to define a distance to target indicator. As outlined earlier, these targets may be defined on the basis of historical,

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Table 1. Ecological proxies for ecosystem functions Ecosystem service

Important ecological parameters

Supply of nutrients Food web including earthworms Primary production Ratio bacteria/fungi (De)nitrification Water regulation Earthworms Abundance and ratio bacteria/fungi pH, content of soil organic matter, groundwater level Structure Earthworms Abundance and ratio bacteria/fungi pH, content of soil organic matter Nematode Channel Ratio Supply of clean Specific activity bacteria and fungi shallow Clean soil (concentration of pollutants lower than a maximum concentration) groundwater Extent of leaching of nitrogen, phosphate, and halogenated pollutants (EOX) Activity of the nitrogen cycle Supply of clean Amount and biodiversity of bacteria and fungi deep groundwater Clean soil Extent of washout of nitrogen and phosphate Pest control in Plant Parasitic Index of nematodes agriculture Abundance and ratio bacteria/fungi Mycorrhiza fungi Changeability of Diversity of soil organisms soil use Concentrations of nitrogen and phosphate in the soil Resilience and Diversity (within functional groups) resistance

Table 2. Ecosystem services with different types of land use Siapply of

mutrients

WaterStructure regulation

Supply of clean shallow groundwater

Supply of clean deep groundwater

Pest Changeregulation ability in in agrisoil use culture

Resistance and resilience

Grassland Arable land Nature Urban area

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

X

Rural area Paved soil

X

X

X

X

X

X

X

X X

X

X X

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geographical, desired or theoretical references. Such references may be positive, based on the soil quality in nature areas or on organic farms, or negative, based on soil quality of heavily polluted sites. The indicator may have an outline as illustrated in Figure 4. A target situation is a subjective, political decision. It will be based on soil use and soil type. This soil type may be deduced from a weighed integration of data relating to abiotic conditions such as soil type, diversity of specific organisms, abundance of specific organisms, process rates, and soil use. These quality criteria may be general, as is the case with intervention values for severe soil pollution, or local, as used in the TRIAD approach for location-specific risk assessment. 10.9.6 Bottlenecks in monitoring terrestrial ecosystems

All ecological assessment systems are based on extensive monitoring activities and databases. The set up of terrestrial systems differs widely from aquatic systems or nature. Nature areas are owned by the state, nature organisations, landscape conservators as well as large-scale owners (such as state forestry services, provincial and regional nature trusts). The government and district water boards generally own water. It is relatively simple to make management and monitoring agreements with such owners, and this has been done successfully over many decades. Soil, however, has a large variety of owners with different interests, especially outside nature areas. To be able to agree on a systematic monitoring system on a scale comparable with the scales of water and nature monitoring a management structure must be developed. Farmers own a large proportion of the land. They are used to soil being sampled to determine of soil quality for fertilization and cultivation purposes and to determine the presence of pathogens. So they are likely to agree to other monitoring. Like

A = negative reference B = actual situation C = target situation D = positive reference

Figure 4. The outline of an ecological quality indicator

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monitoring activities in aquatic systems and nature areas, data must be collected in a national database. Clear arrangements are necessary, like in the nature and water world, to assess the ecological quality of soil in detail. Here there is a noticeable lack of any co-ordinating framework or awareness of common interest amongst the different types of landowners as far as the maintenance of good soil and groundwater quality are concerned. 10.10 Proposed monitoring activities based on earlier considerations Data collection ought to result in robust parameters based on solid methods. Chemical and physical parameters of soil are more constant in time than biological measures. Some populations may fluctuate strongly according to season. This has to be taken into account in monitoring. It must be very clear when and how monitoring should take place, and also the way samples are handled should be standardized. The proposal for guiding (proxies) monitoring activities is shown in Table 1, and some ecological aspects are mentioned here. Soil biodiversity parameters form a special aspect, because of changes during the course of the year. Three different aspects of soil biodiversity have to be taken into account activity, diversity and abundance. In the ecological capital of the soil the relative amount of microbial biomass such as bacteria and fungi give a good indication of the activity and stability of the ecosystem. When monitoring activity takes place on a large scale, it is possible to get these measurements well standardized and automate the process using standardized microscope and computer equipment. Genetic and physiological characterization of microbial communities provides an opportunity for distinguishing between different communities, although it does not give direct information on trends in biodiversity. A diversity of organisms in the soil might be better monitored using higher organisms such as nematodes and earthworms that are extremely indicative of soil status. Nematodes are present in high numbers, have high species diversity and relative and absolute amount provides good information on the diversity and stability of the ecosystem. The organisms are easy to handle and there are many companies able to carry out routine taxonomic work. At the moment there are several different groups working on developing genetic tools to identify nematode species. This will provide information relevant to the genetic diversity mentioned in the biodiversity treaty, and help to simplify taxonomic work. Earthworms may be present in large numbers, and their diversity can be easily characterized. They are also easy to count. They are very appealing to

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farmers, who know that earthworms are important for the structure of the soil. Farmers also have the (essentially good) idea that the more earthworms the better, so that they can see with their own eyes and feel with their own hands whether management practices are having a positive or negative impact on the soil. Nature conservationists see earthworms as the main food source for many above ground animals such as birds and badgers. To ecotoxicologists earthworms are useful organisms because they provide evidence of whether or not pollutants are present in a form that soil organisms experience as negative. To the suppliers of organic matter earthworms play important roles as far as the degradation and transport of organic matter in the soil is concerned. The interpretation of the monitoring data on microbial biomass, nematodes and earthworms will be sufficient to derive information on the status and trends in the ecological quality of the soil. Also information about stability in the form of diversity within functional groups can be derived from such data. In addition, all kinds of indices that may indicate ecological soil quality, such as maturity index, plant parasite index, nematode-channel index, and the ratio of epigeic to endogeic earthworms can be compiled. 10.11 Quality criteria based on the above considerations It demands knowledge and experience to deduce quantitative quality criteria for ecological soil quality. The number and diversity of organisms differs from one type of soil to another and is dependant on the use being made of the soil. Schouten et al. (2004) provide an example of the number of genera of nematodes present in different types of soil (Figure 5). Similar data can be found for other organisms, for example, micro-organisms (Bloem and Breure, 2003) and enchytreaids and earthworms (Didden, 2003). Also the intensity of management influences biodiversity, as was clearly shown by Mulder et al. (2003). These observations lead to the conclusion that a local reference or target is the most useful way to derive quality criteria. In this approach, the targets are derived from a local reference site that is subsequently compared with the location where quality criteria are required. There are ways to predict the diversity and abundance of soil organisms in relation to management measures and abiotic conditions (Mulder et al., 2003). This approach can be used to derive a target and quality criteria from a local reference location and to provide opportunities for deducing management measures from the target itself. The approach used consisted of generalised linear modelling of the biotic and abiotic monitoring data. This approach requires a considerable amount of monitoring data. A new challenge is the

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Number of genera per location type

Average number of genera per location

(20 sites per land use type)

(20 sites per land use type)

Figure 5. Biodiversity ofnematodes in different combinations of land use and soil type (modified after Schouten et ah, 2004)

development of criteria to determine the stability of the food web, based on a limited amount of monitoring data from a site. Earlier calculations of the stability of soil food webs are based on very extensive monitoring activities (De Ruiter, 1993). Intelligent statistical techniques might make it possible to assess the stability of the soil food web based on a number of monitoring data that can be obtained from a regular monitoring programme. 10.11.1 International Developments

Within EU, OECD, FAO and several European countries there are various initiatives to develop an ecological characterization of the soil. The FAO has initiated the International Soil Biodiversity programme. The FAO workshop in Londrina (2002) agreed on a strategy for the implementation of two main objectives. Firstly, promoting awareness, knowledge and understanding of the key roles of functional groups and of the impacts of diverse management practices in different farming systems and agro-ecological and socio-economic contexts. Secondly, promoting ownership and adaptation by farmers of integrated soil biological management practices as an integral part of their agricultural and sustainable livelihood strategies. This soil biodiversity initiative fits within FAO's activities on agricultural biodiversity, which focus on the variety and variability of animals, plants, and microorganisms important for food and agriculture. The programme has a spatial, temporal and scale dimensions. It covers the wide variety of genetic resources

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and species used directly or indirectly for food and agriculture and in the production of fodder, fibre, fuel and pharmaceuticals. It also includes the diversity of species that support production, wider agro-ecosystems and the diversity of the agro-ecosystems themselves (FAO, 2002). The OECD is working in close co-ordination with FAO on a set of indicators on agri (not agro)-biodiversity. In defining the framework for such indicators they describe the agri-ecosystem base as consisting of production species (crops and livestock) and production support species (soil biodiversity) (OECD, 2003). In 2003 OECD organised a workshop in Rome, specifically on the development of soil biodiversity indicators. Fifty papers were presented on the development of soil biodiversity indicators in OECD countries. Monitoring programmes exist in Germany, Italy, France, New Zealand and initiatives are being taken in several other countries. At present the EU is working on the development of a Europe-wide soil policy that focuses on the sustainable use of the soil. Recently, a European soil initiative was launched with the intention of contributing to the creation of a soil-monitoring directive for the whole EU. One of the important items in this directive was soil biodiversity. This will have to be taken up in the monitoring directive, and its protection will be incorporated into soil policy because of its intrinsic value and role in the functioning of the soil. In 2005 the European Commission must present a thematic strategy for soil protection to strengthen current policies and define problems as well as qualitative and quantitative objectives. Also it should not only indicate the means by which these objectives can be achieved, but also provide a timetable and general principles for evaluation and monitoring. It will focus on: " protecting soil in its role in storing CO2, securing water resources and preserving biodiversity; protecting soil for the sustainable production of food and renewable resources". Soil protection is seen as a precondition for achieving, among other things, the objectives of the Water Framework Directive in (as far as preventing diffusion of pollution is concerned) the Habitats Directive in relation to soil biodiversity, and the Kyoto Protocol in terms of the capacity of the soil and subsoil to retain CO2. It should also summarise data on complex environmental issues to indicate the overall status and trends of soil biodiversity. 10.12 Implementation in soil management The ultimate challenge of this chapter is practical implementation and here simplicity and ecological relevance are major problems. Too much attention to simplicity will negate ecological relevance and lead to a weakening of

