Function of Soils for Human Societies and the Environment (Geological Society Special Publication No. 266)

Function of Soils for H u m a n Societies and the Environment

The G e o l o g i c a l Society of L o n d o n

Books Editorial Committee Chief Editor BOB PANKHURST (UK)

Society Books Editors JOHN GREGORY (UK) JIM GRIFFITHS (UK) JOHN HOWE (UK) PHIL LEAT (UK) NICK ROBINS (UK) JONATHAN TURNER (UK)

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FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266.

JANSA, J., WIEMKEN,A. & FROSSARD,E. 2006. The effects of agricultural practices on arbuscular mycorrhizal fungi. In: FROSSARD, E., BLUM, W. E. H. & WARKENTIN, B. (eds) Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 89-115.

GEOLOGICAL SOCIETY SPECIAL PUBLICATION NO. 266

Function of Soils for Human Societies and the Environment

EDITED BY E. F R O S S A R D ETH Zurich, Switzerland W. E. H. B L U M University of Natural Resources and Applied Life Sciences (BOKU), Austria and B. E W A R K E N T I N Oregon State University, USA

2006 Published by The Geological Society London

THE

GEOLOGICAL

SOCIETY

The Geological Society of London (GSL) was founded in 1807. It is the oldest national geological society in the world and the largest in Europe. It was incorporated under Royal Charter in 1825 and is Registered Charity 210161. The Society is the UK national learned and professional society for geology with a worldwide Fellowship (FGS) of over 9000. The Society has the power to confer Chartered status on suitably qualified Fellows, and about 2000 of the Fellowship carry the title (CGeol). Chartered Geologists may also obtain the equivalent European title, European Geologist (EurGeol). One fifth of the Society's fellowship resides outside the UK. To find out more about the Society, log on to www.geolsoc.org.uk. The Geological Society Publishing House (Bath, UK) produces the Society's international journals and books, and acts as European distributor for selected publications of the American Association of Petroleum Geologists (AAPG), the Indonesian Petroleum Association (IPA), the Geological Society of America (GSA), the Society for Sedimentary Geology (SEPM) and the Geologists' Association (GA). Joint marketing agreements ensure that GSL Fellows may purchase these societies' publications at a discount. The Society's online bookshop (accessible from www.geolsoc.org.uk) offers secure book purchasing with your credit or debit card. To find out about joining the Society and benefiting from substantial discounts on publications of GSL and other societies worldwide, consult www.geolsoc.org.uk, or contact the Fellowship Department at: The Geological Society, Burlington House, Piccadilly, London WIJ 0BG: Tel. +44 (0)20 7434 9944; Fax +44 (0)20 7439 8975; E-mail: [email protected]. For information about the Society's meetings, consult Events on www.geolsoc.org.uk. To find out more about the Society's Corporate Affiliates Scheme, write to [email protected]. Published by The Geological Society from: The Geological Society Publishing House, Unit 7, Brassmill Enterprise Centre, Brassmill Lane, Bath BA1 3JN, U K (Orders: Tel. +44 (0)1225 445046, Fax +44 (0)1225 442836) Online bookshop: www.geolsoc.org.uk/bookshop The publishers make no representation, express or implied, with regard to the accuracy of the information contained in this book and cannot accept any legal responsibility for any errors or omissions that may be made. 9 The Geological Society of London 2006. All rights reserved. No reproduction, copy or transmission of this publication may be made without written permission. No paragraph of this publication may be reproduced, copied or transmitted save with the provisions of the Copyright Licensing Agency, 90 Tottenham Court Road, London W I P 9HE. Users registered with the Copyright Clearance Center, 27 Congress Street, Salem, MA 01970, USA: the item-fee code for this publication is 0305-8719/06/$15.00.

British Library Cataloguing in Publication Data A catalogue record for this book is available from the British Library. ISBN 10:1-86239-207-2 ISBN 13:978-1-86239-207-6 Typeset by Type Study, Scarborough, UK Printed by Cromwell Press, Trowbridge, UK Distributors North America For trade and institutional orders: The Geological Society, c/o AIDC, 82 Winter Sport Lane, Williston, VT 05495, USA Orders: Tel. +1 800 972-9892 Fax +1 802 864-7626 E-mail [email protected] For individual and corporate orders: A A P G Bookstore, PO Box 979, Tulsa, OK 74101-0979, USA Orders: Tel. +1 918 584-2555 Fax +1 918 560-2652 E-mail [email protected] Website http://bookstore.aapg.org India Affiliated East-West Press Private Ltd, Marketing Division, G-l/16 Ansari Road, Darya Ganj, New Delhi 110 002, India Orders: Tel. +91 11 2327-9113/2326-4180 Fax +91 11 2326-0538 E-mail [email protected]

Contents BLUM,W. E. H., WARKENTIN,B. E & FROSSARD,E. Soil, human society and the environment FELLER, C., MANLAY,R. J., SWIFT,M. J. & BERNOUX, M. Functions, services and value of soil organic matter for human societies and the environment: a historical perspective LEIFELD, J. Soils as sources and sinks of greenhouse gases

23

BERGSTROM, L. E & DJODJIC, E Soil as an important interface between agricultural activities and groundwater: leaching of nutrients and pesticides in the vadose zone

45

Dosso, M., PHILIPPON, O. d~zRUELLAN,A. Understanding of a soil system derived from a single bed-rock, for improved vineyard management in Southern France

53

LUSTER, J., ZIMMERMANN,S., ZWICKY,C. N., LIENEMANN,P. 8z BLASER, P. Heavy metals in Swiss forest soils: modification of lithogenic and anthropogenic contents by pedogenetic processes, and implications for ecological risk assessment

63

BA~UELOS, G. S. & LIN, Z.-O. Reuse of agricultural drainage water in central California: phytosustainability in soil with high levels of salinity and toxic trace elements

79

JANSA,J., WIEMKEN,A. • FROSSARD,E. The effects of agricultural practices on arbuscular mycorrhizal fungi

89

BURGHARDT,W. Soil sealing and soil properties related to sealing

117

WELLS, E. C. Cultural soilscapes

125

LANDA, E. R. From agricultural geology to hydropedotogy: forging links within the twenty-first-century geoscience community

133

HAZELTON,P. A. Australian examples of the role of soils in environmental problems

141

MONTANARELLA,L. Policies for a sustainable use of soil resources

149

INACIO, M., PEREIRA, V. & PINTO, M. Assessing anthropogenic inputs to soils by comparing element contents and their spatial distribution in O- and A- horizons

159

MENZI, H. & GERBER, P. Nutrient balances for improving the use-efficiency of non-renewable resources: experiences from Switzerland and Southeast Asia

171

UGOLINI,E C. & WARKENTIN,B. P. Perspectives on the relationship between soil science

183

and geology Index

191

Soil, human society and the environment W. E. H . B L U M 1, B. R W A R K E N T I N

2 & E. F R O S S A R D

3

1University o f Natural Resources and Applied Life Sciences (BOKU), Vienna, Austria (e-maik winfried, blum@boku, ac.at) 2Oregon State University, Corvallis, USA 3ETH Zurich, Eschikon-Lindau, Switzerland Abstract: Soils, forming the top layer of the Earth's crust, are a mixture of mineral particles, organic matter, water, air and living organisms. Processes between these components perform vital functions within ecosystems. The soil forms an interface between the geosphere, the biosphere, the hydrosphere and the atmosphere, and is a largely non-renewable resource. Ugolini & Warkentin show the fruitful relationships which geology and soil science have established since the birth of soil science, and how these two disciplines together could contribute to solve future problems. The dynamic soil system delivers functions and services vital for human societies and the environment. Soil is the basis for food and biomass production, and plays a central role as a habitat for biota and as a gene pool. Moreover, it stores, filters, buffers and transforms a large variety of substances, including water, inorganic and organic compounds, and is a major sink and source for greenhouse gases. Soil provides raw materials for human use. It also serves as the basis for human activities (landscape and heritage) and for our technical and socio-economic infrastructure, delivering materials for their implementation and maintenance.

Soil formation Soils can be visualized using in a three-dimensional cross-section of the uppermost crust of the Earth, subdivided into different soil horizons (Figs 1 & 2). Soils are produced by physical, chemical and biological weathering processes, starting from solid unweathered rock or loose rock material such as gravel or sand. Chemical weathering is aided by solar energy from radiation and water, in which CO2 and other atmospheric gases are dissolved, forming acid solutions. Physical weathering processes are based on frost and thaw cycles, direct radiation, and t e m p e r a t u r e changes, as well as mechanical transport by water, ice or wind. T h r o u g h those mechanical c o m m i n u t i n g processes, the surface area of the rock material is increased, and space is provided for chemical processes, which are mainly reactions at particle surfaces. As soon as the first w e a t h e r i n g products, such as clay minerals and oxides are present, biota start to develop, initially forming a very sparse, but later a dense surface vegetation cover, which converts solar energy into biomass. These organic substances are cycled back to the ground after decay, where they are converted into soil organic matter by physical bioturbation and biological/biochemical mineralization and immobilization processes.

T h r o u g h these processes, minerals and organic matter are mixed, forming a soil horizon that contains a high amount of organic matter. The diversity and numbers of soil biota increase with time. These processes continue to increase the weathering depth of soil and form a new substrate, which is totally different from the rock p a r e n t m a t e r i a l ( D u c h a u f o u r 1997; Scheffer & Schachtschabe12002; Sumner 2002). The weathering rate and type are irreversible soil processes that leave markers in the properties of soil horizons. By studying the properties of these diagnostic horizons, soil scientists gain clues about the processes through which that specific soil was formed. No two soils are likely to be the same, because weathering depends upon the m a n y physical, chemical and biological factors interacting with the specific parent material. The soil management, superimposed on the i n h e r i t e d properties, creates an additional diversity. These processes determine the biological habitat, resulting in biodiversity.

The pore systems of soils Mechanical pressure exerted by roots, desiccation of areas a r o u n d roots by plant wateruptake, and general wetting and drying, create pores of different sizes. Figure 3 shows macropores, larger t h a n 50 gm in diameter; small

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soilsfor Human Societiesand the Environment. Geological Society, London, Special Publications, 266, 1-8. 0305-8719/06/$15 9 The Geological Society of London 2006.

2

W.E.H. BLUM E T A L .

Fig. 1. Deeply weathered soil (Ferralsol/Oxisol) in southern Brazil (photo by Blum).

macropores, 10-50 pm in diameter; medium pores, 10-0.2/am in diameter; and small pores of less than 0.2 pm in diameter. Pores of different sizes have very different functions in soil processes (Table 1).

Water and solutes are freely transferred through macropores, and water is stored and retained against gravity in the medium pores. In the small pores, the energy of water retention is too high to allow extraction by plant roots; the large and small macropores provide both the space for the growth of plant roots and a habitat for soil biota. The medium pores serve mainly as a habitat for microbial activities, due to the availability of water and air. Small pores are difficult to access by roots or by soil organisms. The water in these pores serves mainly as a medium for forces binding small soil particles, e.g. clays, together. This is the lower level of particle aggregation in a hierarchy that produces the visible soil structure or 'soil architecture'. Soil processes occur in the pore space, schematically depicted in Figure 4. The pores form a continuum of sizes, with different materials such as humic substances, clay minerals, carbonates or oxides constituting the pore walls. All these soil components bear electric charges on their surfaces. Reactions such as sorption/desorption and precipitation/ dissolution occur between the contents of the pores and the pore walls. Living organisms, fungi, bacteria, and other biota, up to earthworms and plant roots, actively participate in the soil processes, by absorbing nutrients, shedding dead tissue, and exuding low- or high-molecular-weight compounds (ions, organic acids, polysaccharides, amino acids, proteins, phenolic compounds, antibiotics) that can significantly change the properties of their local soil environment. The total inner surface of a soil can be estimated on the basis of its constituents. A soil volume of 1 ha (100 m x 100 m) and 20 cm of

Fig. 2. Soil as a three-dimensional cross-section of the uppermost crust of the Earth, subdivided into horizons (Schroeder & Blum 1992).

SOIL, HUMAN SOCIETY AND THE ENVIRONMENT

3

Fig. 3. Pore system of the B-horizon of an Amazon Oxisol (Brazil), observe the scale in pm (Blum 2002). Table 1. Classification of pore sizes in soils and their physical and biological characteristics (Blurn 2002) Pore size

Hydraulic conductivity water retention capacity (in pF = log cm water column)

Biological activities

Large macropores >50 pm diameter Small macropores 50-10 pm diameter Medium pores 10~).2 ~m diameter

Excellent: 0-1.8 pF Good: 1.8-2.5 pF Very limited: 2.5-4.2 pF Water is retained against gravity None: >4.2 pF Water is retained against plant root extraction forces

Large plant roots; soil macrobiota Small plant roots; fungal spores Bacteria; fungal hyphae

Small pores <0.2 pm diameter

depth, with a bulk density of 1.5 t m -3, forms 3000 t of soil material. Assuming 20% of clay minerals, with an average 2 0 0 m 2 surface area g-l, these 600 t of clay minerals have a total surface area of 120 000 km 2. If we further assume that this soil contains 3 % organic matter with 1000 m e surface g-l, the 90 t of organic substances have a total surface area of 90 000 km a. Therefore, the total inner surface of such a soil volume (without counting oxides and other types of minerals) is 210 000 km 2. For comparison, the total land surface of Italy (301 225 km2), is contained in a soil volume of 1 2 0 m • 1 2 0 m (c. 1 . 4 3 h a ) and 2 0 c m of depth.

Enzymes; no space for plant roots or soil organisms

E v e n assuming that only part of these inner surfaces are accessible through the pore system, the large inner surface of the soil becomes one of the very specific and unique characteristics of soil, distinguishing it from stone and rock material. A further specific characteristic of soils is the association between organic compounds, oxides and clay minerals, which provides highly reactive soil particles. Finally, all these surfaces p r o v i d e living conditions for an e n o r m o u s variety of soil organisms, which are actively participating in the processes occurring on these inner soil surfaces. These processes occur in the upper layers of the soil, between the soil surface and the water

4

W.E.H. BLUM E T A L .

Fig. 4. Schematic drawing of a pore system as a continuum of different pore sizes with different materials forming the pore walls (Blum 2002). table, which are known as the vadose zone. This zone is only partly saturated with water; the remainder of the pore space is filled with air. It is through this vadose zone that water moves down to the water table, transporting soluble and particulate materials which can pollute groundwater. The vadose zone is biologically active because the soil pores contain both air and water. Vadose zone studies of these processes are carried out by multidisciplinary groups of soil scientists, hydrologists, geologists and environmental engineers. The e n v i r o n m e n t a l quality standards and laws of different countries further these studies, in order to provide a high-quality e n v i r o n m e n t for humans. In their paper, Bergstr6m & Djodjic show the importance of soil porosity for p h o s p h a t e and pesticide transport through soils, and for pesticide degradation. Dosso et al. show the importance of soil distribution in a landscape to the movement of water and copper, relating these movements to vine mortality.