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ecological perspective and integrity. Because of the lack of knowledge and experience, it is impossible as yet to provide a simple validated and calibrated set. Therefore, once a set has been recommended it must be improved as soon as more experience is gained. Monitoring should follow the TRIAD approach, which uses multiple lines of evidence to deduce the ecological quality of the soil, and the MOVE approach, which is the statistical extrapolation of monitoring data. Minimum or generic set In order to present soil biology status two aspects of the TRIAD approach can be used: biological field observation and availability of contaminants. The field characterizations are plants and biota, such as the nematode community, microbial biomass and activities, and the ratio of fungi to bacteria. If contamination is expected, then the available fraction of these contaminants should be measured and monitored. For natural recovery or development, the vegetation approach might be based on the models MOVE and DEMNAT. Characterization of the soil type (sand, clay, peat an their organic matter and nitrogen content), water supply (three regimes) and water quality (nutrient status, salt status and pH) as well as the presence of seedbanks is necessary. Those measurements can all be performed by research institutes such as Alterra and some departments in Wageningen and other universities, or in laboratories equipped to deal with routine soil matters. This general set can be applied in city green areas or agricultural areas developing towards a more extensive land use, and in nature areas. Ecological assessments by experts, should be supported by statistics. Therefore, every location should be "examined" in fivefold. When the location is large it should be divided into sub-locations of 5x5 kms. Tailor-made set On top of the minimum set, aspects of functioning and stability should be added to include determination of netto primary production and stability based on various food webs. Such measurements can be carried out by the research departments of universities that are equipped for this task. There is no routine database. These approaches should be used for areas that are large in volume and which have been analysed in relation to ecological risk over a period of years.

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References Bal, D., H.M. Beije, Y.R. Hoogeveen, S.R.J. Jansen and P.J. van der Reest, 1995. Handboek Natuurdoeltypen in Nederland. IKC Natuurbeheer, Ministerie van LNV, Wageningen. Bloem, J. and A.M. Breure, 2003. Microbial indicators. In: Markert, B.A., A.M. Breure and H.G. Zechmeister (Eds). Bioindicators & Biomonitors. Principles, Concepts and Applications, pp 259-282, Elsevier, Oxford. Breure, A.M., AJ. Schouten and M. Rutgers, 2002. Het bodemleven als indicator voor duurzaam bodemgebruik. Bodem 9, 149-151. Breure, A.M., M. Rutgers, J. Bloem, L. Brussaard, W. Didden, G. Jagers op Akkerhuis, Ch. Mulder, AJ. Schouten and HJ. van Wijnen, 2003. Ecologische kwaliteit van de bodem. RIVM rapport 607604005. Chapman, P.M., 1986. Sediment quality criteria from the sediment quality triad: An example. Environ. Toxicol. Chem. 5, 957-964 Costanza, R., 2000. Social goals and the valuation of ecosystem services. Ecosystems 3, 4-10. De Heer, M., R. Alkemade, M. Bakkenes, M. van Esbroek, A. van Hinsberg and D. de Zwart, 2000. MOVE, nationaal model voor de vegetatie, versie 3. De kans op voorkomen van ca. 900 plantensoorten als functie van 7 omgevingsvariabelen. RIVM Rapport 408657002. De Ruiter, P.C., Van Veen, J.A., Moore, J.C, Brussaard, L., Hunt, H.W. 1993. Simulation of nitrogen mineralization in soil food webs. Plant Soil 157, 263-273 Didden, W., 2003. Oligochaeta. In: Markert, B.A., A.M. Breure and H.G. Zechmeister (Eds). Bioindicators & Biomonitors. Principles, Concepts and Applications, pp 555-576, Elsevier, Oxford. Doran, J.W. and T.B. Parkin, 1994. Defining and assessing soil quality. In: Doran, J.W., D.C. Coleman, D.F. Bezdicek and B.A. Stewart (eds). Defining Soil Quality for a Sustainable Environment, 35. American Society of Agronomy Special Publication, Madison, WI, pp. 3-21. FAO (United Nations Food and Agriculture Organization), 2002. Report of the international technical workshop on biological management of soil ecosystems for sustainable agriculture. Organized by EMPRABA-Soya and FAO, Londrina, Brazil, 24-27 June 2002 Hanski, I., 1997. Be diverse, be predictable. Nature 390, 440-441. Hueting, R. and L. Reijnders, 1998. Sustainability as an objective concept. Ecol. Econ. 27,139-147. Lancaster, J., 2000. The ridiculous notion of assessing ecological health and identifying the useful concepts underneath. Human Ecol. Risk Assess. 6, 213-222. Lawton, J.H. and V.K. Brown, 1994. Redundancy in ecosystems. In: Schulze E.D. and H.A. Mooney (eds). Biodiversity and ecosystem function. Springer, Berlin. Lorenz, CM., 1999. Indicators for sustainable river management. PhD Thesis Vrije Universiteit Amsterdam. Mulder, Ch., D. de Zwart, H.J. van Wijnen, A.J. Schouten and A.M. Breure, 2003. Observational and simulated evidence of ecological shifts within the soil nematode community of agro-ecosystems under conventional and organic farming. Funct. Ecol. 17, 516-525. OECD, 2003. Agriculture and Biodiversity. Developing Indicators for Policy Analysis. Proceedings from an OECD expert meeting Zurich, Switzerland, 2001. Polis, G.A., 1998. Stability is woven in complex webs. Nature 395, 744-745. Posthuma, L., G.W. Suter and T.P. Traas, 2002. Species sensitivity distributions in ecotoxicology. Lewis Publishers, Boca Raton, 587 pp.

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Roots, E.F., 1996. Environmental Information - autobahn or maze. In: Schroder, W., O. Franzle, H. Keune and P. Mandry, (eds). Global monitoring of terrestrial ecosystems. Ernst & Sohn, Berlin, pp. 3-31. Rutgers, M., J.H. Faber, J.F. Postma and H. Eijsackers, 2000. Site-specific ecological risks: A basic approach to the function-specific assessment of soil pollution. Reports of the Programme on Integrated Soil Research. Vol 28, Programme Bureau Soil Research, Wageningen. Schaminee, J.H.J., A.H.F. Stortelder and V. Westhoff, 1995. De vegetatie van Nederland: deel 1: inleiding tot de plantensociologie - grondslagen, methoden en toepassingen. Uppsala, Leiden: Opulus press, 296 p. Schouten, A.J., Brussaard, L., De Ruiter, P.C., Siepel, H., Van Straalen, N.M. (1997) Een indicatorsysteem voor de life support functies van de bodem in relatie tot biodiversiteit. RIVM report 712910005, 90 pp. Schouten, A.J., J. Bloem, A.M. Breure, W.A.M. Didden, M. van Esbroek, P.C. de Ruiter, M. Rutgers, H. Siepel and H. Velvis, 2001. Pilot project Bodembiologische Indicator voor Life Support Functies van de bodem. RIVM rapport 607604001. Schouten, T., A.M. Breure, Ch. Mulder and M. Rutgers, 2004. Nematode diversity in Dutch soils. From Rio to a biological indicator for soil quality. Nematology Monographs & Perpectives, in press. Stenberg, B., 1999. Monitoring soil quality of arable land: microbiological indicators. Acta Agric. Scand., Sect. B, Soil Plant Sci. 49,1-24. STOW A, RIZA, RIVM, IPO, 2001. REBEWA Een raamwerk voor Ecologische Beoordeling van Watersystemen, Onderzoeksnota. Van Straalen, N.M., 2002. Assessment of soil contamination - a functional perspective. Biodegradation 13, 41-52. VROM (Netherlands Ministry of Spatial Planning, Housing, and the Environment) (2002a): Besluit Locatiespecifieke Omstandigheden. Staatsblad van het Koninkrijk der Nederlanden 23-4-2002, nr. 192. VROM (2002b): Regeling Locatiespeciefieke Omstandigheden. Staatscourant 10-10-2002, nr. 195. VROM (2002c): Rectificatie Regeling Locatiespeciefieke Omstandigheden. Staatscourant 4-12-2002, nr. 234 WCED (World Commission on Environment and Development), 1987. Our common future. Oxford University Press, Oxford, UK. Witte, J.P.M., Groen, C.L.G. and J.G. Niehuis, 1992. Het ecohydrologisch voorspellingsmodel DEMNAT-2; conceptuele modelbeschrijving. Onderzoek effecten grondwaterwinning 1. RIVMrapport 714305007, ISBN 90-6960-30-7.