Organic matter in soils Organic matter is the most active component of soils, with continuous additions and continuing decomposition. The additions are raw plant material from shoots and roots, as well as animal material. The decomposition products are carbon dioxide, plant nutrient elements, and humus, which decomposes at a slow rate. This humus is chemically b o n d e d to clay-sized m i n e r a l particles forming the functioning matrix of the soil. The total soil organic matter content is considered one of the most important parameters when rating the quality

of agricultural soils. In their paper, Feller et al. p r o p o s e a historical review of the role of organic matter in agricultural soils and of the role of soils as a source and sink of greenhouse gases. Leifeld presents global data on the greenhouse gases, describes how their production is affected by soil properties and management, and shows the potential of mitigation practices to decrease their concentration in the atmosphere. The energy source for soil biota is organic material, produced by the conversion of solar energy into plant biomass (Blum 2001).

Soil biota The more organic material there is, and the more nutrients it contains, the more biota can live in a soil, if both air and water are present. A soil volume with a surface of 1 ha and a depth of 30 cm may contain about 10 t of bacteria and actinomycetes, 10 t of fungi, 4 t of earth worms and 1 t of other soil animals, such as nematodes and a large variety of insects, making a total of 25 t of biomass. In contrast, only 1.0-1.5 t of farm animals can be nurtured at the top of the soil, assuming two to three livestock units of 500 kg each per hectare.

Ecological soil functions The soil pore space, the chemical reactivity of the finely divided soil particles, and the presence of numerous living organisms allow soils to perform three main ecological functions. (1) Soils support plant growth, providing food and fodder to humans and animals.

SOIL, HUMAN SOCIETY AND THE ENVIRONMENT (2) They have the capacity to filter, buffer and transform materials between the atmosphere, the plant cover and the water table. They strongly influence the water cycle at the Earth's surface, as well as the gas exchange between terrestrial and atmospheric systems. These processes are shown in Figure 5, indicating the filtration of solid and liquid compounds in the pore space (a mechanical process), and the buffering capacity through absorption and precipitation of all kinds of organic or inorganic compounds (physico-chemical reactions on the inner surface of the soil). Last but not least, soils have a microbiological and biochemical capacity for transformation, through the alteration and decomposition of organic materials by mineralization and hydrolytic processes. As long as the soil can maintain these capacities, there is little danger that solid, liquid or gaseous, inorganic and organic compounds, e.g. pollutants, will reach the soil solution. However, if these functions can not be fulfilled any

(3)

5

more, higher quantities of heavy metals will reach the soil solution and be transferred to the plant, contaminating the food chain, or being leached beneath the profile and thus contaminating the groundwater. Luster et al. studied the effect of soil formation on heavy-metal transfer from forest soils to the groundwater, whereas Bafiuelos & Lin demonstrate how the filtering and buffering functions lost by soils as a result of intensive irrigation, with respect to the transfer of selenium and other salts of geological origin, can be partially recovered by appropriate agro-ecosystem management. The third important ecological function of the soil is as a biological habitat and a gene reserve for a large variety of organisms. The work conducted by a soil microbiologist, Selman Waksman and his team at Rutgers University in the first half of the twentieth century, resulted in the discovery of streptomycin, a very efficient antibiotic produced by Streptomyces griseus, a drug which is still widely used today. This work,

Fig. 5. Filtering, buffering and transformation processes in soils, protecting the food chain and the groundwater (Blum 2002).

W.E.H. BLUM ETAL.

6

which was honoured by the Nobel Prize in 1942, illustrates how important soil organisms can be in our daily life. Although it is difficult to use the concept of species when dealing with micro-organisms, the number of bacteria species present in one gram of soil can vary between 103 and 104, while the number of fungal species can vary between 10 and 102. However, the function of most of these micro-organisms remains unclear. This is partly due to the difficulties we have in studying them (95% of soil microorganisms cannot be cultured). As an example, Jansa et aL discuss the functional diversity of arbuscular mycorrhizal fungi (a group of soil fungi that live in symbiosis with the roots of most plants), and show how these are affected by agricultural management. Soil as cultural h e r i t a g e The chapter by Wells describes the functions of soil as an earth cover that protects and preserves the physical artifacts of our cultural heritage. Soil also has more general cultural functions. Soils are a part of us, as we are part of soils - part of the cultural landscape of our minds as well as of the physical world around us. Schama (1996) explores these ideas in his book Landscape and Memory. An attachment to 'home soils' or a 'sense of place' is a cultural attribute developed more strongly in certain peoples or in certain individuals. PreColumbian peoples of North America had a strong sense of belonging to the Earth. Their creation myths have people coming from the Earth, rather than from the sky as in many Western myths. A 'mother earth' concept was also a part of myths in pre-Greek societies. In 1998, representatives of world religions and cultures met with soil scientists in the Alsace region of France, to discuss the relationship between soil and world religions and cultures (Lahmar & Ribaut 2001). Artists often view soils as part of the cultural landscape. The French painter Jean du Buffet, half a century ago, in his phase of soil paintings, showed how he saw soils. He painted soils on a micro-scale; some of his works are markedly similar to the thin sections which soil scientists use to study the arrangement of soil components. He stated, 'The things we truly love, the basis of our being, we never look at. We are unconscious of them because they are so familiar.' More generally, soil was one of the important factors determining human migrations, affecting why people moved and why they settled, what lifestyles were possible for them, and how they

used their resources. This is the realm of soil geography. Soil: a t h r e a t e n e d r e s o u r c e Soil conservation is a major concern for the sustainable use of soils by society. Soils are primarily used for agricultural production, but also for building sites and the extraction of resources, such as gravel. Burghardt shows that an urban soil can retain some of its functions, and argues for a more reasonable use of the soil resources in urbanized areas. Degradation through desertification is an issue on a broad scale, with global effects. Loss of soil functions, due to compaction of surface soil by agriculture; or heavy-metal pollution in industrial areas, may represent a more local problem. Soil degradation has increased dramatically over the last few decades. Lal (1994) estimates that between 1945 and 1996, the percentage of soil surface that was moderately, strongly or very strongly degraded worldwide reached 10.5% (1 215 106 ha), with 16.7% in Europe (158 106 ha) and 14.3% in Africa (320 106 ha). While continuing soil degradation will not immediately endanger total worldwide food production (Penning de Vries 2001), the situation is worse in developing countries, where soil is often one of the only assets of resourcepoor farmers. In these areas, any increase in soil degradation will result in rising poverty and food insecurity. Soil degradation has many causes. Agricultural management systems not adapted to their environment lead to erosion, soil structure deterioration from compaction; salinization; loss of organic matter and of biodiversity; nutrient mining and the input of toxic minerals and organic compounds. Although excess accumulation of nutrients, such as nitrogen or phosphorus, may not threaten soil fertility, the excess phosphorus in the soil may be lost to surface water or wind erosion, and excess nitrogen may be leached into the groundwater, leading to water- and air-quality degradation. Bergstr~m & Djodjic discuss the leaching of nitrate and phosphate from agricultural soils. Our entire society contributes to soil degradation or destruction, through careless disposal of waste in the soil, or the emission of pollutants into the air, which enter the soil at some stage. Soil disturbance during the construction of buildings leads to increased erosion, with the soil moving to places where it interferes with drainage and causes turbidity in surface water. In Switzerland, for example, i m 2 of soil disappears each second under construction, 10 000 ha are considered to be polluted and 10

SOIL, HUMAN SOCIETY AND THE ENVIRONMENT to 40% of the agriculturally cropped surface is threatened by erosion (Stamp et al. 2002). Soil m a n a g e m e n t systems have changed, in order to take advantage of erosion on a geological scale. While removal of topsoil from a field decreases yields, this topsoil may be transported to flatter land, more suitable for agricultural crops. Nevertheless, erosion also has a lot of offsite effects which must be controlled, such as water pollution and mud accumulation on road and residential areas. Many methods have been tested to decrease soil loss by erosion or loss of soil functions by degradation. The application of these methods involves changing soil m a n a g e m e n t practices. The adoption of control practices has become a social or cultural issue, through education and government policy. Montanarella presents some policies which have been a d o p t e d for soil protection, as well as the E u r o p e a n strategy for soil protection. Landa shows the importance of education in soil science, and how an alliance with geology could contribute to improving education in the earth-sciences, and Hazelton explains why public awareness of the importance of soils should be increased. Finally, besides policies and education, it is also important to ensure that soil use is sustainable. Imicio et aL, and Menzi & Gerber propose appropriate tools for this purpose, lmieio et aL propose low-density geochemical exploration to assess changes in heavy-metal concentration in upper soil horizons at a country scale. Menzi & Gerber suggest the use of nutrient balances to assess, on a farm scale, w h e t h e r n u t r i e n t d e p l e t i o n is occurring or whether the nutrient load is excessive. By doing so, they increase the awareness among farmers about the use of non-renewable resources, such as phosphate. In the longer term, climate change might also affect soil d e v e l o p m e n t and soil properties. Increased temperatures and changes in water dynamics in the colder regions (in mountainous areas and at high latitudes), will lead to increased organic-matter mineralization, plant growth and weathering rates (Robert 1999). While changes in precipitation regimes, such as those observed in Switzerland, with m o r e rainfall in winter and less in summer, will also increase erosion and nutrient losses by leaching, runoff and gaseous losses. Altogether, we urgently need to limit soil degradation by implementing pro-soil national policies; by improving education about soil for all stakeholders involved with soil, and by improving public awareness of soil.

7

References BAlqUELOS, G.S. & LIN, Z.-Q. 2006. Reuse of agricultural drainage water in Central California: phytosustainability in soil with high levels of salinity and toxic trace elements. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN,B.P. (eds) Function of Soils for

Human

Societies

and

the

Environment.

Geological Society, London, Special Publications, 266, 79-88. BERGSTROM, L.E & DJODJIC, F. 2006. Soil as an important interface between agricultural activities and groundwater - leaching of nutrients and pesticides in the vadose zone. In: FROSSARD,E., BLUM, W.E.H. & WARKENTIN,B.P. (eds) Function of Soils for

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Geological Society, London, Special Publications, 266, 45-52. BLUM, W.E.H. 2001. The energy concept of soils. In: Functions of Soils in the Geosphere-Biosphere Systems, MAX Press, Moscow, 20-21.

BLUM,W.E.H. 2002. Soil pore space as communication channel between the geosphere, the atmosphere and the biosphere. 17th World Congress of Soil Science, 14-21 August 2002, Bangkok~Thailand.

CD-ROM. Transactions, 2014, Abstracts, Volume V, 1942, IUSS, Vienna, Austria. BURGHARDT,W. 2006. Soil sealing and soil properties related to sealing. In: FROSSARD,E., BLUM,W.E.H. & WARKENTIN, B.P. (eds) Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266,117-124. Dosso, M., PmLIPeON,O. & RUELLAN,A. 2006. Understanding of a soil system derived from a single bedrock for improved vineyard management in Southern France. In: FROSSARD,E., BLUM,W.E.H. & WARKENTIN, B.R (eds) Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 53-61. DUCHAUFOUR, PH. 1997. Abr~g~ de P~dologie-Sol, V~getation, Environnement (5e 6dition). Masson, Paris. FELLER, C., MANLAY,R.J., SWIFT,M.J. & BERNOUX,M. 2006. Functions, services and value of soil organic matter for human societies and the environment: a historical perspective. In: FROSSARD,E., BLUM, W.E.H. & WARKENTIN,B.P. (eds) Function of Soils for

Human

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Environment.

Geological Society, London, Special Publications, 266, 9-22. HAZELTON,RA. 2006. Australian examples of the role of soils in environmental problems. In: FROSSARD, E., BLUM, W.E.H. & WARKErCrIN, B.P. (eds) Function of Soils for Human Societies and the Environment. Geological Society, London, Special

Publications, 266, 141-147. INACIO, M., PEREmA,V. & PINTO,M. 2006. Assessing anthropogenic inputs to soils by comparing element contents and their spatial distribution in O and A horizons. In: FROSSARD,E., BLuM,W.E.H. WARKENTIN, B.E (eds) Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 159-170.

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JANSA, J., WIEMKEN, A. & FROSSARD, E. 2006. The effects of agricultural practices on arbuscular mycorrhizal fungi. In: FROSSARD,E., BLUM,W.E.H. ~i; WARKENTIN, B.R (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 89-115. LAHMAR, R. 8z RIBAUT, J.-P. 2001. Sols et SocidtOs Regards Pluriculturels. Editions - Diffusion. Charles L6opold Mayer, Paris. LAL, R. 1994. Sustainable land use systems and soil resilience. In: GREENLAND, D.J. ~; SZABOLCS, I. (eds) Soil Resilience and Sustainable Land Use. CABI, Wallingford, UK, 41-68. LANDA, E.R. 2006. From agricultural geology to hydropedology: forging links within the 21stcentury geoscience community. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN,B.E (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 133-140. LEIFELD, J. 2006. Soils as sources and sinks of greenhouse gases. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN, B.R (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 23-44. LUSTER, J., ZIMMERMANN,S.K., ZWICKY, C.N., LIENEMANN,P. d~zBLASER,P. 2006. Heavy metals in Swiss forest soils: modification of lithogenic and anthropogenic contents by pedogenetic processes and implications for ecological risk assessment. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN,B.E (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 63-78. MENZI, H. & GERBER, P. 2006. Nutrient balances for improving the use-efficiency of non-renewable resources: Experiences from Switzerland and South-East Asia. In: FROSSARD, E., BLUM, W.E.H. WARKENTIN, B.E (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 171-181.

MONTANARELLA, L. 2006. Policies for a sustainable use of the soil resource. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN,B.R (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 149-158. PENNINGDE VRIES,EW.T. 2001. Food security? We are losing ground fast! In: Crop Science: Progress and Prospects. NOSBERGER,J., GEIGER, H.H. & STRUIK, EC. (eds) CABI Publishing, Wallingford, UK, 1-14. ROBERT, M. 1999. Impact des changements climatiques sur l'6volution des sols et cons6quences sur le bilan hydrique. Comptes Rendus de l'Acad~mie d'Agriculture de France, 85, 35-44. SCHAMA,S. 1996. Landscape and Memory. Fontana Press, London. SCHEFFER, E & SCHACHTSCHABEL,P. 2002. Lehrbuch der Bodenkunde (15. Aufl.), Spektrum Akademischef Verlag GmbH, Heidelberg, Berlin. SCHROEDER,D. & BLUM,W.E.H. 1992. Bodenkunde in Stichworten (5. Aufl.), Gebrtider Borntrager Verlagsbuchhandlung, Stuttgart. STAMP, P., LIEDGENS, M., OBERSON, A., SCHMID, J.E., SOLDATI, A. & FROSSARD, E. 2002. Nachhaltiger Ackerbau in der Schweiz - eine Gratwanderung zwischen Ern~ihrungssicherung und Umweltschutz. Neujahrsblatt der Naturforschenden Gesellschaft in ZUrich, NGZ auf das Jahr 2003, Sttick 205. SUMNER, M.E. 2002. Handbook o f Soil Science, CRC Press, Boca Raton. UGOLIN1, E & WARKENTIN,B.E 2006. Perspectives on the relationship between soil science and geology. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN, B.E (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 183-190. WELLS, E.C. 2006. Cultural soilscapes. In: FROSSARD, E., BLUM, W.E.H. & WARKENTIN, B.P (eds) Function o f Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 125-132.