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Chapter 11

SYNTHESIS FOR SOIL MANAGEMENT P. Doelman

Abstract

This synthesis aims to bridge the gap between scientific knowledge on soil and soil policy. Current scientific understanding of the way in which soil ecosystems function is discussed in relation to practical strategies that can contribute to the preservation of soil vitality. Soil is a vital component in the functioning of ecosystems and a vital resource for all types of land use. Soil vitality has been defined in many ways and such definitions contain elements of subjectivity. Soil is a function of parent material, climate, time, vegetation, relief, soil fauna and soil microflora. In a vital soil, there is a continuous stream of energy and nutrients between the formation of organic matter (primary production) and its transformation (secondary production). Each author in this volume shares a sense of astonishment at the abundance and functional equivalents within soil microflora, soil fauna and soil vegetation. Spatial heterogeneity has been identified as the key to rich biodiversity and the stable functioning of the total system. Taking into account the definitions provided for ecology and soil ecosystems, monitoring of soil vitality or soil ecosystem functioning should be based on biological, chemical and physical soil properties. Physical aspects are compaction, soil type and habitat structure. Chemical aspects are carbon-, nitrogen- and phosphorous content. Biological aspects are species diversity, functional community diversity, and functional food web diversity (stability). In practice as well as in science, the integration of these three is rare. Soil physical features have seldomly been determined in relation to ecological risk. When a soil adapts to changes it shows its vitality. Adaptation to gradual changes results in the same finetuning. Sudden changes lead to adverse impacts, resulting in another functioning, and eventually another kind of stability. The levels of stability reached after disturbance or due to disturbance, may vary. A large number of functional equivalents with different sensitivities towards disturbance is a good guarantee for the maintenance of vitality. For monitoring soil functioning generic as well as tailor-made monitoring sets have been suggested. Their specific application depends on the hypothesis of the development of the area mentioned, the ambition. Whether we want "functioning" or whether we want "higher species diversity or higher animals" is a question that should be answered by management. Then a relevant monitoring set can be designed.

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11.1 Introduction This synthesis is an attempt to create a bridge between the complexity of scientific knowledge on all aspects and consequences of soil ecosystem functioning and the desire for simplicity in soil policy and management. It tries to address the issue of how to apply current scientific understanding of soil ecosystem functioning in practise. It takes into account the opinions of terrestrial ecologists, the warnings of environmentalists, and the views of aquatic ecologists who face similar issues. It combines and evaluates various overviews to develop practical monitoring that preserves soil vitality. By definition synthesis is (i) the action of putting together, (ii) the formation of a whole out of parts or elements, and (iii) the deductive reasoning by combining principle theses and their anti-theses towards a higher stage of truth (Gove et al., 1976). These three aspects will enable an objective and straightforward monitoring approach as well as a subjective perception of vitality of soil. "Vital Soil" relates natural soil functions to human-oriented utility functions and tries to answer questions about the consequence this may have for politics, practice and people. In a mechanistic way this is depicted in Figure 1. In this figure the scientific information on the functioning of vital soil flows to policy and practice and then practice and policy return questions to science. Soil is a function of parent material, climate, time, vegetation, relief, soil fauna and soil microflora. There is a continuous stream of energy and nutrients due to the ability of plants to combine solar energy and anorganic molecules to form organic matter. This energy is released as a result of plant activity when consumed by animals. Because of the variety of conditions, soil types, plants and animals, the diversity in terrestrial eco-systems is large but energy flow is based on the same principle outlined in the left wheel of Figure 1. Soils are subjected to and respond continuously to change and natural stress. For centuries, soils have been sinks for many types of contaminants and other human induced influences or disturbances. The question for soil scientists and soil managers is how vital is a soil ecosystem? For this to become the central question some other questions should be answered first. Besides sunlight, water and organic matter, what other characteristics determine the quality of the soil? Has the understanding of the many interactions that occur in the soil gone far enough to answer questions such as what are the minimum levels of diversity, nutrient turn-over and soil structure required in order to prevent soil systems becoming irreversibly damaged, or to secure their recovery? Legislation to preserve soil and the demanded controlling actions that can be implemented in practice is derived from policy. This policy must, in turn, be

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Figure 1. The functioning of soil ecosystems in relation to political and practical judgement. Primary production is plant growth; plants return to soil or are eaten by herbivores and omnivores; herbivores and omnivores return to soil or are eaten by carnivores; ultimately organic matter returns to soil as dead material and is recycled by the soil microflora and soil fauna to nutrient minerals, so plant growth can continue

rooted in actual knowledge. When developments occur in scientific knowledge, practice and policy must adapt to these new insights and they should be reflected in revised legislation. When new developments occur in policy, scientific questions may change. When new developments occur in practice, policy and science may be affected. The sequence science —* policy —»• practice is based on the assumption that science is ultimately objective and a guide for decision making in human society. In Section 2 of this chapter the history of human concern for the way in which the ecosystem functions and how, over the years, ecology and ecosystems have been defined, is examined briefly. Views on ecosystem quality from aquatic scientists may prove useful in terrestrial science. The evaluation of function, value and properties of the interaction between soil science disciplines is given in Section 3. The discussion and the integrated synthesis with general and tailor-made monitoring options are presented in Section 4 including their opportunities and limitations.

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11.2 Historical consciousness For many years, there was considerable concern about the functioning of wild life in aquatic and terrestrial ecosystems from an emotional as well as a scientific point of view. Ultimately this led to attempts to define ecology, ecosystems, ecotoxicology, and the health and quality of ecosystems. Historical and scientific lessons are instructive for this synthesis and are included here. 22.2.2 Environmental concern

There have been environmental concerns since industrialization began. The book Silent Spring (Carson, 1962) questioned the mismanagement of pesticides that lead to the accumulation of DDT in food chains: birds of prey became well known victims. The book The Limits to Growth (Meadows, 1972) focused on the possible negative impacts of population growth, over-consumption and urbanization on natural resources while in the book Extinction (Ehrlich and Ehrlich, 1981) questioned the relationship between biodiversity and the felling of forests and the ploughing and paving of the surface of the earth. In the book The End of Evolution (Ward, 1994) the dominant influence of mankind in the biosphere as a whole is questioned. These critical explorations and the issues raised have contributed to more awareness and have helped ensure that environmental legislation and standards to preserve the quality of air, water and soil are maintained. Many of those standards meant to guarantee quality have been disputed since. 22.2.2 Attempts to define ecosystems and ecosystem health and quality

The notion "environment" Linnaeus (1707-1778) founded the system for classifying living organisms and established the value of describing in an exact manner. Darwin (1809-1882) proposed the theory of evolution by natural selection. Tansley (1935) defined the word ecosystem by suggesting that organisms have to be considered, not separately, but in relation to their specific environment. Odum (1971) defines ecology as the relationship between plant, animal and environment. Ecotoxicology is concerned with the toxic effects of chemical and physical agents on living organisms, especially on populations and communities within defined ecosystems; it includes the transfer pathways of those agents and their interactions with the environment (Butler, 1978). In classical terrestrial ecological field studies, from the 1920s to the present day, the environment of species is generally described in terms of rainfall, temperature, daylight, biotope structure and soil type. Investigations into

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population dynamics or community dynamics were mainly carried out in the context of food and prey-predator relations with little attention to any description of soil habitat. Habitat is defined as a living house or living village (Verstraete and Mertens, Chapter 5) and niche as the relationship between environmental conditions and eco-physiological properties. Detailed descriptions and definitions of the living environment of those areas are scarce in literature. Since the 1960s environment has often only been related to contamination. The notion "health" and "quality" The introduction of the health and quality of ecosystems emerged in response to contamination of the environment by heavy metals and biocides. Contaminants introduced a new or extra environmental factor: environmental pollution. Gove et alv (1976), stated that health is usually related to the normal or proper condition of an organism in relation to disease, recovery or prevention while quality has a degree of intrinsic excellence, a charm applicable to inherent traits. O'Neil et al. (1986) stated that a healthy ecosystem is characterised by the integrity of nutrient cycles, energy flows as well as stability and resilience. There have been many attempts to define soil ecosystem health, soil wealth and soil vitality (Eijsackers, Chapter 1). In assessing the impacts of contamination the first tools were simple doseeffect measuring methods under defined conditions for single species, or easy to measure activities such as soil respiration and soil enzyme levels. Within the more holistic approach, phenomenon such as resistance, resilience, stress, elasticity, inertia, integrity and stability have been mentioned in various earlier chapters. Techniques or methods to quantify these properties are lacking. In the field of aquatic research these holistic properties are also mentioned as part of scientific research, but are not used in aquatic environmental management (Vigerstad and McCarthy, 2000). Practical water management deals with quantitative properties such as assimilation capacity (primary production) and nutrient cycling and phosphor loading and species diversity in relation to functioning. Furthermore it deals with properties such as aesthetic value, recreational quality, market value such as tourism, and susceptibility to events, relating the perception of quality to enjoyment and use. This corresponds with the functional approach of soil protection. Lancaster (2000) states that terms such as ecosystem quality or health are used by people or agencies concerned with the way anthropogenic factors impact on ecosystems, but who are unaware of underlying ecological concepts. She condemns such an approach because there are no objective definitions of

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health or methods for defining degrees of health. Vigerstad and McCarthy (2000) emphasize that observation of quality or health should be quantitative, indicating that a defined method or technique should be applied. Hence to prove presence or any alteration of any kind of soil life, data obtained by objective, hall-mark equipment are necessary. This also accounts for extraterrestrial life, as Kenneth Nealson, the conductor of the American experiments to look for microbial life on Mars stated: "define and restrict to quantifying features/arrangements, measurable in the in situ situation" (Nealson, 2002). The need to define precisely what terms should be applied is relevant to soil approach. Properties such as primary production (Ernst, Chapter 3), nutrient cycling (Verhoef, Chapter 4; Verstraete and Mertens, Chapter 5), functional diversity (Verhoef; Chapter 4; Verstraete and Mertens, Chapter 5; Van Straalen, Chapter 6), food webs (De Ruiter, Chapter 7) and trophic levels (Van den Brink, Chapter 8) have been studied qualitatively and quantitatively in all kind of soil ecosystems. 22.2.3 Experience in describing in soil health in science and practice

Since the early 1970s, scientific research has been carried out into the effects of soil contamination. Specified dose-effect research provided the bases (Eijsackers, Chapter 1; Breure, Chapter 10), and experiments were interpreted and assessed primarily according to statistical significance. The Dutch Health Council (1988) evaluated all literature data and recommended prognostic tests to anticipate possible effects, and diagnostic tests as to assess the impact of suspected soils (Table 1). For prognosis, sensitive and measurable parameters were recommended such as nitrification, earthworms and millipedes, all on a population level or on a defined process level. As far as diagnose was concerned, the Dutch Health Council recommended, in general, the community approach, indicating that it was important to look at ecosystem level, or at least at sub-ecosystem level. So recommendations to focus at system level have been in existence for more than 15 years. In principle impact assessment should be based on a control measurement. Defining and finding a control or reference ecosystem is virtually impossible according to Lancaster (2000). Kelly and Campbell (2000) agree but offer constructive suggestions: define why, what and how to measure, make choices, separate variability from uncertainty and ensure that observations are carried out at ecosystem level. The absence of data from reference biotopes was strongly felt in the world of soil management. Recommendations were made that reference data should be

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Table 1. The proposed ecological parameters to judge the severity of soil contamination in relation to its functioning (Dutch Health Council, 1988) Parameter

Prognostic

Microflora Respiration Nitrification Nitrogen fixation Ratio sensitive/resistant

CO2/O2 in time NO2/NO3 in time Nodules

Diagnostic

Isolate organisms and test

Soil fauna

Earthworms Nematodes Springtails Collembola Insect larvae Protozoa Ants

Growth/reproduction Population growth Population growth/reproduction Leaf consumption Larval growth Population growth Mortality

Biomass per species Feeding groups Composition community Composition community Appearance/morphology Composition community Behaviour

Germination and growth Photosynthesis

Germination and growth Photosynthesis Senescence Mycorrhiza

Mortality/reproduction/growth Mortality/reproduction/growth

Composition community Composition community

Fragmentation

Fragmentation

Vegetation

Seeds Plants Leaves Plantroots Vertebrates

Mice Birds Other parameters

Litterbag

collected to develop soil quality criteria (Breure, Chapter 10). Field observation and bio-assays using the TRIAD approach (Breure, Chapter 10) were derived from Table 1. The surprise, the entanglement, the enigma of an apparent or seemingly overwhelming number of species of soil fauna and soil microflora representatives with seemingly much the same niche, partly restrained soil biologists from going deeper into functional diversity or studying functional equivalences and partly stimulated them (Faber, 1992; Bongers and Bongers, 1998). The practical development aimed at extremely simple tests such as Microtox, where the lighting of bacteria corresponds with stress, since the testing should be reproducible and cheap. Out of many suggested and officially published parameter sets, two test sets that have not been mentioned so far will be referred to here because of their minimum size (Doran and Parkin, 1996) and practical size (Sparling, 2003). Doran and Parkin propose a minimal set

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consisting of physical and biological characteristics. Basically three features are mentioned: organic matter, soil biomass and biomass activity. With these three measurements they claim to cover the relevant part of the process of recycling of elements. Sparling suggests a set of seven characteristics: (i) total carbon, (ii) total nitrogen, (iii) anaerobically mineralized nitrogen, (iv) pH, (v) phosphate, (vi) bulk density, and (vii) macroporosity. Sparling explains his choice as follows: - total organic carbon; carbon retains moisture and nutrients; good for structure, roots and biota; - total nitrogen; nitrogen is essential for plants and animals; anaerobically mineralized nitrogen; a kind of stress for plants (N not available); acidity (pH); plants and animals, each have their own optimum pH range for growth; - phosphorus (Olsen); available phosphorus, adsorbed by clay and organic matter; - bulk density; how densely a soil is packed (percentage water and air; compacted soil has a poor root system (no mycorrhiza)); - macro porosity; large pores with a diameter greater than 60 [am; smaller pores reduce aeration: less N-fixation. There are many very extensive sets recommended (Verhoef, Chapter 4; Verstraete and Mertens, Chapter 5; Breure, Chapter 10). The uniqueness of Sparling's set is its restricted size, its payability and the environmental description of the soil in terms of bulk density and macro porosity. These two factors provide intrinsic protection (or limiting) conditions for fauna, since pore size can relate to the size of biota. Sparling's explanation is: "when soil conditions and soil functioning is fine, why bother and spend more money on extra indirect measurements such as biodiversity" (Sparling, 2003b). 11.3 Characterization of vitality of soil by synthesizing aspects The three key elements of soil vitality are (i) dynamics and mineralization of organic matter, (ii) soil structure formation and maintenance, and (iii), support and control of plant production and species diversity. The conceptual approaches of these elements were introduced by Eijsackers (Chapter 1) and broadly covered from various angles of soil functioning in Chapters 2 to 8, which can be divided into two coherent clusters. The chapters on the formation of soil (Van Breemen), plant growth and the formation of organic material (Ernst), and the recycling of organic matter (Verhoef) is the first cluster. The

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second cluster has its coherence in the soil biota. Soil microflora (Verstraete and Mertens) and soil fauna (Van Straalen) are exhibited as sub-groups, with intrinsic connections to food webs, issues of stability (De Ruiter) and as food chains (Van den Brink). The fate of contaminants (Peijnenburg) and the factual approach of monitoring (Breure) can be regarded as a third omnibus cluster. 11.3.1 Linking soil formation, vegetation and organic matter, and soil biota

The origin of the functions of the soil and its properties is determined by the soil itself, together with primary production and secondary production processes. The overviews by Van Breemen (Chapter 2) on soil formation; by Ernst (Chapter 3) on vegetation and organic matter, and by Verhoef (Chapter 4) on the activity and structure of soil biota are the supporting cornerstones of soil ecosystems (Figure 2). Alterations of the soil's parent material are slow and can be seen only after decades or millennia. The soil forming factors are parent material, topography, climate, biota and time. A set of physical, chemical and biological factors create a particular soil. Weathering is the chemical and mineralogical transformation to solutes and the solid residues, the mineral fraction. The physical ripening and structural formation of the soil is based mainly on the removal of water, but other factors including temperature also play a role. Chemical ripening is mainly based on changes in pH. Biological ripening of the crust is mainly determined by engineers or habitat builders such as micro-organisms, soil

Figure 2. General supporting cornerstones of the functioning of soil ecosystems

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fauna and plant roots. Van Breemen refers to them as ecological engineers. Ecological engineering roles have been described in detail by Ernst such as the role of root and mycorrhiza and root exudates resulting in a root layer, with a very high level of biological activity. The top soil is a living system of surprising diversity. The cradle of this diversity is the enormous spatial and temporal heterogeneity of environmental conditions, as Verhoef emphasizes. The coherence between the three chapters is clear. The views of the authors on the effects of soil threats such as contamination or physical disturbances differ however. The view of Van Breemen is more or less the typical view of a geologist in which a fast process takes place in ten years, whereas for the soil biologist a fast process takes place within weeks. Van Breemen is laconic on the disturbance of soils, since it has been going on for ages in the Netherlands, and in all highly populated countries. Ernst has an optimistic view on the impact and on soil sustainability due to various adaptation phenomena. Verhoef is especially concerned with the fact that new stress added to existing stress may be fatal to the vitality of soil. Ernst indicates the capacity of living organisms to adapt to extreme habitats. His view is that, in general, plants and micro-organisms, and to an essentially lesser degree animals, have sufficient genetic potential to colonize every type of soil as long as water is available and temperature and radiation are in the range of biological processes. When natural changes or contamination impacts are slow, adaptation will occur. Under extreme conditions such as when saline or heavy metals are present, the biodiversity of the vegetation and soil fauna may be very low. Despite that low production, the soil is still vital since there is energy and nutrient flow. However, any evaluation of soil vitality first has to define the scope of activities. Principles of adaptation to slow alterations in the environment are given in Figure 3, which explains that plants adapt or acclimatize either by becoming tolerant and resistant, or by using or avoiding particular elements. Ernst's views on the sustainability of micro-organisms and animals are based on these principles, as shown for plants. Verhoef states that disturbance by heavy metals or eutrophication is best measured by the species diversity of the soil fauna. These disturbances lead to loss of diversity. Knowledge about the cause of a disturbance, determines the aim and direction of monitoring. Verhoef provides a clear set of monitoring parameters, however, he does not make a choice. In his view, the indication of soil quality in all cases should be physical, chemical and biological. Ernst is more outspoken as he refers to the Ellenberg-indicator set, and states that a general indicator of soil quality is the change in vegetation composition, being an expression of the soil conditions. Direction of change in vegetation can be predicted by determining the

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Figure 3. Principles of plants adaptation to slow alterations (modified after Ernst, 2003)

extent to which soils are loaded with N, P, S, heavy metals, and H+ (soil acidity). For practical purposes, this can be condensed into three soil types, nutrient levels and hydrology systems. The accumulation in leaf material and litter may be incorporated into the evaluation of soil quality as well as the ratio of fulvic humic acid and the presence of aliphatic humus compounds. Physical impacts on the soil, resulting in soil compaction can be evaluated by root and mycorrhiza development, which is restricted in compacted soil. When combining the different views of the three authors it can be concluded that in managing soil vitality, land use and land structure in the past, present and future has to be defined, as well as the expected impacts of physical, chemical and biological disturbance. 11.3.2 Linking of soil biota connections

This cluster focuses on and takes a broad view of the various aspects of the soil biota. The role of the soil microflora (Chapter 5, Verstraete and Mertens), the applicability of the soil fauna (Chapter 6, Van Straalen), the connected complexity of food webs, their dynamics and stability (Chapter 7, De Ruiter) and the substances and energy transfer along food chains, starting in soils and reaching birds, mammals and amphibians (Chapter 8, Van den Brink), is considered by these authors and their views on monitoring soil vitality are presented. Its coherence is depicted in Figure 4.