Functions, services and value of soil organic matter for human societies and the environment: a historical perspective C. F E L L E R 1, R. J. M A N L A Y 2'3 , M. J. S W I F T 3,4 • M. B E R N O U X

3

1Institute for Research and Development (IRD, ex-ORSTOM), UR179, BP 434, 101 Antananarivo, Madagascar (e-mail: [email protected]) 2Institute o f Forestry, Agricultural and Environmental Engineering (ENGREF), BP 44494, F-34093 Montpellier Cedex 5, France 3IRD, UR179, BP 64501, F-34394 Montpellier Cedex 5, France 4Tropical Soil Biology and Fertility Institute o f CIAT (TSBF-CIAT), PO Box 30677, Nairobi, Kenya Abstract: Soil organic matter (SOM) contributes significantly to the chemical, physical and biological ecosystem functions of soil. It influences on plant growth, thus contributing to agricultural production, and performs environmentally valuable services such as carbon sequestration, regulation of the water cycle and detoxification of pollutants. Identification of the functions and services provided by SOM has a long and tumultuous history of scientific discoveries and struggles against false assumptions. This work reports the major steps of this history, with emphasis on two services secured by SOM: (1) the role of SOM in plant production and its connection to soil fertility and thence to the sustainability of cropping and farming systems; and (2) the recognition and assessment of the contribution of SOM to climate-change regulation. Finally, the work explores how SOM, as a multifunctional resource, may be allocated an economic value as a way of promoting its conservation.

It remains difficult to decide w h e t h e r soil resources should be seen as renewable. However, archaeologists have identified examples of h u m a n societies that have been brought to the limit of sustainability by soil depletion, even resulting in some cases in the decline and fall of their civilization ( H y a m s 1976; Olson 1981). Today, the role of soil organic m a t t e r in controlling the capacity of soil resources to deliver agricultural and environmental services and sustain h u m a n societies at both local (e.g. fertility maintenance) and global (e.g. mitigation of atmospheric carbon emissions) scales is well established (Tiessen et al.

1994; Syers & Craswell 1995; Wolf & Snyder 2003). Soil organic matter (SOM) contributes to a range of functions that can be connected to goods and services at the ecosystem level (Table 1). Scientific recognition of the relationship between SOM, the sustainability of h u m a n activities and the state of the environment, and its implications for farming practices and landuse m a n a g e m e n t options, have fluctuated over time. This is in part due to continuing difficulty in adequately distinguishing and defining S O M functions and quantifying their values in terms of the environmental services that they provide, and in part to erroneous attributions of the

Table 1. Mainfunctions, 'goods' and "services"of soil organic matterat the ecosystem level Functions

Ecosystem goods and services

Nutrient reserves for plant and soil biota (decomposition, mineralization processes) Energy reserve for soil biota; micro- and macrohabitat building Aggregate formation and stabilization

Nutrient storage and availability; chemical fertility

Decomposition, sorption; elemental transformation Sink/source of greenhouse gases

Regulation of biological populations, including diseases and pests; biodiversity Regulation of water flow and storage, and regulation of soil and sediment movement Detoxification of chemical and biological pollutants (including water purification) Regulation of atmospheric composition and climate

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soilsfor Human Societiesand the Environment. Geological Society, London, Special Publications, 266, %22. 0305-8719/06/$15 9 The Geological Society of London 2006.

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C. FELLER ETAL.

agro-ecological functions of SOM that have been made at various times. The objective of this work is not to give a detailed review of the present knowledge of the functions and services provided by SOM to human societies and the environment, but to present a brief history of the construction of concepts and tools that lead to the perception of the interconnection between SOM, soil fertility and ecosystem sustainability. We also briefly address ways in which society has begun to attribute both ecological and economic value to SOM. The focus of this paper is on two broad functions of SOM: the first is productive and relates to ecosystem fertility and farming sustainability; the second is environmental and deals with the control of the greenhouse effect, climate change and soil carbon (C) sequestration.

Agronomic functions of soil organic matter and their connection to cropping and farming sustainability: a complex and tumultuous history F r o m antiquity to the eighteenth century Concepts of plant nutrition varied greatly during antiquity (Browne 1944). Among the Greek philosophers in the period from 640 to 435 Bc, sources of plant material were: water for Thales of Miletus, c. 625 to c. 546 BE; air for Anaxinemes (c. 585 to c. 525 Bc); fire for Heraclitus (c. 535 to c. 475 Bc); and for Empedocles all four of the basic elements (earth, water, air and fire). Aristotle (384-322 BE) used these precedent theories to establish the general 'Four Elements Theory', whereby the union of the four elements in the soil enables minute particles to sustain plant nutrition. As a consequence, all plant material was thought to originate from soil. According to Aristotle soil fertility - like anything in nature - would also be driven by the 'four qualities' - warmth, coldness, humidity and dryness. Plants were also assumed to feed on organic material of related nature: for instance, olive stones were fed to olive trees, and vine shoots to vines, to sustain plant production. Such beliefs and Aristotle's theory were still influential during the Middle Ages. Palissy (1510-1589), whose theory of 'salts' was published in 1580 (in Palissy 1880) is generally considered by historians of soil science to be a major forerunner of the mineral theory later established by Liebig (1803-1873) (Liebig 1840); however, since Palissy's definition of'salt' is not strictly mineral, this opinion is question-

able (Feller et al. 2001, 2003). In the seventeenth century, Van Helmont (1577-1644), among others, took up Palissy's ideas about the role of soil as a simple source of water and mineral nutrients for the plant (Boulaine 1989). In 1699, Woodward (1665-1728) showed that something from the earth, other than water, was important to plant growth. During the eighteenth century, 'humus' was often understood to be synonymous with soil, and many theories and agricultural practices (e.g. soil tillage; Tull 1733) about plant nutrition were based on the belief that plants relied directly on humus for their own carbon supply. S O M as a possible source for plant carbon nutrition (1800-1900). By the end of the eighteenth century, Hassenfratz (1755-1827) still asserted, but without referring to experimental facts, that a fraction of humus in the form of soluble carbon was directly assimilated by plants (carbon heterotrophy) and was the almost unique source for plant carbon nutrition (Feller et al., 2001). But during the same period several authors - all cited in Bourde (1967), for example, Priestley (1733-1764) (Priestley 1777), Ingen-Housz (1730-1799) (Ingen-Housz 1780), Senebier (1742-1809) (Senebier 1782) and de Saussure (1740-1799) (de Saussure 1804) partially refutated these theories by demonstrating experimentally both the gaseous origin of carbon and the role of light during photosynthesis. Nonetheless, de Saussure still considered that a small part of plant material could derive from soluble humus. Contradictory debates arose on the subject, but many agricultural scientists shared an intermediary point of view and assigned functions in plant nutrition to both SOM and air. This was in particular the case for the famous German agronomist Albrecht Daniel Tha~r (1752-1828), known for the 'theory of humus' developed in his seminal book Principles of Rational Agriculture (Tha~r 1809). Tha~r's Principles. Tha~r's Principles contains some unverified theoretical concepts of plant nutrition that served as a basis for the first rational and systematic approach to fertilization within the context of sustainable cropping practices (de Wit 1974; Feller et al. 2003). Tha~r's theory of humus (1809) integrated an analysis of the management of soil fertility and the concept of sustainability that deserves particular attention. Tha~r's book was released in the midst of a period of controversy about whether the soil or the atmosphere was the actual source of carbon used by plants. Tha~r did not deny that

SOM FUNCTIONS: A HISTORICAL PERSPECTIVE atmospheric CO2 could be a carbon source for the plant, but since this compartment seemed unlimited, he considered soil humus and its management as the main limiting factor of plant carbon nutrition. According to Tha~r: (1) the majority of plant dry-matter derives from the 'soil nutritive juices' contained in the fraction of soil humus that is soluble in hot water; and (2) plant demand for 'juices' is selective and varies with the species cultivated. Therefore management of soil fertility must be based on the management of the soil humic balance as well as on that of the crop succession. Although incorrect, these theoretical assertions encompassed the whole soil-plant system and were used to support the first quantified, complex but complete system of analysis for the diagnosis and prediction of fertility (Feller et al. 2003). This is certainly the first example of real concern about farming sustainability, and what is more, it is based on organic practices. Tha~r's analysis also included an economic approach (see p. 18). Conceptually, Tha~r's approach to fertility encompassed the plant-soil system as well as cropping patterns and rotations. In so doing he introduced and discussed modern agricultural issues, such as the identification of soil-quality indicators, systematic analysis and the agroeconomic sustainability of farming systems. His work seriously influenced the thinking of his peers during the first half of the nineteenth century. If Tha~r had focused on mineral rather than organic budgets, he would probably have been regarded as the founder of Western scientific agriculture. From the 'mineral theory' to the new concept o f bioavailability and the indirect role of S O M in mineral nutrition. Although Liebig took many of his ideas from the work of Sprengel (1787-1858) (Sprengel 1838, in van der Ploeg et al. 1999), his authoritative text Die organische Chemic in ihrer Anwendung auf Agrikultur und Physiologic (1840) is often considered as the first demonstration, based on scientific experiments, of the origin of plant dry-matter from mineral compounds. Carbon was described as being derived from carbon dioxide; hydrogen from water; and other nutrients from soluble salts in soil and water. Since Liebig's synthesis accounted rather satisfactorily for the fertilizing effect of mineral inputs, it provided the basis for modern agricultural sciences. Liebig promoted the use of mineral fertilizers to compensate for soil mineral depletion, and his work paved the way for recommendations for the massive use of chemical fertilizers in cropping systems, and the abandonment of organic or organomineral

11

fertilization. Nonetheless, Liebig, as 'one of the last "complete" men among the Great Europeans' (Hyams 1976), was himself an advocate of mixed fertilization. In the sixth volume of his exhaustive Cours d'Agriculture, Gasparin (1783-1862) (Gasparin 1860) took a similarly moderate position: he included organic and chemical fertilizers in the same category, but was already emphasizing the low economic cost of organic fertilizers produced on the farm. In fact, the limited references to chemical fertilizers in Gasparin's textbook are certainly due to the limited production and use of inorganic fertilizers before the 1880s (Boulaine 1989; Smil 1999). Finally, direct but very limited absorption of some organic compounds by plant roots was to be demonstrated in the early twentieth century (Acton 1899; Maz6 1899, 1904, 1911; Laurent 1904; Cailletet 1911; Knudson 1916, all cited by Waksman 1938). Today, the importance of humic substances for enhancing absorption of mineral, organic or organomineral phytohormones is well recognized (Chen, 1996; Chen et al. 2004).

From agronomy to ecology The mineralist approach to the management of soil fertility reached its apogee in the thirty-year period following the Second World War, with the establishment of high-input, subsidized agriculture in Europe and North America, and the huge fertilizer-driven increases in the production of Green Revolution cultivars of rice, wheat and maize in South and South East Asia and parts of Latin America (Pinstrup-Anderson & Hazell 1985). As a result, interest within the conventional agricultural research community in managing SOM as a source of fertility declined even further. In contrast, the same time period, however, saw the rise of a number of other scientific initiatives that resulted, in the last two decades of the century, in renewed and scientifically based interest in managing SOM under the rubric of 'sustainable agriculture'. These approaches can be seen as being derived substantially from two convergent sources: firstly, developments in ecosystem science, including improved scientific capacity for the study of SOM and associated aspects of nutrient cycling; and, secondly, concerns about environmental degradation and the loss of ecosystem services, and the expression of these concerns in the rise of the organic farming movement. The 'healthy' function of S O M and the organic farming movement. The concerns about the impacts of high-input agriculture expressed by

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C. FELLER ETAL.

the formal scientific sector were probably less influential in triggering of new interests in the management of SOM as a component of soil fertility than those derived from the rise of 'alternative' farming practices under the rubric of 'organic agriculture'. Predating the conceptual debate about the substitutability of organic amendments by chemical fertilizers (Smil 1999; Rigby & Caceres 2001), societal criticisms concerning the sustainability of intensive farming arose as early as the 1930s, leading to the formulation by Balfour (1944) of a hypothesis of the link between the decline in soil fertility, the quality of the human diet and the state of human health. Concerns about the loss of biological function and decline in fertility in cropped soils that are managed without returning organic matter to the soil, date back to ancient times, but the lack of sound principles of soil ecology diminished their impact on scientific thinking. Steiner's lectures (1924) provided the foundation for biodynamic agriculture. The scientific basis of Steiner's lectures and of the publications of his followers (e.g. Pfeiffer 1938) was poor, as they referred to both holistic and cosmogonic concepts (i.e. interrelations between the stars, soil and geochemical elements, plants, animals and humans) as the basis for a new kind of agriculture that excluded the use of any chemical input. The most influential - and at least rational publications on modern organic farming are those from Howard, Balfour and Rodale (Howard 1940; Balfour 1944; Rodale 1945; Howard 1952); for a more detailed review of the history of organic farming, see Scofield (1986) and Lotter (2003). The main objective that they shared was to improve soil, plant, animal and human health by the biological management of soil fertility. Two fundamental aspects of the organic farming philosophy put SOM at the heart of cropping sustainability: the holistic paradigm and the Law of Return. -

The holistic paradigm. In The Living Soil, Balfour (1944) presented the quintessence of the philosophy of organic farming. Her leading hypothesis was that the reason for the obvious - according to her criteria - decline in the health of the human race was the decrease in plant health, itself a consequence of the decline in the health of the soil. The philosophy of organic farming is fundamentally holistic and perceives all life,all creation as being inextricablyinterrelated, such that something done or not done to one member,part or facet will have an effect on everythingelse. (Merrill, 1983) This is best illustrated by the biotic pyramid of Albrecht (1975, cited in Merrill 1983).

This pyramid is made of several layers, with the soil as the basement and humans at the top of the pyramid. Within this scheme, any degradation of soil quality can threaten civilization and even humankind itself - hence the need for careful soil husbandry.

The Law of Return. Another principle of organic farming is the Law of Return. It stems from the concept of the 'Living substances cycle', which originated in antiquity and reappeared in treatises on agriculture in the sixteenth and seventeenth centuries. The breaking of this principle is one of the factors suggested in several historical records where collapses of civilizations have been attributed to failures of their agriculture, and it still underpins present critical issues in urban waste recycling (Magid et al. 2001). According to this principle, life can be maintained only if living beings, or at least the residues of their activity and body decomposition, are cycled at each step of the biotic pyramid. A crucial process is thus the establishment of organic flows to the soil to maintain its fertility. Since this return is SOMmediated, Balfour (1944), and above all Rusch (1972), adopted a sceptical position towards what they termed Liebig's 'rather naive theory', and developed a partly rigorous (Balfour and Howard), partly ideological (Rusch) analysis of the agro-ecological role of SOM. Howard's opinion, as expressed in his The Soil and Health (1952) matches Balfour's holism. His more precise causal interpretation of the relation between soil, plant, animal and human health is anchored in the idea of the cycling of proteins among living beings, and their quality of protein. Even if his opinions were partly ideological, Howard (1940, 1952) did publish rigorous and famous technical handbooks for the production of compost, which he termed 'manufactured humus'. For the past 10 years, in scientific community, there has been a renewed interest in holistic approaches to soil management, as evidenced by the proliferation of scientific meetings, research programmes (and, of course, the consequent publications) on the topics of 'soil health', 'soil quality' indicators and 'sustainable soil management'. SOM (total or in compartments) and soil biota are invariably key parameters of these initiatives (Lavelle & Spain 2001; Doran 2002). Towards ecological agriculture. Setting aside the ideological elements, there is clearly a degree of convergence between some of the holistic principles of organic agriculture and those of ecosystem science. This convergence

SOM FUNCTIONS: A HISTORICAL PERSPECTIVE has been embraced in the developing concepts of 'sustainable' or 'ecological' agriculture. The term 'sustainable development' came to global attention with the publication of the report of the World Commission on Environment and Development (WCED 1987), where it was defined as 'development that meets the need o f present generations without compromising the ability o f future generations to meet their own needs'. This obvious congruence with the

environmental concerns about the impact of intensive high-input agriculture, coupled with the failure to achieve persistent and consistent results in many parts of the world, notably Africa, stimulated a substantial effort to find sustainable means of agricultural production (Conway & Barbier 1990). This focus naturally fell upon the use of renewable natural resources; in the case of soil-fertility management this resulted in fresh attention being given to the management of organic matter and biological processes (Scholes et al. 1994). One of the key features of sustainable soil practice is the return to managing soil fertility through the combination of organic matter (crop residue, compost or manure) and mineral nutrient inputs (Pieri 1992). This rediscovery of the benefits of the ancient concept of integrated nutrient management became the mainstay of soil-fertility management at the turn of the twentieth century (Mokwunye & H a m m o n d 1992; Palm et al. 1997), and maintenance and/or improvement of SOM status is central to the philosophy. Management of organic inputs has been able to draw on the knowledge gained from ecological studies of decomposition processes, nutrient cycles and nutrient balances (Myers et al. 1994; Cadisch & Giller 1997; Palm et al. 2001). Similarly, the management of SOM has been enhanced by the application of the knowledge embedded in different simulation models, with a particular focus on manipulating the labile pools, while seeking to maintain or build up the stable SOM fractions (Vanlauwe et al. 1994). The major scientific challenge remains how to extend the ecological principles beyond the manipulation of the plant component (with the consequent indirect influence on the soil biota, decomposition processes and SOM dynamics) to more direct manipulation of the soil biota (Swift, 1998). Successes obtained with the dinitrogen (N2) fixing bacteria (Giller 2001) have, however, still to be matched in other groups of soil biota. Since Odum's strategy of ecosystem development (1969), general conceptual advances have stressed the aptitude of ecosystems - based on their internal organization - to escape the constraints of the abiotic environment by

13

building biotic buffers or even modifying abiotic factors (Perry et al. 1989). In terrestrial ecology, SOM has been recognized as a pivotal factor buffering climate and soil constraints and establishing close links between plants and soils from the perspective of ecosystem rehabilitation (Perry et al. 1989; Aronson & Le Floc'h 1996). The contradiction that appeared subsequently between the role of SOM as a source of nutrients requiring its decomposition and its structural role in improving soil physical and chemical properties and stabilizing the plant-soil interactions - has been underlined by de Ridder & van Keulen (1990). In fact, recent applications of the thermodynamic theories of open systems kept far from their equilibrium, such as a soil ecosystem, may have at least partially solved this contradiction (Odum et al. 2000). They suggest that soil structure and organization can be largely controlled by soil biota at the cost of energy - mostly carbonmediated - dissipation, thus implying SOM recycling (Perry et al. 1989). The treatment of SOM as a dynamic, biologically regulated pool of carbon and nutrients in science-based sustainable agriculture converges with Balfour's definition of soil fertility in organic agriculture as 'the capacity o f soil to receive, store and transmit energy' (Balfour 1976, in Merrill 1983). Increased promotion or adoption worldwide of precision agriculture, agroforestry (Steppler & Nair 1987; Ewel 1999), and of composting, mulching, direct sowing, reduced tillage and cover cropping (CIRAD 1999; Erenstein 2003) testifies to the renewed recognition of the scientific value of integrated SOM management for the development of sustainable cropping patterns. Similarly the incorporation of ecological concepts into modern agriculture, although slow, represents a return to principles that were generally widespread before the mineralist era and which were derived empirically from observation of nature, many of which have been retained in traditional indigenous knowledge in many parts of the world (Altieri 2002; Jackson 2002; Tilman et al. 2002). This progress has been documented recently in a book (McNeely & Scherr 2002) which celebrates the achievements of what they term 'Ecoagriculture'.

A n environmental function of SOM: control of the greenhouse effect and carbon sequestration Beyond its role in nutrient cycles, SOM has also come to be valued for its influence on a wide range of so-called ecosystem services. These

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C. FELLER ETAL.

include structure-related features such as water storage and availability and resistance to soil erosion, as well as the energy contributed to supporting the biomass and diversity of the soil biota and their actions as biological control agents and regulators of soil and water pollution (Swift et al. 2004). Nonetheless, these benefits depend on balance. High levels of SOM can also create negative effects such as excessive nitrate production, and application of large amounts of organic matter can result in the build-up of pests (Chikowo et al. 2004). In recent years, increasing attention has been given to the p o t e n t i a l of SOM for carbon sequestration. Concerns about increasing a t m o s p h e r i c g r e e n h o u s e gas c o n c e n t r a t i o n s (GHG, mainly CO2, CH4 and N20 ) and global warming and climate change, have raised questions about the role of soils as sources or sinks of carbon (Houghton 2003). The terms 'sequestration' and 'C sequestration' were proposed to define the aptitude of terrestrial ecosystems to act as sinks for these GHG. Key aspects of the history of the appearance and significance of the term 'carbon sequestration' and of methods to estimate it in soil at different scales in time and space as well as those procedures used to measure CO2 fluxes in the soil-plant system, are presented briefly below.

A p p e a r a n c e o f the terms "carbon sequestration" a n d 'soil c a r b o n sequestration' A search of the ISI-Web of Science database for the 1945-2005 time period suggests that the first occurrence of the linked terms 'soil' and 'carbon' and 'sequestration' dates from only 1991, but the n u m b e r of references has increased rapidly over the past 15 years (Bernoux et al. 2006) (Table 2). The concept of soil C sequestration is thus relatively new. Most definitions of C sequestration, whether soil specific or not, refer simply to CO2 removal from the atmosphere and storage in an organic form in the soil or plant compartments. But methane ( c n 4 ) and nitrous oxide (N20) are also involved in exchanges between the s o i l - p l a n t system and the atmosphere. The Kyoto Protocol includes an inventory of all sources and sinks of these gases. Net G H G emission calculations of the signatories of the U n i t e d Nations F r a m e w o r k C o n v e n t i o n on Climate Change (UNFCCC) are expressed for all the gases in equivalents of CO2 after application of a conversion factor, the global warming potential (GWP) of each gas. Current conventions yield a 100-year-GWP value of 23 for CH 4 and 296 for N20. A recent review by Six

Table 2. Number of references indexed in the ISI-Web of Science (1945-2005) returned by combining the queries (1) 'soil' AND 'carbon ', and (2) 'soil' AND 'carbon' AND "sequestration', in the "topics' and 'title' (between brackets, Query 2 only)fields. Updated from Bernoux et al. (2006) Period

Number of references returned by query 1 = 'soil' AND 'carbon'

1945-1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 Total (1940-2005)

Query 2:Query I ratio (%0)

2 = 'soil' AND 'carbon' AND 'sequestration'

719 643 694 816 908 985 1220 1398 1520 1568 1618 1727 1850 2136 2142 2611

0 1" 5 14 7 21 24 36 47 42 78 107 153 217 174 255

(2) (3) (3) (8) (14) (15) (33) (17) (26)

0 1.6 7.2 17.2 7.7 21.3 19.7 25.7 30.9 26.8 48.2 67.5 82.7 101.6 81.2 97.6

21555

1181

(126)

54.8

* Thornley et al. (1991); * Dewar and Cannell (1992).

(1') (1) (1)

SOM FUNCTIONS: A HISTORICAL PERSPECTIVE

et al. (2002) illustrates the importance of those considerations. The authors found that: (1) in both tropical and temperate soils, C levels increased in no-tillage (NT) systems as compared to those under conventional tillage (CT); but (2), in temperate soils average N20 emissions increased substantially under NT as compared to CT; and (3) the increase in N20 emissions (when expressed on a C-CO2 equivalent basis) lead to a negative total GWP, even if positive C storage was observed in the soil. Even N fertilization in an organic form is an N20 emission hazard (Flessa et al. 2002; Giller et al. 2002; Millar et al. 2004). From these considerations, it appears clear that a concept of 'soil carbon sequestration' must not be restricted to a mere quantification of C storage or CO2 balance. All G H G fluxes must be computed at the plot level in C-CO2 or CO2 equivalent, incorporating as many emission sources and sinks as possible across the entire soil-plant system. Therefore, Bernoux et al. (2006) propose a new definition for C sequestration, applied to the soil or soil-plant-system: 'Soil carbon sequestration' or 'Soil-plant carbon sequestration', for a specific agro-eeosystem, in comparison with a reference one, should be considered as the result, for a given period of time and portion of space, of the net balance of all GHGs, expressed in C-CO2 equivalent or CO2 equivalent, computing all emissions sources at the soil-plant-atmosphere interface. The confusion (as is often the case) between the notion of 'SOC (SOC, soil organic carbon) storage' (C stored in the soil whatever its origin) and 'soil C sequestration' (GHGs, expressed in equivalent C-CO2 stored in the soil and originating from the atmosphere) can thus be avoided.

The first m e a s u r e m e n t s o f soil C 0 2 concentration and fluxes The discovery of carbon dioxide is attributed to Joseph Black (1728-1799), who published his thesis in 1754. Black named it 'fixed air', for it was emitted during heating and decomposition of calcium or magnesium carbonates.

The first in situ and in vitro measurements of soil COe concentrations. The first in situ measurements of soil CO2 concentrations were made by Boussingault & Levy (1852, 1853) at depths ranging from 40 to 240 cm. Using sophisticated equipment to avoid contamination of soil CO2 by atmospheric CO2, these authors showed that CO2 concentrations in soils without farmyard manure (FYM) application were 22 to 23 times

15

higher than in the atmosphere, and that applying FYM could increase this concentration by 245-fold. Wollny, in his book on SOM decomposition (1902) inventoried the effect of different environmental factors on soil CO2 concentrations. He demonstrated the positive effects of soil temperature and humidity on CO2 emissions, and showed that any agricultural or environmental factor that influences soil temperature and humidity has an effect on CO2 fluxes. According to Waksman (1938) the first measurements of soil CO2 emissions in laboratory and controlled conditions were done by Ingen-Housz (1794-1796), who demonstrated the effect of organic inputs and the importance of oxygenation, temperature and humidity. As early as 1855, Corenwinder (1855; 1856) used equipment very similar to today's respirometry apparatus. The technique was largely used by D e h r r a i n & Demoussy (1896a, b), Wollny (1897) and Stoklasa and Ernest (1905). The latter proposed the measurement of CO2 evolution as an indicator of the availability of SOM. In 1920, L e m m e r m a n n & Wiessmann (1920) proposed a mathematical model for CO2 production under laboratory and controlled conditions as an exponential function of time and of the initial concentration in soil CO2.

Measurements of COe fluxes at the soil-plantatmosphere level Lundeg&rdh's studies at the plot scale. The main forerunner to modern measurements at this scale was the Danish ecophysiologist Henrik Lundeghrdh (1888-1969). His summarized biography was recently published by L a r k u m (2003). Between 1924 and 1930, Lundeghrdh published considerable data on CO2 fluxes at the soil-plant interface in two books (1924,1930) in German and a large paper (1927) in English. In these three publications, Lundeg~rdh reports an impressive amount of quantitative data on in situ CO2 fluxes between atmospheric, plant and soil components. Data were collected using instruments for the sampling of the soil atmosphere (equivalent to our present static chamber) or continuous monitoring of CO2 fluxes at the plant or atmosphere level. In his 1927 publication, he even described field equipment and experimental designs completely equivalent to the present-day 'free air CO2 experiment (FACE)', which are among the most sophisticated experiments for the study of CO2 fluxes in the field. F A C E experimenters, however, seldom refer to Lundeg~rdh's remarkable pioneering work. Lundegg~rdh's findings are also notably close to present data concerning

16

C. FELLER ETAL.

C O 2 fluxes for the soil-plant system, and can be summarized in the following points:

(1) an increase of 0.01-0.32% in the atmospheric CO2 concentration can change drastically the plant C assimilation rate, which is dependent on illumination and temperature; (2) large monthly and interannual variations in air and soil CO2 concentrations can be observed, for a given location and soil; (3) soil CO2 emissions vary depending on soil type, and range from 1.25 to 23.4 kg CO2 ha q hq; (4) soil CO2 production decreases strongly with depth. His work on the effect of soil management on CO2 fluxes showed that: (1) the relationship between CO2 emissions and SOM content is not completely direct; (2) organic inputs lead to large and persistent (up to a year long) increases in CO2 emissions; (3) mineral fertilization significantly increases soil CO2 emissions, due to a priming effect on SOM mineralization, and this increase contributes indirectly to better plant C nutrition, in addition to the positive fertilization effect on plant productivity. All these questions are currently topical with regard to quantitative knowledge about the effect of land use and land-use change on the global CO2 balance.

From the square metre scale to the hectare scale. The eddy covariance (or eddy correlation) technique is commonly used for the estimation of CO2 fluxes in continuous natural or cultivated agro-ecosystems at the plot (>1 ha) scale. A recent and exhaustive historical review of the results obtained by this new approach was published by Baldocchi (2003) and need not be repeated here. This technique can also be applied at the cultivated plot scale (100 me), and was used by Reicosky et al. (1997) for the study of the effect of soil tillage on soil CO2 fluxes.

A s s e s s m e n t o f soil C stocks and dynamics at different scales The dynamics of the soil compartment are heavily implicated in the impacts of land use, land-use change and forestry (LULUCF) on the atmospheric G H G budget. The soil may act as a sink (by SOC accretion and CH4 absorption)

or a source for C-CO2 in the medium term (0-50 years). There has thus been a growing need: (1) to quantify present SOC stocks at different spatial scales, from the plot-scale to continent-wide); and (2) to predict its dynamics in response to LULUCF, using simple and robust mathematical models.