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Food chains higher animals (Van den Brink)

Soil microflora (Verstraete and Mertens)

Food web principles (De Ruiter)

Figure 4. Combined broad views on interaction patterns of soil biota

Function and value Verstraete and Mertens state that most chemical transformations in the soil involve the active contribution of micro-organisms, that play key roles in human, plant and animal health and therefore in environmental health. They indicate the various processes in the N- and P-cycle in relation to functional diversity. N-fixation and nitrification are based on small species diversity, and organic nitrogen mineralization is based on a large functional diversity. The application of molecular techniques has increased the possibility of investigating the presence of non-culturable microbes. In a table they summarize the key parameters to be monitored and they indicate their preference. Van Straalen states that below ground diversity is essential for above ground ecosystem function. The function of the soil fauna is a catalyzing one leading to the recycling of elements but also to vegetation diversity and succession. Due to their abundant presence, the soil fauna can be seen as the ecological insurance of the soil ecosystem. In an extensive table, Van Straalen summarizes the most sensitive soil fauna representatives (invertebrates) for the various groups of contaminants. He provides an overview of soil quality indices, to be applied in monitoring the degree of ecological recovery or decline. Monitoring indices in time is worthwhile when data from reference locations are lacking. De Ruiter states that the ability of soil ecosystems to stay in balance is the

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key of soil vitality. He describes interaction patterns and strengths between soil faunal predators and microbial and soil faunal preys. He emphasizes the function as prey and as predator among the invertebrates. The classification of invertebrates, based on food choice such as soil bacteria, fungi, specific mites, etc., provides insight into the functional diversity. The broader the functional diversity, the higher the stability of the whole community and the less vulnerable it is to disturbances. The food webs can be large, but the food pyramids are mostly three or four levels high: organic matter, microbes, microbe-consumers and fauna consumers. Van den Brink states that higher organisms, such as birds, mammals and amphibians in their various habitat and landscape structures, depend in many different ways on soil food webs. In food chains the quality of soil can be reflected in a very sensitive way. He refers to direct and indirect effects, resulting in concentrations of contaminants in higher animals, one or two orders of magnitude higher than in soil. Van den Brink exceeds the soil food webs by taking into account higher animals such as mice or birds feeding on invertebrates and those animals consumed by birds of prey. He provides impressive quantitative data on various food chains, but he warns us about the complexity of correct monitoring and explains the many pitfalls that can be expected. Monitoring properties The views of the different authors on how to monitor soil biota complement each other. The various components of soil quality include physical properties such as soil structure, aeration and hydrology; chemical properties such as organic matter and nutrient status; and biological properties such as the abundance and activity of soil organisms. The characteristics to be monitored are (i) materials, (ii) energy, (iii) nutrients, (iv) growth and decay rates in soil organisms and biodiversity, (v) soil community structure and ecological soil processes, and (vi) community food webs and food chains. Species diversity and especially functional diversity makes a functioning community more stable and secure against disturbances. Verstraete and Mertens assert there are too many uncertainties about degradation pathways under certain conditions and, for this reason control activities such as measuring degradation rates and conditions are important. The possible occurrence of harmful intermediate compounds of degradation processes (metabolites) is a concern. The fact that the public and politics are unaware of the importance of soil vitality is a major threat. Van Straalen indicates that most is known about the impact of heavy metals

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for which arthropods are less susceptible than earthworms. The compositions of the nematode and earthworm species communities are excellent tools to observe soil quality since it combines species diversity with functional aspects such as feeding strategy. In general, according to Van Straalen, the characteristics to be monitored and to be studied in relation to ecological threat or recovery are (i) species richness and diversity, (ii) distribution of numbers over species, (iii) the distribution of body-size over species, (iv) classification of life cycle attributes, (v) classification according to eco-physiological preferences, and (vi) the structure of the food web. De Ruiter states that the effects of contaminants on community structure and ecosystem function have hardly been studied and monitored. The knowledge necessary to enable the adequate measurement and monitoring of soil vitality is available: divide species into groups sharing the same prey and predators; quantify the biomass of the functional groups; and quantify the feeding rates among the functional rates. He shares Van Straalen's view that the most knowledge and experience available is on nematodes. The relevant characteristics of nematodes are given in Figure 5, just as an example how other groups may be classified. Resistance or adaptation of a microbial community may be coupled with a changed strategy towards energy and decomposition management (Doelman, 1986). A loss of functional degradation capacity has been shown under contaminated conditions (Doelman et al., 1994). However, its meaning has never been clarified, although a rapid method has been developed to measure heavy metal induced shifts (Rutgers et al, 1998). The Pareto-approach may be useful in assessing that meaning and is recommended by Verstraete and

Figure 5. Nematode characterization on two aspects offunctional diversity

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Mertens for monitoring soil vitality. It is a new strategy to study effective associations of micro-organisms. In soils, most of the microflora are passive or dormant. The addition of rather easily degrading organic matter usually activates a few percent of the microflora. It would appear, therefore, that there is a huge buffering capacity always on standby and normally hardly involved. This suggests that 80% of the task requires - for reasons of reservation and stability - at the most 20% of this diversity. The question is what do these shifts in active/passive population functioning ratios mean. My personal opinion is that this may lead to a fundamental research on microbial interaction or communication. Monitoring actual functional diversity in relation to potential functional diversity is a good warning system although it probably does not answer the fundamental question of what is the minimum required number of species necessary for functioning. This approach needs a combination of techniques that are as yet not fully available and involve a combination of old fashioned and recent genetic techniques such as (i) Most Probable Number counting (MPN) on specific substrate, (ii) molecular techniques, based on fluorescense in situ hybridization (FISH), (iii) thymidine incorporation, and (iv) direct (Fluorescense Microscopy) and (v) indirect measuring of total biomass by fumigation/respiration or (vi) adenylate energy charge (AEC). This may result in an expensive monitoring set. For soil fauna representatives the Pareto approach can also be applied. For nematodes the knowledge is available and De Ruiter (Chapter 7) provides applicable data for the predator and prey functions of many different soil fauna representatives. 22.3.3 Synthesized recommended monitoring programme

The authors of Chapters 2 - 8 were challenged to suggest solid monitoring options for practical situations. They were invited to move from fundamental knowledge to practical issues. Together with the information on the nature and behaviour of contaminants in soil (Peijnenburg, Chapter 9) and on monitoring (Breure , Chapter 10) the synthesis of key characteristics and monitoring schedules, called the basic commandments, is given in Table 2. The key characteristics are physical, chemical and biological. However, a precise indication of the physical aspects is almost completely missing. Ernst (Chapter 3) mentions the composition of the soil and groundwater. Van den Van Brink (Chapter 8) mentions landscape- and habitat-structure, and Breure (Chapter 10) mentions the physical characteristics without further defining them.

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P. Doelman

Table 2. The basic commandments: the summarized key characteristics and monitoring suggestions from Chapters 2-10 Cluster y^~\ / primuy^y 1

I

production

\

/ /

E ( m.D ^ - ^ \

^ \

Key characteristics - Dynamics organic matter - Soil structure

/~~^ Seeond«^\

/

production

\

(ve,i,oef,

)

„

.

.

.

.

- Species diversity

\ ;/

^ v I Soil ecosystem

Soil

\

biota

j

(Peijnenburg)

^-

""

I

(Breure) ^^

/

.

*

Chemical: - organic matter / nutrient status Biological: - activity and abundance (functional diversity)

" Microbial transformation . Sorption

Functioning: - Energy turnover - Nutrient recycling - Nitrification / N-fixation - Soil respiration ^-^

-.

- Growth rates Diversity: - Community structure - Species and body size distribution Functional diversity: - Indices Maturity Index Fungi / bacteria Oligotrophes / copiotrophes Change in catabolic genes - Mycorrhiza - Life cycle species - Ecophysiological preference species - Community food web structure - Predator-prey roles - Food chains - Biomarkers - Total concentration - Extractable concentration

r

/

^~~\

Monitoring

Physical: - soil structure / aeration / hydrology; " la n dscape-and habitatstructure -^^

(Vlns len>

-^

/"""

j

" Leaf litter composition - Roots and mycorrhiza

/^ Fate of * > \ ( contamination \

(

•1

1

1 u«**> \ J j | 1 pZ!ipiS j 1 J

* ^

. .

water

1 Foodch,to, I uS"°Mnt| / —v"—^J———A | mfc*Son / son A s°" \ L (Verslraele and

_-,

- Composition soil and ground

\y^

^ / |*— \ I

Monitoring - Mineralization organic matter - Fauna diversity

- Chemical/biochemical stabilization - Uptake plants - Uptake fauna Starting points:

- Chemical

- Political -Technical and statistical - Standardization

- Physical -Managerial - Biodiversity

N

) -"'^

j

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Current practice of assessing contaminated soil Policy considers contaminated soil along the domains: human health, ecological health and the risk of spread. For site specific ecological risk assessment a socalled TRIAD approach has been proposed in the Netherlands. It aims at providing evidence in a (i) chemical, (ii) toxicological, and (iii) ecological way: (i) total concentration of contaminants; (ii) bio-assays (with soil samples from the suspected site, measuring impacts on growth/reproduction of earthworms, on lettuce germination and microtox, etc.); (iii) ecology (field data from the suspected site on vegetation, nematodes (community composition in relation to soil type), earthworms (community and accumulation of contaminants), decomposition of bait lamina as cellulose in soil, etc.); This TRIAD approach appears to be simple but it partly agrees with the views of soil scientists. No physical characterization is involved and hardly any chemical characterization. Ecology is mainly based on nematodes, earthworms and vegetation. Bio-assays as toxicology experiments on certain species are mentioned as having diagnostic value. However, soil bio-assays have a potential value rather than a realistic value. This is due to the pre-treatment of the soil sample as homogenization, artificially increasing the availability of the contaminants. Rombke et al. (2003) provide an extensive overview on the feasibility of bio-assays in soil. In the written publication Rombke does not discuss the limitations of over-exposure, in oral presentations he does (Rombke, 2003). The technical collecting of TRIAD data is a simple step in comparison to explaining the meaning and the consequences for policy. Breure (Chapter 10) indicates the various steps to be taken before technical data, such as ecological soil risk data, are settled in policy and legacy and licensing acts. Technical practice "in the field" aims at simple, reproducible and payable methods (Eijsackers, Chapter 1; Breure, Chapter 10). Reviewing doses effect data with single species in the context of the function of those species, combined with functional diversity considerations (Chapter 1-8) should lead to far more attention to the field data aspect of the TRIAD approach in practice. 11.4 Considerations on key characters and monitoring of vital soil In this concluding part the different views are brought together, replenished with clarifications from the literature available and guided by my own view. It begins with noting the difficulty of defining and quantifying the various