Evaluation of SOC stocks. The content of OC, OM or humus in soil was determined as early as the beginning of the nineteenth century, as ThaEr's Humus Theory (1809) shows. The emergence of the soil C sequestration issue has resulted in a large effort to acquire databases of SOC stocks at scales from the plot to the globe. Table 3 summarizes the historical data on the evaluation of SOC stocks at the global scale. The first publication was probably that of a geologist, Lyon, as early as 1915 (Rubey 1951). His study was based on nine soil profiles only (for the whole planet!) but his estimate (710 Gt C for the 0-100 cm layer) was reasonably close to Batjes's modern (1996) result (based on 4353 soil profiles) of 1500 Gt C for the same depth. Similarly, the global estimates of Waksman (1938) of 400 Gt C, for the upper 30 cm of the soils are also close to that of Batjes' (1996) 684-724 Gt C for the same soil layer. The need to model SOC dynamics. The first qualitative approach for modelling SOM dynamics was by H. B. de Saussure in his famous Voyages dans les Alpes, w (1780-1796). Extracts were republished by his son, N. T. de Saussure, in his book Recherches Chimiques sur la VOg~tation (1804). They were based on observations made by his father during a journey traversing the plain between Turin and Milan - a region cultivated since antiquity. These observations and reflections can be summarized as follows: (1) since no continuous accumulation of SOM occurs, even with continuous organic inputs, some of these inputs must be destroyed; (2) the amount which is destroyed must, to a certain extent, be proportional to the absolute existing amount; (3) limits to SOM accretion must vary depending on climate, the nature of the bedrock, vegetation, the cropping system and the fertility of the land; (4) even if all the conditions are favourable to SOM accumulation, there must be a maximum for the thickness of the humus layer, beyond which destructive causes equal productive ones.

SOM FUNCTIONS: A HISTORICAL PERSPECTIVE

17

Table 3. Publications including an evaluation o f S O C stocks at the global level

Authors

Year

Number of profiles

Results for soil profile (Gt C) Soil layer 0-100 cm Litter included: yes/no

Waksman Rubey* Bohn Bohn Post et aL Eswaran et al.

1938 1951 1976 1982 1982 1993

Sombroek et al. Eswaran et al.

1993 1995

n.d. 9 c. 200 187 2696 1000 (world) + 15 000 (USA) 400 1000?

Baqes

1996

4353

Jobb~gy&Jackson

2000

2721

3000 2220 1395 1576

(n.d.) (y) (n) (n)

1220 1576

(n) (n)

1462-1548

(n)

1502

(n)

Other

Depth in cm

400 709

(0-30) (0-?)

652 927 684-724 2376-2456 1993 2344

(0-25) (0-50) (0-30) (0-200) (0-200) (0-300)

n.d.: not determined. * Rubey (1951) used SOC contents for 9 main soil types published by Twenhofel (1926) based on values reported by Lyon et al. (1915). H. B. de Saussure's conclusions, completely ignored by historians of soil science, thus convey the basic equilibrium concepts utilized by modern mathematical models of SOM dynamics. Yet it was not until 137 years later that a mathematical formulation of SOM (C or N) dynamics for the decrease in organic N content with cultivation was to be expressed by Jenny (1941), followed by the more general model of SOM dynamics proposed by H6nin & Dupuis (1945). Many models have now been published and are being used (Smith et al. 1997). The most famous ones are probably the RothC model of Jenkinson & Rayner (1977) and the Century model of Parton et al. (1987). These models were designed to run at the plot level. Coupling them with geographical information systems (GIS) in order to simulate changes in SOC storage at scales from the plot to global, is the ongoing challenge faced by investigators of global change. Towards an e c o n o m i c value for S O M ?

The range of benefits in terms of 'goods' and 'services' provided by the SOM functions (Table 1), indicates that this resource is of great value to humans. This value is, however, only poorly comprehended by society in general one of the major reasons being that it is not generally expressed in cash terms. In recent years, economists have intensified attempts to

provide economic values for natural resources. In most cases, these values must be attributed indirectly, on the basis of the 'support service' provided for marketable products, rather than a direct (i.e. cash) value. Nonetheless, soil organic carbon has finally come to have a recognizable direct value. Because of the regulations requiring public and commercial organizations to reduce their contributions to global climate change, mechanisms have been sought for sequestering carbon in vegetation or soil as described on p. 14. Institutions which are net producers of carbon-based greenhouse gases have therefore entered into trading agreements with institutions that are able to sequester carbon. Carbon traded in this way is currently offered on the world markets for between 5 and 20 US$ t -1. The viability of this market for sequestrable carbon remains to be proven, however. The economic implications of locking farmers and other land users into long-term storage of carbon at relatively high levels, which may exclude a range of potential land-management practices, are far from apparent. Trading in carbon over the long term is also dependent on the acceptance by all parties of methods for measuring and monitoring carbon change. Such standards are not yet in place, and indications are that the costs may render the trading uneconomic (Smith 2004). Apart from the vital issue of sensitivity for tracking presumably low and slow variations in soil C density, one

18

C. FELLER ETAL.

requirement for establish monitoring systems for soil carbon, whether for its role in G H G mitigation or for other environmental services, is to have acceptable methods for establishing thresholds or boundaries as minimum - and perhaps also maximum - values for SOM content, values below which agricultural and ecosystems services cannot be achieved. The above review of methods for measuring soil carbon indicates that this still remains a major challenge. One recent attempt to achieve this was made by Feller (1995) for annual cropping systems on low-activity clay (LAC) soil in the sub-Saharan West Africa (sahelo-sudanian region). This resulted in a simple equation for calculation of threshold values of SOM content, expressed on a SOC basis of the 0-10 cm layer in relation to soil texture (percent clay + silt content, c + s, in g 100 g-1 soil), i.e.: SOC (g kg -1 soil) = 0.32 (c + s %) + 0.87 (r = 0.97, n = 15) Below this threshold soil physical, chemical and biological properties are very low and plant yield severely inhibited. The value of SOM goes beyond its significance in mitigating climate change. The other benefits do not, however, carry such a recognizable market value. Most would probably agree that the greatest benefit of SOM derives from its contribution to soil fertility and thence to the production of food and fibre. These contributions are indirect, in the sense that SOM is not itself a product, but has properties (as a source of nutrients, through improving the water storage capacity of soil, etc.) that contribute to enhanced crop production. Economic methods exist for making estimates of such indirect values (e.g. see Perrings 1995). Such estimates remain to be made for SOM, but the principles seem to be well established. The basis lies in calculating credible contributions of SOM to services that can be given a value, e.g. crop production. This is not a trivial task, but it can be approached through crop production models. These calculations are very common at the farming-system level, but have not generally been used for the purpose of estimating the SOM value. These approaches are, however, not new. In the nineteenth century, the quantification of the economic value of SOM for fertility was a large component of Tha6r's system, with the humus theory being used as a tool to predict not only the productivity but also the cost/benefit of SOM management. For instance, rye productivity was used as an indicator for the biological and economic evaluation of different management models proposed by Tha6r (Feller et al. 2003). This analysis included all the costs

(labour, space, care of animals) of organic maintenance of fertility, based on fallowing and manure application. Similar modelling approaches could be used to estimate the value contributed by SOM to other ecosystem services, such as its contribution to erosion control. A number of well-known biophysical erosion models include an economic output (e.g. the E P I C model, Williams 1989; Jones et al. 1991). A major challenge in this work is to quantify the SOM contribution to processes that are also influenced by many other factors. Values can also be given to resources for their future (sometimes known as optional or serependic) values, i.e. their potential to yield benefits in the future as well as any realized today. Thus, a value might be attributed to retaining SOM at a level which ensures its ability to store additional carbon in the future, i.e. to pay today to avoid jeopardizing future utility. Despite its obvious functional significance and importance, it is difficult to assign a cash value to SOM from the benefits that it provides. Alternatively the existence or non-use value is the value that we may be prepared to give something, simply on the basis that we welcome the knowledge that it exists. In the same way, people express their 'willingness to pay' for the maintenance of key species of the world's biodiversity, such as pandas and rainforest trees, they may also be willing to do so for the existence of a living healthy soil. There may be people who simply like to see and to smell a beautiful humus horizon in the forest, or get pleasure from the view of furrows in the cultivated fields after tillage. The Flemish painter, Brueghel the Elder, enjoyed painting cultivated landscapes exhibiting, for instance, his appreciation of a well-tilled soil in the Fall o f Icarus. Philosophers and writers on nature, such as those involved in the organic farming movement (e.g. Steiner 1924, or Rusch 1972) have attributed a clear 'existence value' to SOM by considering humus as one of the 'principles of life'.

Conclusions As shown above, there is a long history of scientists' engagement in the study of SOM, SOC or the C cycle, as a consequence of their being convinced of its functional value. There are many cases, often forgotten, of perceptions that prefigure present-day concepts accepted as essential for sound management of natural resources, such as that of sustainability (e.g. Tha~r and his system for predicting the sustainability of a farming system). Moreover, these ideas have often been based on the development of new approaches (such as

SOM FUNCTIONS: A HISTORICAL PERSPECTIVE

modelling) and tools (such as FACE-type experiments) which are readily recognizable by present-day scientists. Today's agronomists and ecologists are concerned about the impacts that h u m a n activities have on SOM. It is now generally accepted by scientists that loss of SOM is one of the major factors leading to d e g r a d a t i o n of e c o s y s t e m services and loss of ecosystem resilience. In many countries, however, conflicts have arisen, b e t w e e n policies for ecosystem protection that e m b r a c e sustainable soil m a n a g e m e n t , and those t a r g e t e d at agricultural d e v e l o p m e n t . These conflicts are often blamed on the ignorance of decision-makers, but scientists must accept that they have an equal responsibility to ensure that their k n o w l e d g e is shared in an accessible way: society is unlikely to embrace these issues unless it is c o n v i n c e d of the e c o n o m i c value of SOM. The k e y to this persuasion rests on our capacity to demonstrate that SOM is a major and essential c o m p o n e n t of ecosystem functions and services and must be conserved and sustained by appropriate ecosystem m a n a g e m e n t practices. The authors acknowledge helpful reviews by Ken Giller and Roel Merckx. This work received financial support from the Institute of Research for Development (IRD) and the Institute of Forestry, Agricultural and Environmental Engineering (ENGREF).

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SWIFT, M.J., IZAC,A.M.N. & VAN NOORDWIJK,M. 2004. Biodiversity and ecosystem services in agricultural landscapes - are we asking the right questions? Agriculture Ecosystems and Environment, 104, 113-134. SYERS, J.K. • CRASWELL, E.T. 1995. Role of soil organic matter in sustainable agricultural systems. In: LEFROY,R.D.B., BLAIR, G.J. & CRASWELL,E.T. (eds) Soil Organic Matter Management for Sustainable Agriculture: a Workshop Held in Ubon, Thailand, 24-26 August 1994, 7-14. THAi~R, A. 1809. Grundsi~tze der rationnellen Landwirtschaft (1809-1812). Realschulbuch, Berlin. THORNLEY, J.H.M., FOWLER, D. & CANNELL, G.R. 1991. Terrestrial carbon storage resulting from CO2 and nitrogen fertilization in temperature grasslands. Plant Cell and Environment, 14, 1007-1011. TIESSEN, H., CUEVAS, E. & CHACON, P. 1994. The role of soil organic matter in sustaining soil fertility. Nature, 371, 783-785. TILMAN,D., CASSMAN,K.G., MATSON,EA., NAYLOR,R. & POLASKY,S. 2002. Agricultural sustainability and intensive production practices. Nature, 418 (6898), 671-677. TULL, J. 1733. Horse-hoeing husbandry or an essay on the principles of tillage and vegetation. London. TWENHOFEL,W.H. 1926. Treatise on Sedimentation, 1st edn. Williams & Wilkins, Baltimore. VAN DER PLOEG, R.R., BOHM, W. & KIRKHAM,M.B. 1999. On the origin of the theory of mineral nutrition of plants and the Law of the Minimum. Soil Science Society of America Journal, 63, 1055-1062. VANLAUWE, B., DENDOOVEN,L. & MERCKX, R. 1994. Residue fractionation and decomposition: the significance of the active fraction. Plant and Soil, 158, 263-274. WAKSMAN, S.A. 1938. Humus. Origin, Chemical Composition and Importance in Nature. 2rid edition, revised. The Williams and Wilkins Company, Baltimore and London. WCED 1987. Our Common Future. Oxford University Press, Oxford, UK. WILLIAMS,J.R. 1989. EPIC: the Erosion-Productivity Impact Calculator. In: CLEMA,J.K. (ed.) Proceedings of the 1989 Summer Computer Simulation Conference, USA, 676-681. WOLF, B. 8z SNYDER,G.H. 2003. Sustainable Soils. The Place of Organic Matter in Sustaining Soils and Their Productivity. Food Products Press, New York, 352 pp. WOLLNY, E. 1897. Die Zersetzung der Organischen Stoffe und die Humusbildung mit Rgtchsicht Bud die Bodenkultur. Heidelberg (quoted byWaksman, 1938). WOLLNY, E. 1902. La Ddcomposition des Mati~res Organiques et les Formes d'Humus clans Leurs Rapports avec l'Agriculture (Organic matter decomposition and humus forms in relation to agriculture). Berger-Levrault et Cie, Paris and Nancy, 657 pp. WOODWARD,J. 1699. Some thoughts and experiments concerning vegetation. Philosophical Transactions of the Royal Society, 21, 193-227.

Soils as sources and sinks of greenhouse gases JENS LEIFELD

Research Station Agroscope Reckenholz-Taenikon ART, Air Pollution~Climate Group, Reckenholzstrasse 191, 8046 Ziirich, Switzerland (e-mail: jens. leifeld@art, admin, ch) Abstract: Soils annually emit between 6.8 and 7.9 Gt C O 2 equivalents, mainly as C H 4 from intact peatlands and from rice agriculture; as N 2 0 from unmanaged and managed soils; and as CO2 from land-use change. Methane emissions attributable to other wetlands add another 1.6-3.8 Gt CO2 equivalents. From a global standpoint, N20 from unmanaged soils and C H 4 from peatlands and other wetlands make soils naturally net greenhouse gas emitters. In addition, the storage of carbon in soils and the fluxes of C H 4 and N 2 0 have been changed by anthropogenic effects towards emission rates 52 to 72% above those under natural conditions before the dawn of intensive agriculture and land-use change. Land-use changes on mineral soils induced most of the recorded losses of soil organic matter (SOM), but there is evidence that proper agricultural management of soil resources is able to recover some of these losses and to maintain soil functions. However, the discrepancy between so-called 'sequestration potentials' and the measures already adopted is amazingly large. Globally, only about 5% of the cropped areas is managed according to practices such as no tillage or organic farming. The contribution of soil loss by erosion, desertification and sealing to global oxidative SOM losses is uncertain; however, in the case of soil erosion, it is considered to be a major factor in global SOM decline. Mitigation options calculated for SOM restoration, reduced CH 4 and N20 emissions are able to alleviate mean annual emissions by 1.2 to 2.9 Gt CO2 equivalents, mainly as a result of carbon sequestration, which is the most efficient measure for the next few decades. In the longer term, however, the large potential for reducing C H 4 and N 2 0 emissions outweigh the finite capacity of soils to recover C. Integrated assessment of net greenhouse-gas fluxes is key for evaluating management practices aimed at reducing overall emissions. From the viewpoint of climate change and taking into consideration the mean fluxes of CO2, CH4 and N20 , peatland protection is more favourable than peatland cultivation in the long term. The most important gaps in our understanding appear to be with regard to estimating fluxes along with soil erosion and desertification processes, in the extent of peatland cultivation; the role of black carbon formation, natural 'background' sequestration rates of undisturbed soils; and the net response of soils, particularly in cold regions, to global warming. With regard to the societal perception of soil contributing to the global cycling of greenhouse gases, it is important to emphasize that significant proportions of the emissions are inevitably linked to intensive agriculture.