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integrated parts in a soil ecosystem. The holistic, appealing and perceptual notions such as resilience, resistance, robustness, stability, equilibrium and disturbance will be bent to practical approaches which can be system oriented with a desired function and/or a desired food web and/or a desired "pet" animal. 11.4.1 Holistic approaches

What all authors share is their fascination with the abundance of soil life and the seemingly immoderate excess of functional equivalents within soil flora and soil fauna. They indicate that there is not a quantitative overview of heterogeneity in soil habitats, the living areas of species and communities, and that there is limited knowledge about the eco-physiological properties and potentials of the many species and communities. There is consensus on the view that many interactions are still unknown or hidden, but their views on the consequences of disturbances vary. The impact of disturbance on the stability of an ecosystem is given in Figure 6. It illustrates that after a temporary disturbance the same or a new equilibrium will appear. Whether we appreciate the features of this equilibrium may also depend on perception. Disturbance and adaptation towards a new stability is common in soil ecosystems. Eijsackers (Chapter 1) discusses natural fluctuation with large and small amplitudes and their consequences for resilience. His view is that eventually the equilibrium has to return to the original level of stability. He

Figure 6. Some disturbance notions in soil ecosystems (modified after Aber and Melillo, 1991, by introducing a larger variety of new equilibria)

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relates the time function to the restabilization of specific parts of the soil ecosystem. In the case of a particular type of vegetation decades may be needed for recovery to take place and to return to its equilibrium whereas in the case of soil profiles the process may take centuries. Van Breemen (Chapter 2) reviews along the geological time-table (extremely long time periods), observing that somehow there will always be a new equilibrium at any possible level. Ernst (Chapter 3) strongly emphasizes the factor of time in relation to disturbance time, stating that when the changes are slow and mild there is acclimatization or an establishment of a new equilibrium on a different level of functioning such as in areas heavily polluted with heavy metals. Verhoef (Chapter 4) observes that due to limited knowledge about interactions it is difficult to draw any final conclusion at this time on disturbances, but that he is concerned about ecosystem sustainability. A pessimistic view is expressed by Verstraete and Mertens (Chapter 5) where they warn about a possible collapse. Van Straalen (Chapter 6) does not express a view on new or old equilibria, but he emphasizes the difference in species sensitivity to gradual disturbances. He indicates ways of determining the number of equivalent functioning organisms within a community by determining their life cycle strategy, including their ecophysiological preference. This may explain stability and reactions to disturbances. De Ruiter (Chapter 7) explains and unravels soil stability in relation to soil vitality. He demonstrates the basic principles of the stability of soil biodiversity within three trophic levels. He refrains from evaluating the various levels of stability or new equilibrium. Van de Brink (Chapter 8) relates the stability and sustainability of larger habitat structures to his food chain approach. Any new equilibrium affects the food chain of higher animals. In soil ecology literature there are fascinating examples of "incredible equilibria", of soil ecosystems functioning. Two of them are mentioned briefly. Wall and Virginia (1999) describe the functioning of a simple food web in the Antarctic Dry Valley (cold and windy): algae as primary producers, mineralization (recycling) by bacteria and fungi and three nematode species: Scottnema as fungivore and bacterivore, Plectus as bacterivore and Eudorolaimus as omnivore-predator. Their habitat is a high surface roughness causing a relatively calm layer above the soil, protecting it from being blown away. Another example comes from the Negev desert (dry and warm) where algae and one micro arthropod create an equilibrium, a new ecosystem (Shachak, 2003). Organic matter, water and shelter seem to be the basis for invasion and creation. These two examples are extremely susceptible to disturbances since their functioning is based on limited diversity.

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Soil functioning and species diversity The relationship between soils as a functioning system and species diversity has in principle four styles such as "redundancy", "rivet", "predictable change" and "idiosyncratic" (Ritz and Griffith, 2001). Redundancy is the superfluous, the excess, of organisms capable of carrying out the same performance. The hypothesis is that as long as all functional groups are represented, system functioning is independent of diversity. Rivet is that all species make a significant contribution to function and that a decrease in diversity leads to a progressive decline in function. Predictability change is that a change in diversity results in an alteration in function, which is predictable given knowledge about the trophic structure, biomass and the principle relationships between the biotic components within a system. Idiosyncratic is the hypothesis that it is all unpredictable. So perception, besides knowledge, may determine the view on functioning. Recolonization/restoration As long as the disturbance that takes place is mild and slow, acclimatization will take place on the same or another level of functioning and of diversity. When there has been a fatal impact and complete elimination of the biota the physical and chemical status of the soil determines revitalization. The possibility of re-introduction from outside sources is important. Active mobility, passive co-transport and the size of the empty habitat are important factors in re-introduction. Smaller animals immigrate more easily. Their sources or 'refugia' may be storage sources such as seed from deeper layers. The capacity to survival for prolonged periods thus becomes important. Nelumbo nucifera seed survived in peat for 800 years (Chaney, 1951). Micro-organisms may survive in resting spore stages for decades and certain nematodes and micro arthropods can survive for months (Goodey, 1923). Besides new residents entering the habitat, there may be habitat changes caused by the re-creation of new physical conditions for equivalent functioning (eco-engineering). Scientific knowledge and practical experience have shown mankind to be able to exploit soil for desired biological soil functions. Agriculture is such an example. For certain desired functions the prerequisites are physical and chemical soil characteristics. For certain desired diversity or key-higher animals time is important so the right food webs and appropriate food chains can be developed in the demanded landscape structure (the ambition). The role of soil contaminants in relation to function and diversity has been addressed in all chapters. In addition Peijnenburg (Chapter 9) has shown how and over which period of time the soil may eliminate these contaminants. So the actual

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monitoring of soil vitality depends on the questions raised, either towards functioning or towards diversity. 11.4.2 Practical approaches

Physical, chemical and biological soil characteristics are the basic elements of a proper assessment of vital soil. These cornerstones have to be considered in an integrated manner as the definition of ecology commands, referring to Tansley (1935) and Odum (1971) and illustrated in Figure 7. All views in the preceding chapters subscribe to this principle message. In addressing characteristics, the biological cornerstone is mentioned most. The applicable monitoring items are diversity of vegetation and of earthworms, micro arthropods and nematodes, microbial biomass and nitrification. Soil chemistry is addressed as organic matter, N- and P-content and groundwater quality in relation to its nutrient, salt and acidity level. These aspects are measurable on a routine basis and therefore practically applicable. Soil physics is hardly applied at all in present day soil ecological risk assessment. Soil structure, aeration and hydrology have been mentioned, can be measured, but are hardly applied. This lack of physical characterization is surprising. Van de Brink (Chapter 7) emphasizes the landscape structure being indicative for the kind of higher animals present. Both Verhoef (Chapter 4) and Verstraete and Mertens (Chapter 5) comprehensively praise the role of spatial heterogeneity (habitats, living areas, living streets) as the prerequisite of biodiversity.

Figure 7. The basic triangle to monitor soil vitality

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P. Doelman

Moreover, they recognize homogenizing and paving of soils as threats to soil biodiversity, but they do not mention soil structural units as the ratio between solid, liquids and gas phases as Water and Oades (1991) have done. The pore size determines the presence possibilities for the various representatives of soil biota. The relation between pore size and biota is shown in Table 3. Therefore I suggest that monitoring of compaction and pore size (actual soil structure!), as proposed and carried out on a routine bases by Sparling (2003), should be one of the three cornerstones in the practical set. 11.4.3 General and tailor-made approaches

Present insights into the functioning of soil ecosystems point to (i) the need for more attention to functional diversity within the community structure than monitoring specific species and (ii) to the (so far neglected) physical characterization of the soil and other environmental factors. The single species approach to sensitive organisms, as has been applied for many years, should be integrated into functional approaches. Resistant species may take over the function of eliminated sensitive species. Consequently the bio-assay, as suggested cornerstone in the TRIAD approach, will be less relevant since it mostly concerns one species. Moreover, there are many promising techniques to determine the bio-available fraction of contaminants in soil (Peijnenburg, Chapter 9). The general monitoring set is already given in Table 2, several parameters of which still have to be included in standard laboratory tests. The actual situation of a soil (its present use) and the intended function determine the type of tools that should be used. In Table 4 specific monitoring sets for contaminated soils are given. A tailor-made set for the ecological

Table 3. Pore size in relation to biota (modified after Waters and Oades, 1991) Pore function Micropores Adsorbed Intercrystallinic Water Mesopores (plant available Water) Macropores (aeration Fast drainage)

Biota

Organic molecules Numic substances Viruses Bacteria Fungal hyphae Root hairs Roots Mesofauna Worms Moles

Metres = 1O10 (1 Angstrom) = 10-9 (1 Nanometer) = 10-8 = io-7 = 1CH> (1 Micrometre) = 10-5 = 10-4 = 10-3 (1 Millimetre) = 10-2 = 10-1 (10 Centimetre)

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Table 4. Specific monitoring sets attuned to various contamination compound groups

Heavy metals Chemical litter composition Bacterial biomass Bact. ratio sensitive/resistance Nematodes Earthworms: species diversity and bio-accumulation Soil fauna Food web accumulation Extractable level heavy metals

Carbon content N content Soil type Soil compaction (structure) Soil pH, redox Landscape structure Nutrients Phosphate Vegetation Soil water composition: Ec, pH, Nutrients Mycorrhiza Bact. ratio copio-/oligotrophic Nematodes

POPs (+ PAHs) Bacteria: catabolic genes Biodegradation rate Earthworms: species diversity and bio-accumulation Food web accumulation Extractable level POP and PAH

recovery of a fallow soil area or a former paved area will differ from contaminated areas. These are brownfields such as large waste deposit areas, former metal smelting areas, gaswork areas, dredget sediment material areas and eutrophic areas. These areas should be considered super-tailor-made when it comes to determining ecological threat or ecological recovery. It should be realized that alteration may proceed slowly and this will again involve longterm monitoring in which monitoring frequency might be reduced and a number of parameters changed or deleted during time. A general view is that monitoring references should be used as kind of check or control. My opinion is that real control areas do not exist. However, for certain combinations of soil type and land use reference values do exist such as nitrification, soil biomass, nematodes, earthworms, enchytraeids and mites (Schouten et al., 2004; Breure, Chapter 10). When reference cannot be applied the degree of recovery or extinction can be shown by internal indication references such as nematode maturity index, predatory mite maturity index, biological soil quality index (Van Straalen, Chapter 6), the relative number catabolic genes, ratio oligotrophes/capiotrophes (Verstraete and Mertens, Chapter 5). The practical consequences of monitoring in relation to numbers (replicates) and frequency are mentioned by Breure (Chapter 10). What he recommends will be very expensive, since he aims at statistical significance. In my view, two- or threefold observations (monitoring actions) once a year or once in two years might be technically and financially manageable.

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11.5 Final remarks Soils are vital and their degree of vitality can be reflected in the way they function. Knowledge and experience necessary to make a soil function has been around for a thousand years. The sequence of achievement should be source pathway - desired unit (ambition level). There is sufficient scientific knowledge and field experience to answer all questions from policy and practice about the functioning of soil. Government and soil managers need to be clear about their desired unit, whether they want function or biodiversity. Then scientists should clarify the possibilities and prerequisites (the source and the pathway). For plants, soil fauna and soil microflora these pathways are far clearer than for higher animals. References Aber, J.D. and J.M. Melillo, 1991. Terrestrial Ecosystems. Sounders College Publishing. USA, p 67. Anderson, J.M., 1977. The organization of soil animal communities. In: Lohmm, U. and T. Persson, Eds. Soil organisms as components of ecosystems. Proc 6th Int. Zool. Coll. Ecol. Bull., Stockholm, Vol. 25:15-23. Anderson, J.M., 1975. The enigma of soil animal species diversity. In: progress in Soil Zoology, J. Vanek, ed., pp 51-58. The Hague: Junk BVZz. Andren, O., J. Bengtsson and M. Clarholm, 1995. Biodiversity and species redundancy among litter decomposers. In: H.P. Collins, G.P. Robertson and M.J. Klug (eds). The significance of regulation of soil biodiversity 141-151; Kluwer Acad. Publ. Netherlands. Beare, M.H., D.C. Coleman, D.A. Crossley Jr., P.F. Hendrix and E.P. Odum, 1995. A hierarchical approach to evaluating the significance of soil biodiversity to biogeochemical cycling. Plant Soil 170:5-22. Bongers, T. and M. Bongers, 1998. Functional diversity of nematodes. Appl. Soil Ecol. 10: 239-251. Brussaard, L., T.W. Kuyper, W.A.M. Didden, R.G.M. de Goede and J. Bloem, 2004. Biological soil quality from biomass to biodiversity - Importance and resilience to management stress and disturbance. In press. Butler, G.C., 1978. Principles of ecotoxicology. Ed. Butler., Scope 12. Chaney, R.W., 1951. How old are Manchurian lotus seeds?. Garden J., N.Y. Botan. Gardens 1:137. Doelman, P. 1986. Resistance of soil microbial communities to heavy metals. In: Microbial Communities in soil, pp 369-383. Eds. Jensen, V., A. Kjoller & L.H. Sorensen. Doelman, P., E. Jansen, M. Michels and M. van Til, 1994. Effects of heavy metals in soil on microbial diversity and activity; the sensitivity/resistance index, an ecologically relevant parameter. Soil Biol. Fertil. Soils 17: 177-184. Doran, J.W and T.B. Parkin, 1994. In: Doran, J.W., D.C. Coleman, D.F. Bezdicek and B.A. Steward (eds). Defining soil quality for a sustainable environment. SSSA Spec Publ. 35 SSSA Madison, WI. USA. Dutch Health Council, 1988. Ecological risk-evaluation of components. Eijsackers, H., 2001. A future for soil ecology Connecting the system levels: moving from genomes to ecosystems. Eur. J. Biol. 37: 213-220.

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Ellenberg, H., 1979. Zeigerwerte der Gefasspflanzen Mitteleuropas. Scripta Geobotanica, 9. Ehrlich, P. and A. Ehrlich, 1981. Extinction, the causes and consequences of the disappearance of species; Ballantine books. New York. Ernst, W.H.O. 2002, Living at the border of life. Retirement Oratio Vrije Universiteit, Amsterdam. Faber, J.H., 1992. Soil fauna stratifications and decomposition of pine litter. PhD-thesis Vrije Universiteit Amsterdam. ISBN 90-9005091-0. Goodey, T., 1923. Quiescence and reviviscense in nematodes with special reference to ylenchus triotici and tylenchus dispar. J. Helminthol. 1: 47-58. Gove, P.B. and the Merrian-Webster editorial staff, 1976. Websters third international dictionary; unabridged. Publishers G.&C. Merriam Company, Springfield, Massachustts, U.S.A. Kelly, E.J. and K. Campbell, 2000. Separating variability and uncertainty in environmental risk assessment-making choices. Human Ecol. Risk Assess. 6 (1) 1-13. Lancaster, J., 2000. The ridiculous notion of assessing ecological health and identifying useful concepts underneath. Human Ecol. Risk Assess. 6 (2) 213-222. Meadows, D.L., A. King, S.O. Kitai, A. Peccei, E. Peslel, H. Theimann and C. Wilson, 1972. The limits to growth. Universe Books New York. Nealson, K.H., 2002. Discussion, Symposium Subsurface Microbiology, Copenhagen. Odum, H.T., 1971. Fundamentals of ecology. Philadelphia PA W.B Saunders. O'Neil, R.V., D.L. DeAngeles, J.B Waide and T.F.H Allen, 1986. A hierarchical concept of ecosystems. Princeton University Press Princeton NJ, 263 pp. Ritz, K and B.S. Griffiths, 2001. Implications of soil biodiversity for sustainable organic matter management. In: Rees, R.M., B.C. Ball, CD. Cambell and S.A. Watson (eds). Sustainable management of soil organic matter. CAB International, pp 343-356. Rombke, J. and J. Weeks, 2003. The feasibility of bioassays in site-specific ecological risk assessment (SS-ERA). 8th Int. FZK/TNO Conference on Contaminated Soil, Conference Proceedings, Cdrom. pp. 3602-3605. Rombke, J., 2003. Personal discussion after the presentation. Rutgers, M., I.M. Verlaat, B. Wind, L. Postuma and A.M. Breure, 1998. Rapid method for assessing pollution-induced community tolerance in contaminated soil. Environ. Toxicol. Chem. 17: 2210-2213. Salomons, W. and W.M. Stigliani (Eds), 1995. Biogeodynamics of pollutants in soils and sediments. Springer -Verlag Berlin Heidelberg. Shachak, M., 2003. Lecture: Current themes in Ecology. Wageningen. Sparling, G.P., L. Lilburne and M. Vojvodic-Vukovic, 2003. Provisional targets for soil quality indicators in New Zealand; Palmerston North N.Z.: Landcare Research New Zealand. Sparling, G.P. 2003b. Discussion. Course Soil Ecology, Wageningen. Tansley, A.G., 1935. The use and abuse of vegetational concepts and terms. Ecology 16 (3): 284-307. Vigerstad, T.J. and L.S. McCarty, 2000. The ecosystem paradigm and environmental risk management. Human Ecol. Risk Assess.: 6 (3) 369-381. Wall, D.H. and R.A. Virginia, 1999. Controls on soil biodiversity: in sight from extreme environments. Appl. Soil Ecol. 13:137-150. Ward, P., 1994. The end of evolution, a journey in search of clues to the third mass extinction facing planet Earth. Bantam Books. New York Toronto London Sydney Auckland.

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Waters, A.G. and J.M. Oades, 1991. Organic matter in water-stable aggregates. In: Advances in Soil Organic Matter Research: The Impact on Agriculture and the Environment. Wilson, W.S. (ed.). The Royal Society of Chemistry, Cambridge, pp 163-174.