A t m o s p h e r i c c o n c e n t r a t i o n s of the greenh o u s e gases ( G H G s ) carbon dioxide (CO2), nitrous oxide (N20), and m e t h a n e (CH4) are increasing significantly due to h u m a n activity, accounting for m u c h of the h y p o t h e s i z e d a d d i t i o n a l a n t h r o p o g e n i c g r e e n h o u s e effect (IPCC 2001). Most of this increase is assigned to CO2, and the relative importance of C H 4 and N 2 0 in terms of global warming is d e t e r m i n e d by the m e a n residence time of these gases in the atmosphere and their absorptivity for infrared radiation, resulting in calculated globalwarming potentials of CH4 and N 2 0 relative to CO2 of 23 and 296 for a 100-year time horizon (kg -1, IPCC 2001). F r o m the radiative forcing of 2.43 W m -2 b e t w e e n 1850 and 1990, 1.46 W m -z, 0 . 4 8 W m -2 and 0.15 W m -2 stem f r o m the atmospheric increase in CO2, CH4 and N 2 0 ,

respectively (IPCC 2001). The rest is attributed to halocarbons. Soils are key elements in the global transformation, processing, and turnover of CO2, N 2 0 and C H 4. They not only harbour micro-organisms as the m a i n actors of various t r a n s f o r m a t i o n processes, they also store organic and inorganic carbon and are regarded as a potential 'sink' for CO2. Soils also provide the space for agricultural and other land-use activities as important factors driving the atmospheric increase in GHGs. Soils serve as both sources and sinks of the major G H G s - both naturally and under the influence of anthropogenic activities. The goal of this paper is to review, how and to what extent the soil-related sources and sinks of G H G s have b e e n p e r t u r b e d historically by direct h u m a n - i n d u c e d activities and what the

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENT~N,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 23-44. 0305-8719/06/$15 9 The Geological Society of London 2006.

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J. LEIFELD

current challenges are. This is achieved by analysing (1) the most important changes in land-use and management with respect to soils and GHGs; (2) the extent of global soil loss and its impact on the cycles of CO2; and (3) the potentials of emission reductions and sequestration. A brief introduction to the processes of G H G formation is given in the following section.

Production, consumption, storage and emission of greenhouse gases in soil The processing and transformation of greenhouse gases are mostly microbially mediated, and the soil architecture is thus the living space in which these processes occur. Organic carbon can be regarded as a repository for carbon dioxide fixed via photosynthesis, and is part of the solid and liquid phase in the soil. Together with inorganic carbon derived from parent material or precipitated via desiccation of porewater rich in Ca 2+ and HCO3- during capillary rise in arid and semi-arid regions, carbon stored in soil is among the most important C pools on the global scale, exceeding that in vegetation and the atmosphere by a factor of two (Bolin & Sukumar 2000). In contrast, production, consumption and emission are the main processes related to nitrous oxide and methane, and storage of these gases in the soil pore space is limited. Measurements of the soil's C balance are carried out mainly by (1) direct measurement of the soil organic carbon (SOC) content, bulk density, and stone content for a particular depth and at different points in time; (2) integrated net CO2 exchange over larger areas (e.g. at the field level) by using Eddy-covariance techniques; and (3) various chamber techniques to quantify the CO2 efflux from soil. The first technique gives a direct estimate of potential gains and losses; however, it requires large sample numbers, owing to the heterogeneity of SOC in the field and the comparable small rates of change (e.g. P. Smith 2004a). Eddy-covariance quantifies the bidirectional CO2 flux over a field with high resolution in time, but lacks differentiation between autotrophic and heterotrophic processes, and hampers an assignment of fluxes to soil. It thus needs to be complemented by additional methods (e.g. Ehman et al. 2002). For measuring soil respiration as well as soil CH 4 and N20 fluxes, chamber techniques are most frequently used. As for SOC inventories, CO2 chamber measurements are plagued by high spatial - and in addition - also temporal vari-

ability of fluxes (e.g. Davidson et al. 2002). This variability is the cause of much of the error documented for chamber-based N20 and CH 4 fluxes. Soil organic m a t t e r ( S O M )

This is the product of residue decomposition and consists of a variety of constituents, some of which have not yet been identified chemically. In undisturbed ecosystems, the soil carbon stock, just like the vegetation carbon stock, is considered to be almost in a steady state between inputs (e.g. as plant residues) and outputs (e.g. as CO2 from respiration) of carbon. This does include episodic fluctuations (e.g. flooding, storms, vegetation fires), whose effect on the carbon stock can still be described by a steady-state model if the corresponding time span is allowed to be long enough to capture those fluctuations (e.g. KOrner 2003). With respect to soil C, a steady state can also be reached in managed or otherwise altered ecosystems if the strength of the disturbing force (e.g. frequency and intensity of tillage) remains constant. Under such conditions, the system adapts to the new situation (this often takes decades to centuries) until inputs again equal outputs. Under steady-state conditions, the relationship between SOC stock, input, output, and turnover is: C stock (t ha -1) = input rate (t ha -1 a-1) x turnover time (a) and Input rate (t ha -1 a -a) = output rate (t ha -a a -]) Evidently, a change in the C stock is attributed to (1) a change in the input or output rate (e.g. by changing the rate of residue input in agriculture or changing the output by erosion) or (2) a change in the turnover time (e.g. by changing tillage practices, which affects aeration and moisture as important abiotic forces influencing microbial activity). A significant increase in the C stock over time, namely carbon sequestration, is always associated with a non-steady-state system, as is the significant decline of the C stock. The reason that increased residue inputs do not infinitely increase the C stock is given by the soil's limited potential to store C. The state of maximum C storage is termed 'saturation', and the difference between the steady state and saturation with respect to soil C storage has been addressed by Six et al. (2002). The main reasons why SOM is accumulating in soil until it reaches a steady state can be explained by the ability of a mineral soil to protect organic materials against biochemical attack by several

SOILS AND GREENHOUSE GASES physico-chemical mechanisms, together with the selective and relative enrichment of substrates that have a limited availability for microorganisms (Sollins et al. 1996). In combination, these mechanisms probably cause an energetically unfavourable situation, where the costs for the production of enzymes and locally low substrate concentrations offset the benefit for the organisms. An additional stabilization pathway involves vegetation fires and the accompanying formation of charred products (black carbon, BC), which probably contribute to the accumulation of more resistant organic materials. Charcoal has been shown to be much more resistant to microbial degradation compared to non-charred precursors (Baldock & Smernik 2002). In soils exposed to periodic or permanent waterlogging or water saturation, and in the interior of soil aggregates, oxygen deficiency and the accompanying decelerated decomposition due to thermodynamically more unfavourable reaction mechanisms, may cause accumulation of SOM, which, in the case of organic soils, may equal thousands of tons per hectare. The agricultural use of peatlands is almost unavoidably associated with a loss in soil C, because oxygen deficiency caused by water saturation is the major mechanism leading to the accumulation of peat, and its direction is reversed by drainage and subsequent peat oxidation. Hence, rates of oxidative peat decomposition can be affected mainly by regulation of the water table (Augustin et al. 1996; Kasimir-Klemedtsson et al. 1997). Nitrous

oxide

This is an intermediate product of two main opposing processes: denitrification and nitrification. It occurs naturally in almost all ecosystems. In the enzymatic reduction chain of nitrate to nitrogen, nitrous oxide is produced together with NO and N 2 as the two other products of denitrification by various micro-organisms. A reaction of oxygen as an electron acceptor with an organic or inorganic electron source has a more negative AG, and is thus thermodynamically more favourable than a reaction with nitrate. Therefore, denitrification particularly occurs under anoxic conditions. However, nitrous oxide emissions have also been measured in aerated soils (Mtiller et al. 2004). These emissions are probably caused by anoxic conditions in the micropores that have been caused by limited oxygen diffusion or high oxygen consumption rates. Nitrification is the oxidation of ammonia to nitrate along a redox

25

gradient that converts nitrogen from the N(-III) to the N(+V) state, and is favoured under oxic conditions. Like denitrification, nitrification is associated with both chemo-autotrophic and heterotrophic organisms (Paul & Clark 1996). Because the reaction rate is proportional to the concentration of the reactants, the availability of either nitrate or ammonia together with organic matter for heterotrophic or CO2 and oxygen for autotrophic transformations mainly determines the production rate of nitrous oxide. The efflux from soil is further modified by the diffusivity and the diffusive path length, because uptake of nitrous oxide in the upper few centimetres derived from deeper horizons can be quantitatively important (Neftel et al. 2000). Not surprisingly, nitrous oxide emission rates from soils typically peak after fertilization events (i.e. when there is a surplus of mineral N), and after rainfall (increasing oxygen deficiency and increasing microbial activity; Ball et al. 1999; Simek et al. 2004). Methane

Methane in soils is a product of biochemical reduction processes and can typically be found in anoxic waterlogged ecosystems having redox potentials below -200 mV, e.g. floodplains, peatlands, marshes, and in lowland rice-cropping areas. Because of their oxygen limitation and the accompanying accumulation of organic debris, wetland soils producing methane are often rich in organic matter. Methane is produced by methanogenic bacteria along two different pathways: acetoclastic methanogenesis, and H2/CO2 methanogenesis. During acidogenesis, anaerobic decomposition of organic matter produces intermediates such as lowmolecular-weight fatty acids, acetate, H2 and CO2. In sulphate-rich sediments (e.g. marshes), sulphate serves as the final electron acceptor, and the organic material is finally mineralized to CO2 and H2S. If sulphate is absent (either consumed or naturally in low abundance, e.g. in freshwater systems), the intermediates of anaerobic decomposition are processed further by methanogenesis. During acetoclastic methanogenesis, acetate is converted to CO2 and CH4, and during H2/CO2 methanogenesis, H2 and C O 2 react to form C H 4 and H 2 0 . The latter pathway is also responsible for C H 4 emissions from ruminants. If this methane enters oxic soil horizons with higher redox potentials, it can be oxidized to CO2 and water by a number of bacteria, so its warming impact is reduced to a large extent. Because of this important transformation, it is

26

J. LEIFELD erosion, sealing or physical deterioration (e.g. Van-Camp et al. 2004). The conversion from forest to agricultural land began with the onset of the Neolithic age around 7000 years BP and became far more noticeable after the arrival of Bronze-age ploughs between 6000 and 5000 years BP (Ruddiman 2003, and references cited therein). Yet by 2000 years BP, agricultural land-use had spread throughout large areas of Eurasia, North Africa, Eastern China, India, Central- and South America (Roberts 1998). Along with the rapid increase in the human population during the nineteeth and twentieth centuries, the agricultural use of soil has increased steadily. FAO estimates for the changes in the area of land under different uses, during the 1960 to 1990s are given in Figure 1. The area occupied by agricultural land has expanded from 4.51 Gha (1 Gha = 109 ha) in 1961 to 5.01 in 2000, and the forest area has been reduced by approx. 0.2 Gha from 4.37 to 4.17 Gha between 1961 and 1994 (FAOSTAT data 2004). Most pronounced is the increase in permanent pastures by 0.33 Gha between 1961 and 2000. At the same time, the arable area has expanded by c. 0.13 Gha. The expansion of arable land in developing countries from 2006 to 2030 is estimated to increase by 25% to 28% for rain-fed and irrigated crop production, respectively (FAO 2003),

worth taking a closer look at the transport mechanism of methane in soils and plants. Methane from soil reaches the atmosphere by three different pathways: ebullition, diffusion, and transport in the aerenchyma of vascular plants. Only in the case of diffusion are the transport rates slow enough to allow a significant proportion of the methane to be oxidized before it enters the atmosphere, while, during convection in the case of gas bubbles or gas transport in the plant aerenchyma, the rate of decrease relative to that of transportation is considered to be small. Oxidation is also an important process of methane consumption in aerobic soils. Its rate is highly dependent on climatic conditions and soil management and is greatest under undisturbed forest, for stillunknown reasons (Robertson & Grace 2004).

Land-use change and agricultural intensification, and their impact on carbon dioxide fluxes L a n d - u s e changes Land-use changes are important driving forces for changes in the soil's state and they lead to many of the important threats to global soil resources, such as loss of organic matter,

5

~

~.~~ .~..~.~ ~'~=J4~E3 -e--Total agricultural area

4

~ Arable land

az (.9

3

- e - Permanent pasture

84

Forests and woodland (FAO Production Yearbook)

2

FAO Global Forest Resource Assessment

1

1961 1965 1969 1973 1977 1981 1985 1989 1993 1997 2001 2005 Year

Fig. 1. Major world land-use categories from 1961 to 2004. Source: FAOSTAT data (2004).

SOILS AND GREENHOUSE GASES with maize being the most important single contributor to this land-use change. According to the FAO Global Forest Resource Assessment (FRA 2000), the forest area decreased from 3.96 to 3.87 Gha between 1990 and 2000. Differences in forest cover estimates are caused by the methodology used (see F R A 2000); however, the trend is consistent. Recent deforestation during the 1990s occurred in the tropical regions (net annual loss 12.3 Mha), while the temperate regions gained forest area (+2.9Mha) by natural expansion, afforestation and reforestation. This represents a general pattern during the twentieth century, where the geographical distribution of land-use change shifted towards the subhumid and humid tropics, accompanied by the still ongoing clearance of large areas of tropical forests and a concomitant increase in tropical pastures (Houghton 2003). Agroforestry, a management system that integrates trees into farmland, is included in the FAO categories 'arable land' and 'forest land', and accounts for approximately 0.4 Gha, mostly allocated in the tropics. Conversion of native grasslands to cropland, managed pastures or managed forest plantations is another important element of global land-use change. Grasslands (tropical savannas and temperate grasslands, both natural and anthropogenic, without semi-deserts) cover approx. 3.5 Gha, or 23% of the global land area (Bolin & Sukumar 2000). For example, grassland conversion has been documented for current land-uses in comparison to the former, potential natural prairie vegetation for large parts of the Central and Western US (Sobecki et al. 2001); for the semi-arid grasslands in China (Zhao et al. 2005); or for South American grasslands (Farley et al. 2004). Cultivation of wetland soils, and of peatland soils in particular, is of major importance for the terrestrial C budget, although its contribution by area is rather small. The estimated total area of peatland soils is 254-480 Mha (Gorham 1991; Paustian et al. 1997; W B G U 1998; Spiers 1999; Moore 2002). According to Spiers (1999) and Moore (2002), the vast majority of peatlands are

27

allocated in the boreal and sub-arctic regions of Canada, Alaska and Russia, in particular. Estimates of the area of tropical peatland range from 30.6 to 45.9 Mha, of which the majority is located in Indonesia (17-27 Mha), Malaysia (1.5 Mha) and China (1-3 Mha) (van Dam et al. 2002). More recent estimates on peatland extension for West Siberia (Sheng et al. 2004) and Indonesia (Page et al. 2002) indicate that the total area of peatlands might even be higher. Thus, peatlands are the most important single category of freshwater wetlands, the latter covering approximately 530 to 570 Mha (Spiers 1999), including swamps as a category similar to peatlands. The share of peatlands drained or used for peat-cutting has been estimated to be 11.5 Mha and 4.4 Mha, respectively (Gorham 1991), and 39 Mha for both drained and cut peatlands by Paustian et al. (1997). O E C D (1996) figures on the drainage of wetlands in 1985 are 26% worldwide (56-65% of the available wetlands had been drained for intensive agriculture in Europe and North America; the figures for tropical and subtropical regions were 27% for Asia, 6% for South America and 2% for Africa). With respect to an estimated total area of 860 Mha of wetlands in the OECD study and the figures on peatland areas amounting to approximately 400 Mha, it is possible that the drainage and/or cutting of peatlands exceeds the values of 15.9-39 Mha cited above.

Land-use changes and impacts on soil-derived C02 fluxes Losses and gains of SOC are closely linked to the bulk CO2 flux attributable to land-use change. For the period 1850-1980, the annual net CO2 flux from land-use change was almost identical for the tropics (tropical Asia, America, Africa) and the non-tropics (Canada, US, Europe, former Soviet Union, China Pacific, North Africa and the Middle East); however, the ratio of CO2 tropics to CO2 non-tropics has changed considerably during the 1980s and 1990s (Houghton 2003; Table 1). For the 1980s

Table 1. Estimates of the main annual sources (+) and sinks (-) o f carbon resulting from different types of land-use change and management, since 1850

Region Tropical areas Other areas (non-tropics) Source contribution from the tropics (%) Based on Houghton (2003).

1850-1980 (Gt C) +56.7 +57.6 50

1980-1989(Gt C a-~) +1.93 +0.06 97

1990-1999(Gt C a-1) +2.20 -0.02 100

28

J. LEIFELD

and 1990s, annual emissions from land-use change have been estimated to be 7.3 + 2.9 and 8.1 _+ 2.9 Gt CO2, respectively. During the period 1850-2000, global CO2 emissions from land-use change were estimated to be 572 ___ 202 Gt CO2 (Houghton 2003) or 3.8 + 1.4 Gt CO2 annually - 87% of which is derived from forest areas and approximately 13% from cultivation of mid-latitude grasslands (Bolin & Sukumar 2000, and references cited therein). In addition to these more recent developments, Ruddiman (2003) also estimated the C lost by the early forest clearances since the Neolithic age. This estimate of 822-914 Gt CO2 even exceeds the loss induced by land-use changes since 1850, but its magnitude has been questioned (Joos et al. 2004). What are the implications of these land-use changes for the soil's function as a source and sink of CO2? The global amount of SOC (from 0 to 100 cm depth) has been estimated to be 1200-1600 Gt (Batjes 1996; Paustian et al. 1997; Jobbagy & Jackson 2000). Based on the difference between precultivated and current SOC stocks in cultivated soils, Paustian et al. (1997) estimate a global loss of 40-60 Gt due to cultivation, from which the major part (approximately 80%) is derived from aerated mineral soils and the remainder (approximately 11 Gt) is from the drainage of 'wetland soils', that is, Histosols and Gleysols (FAO). While the SOC loss estimated so far only accounts for a small fraction of the global SOC stock (2.5 to 6%), it contributes one-third to the estimated C loss of 156 + 55 Gt caused by land-use change since 1850. Two-thirds are attributable mainly to deforestation. In many parts of the world, forest fires (whether ignited by humans or caused naturally) are taken as the cue for land-use change to agricultural or pastoral uses. While the CO2 released by vegetation burning has been considered for global estimates (Houghton 2003), the net effects on the soil's C balance are hardly ever addressed. Potentially, the formation of black carbon (BC) is another quantitatively important budget component accompanying vegetation fires. Its net effect on the global C budget is still poorly understood. The annual rate of BC formation has been estimated to account for 0.05-0.2 Gt C, 80% of which has been assigned to vegetation fires (Kuhlbusch 1998). Most of these fires are human-induced and can also contribute to large CO2 emissions from burning peatlands (e.g. Page et al. 2002). Black carbon is assumed to represent between 1 and 6% of the total SOC, with much higher concentrations in regions

affected by frequent vegetation fires (GonzalezPerez et al. 2004). Estimates of the amount of soil organic carbon stored in peatlands vary widely between 225 Gt C (643 t ha-l; IPCC 2000), 43.6 Gt C (1120 t haq; Paustian et al. 1997, only for artificially drained peatlands), 455 Gt C (1300 t haq; recalculated from Moore 2002, only for temperate peatlands), and 541 Gt C (Bergkamp & Orlando 1999; tropical peatlands 1700-2880 t ha q, boreal and temperate peatlands 1314-1315 t haq). Data from K. A. Smith et al. (2004) for the West Siberian Lowland indicate mean C stocks of approximately 1200 t ha -1. For Swiss temperate fens, Leifeld et al. (2005) estimated a mean C stock of 1600 t ha -1 for intact peatlands, and 710-1050t ha q for peatlands drained several decades ago. On the basis of an assumed global C stock in peatlands of between 225 and 541 Gt, the loss of 11 Gt is equivalent to 2-4.9% of this reservoir. In some regions, the actual loss induced by peatland cultivation and peat excavation might significantly exceed this percentage. Considering that the previous area of intact peatlands in Central Europe of 5.9 Mha has been reduced to a current value of 0.29 Mha (Succow & Joosten 2001), and assuming C stocks of between 1300 and 1600 t ha q for intact - and of between 710 and 1120 t ha q for cultivated - peatlands, the historical peatland cultivation in Central Europe alone accounts for a C loss of 0.68-3.9 Gt C, or 7-51% of the former peatland C stock. To attribute the SOC loss discussed above to particular regions or land-use types is challenging, chiefly because of the regional imbalance in available soil data. In temperate regions, C losses resulting from the conversion to agriculture have been documented frequently (e.g. Cihacek & Ulmer 1997; W. N. Smith et al. 1997; Lal et al. 1998; Mikhailova et al. 2000). As pointed out above, tropical forest clearance and conversion to pasture is currently the quantitatively most important land-use change globally. This land-use change is accompanied by a strong reduction in biomass C, but not necessarily in soil C. In a review of 115 studies from 300 datapoints worldwide, Conant et al. (2001) found significant evidence that, during the conversion from tropical forest to pasture, the soil carbon stock increases more often than it decreases. A similar result has been reported by Guo & Gifford (2002), who reported significant increases in the SOC stock after conversion from forest to pasture for many paired plots, particularly in regions with high precipitation. For the Ecuadorian Andes, Farley et al. (2004) reported a significant decline in SOC in the A

SOILS AND GREENHOUSE GASES horizon of pine plantations for all age classes (five to 25 years) after conversion of the native tussock grasses, which were not cleared or burned before first planting. Results on increasing SOC after tropical forest clearance and pasture establishment, or decreasing SOC after the introduction of plantations do not exclude a possible decline in SOM after forest clearance (e.g. Tiessen et al. 2003), but indicate that, under proper management, SOC stocks can at least be preserved in many cases. Agroforestry is a suitable alternative that can reduce the negative effects on SOC caused by slash-and-burn agriculture followed by the introduction of arable rotations, particularly in tropical regions. Relative to undisturbed forests, agroforestry systems are able to maintain 80 to 100% of the former C stock and may also improve the productivity of degraded tropical pastures (IPCC 2000). Precipitation is a major driver for both plant productivity and SOC (Jobbagy & Jackson 2000). It also significantly affects the rate and direction of SOC change when land use is altered. In a comparison of six adjacent plant communities used either as grassland or where shrub/woodland encroachment was allowed, along a precipitation gradient in the SouthWestern USA, Jackson et al. (2002) showed that there was a significant negative linear relationship between the change in soil C and the annual precipitation. For sites with annual precipitation above 300 ram, woody encroachment reduced SOC contents (up to 50% of the initial value at the Engeling site with 1070 mm annual rainfall), while, for the dry sites, SOC slightly increased. These data emphasize the need to account for co-variables, when effects of land-use change on SOC stocks are evaluated.

Soil degradation and sealing and its impact on the global soil C storage Severe soil degradation by desertification, erosion or sealing has been recognized as an important threat to the soil reserve (Van-Camp et al. 2004); however, it has been less intensively discussed in terms of the soil's role as a source and a sink of GHGs. In contrast to the land-use issues addressed above, soil loss has an almost irrevocable negative effect on SOC storage. Oldeman et al. (1991) estimated the extent of human-induced soil degradation for the world. Out of four categories of soil degradation, strong degradation (i.e. terrain not reclaimable at farm level), and extreme degradation (i.e. irreclaimable terrain destruction beyond

29

restoration) as topsoil loss induced by water and wind erosion, together account for 170.6 Mha. The three main causative factors for soil degradation, in decreasing order of magnitude, are overgrazing, deforestation and agricultural mismanagement. More recent global data are missing, however, U N E P (2002) states that, despite trends towards soil conservation and resource management, there is no clear indication that the rate of land degradation has decreased. Desertification refers to land degradation in arid, semi-arid and subhumid areas due to anthropogenic activities ( U N E P 1993). It can be exacerbated by long, continuous dry periods (Nicholson 2001). Eswaran et al. (2001) estimated the total area vulnerable to desertification to be 4.3 Gha, of which 0.79 Gha are classified as very highly vulnerable. Desertification not only leads to severe reductions in land productivity of up to 50% (Eswaran et al. 2001), but, like erosion in humid regions, it leads to a decline in the soil-profile thickness and thus to a decline in its potential to store C. The impact of soil erosion on the global carbon budget has recently been discussed by Lal (2003). He estimated that 4.0-4.6 Gt C are eroded annually by water erosion, 0.8-1.2 Gt of which is emitted as CO2 to the atmosphere. The major fraction is redistributed over the landscape and, to a minor extent, buried in depositional sites of terrestrial and aquatic ecosystems. The corresponding area prone to erosion is 1.09 Gha for water erosion - 0.75 Gha of which are severely affected, and 0.55 Gha for wind erosion - 0.30 Gha of which are severely affected (La12003). However, an estimate of the C lost by wind erosion is not provided. According to these figures, of the approximately 13.6 Gha total land area, 12% is affected by erosion in general, including 8% by severe erosion. A simulation on global soil dust emissions estimated the total dust loading to be 1.92 Gt, of which only a minor fraction of 0.12 Gt was attributed to agricultural areas, mostly in semi-arid regions of Africa, continental Asia and Australia (Tegen et al. 2004). On the basis of mean SOC concentrations of roughly 1% for soils in semi-arid regions (Batjes 1996, 1999), the total mass of SOC transported by wind erosion becomes 19.2 Mt, with an anthropogenic contribution of approximately 1.2 Mt C. This indicates the very small share of wind to global SOC erosion. For an area prone to severe wind erosion of 0.55 Gha, a global soil dust emission of approximately 1.92 Gt yields an annual soil loss by wind erosion of 3.5 t ha -1, corresponding to 0.035 t SOC ha -1 a -1. As for water erosion, the fate of the eroded material in

30

J. LEIFELD

terms of its global contribution to the CO2 budget is largely unknown. In addition to the mass loss by erosion and the accompanying reduction in topsoil thickness, the remaining soil is often relatively enriched in sand (Zhao et al. 2005), which reduces further plant growth and its future capacity to protect plant residues entering the soil. Soil sealing by urbanization and road construction can also be regarded as affecting the SOC storage capabilities of soils. Although sealed areas are considered to increase worldwide, owing to changes in population, habits and economic structures, reliable figures on their extent are sparse. Recent global estimates on the extent of urban and built-up areas range from 25.6 to 39.5 Mha (WR12005), but probably underestimate the existing expansion. Germany and Switzerland are considered here as examples to compare the values of WRI (2005) with the corresponding data on urbanization as provided by the national authorities of these countries (BFS 2001; BMU 2002). While the WRI (2005) estimates that 0.027 and 0.84 Mha fall under the category 'urban and built-up' for Switzerland and Germany, respectively, the corresponding national data are 0.26 (including areas not sealed but assigned to urbanization, settlement and road construction) and 2.1 Mha (sealed areas). A conservative estimate, taking only areas into account that are severely threatened or already lost for future C sequestration owing to strong and extreme erosion as well as to sealing, gives a total global value of 196-210 Mha. Soil loss and the concurrent decline in SOC caused by the processes discussed above has undoubted consequences for many soil functions at the affected sites, including the soil's function as a carbon reservoir and a CO2 buffer. The situation is more complex when the global CO2 budget is to be assessed. Lal (2003) points out that the fate of the eroded material with respect to its potential contribution to atmospheric CO2 is largely unknown, and cites ranges from 0% to 100% for the oxidizable fraction. The same holds for SOC stored in soils that are excavated and displaced along with construction activities. Although this material is lost as a possible accumulator for additional C, its original C will most probably not be fully oxidized. After its displacement, soil may be deposited on to existing surface horizons, at depressional sites as colluvium; in river floodplains, estuaries and deltas; and, finally, on continental shelves. Many of these geomorphological landscapes provide ecological conditions that are less favourable for decomposition compared to the place of origin,

and may thus serve as a carbon sink in the medium term. Even under favourable conditions, only a fraction of the SOC is readily decomposable (Kiem et al. 2000). With respect to the global CO2 balance, the amount of soil displaced and the fate of SOC as a part of it can be regarded as largely unknown and are thus a subject of high priority for research.