333

SUBJECT INDEX

A above ground biomass agri-environment animal diversity arbuscular mycorrhiza

44 233 160

58, 59, 63,142,145

B bacteria bacteria/fungi balance and stability balanced population bioaccessible fraction bioavailable fraction biogeochemical cycle biological disturbance biological diversity biomarkers bound residues breeding areas breeding sites

102 297

200, 201, 202, 205, 207, 209, 210, 212 197,199 245, 267, 269, 273, 274 144, 256, 270, 272, 276 37,75 42, 69, 86, 317 70

18,127,144,145,150, 215 15, 275 220 219

c C/N ratio C3-pathway C4-pathway carnivorous catabolic genes cation exchange capacity chemical residues C-horizon climax vegetation CO2 fixation

49, 55, 63, 64, 65, 66, 67, 72 45,46 46 230

127,135,145,147, 322, 329 273 256

64,67 55, 64, 68 46

334

community structure

compacted soil contamination pattern

Subject index

108,114,115,145,163,165,166,167,169,174, 176, 177,179,180, 197, 200, 201, 204, 206, 207, 210, 319, 320, 328 42, 70, 314, 317 228, 235

D decomposition processes degradation rates deposition gradient desertification direct effects dispersion of animals distribution coefficient disturbance indicator diversity dynamic diversity of species dunes

83 262 169 23, 69, 81 163, 215, 217, 218, 220, 222, 223, 232, 233 219 258 101,108 102 151,162, 302 46, 51, 65, 66, 67, 69, 81, 83,148

E ecological insurance ecological quality indicator ecological risk assessment ecological surveys ecophysiological potential ecophysiological preferences ecosystem compartments ecosystem engineers ecotoxicological mechanisms ecotoxicological responses ecotype ecto mycorrhiza Ellenberg indicator value endo mycorrhiza equilibrium partitioning

159,161, 318 200, 298 42, 52, 294, 323, 327 159,162,174,182,183,185 52 159,177 198, 211 107 218 163 49 29, 58, 59, 61, 64, 78, 79, 111 50 60 165, 256, 277

F feeding efficiency

215, 220

Subject index

feeding rates fermentation layer fluctuations in numbers food chain food web approach food web modelling food web relations food web structure fulvic acid functional diversity functional replaceability fungi

335

197, 204, 205, 206, 207, 208, 212, 320 58 1, 6 4,103,165,175, 201, 202, 203, 205, 228, 310, 315, 317, 319, 325, 326 105 205 215, 228 112, 201, 205, 207, 288 57, 63, 64, 78, 86, 87 1,127,133,139, 312, 313, 318, 319, 320, 321, 322, 323, 328 106 102

G G-horizon grassland

67 11,18, 27, 34, 37, 45, 46, 52, 54, 57, 59, 66, 67, 69, 71, 73, 80, 84, 85,103,110,141,179,181, 184, 219, 222, 288

H habitat change habitat function habitat preference habitat structure halorespiration hazardous concentrations heath land heavy metal vegetation herbivorous higher trophic levels humic acid humification humus forms humus layer hydrophobic chemicals

113, 221, 233, 326 198,199 215, 219 215, 219, 220, 221, 222, 231, 307, 325 145,263 5 54, 59, 71 82 202 112,113, 202, 206 46, 57, 64, 76, 86, 87, 255, 268, 317 10,11, 49, 55, 58, 70, 76, 78,117, 251, 255 31, 32, 33, 35, 41, 42, 58, 59, 60, 61, 65, 79, 83, 86 51, 55, 66, 80 248, 257, 261

336

Subject index

I indicator value indirect effects insectivorous interaction strength isotope ratio

41, 42, 52, 55, 70, 86,115, 286, 292, 295, 296 217, 218, 219, 221, 222, 231, 232, 319 229,230 177,197, 206, 207, 208 78

L landscape structure landslides lipophilic materials lithosphere litter layer

220, 319, 326, 327 1, 2, 3 258 42,62 55

M management agreement man-made soils marshy environment metal attenuation metal resistant metal speciation microbial biomass microbial communities microbial indicator moder monitoring options mor mowing frequency mull mycorrhizal fungi

233 70 81 74 47 259, 270 59,103,104,110, 111, 113,117,118,133,135, 147, 281, 299, 300, 303, 327 103,109,117,118,131,138,146,148,150,170, 212, 299 134, 135 32, 33, 57, 58, 59, 60, 61, 77, 80, 84 1, 309, 321 32, 33, 34, 57, 58, 59, 60, 61, 63, 77,178 84 32, 33, 34, 57, 58, 59, 60, 61, 65, 66, 80, 84,178 23, 60, 62, 63, 64, 65, 86,108,128

N natural attenuation netto primary production

132,198, 245, 246, 250, 251, 253, 254, 255, 261, 265 43, 44, 45, 303

337

Subject index

N-fixing species N-input N-loss N-mineralization nutrient status nutrient-rich

141 53,112 53 103,105,135,139,140,141,147, 288 83,198, 276, 303, 319, 322 84

o O-horizon oligotrophic ecosystems omnivorous orchid mycorrhiza organic matter

58, 63, 65, 71 76 208 60 1, 3,10,11,12,16, 22, 25, 26, 27, 32, 33, 35, 36, 37, 41, 42, 49, 63, 67, 70, 74, 83, 86, 87, 99,100, 101,102,103,107, 111, 112,113,116,118,128, 135,140,161,170,173,181,182,197,198,199, 201, 202, 207, 209, 210, 211, 212, 245, 246, 255, 258, 259, 260, 261, 262, 266, 269, 271, 273, 274, 275, 277, 282, 297, 300, 303, 307, 308, 309, 314, 315, 319, 321, 322, 325, 327

over-exploitation

69

P parent material phenomenology photosynthesis physical disturbance phytoremediation pioneer vegetation piscivorous plant available plant diversity plant productivity pore water concentration practical implementation preferential feeding localities prey items

21, 22, 23, 24, 26, 27, 30, 42, 49, 59, 62, 63, 255, 307, 308, 315 1 41, 313 316 52 65,72 230, 232 328 41, 114,161,162 64 257, 270 212, 302 220 215, 216, 219, 220, 222, 231, 232, 236

338

primary production production function

Subject index

2,10,11,18, 41, 87,103, 105, 297, 307, 309, 311, 312, 315 198, 199

R ratio recycling of elements redox conditions redundancy reference sites re-introduced animals relative toxicity factor resilience resistance restoration reversible impacts robustness root exudates rooting depth

49, 57, 58, 80, 86, 87,127,135,146,147, 258, 261, 297, 300, 303, 317, 321, 328, 329 99, 263, 314, 318 111, 259, 263, 274 12,106,161, 295, 326 17, 240 80 168 6,14, 281, 285, 289, 290, 293, 295, 297, 311, 324 6, 47, 61, 72, 74, 75,109, 297, 311, 324, 329 1, 6, 72, 81,179, 283, 294, 326 6 6,324 43, 44, 86, 266, 316 22, 46, 86

S secondary production sensitive species Shannon-Wiener diversity index shaping anthrosols soil acidification soil aggregate stability soil biology soil chemistry soil community soil ecology soil ecosystem function soil erosion soil fauna

2, 307, 315 47,11,163, 328 166,167 68 31, 36, 37, 54, 58, 75,170 142 17,23,303 21, 28, 45, 53, 68, 72, 327 99,108,112,114,115,119,160,197,199, 200, 201, 204, 207, 209, 211, 319 87,161,162,198,199, 211, 325 106,198, 211, 307, 308 2,10, 38, 69, 72, 80,151 12,13,14,15,16, 24, 26, 29, 30, 32, 59, 99,100, 103,104,105,107,108,109, 111, 129,159,161, 163,170, 212, 254, 265, 307, 308, 309, 313, 315,

Subject index

soil fertility soil forming factors soil health soil horizon soil management

soil microflora soil moisture regime soil nutrient regime soil policy soil pores soil profile

soil protection soil protection policy soil quality

soil soil soil soil soil soil

quality diamond quality grid quality indices quality management stability structure

soil sustainability soil texture soil threats

339

316, 317, 318, 319, 321, 324, 329, 330 53, 55, 69,128,161,182,197,198,199, 207, 210, 253, 293 24, 27, 29, 33, 315 3, 41, 85, 86,133,134,135,139,143,144,148, 150, 312 22, 26, 34, 35, 47, 64, 70, 84, 105 21, 37, 38, 41, 87, 99,100,119,150,151,159, 184,198, 209, 210, 211, 212, 277, 285, 287, 302, 312 1, 99,100, 307, 308, 309, 313, 315, 317, 330 49, 50 49, 50 3, 4,17, 283, 284, 302, 307, 308 46, 76, 268 22, 27, 31, 37, 42, 47, 58, 62, 63, 64, 65, 66, 68, 70, 71, 72, 73, 74, 80, 81, 82, 83, 84, 86,178, 179, 325 1, 4, 5, 84,146,150, 210, 302, 311 3, 5, 84, 210, 283 3, 5,17,18, 36, 41, 42, 52, 55, 60, 73, 75, 79, 86, 99,100,101,102,106,114,115,116,117,118, 119,127,132,146,150,151,159,162,174,176, 177,180,182,183,184,197,198,199, 200, 209, 215, 217, 218, 219, 220, 221, 222, 223, 224, 225, 229, 230, 231, 233, 236, 238, 239, 240, 276, 281, 284, 285, 288, 289, 291, 293, 298, 300, 313, 316, 319, 320, 329 182,183,185 49 101,115, 318 150,197,198, 281 325 6,12, 22, 24, 26, 28, 34, 35, 42, 43, 46, 50, 61, 68, 70, 72, 86, 99,102,107, 111, 116,133,175, 198, 288, 308, 314, 319, 322, 328 1, 3, 316 12, 26 15, 316

340

soil type

soil vitality

spatial scale species composition species identity stability

stability indicator structural diversity sustainable soil ecosystem symbiotic types

Subject index

6,16,17,18, 24, 41, 44, 57, 60, 87,105,114, 115,116, 210, 219, 260, 284, 287, 291, 295, 296, 298, 301, 303, 307, 308, 310, 317, 323, 329 1, 42, 85, 86, 87,127,197,198,199, 200, 201, 207, 209, 210, 211, 307, 308, 311, 314, 316, 317, 319, 320, 321, 325, 327 41,107, 219, 224, 225, 228, 237, 239 32, 41,167,168,184 105,106 7, 23, 36, 57, 66, 99,100,101,105, 111, 114, 116,132,142,143,161,162,166,197,199, 200, 201, 202, 203, 204, 205, 206, 207, 209, 210, 212, 281, 282, 283, 288, 290, 293, 295, 299, 300, 301, 303, 307, 311, 315, 317, 321, 324 201 6 201 53

T temporal distribution terrestrial ecosystems time scales trophic biomass pyramids trophic groups trophic levels

232 58, 217, 293, 298, 310 21, 28 197, 207 103,104,105,161, 201, 207, 282 177, 202, 206, 312, 325

V vegetation zones vital soil

65, 67 1, 3, 4, 295, 307, 308, 323, 327