Measures to combat soil degradation and increase C sequestration With increasing awareness of soil degradation, decline and, more recently, the role of soils in the global cycle of greenhouse gases, agricultural policy has begun to support measures for soil conservation and soil fertility improvement. These measures include several conservation tillage practices; conversion of erodible land from active crop production to permanent vegetative cover or to agroforestry; adapted crop rotations and adapted grassland management; and improved fertilization, as well as organic farming practices (La12004a). They may already have led to a deceleration in SOC loss and a reduction in erosion, and a partial regeneration of SOC stocks in many parts of North America, South America and Europe, and are considered to contribute to soil carbon sequestration (e.g. West and Post 2002; Lal 2004a; Holland 2004; P. Smith 2004b). Global estimates on the potential to sequester C in soil by means of reclaiming degraded land, range from 16 to 78 Gt (Paustian et al. 1997; Batjes 1999; Lal 2004b). Batjes (1999) distinguished various scenarios for C sequestration for degraded soils, based on the classification given by Oldeman et al. (1991). He found sequestration potentials between 14 + 7 Gt C (restoration of degraded agricultural land and improved management of arable land) and 20 + 10 Gt C (additional restoration of degraded extensive grasslands, forest regrowth on degraded land) to be realistic. It is worth noting that these sequestration measures have the potential to recover substantial amounts of the 40-60 Gt estimated as humaninduced losses in SOC as a result of cultivation and land-use changes. Although the importance of soil C sequestration as a tool to mitigate global warming should not be neglected, the attainable capacity of the soil as a C sink is most likely to be smaller than the historical losses induced by human activity (Lal 2004b). The effectiveness of no tillage and other soil conservation practices also depends on their continuous application, because even a single ploughing event may lead to the loss of all additionally

SOILS AND GREENHOUSE GASES sequestered carbon (Stockfisch et al. 1999), and probably more importantly, to the mineralization of the sequestered organic N. Conservation agriculture includes any practice which reduces, changes or eliminates soil tillage and avoids burning of residues, thus maintaining sufficient surface residue throughout the year (ECAF 2005). Zero or minimum tillage and direct seeding are important elements of conservation agriculture. The global extent of no tillage has been estimated to be 58 Mha (FAO 2001). The area under other soilconserving practices for arable land is probably much larger. A comparison of the European figures (ECAF 2005) for no tillage (0.96 Mha which correspond well with the data in FAO 2001) with the European figures on total conservation tillage (10 Mha) indicates that, on a global level, the area under conservation tillage is also likely to exceed the 58 Mha given by FAO (2001). Organic farming has also been assumed to contribute to soil conservation (e.g. Shepherd et al. 2002). Studies comparing management effects of conventional v. organic farming often reveal a positive effect of organic farming on SOM storage (Condron et al. 2000; Pulleman et al. 2000; Pulleman et al. 2003), mainly because of a higher share of leys and grain legumes in arable rotations, and higher rates of manure application. These management practices seem to balance out the higher mechanical disturbance that is often needed in organic systems for the purpose of mechanical weeding. Differences in SOM might be smaller when compared to integrated farming systems with similar rotations and less mechanical destruction of the soil's structure. The area under organic farming was estimated to be 2 4 . 1 M h a in 2003, and 26.5 Mha in 2005 (Yussefi & Willer 2004; I F O A M 2005). This area adds to the area under conservation tillage practices, because these two systems typically are mutually exclusive. Conversion of marginal land to permanent vegetative cover in the US is estimated to be

31

c. 14.8 Mha (Lal et al. 1998). The underlying Conservation Reserve Program (CRP) has not only led to a decline in soil erosion, but also to increases in SOC (Gebhart et al. 1994). Agroforestry has been practised traditionally in the humid tropics, and is increasingly being adopted to prevent soil erosion and improve soil fertility (Oelbermann et al. 2004). While those benefits, together with higher biomass C stocks, have been reported frequently, the effects on SOC are not consistent. The net effect will depend not only on crop productivity, and thus residue return, but on competition for water and nutrients between arable crops and trees, and on the particular residue management (Montagnini & Nair 2004). Table 2 is a compilation of the most important data discussed above. Assuming that soil loss mainly occurs in agriculture, strong and extreme erosion concerns 3.5% of that area. In addition, 0.51-0.78% of agricultural soils have been destroyed by urbanization, which adds up to a total of c. 4% of the world's agricultural soils that have been lost or which are severely threatened by human interference. These areas are no longer available for storing biogenic C, thus burdening the remaining area with respect to C storage potentials. In contrast, 5.5% of the arable land (1.7% of the agricultural area) is affected by measures that actively contribute to the protection of soil and soil C. Adding the area under CRP in the US reveals that 2.0% of the agricultural area is given active protection. These data, although still a fragmentary approximation, show the gap between potentials for SOC sequestration and protection, and the current situation.

Agricultural intensification and nitrous oxide emissions The key to understanding global N20 emissions from soil is the close relationship between the

Table 2. Some key figures on the global use o f soils (Mha) Agricultural areas

Areas degraded by erosion

Arable crops*

Pastures

Strong*

Extreme*

1530

3490

170.6

4.7

Urban and built-up areas

25.6--39.5

Soil-conservation areas in cropping systems No tillage

Organic farming

58

26.5

*Includes 130 Mha permanent crops, tFor definitions, see the text. Sources: Land use: FAOSTAT data (2004); erosion: Oldeman et al. (1991); urban and built-up areas: WRI

(2005); no tillage: FAO (2001); organic farming: Yussefi & Wilier (2004) and IFOAM (2005).

32

J. LEIFELD

amount of N cycled in terrestrial ecosystems and the amount of N20 released from them. The factors that affect nitrous oxide emission from the soil are partially reflected in the IPCC guidelines for the calculation of G H G emissions from soil (IPCC 1996), whereby input rates of N into the soil by organic and mineral fertilizers and by fixation and plant residues are multiplied by the corresponding emission factors, in order to calculate the nitrous oxide emissions. Important emission factors for N20, according to the IPCC (1996) are 1.25 + 1% for fertilization, 0.5-3% for the storage of manure, 0-1% for the storage of slurry, and 0.2-12% for leached nitrate-N (the percentage of the applied mineral or organic N for fertilization, percentage of stored manure or slurry-N, and the percentage of leached N, respectively). Under natural or semi-natural conditions, many terrestrial ecosystems, particularly in temperate regions, are N-limited, and losses to adjacent compartments, such as to groundwater or the atmosphere, are small (e.g. Asner et al. 1997). Nitrogen in natural or semi-natural terrestrial systems mainly originates from N fixation and from lightning. Before the disseminating agricultural cultivation of legumes, terrestrial organisms fixed c. 90-140 Mt N per year, and an additional 5-10 Mt originated from lightning (Vitousek et al. 1997). With the advent of commercial ammonia production using the

H a b e r - B o s c h process and the selective breeding of N-fixing crops, global amounts of N entering the soil have increased significantly. Vitousek et al. (1997) estimated the annual fixation by agricultural crops to be 32-53 Mt. The production of N fertilizers peaked in 1996 ( 9 0 M t N); 8 7 M t were produced in 2002 (FAOSTAT 2004). Figure 2 illustrates the chronological progression of the global N production and the yields of major N-fixing crops. Between 1961 and 2002, fertilizer-N production increased by a factor of 6.7, soybean production by a factor of 7.7, and the production of other pulses by a factor of 1.6. Dividing the annual fertilizer-N production by the arable cropping area (Fig. 1) yields a calculated N application of 10 kg N ha -] a-] in 1961, and 62 kg N h a -1 a -1 in 2002. These basic figures indicate a dramatic increase in N20 emissions from managed soils during the twentieth century. Global estimates of fertilizer-induced N20 emissions from agricultural soils are highly uncertain and range from 0.6-14.8 Mt N 2 0 - N a -] (IPCC 2001). Emissions from natural soils, including background emissions from agricultural soils, were estimated to be 6.8 Mt N 2 0 - N a -1 (range 3.3-9.9; IPCC 2001; based on Kreileman & Bouwman, 1994). Some of the N20 from 'natural' soils may be already derived from anthropogenic atmospheric deposition.

250

200

Fertilizer-N

.

150

100

0

~

1961

,

,

i

1

1965

~

~

~

,

1969

~

i

1

~

1973

~

~

,

~

1977

,

~

~

~

,

1981

i

~

,

1985

~

,

~

~

1989

i

~

T

~

1993

,

~

~

~

1997

~

~

~

,

2001

~

,

,

2005

Year

Fig. 2. World production of mineral N fertilizer and selected N-fixing crops from 1961 to 2004. Other pulses include beans, peas, chickpeas, cowpeas, pigeon-peas, lentils, bambara beans, vetches and lupins. Source: FAOSTAT data (2004).

SOILS AND GREENHOUSE GASES The frequently applied emission factors for N fertilization (IPCC 1996; based on Bouwman 1994) for the calculation of national G H G emissions, imply a linear relationship between the amount of N added to the soil and the amount of N20 released from it. In contrast to the IPCC (1996) data, more recent studies suggest nonlinear responses between N20 emission and N application and include additional factors to estimate the release of N20 (Bouwman et al. 2002; Flynn et al. 2005), which may lead to higher or lower emission estimates, depending on management and environmental conditions. By using an empirical modelling approach, which included 846 N20 measurements from all over the world, Bouwman et al. (2002) estimated that the annual emission from fertilized agricultural fields (including those where manure was applied) to be 2.8 Mt N20-N. Kaiser & Ruser (2000) evaluated six long-term plots in Germany and confirmed the range of emission factors as proposed by the IPCC (1996); however, they did not find any significant correlation between N application and N20 emission, and 40% of their observed variability in N20 emissions remained unexplained by Bouwman's model (1994). Kaiser & Ruser assigned the residual variability to crop- and site-specific effects on N20 emission. In this study, approximately 50% of the measured emission occurred during winter, thus emphasizing the need for long-term measurements. Bouwman et al. (2002) showed that N20 emissions, irrespective of climate, soil type and fertilizer type, were not linearly, but disproportionately related to the amount of applied N. Significant factors for the prediction of N20 emissions were the N application rate, as well as fertilizer type, climate, SOC content, soil texture, drainage status, soil pH, and crop type. For global assessments, three findings from their study are of particular importance: (1) mean fertilizer-induced emissions were lower than the values proposed by IPCC (1996), for both mineral N and manure N; (2) grasslands had lower emissions than croplands, apart from poorly drained soils where emissions may exceed those from crop cultivation on well-drained soils; and (3) N20 emissions from tropical and subtropical soils exceeded those from temperate soils. Hall & Matson (1999) discussed N20 emissions after N additions in tropical forests on the Hawaiian Islands. They found higher N20 emissions for P-limited than for N-limited forests, and they cautiously assigned this response to the inability of the P-limited system to retain much of the anthropogenic N. Phosphorus limitation is also a common property of many agricultural

33

tropical soils, because of the prevalence of Fe and A1 oxides/hydroxides and allophane with a high sorption capacity for phosphate, particularly at low pH values. This chemical property, in conjunction with higher temperatures in the subtropical and tropical regions, might contribute to the higher N20 emissions per unit applied N, as found by Bouwman et al. (2002). Draining organic soils typically causes a tremendous increase in N20 emissions. Zeitz & Velty (2002) reported emissions of 0.0-0.8 and 0.3-26.9 kg N 2 0 - N ha q a q for wet and drained fens in Germany, respectively. Augustin et al. (1996) compiled data from Finland and the US, showing N20 emissions from undrained fens to be 0.4-0.5 kg N 2 0 - N ha -1 a q, and 1.2-13.1 N 2 0 - N ha -1 aq after drainage. Similarly, Kasimir-Klemedtsson et al. (1997) reviewed studies from Finland, Sweden and The Netherlands that provided evidence for an increase in N20 emissions from zero to 5.7_+ 3.2 (drained, grassland) and 9.6 +_ 7 kg N 2 0 - N ha-* a -1 (drained, cereals), respectively. In view of the increasing rate of additional N input via fertilizers and biological N fixation, global N20 emissions from soils have probably increased sharply during the last five decades. Projections for the use of fertilizers are based on the assumption of an increase in fertilizer consumption, but at a lower annual rate of approximately 1% than in the past, due to improvements in nutrient use efficiency and a projected slowdown of increases in crop production (FAO 2003). From 2006 to 2030, increases in fertilizer use in the industrial countries, especially in Western Europe, are expected to lag significantly behind increases in other regions of the world, because of changes in agricultural policies. Thus, more detailed knowledge on the controlling factors for N20 emission in subtropical and tropical agricultural ecosystems is needed in order to identify appropriate mitigation options. The FAO report (2003) did not distinguish between different chemical elements; however, N fertilizers, as the single most important nutrient, will probably be used in proportion to this projection. In contrast to FAO (2003), an extrapolation of past trends in fertilizer use projects a much higher consumption. Tilman et al. (2001) predicted a fertilizer-N application of 135 Mt for 2020 and 236 Mt for 2050, corresponding to annual increases of 2.8% and 3.4% relative to 2000, respectively. Their estimate, based on linear regressions between fertilizer consumption and time, global population, and global gross domestic product (GDP) as independent variables, is higher than

34

J. LEIFELD

that of the FAO (2003), mainly because of high projections when using GDP as a predictor for N consumption. Extrapolations using population growth roughly halve the projected increase by 2050, to 1.8% annually. Mitigation strategies should be geared to siteand management-specific demands and cannot be generalized easily. Dinitrogen monoxide is an unwanted excess product of N fertilization and, like ammonia and nitrate, can be decreased by appropriate mitigation. Since a further global increase in N fertilization and biological N fixation is probable, an increase in the N utilization efficiency is a key factor not only to reduce N20 emission, but environmentally harmful N emissions in general. The global N utilization efficiency (i.e. N output from harvest and livestock, divided by N inputs) was approximately 50% in 1996 (Mosier 2002). For German agriculture, Schweigert & van der Ploeg (2002) calculated the annual N surplus from the difference between the N input by commercial fertilizers and imported fodder, and the N output by animal and crop products for the period from 1951 to 2000. This surplus increased from c. 1 0 k g h a q in the 1950s to 1 2 0 k g h a q to 70-80 k g h a q in the 1990s. For farms in England, Leach et al. (2004) reported N surpluses of up to 250 kg ha q a-~ for pig/arable and dairy farms, while N surpluses on arable farms without animal husbandry were 100 kg ha q a q. The corresponding N utilization efficiency was 50% for arable farms, while that for animal husbandry (beef/sheep or dairy) was below or close to 20%. A number of management practices have been identified that can reduce the amounts of fertilizer used and improve the efficiency rate. Among these, timing multiple applications rather than single ones during a growing season, avoiding application of manure/slurry during bare fallow seasons, a balanced fertilization (i.e. considering possible growth limitations induced by shortages in other nutrients), dissolution of N fertilizer in irrigation water, application below the soil surface, and measures subsumed as precision farming (i.e. N application depending on actual crop demands as measured by on-site fluorescence, and on soil characteristics and yield potentials) are important ones (Vitousek et al. 1997; Khosla et al. 2002; Mosier 2002). Like recommendations for the increase in N utilization efficiency, management practices to reduce N20 emissions have been assessed (Dalal et al. 2003). These measures include matching the supply of applied N to the spatial and temporal

requirements of the crops and pastures, splitting fertilizer applications; balancing N with other nutrient supplies; and using cover crops to take up residual mineral N from applications to the preceding crop or from mineralization of N fixed during legume or ley phases. Sehy et al. (2003) showed that site-specific N fertilization within a single field resulted in a 34% decrease of N20 emissions, while maintaining crop yields. In the area on their field where sitespecific differences in soil water constrained yields, higher N inputs were not able to increase crop yields. The adoption of nitrification inhibitors is a chemical tool used to reduce N 2 0 emissions and to improve the efficiency of fertilizer utilization. Nitrification inhibitors are chemicals that reduce the rate at which ammonium is converted to nitrate, by interfering with the metabolism of nitrifying bacteria. The loss of N from the root zone can be reduced by maintaining applied N in the ammonium form during periods of excess rainfall prior to N uptake by crops. Nitrification inhibitors were shown to reduce N20 emissions by 60% and 42% when applied together with cattle slurry and calcium ammonium nitrate, respectively (Merino et al. 2002). Carbon and nitrogen sequestered by reducing the soil tillage or by land-use changes require long-term conservation measures to avoid the release and reflux of the elements to the atmosphere and other ecosystem compartments. Ploughing of grassland soil and cultivation of bare soils leads to temporarily increased N20 emissions (Baggs et al. 2000). Increased nitrogen mineralization induced by tillage is typically more difficult to manage because microbes are also active during phases when bare soil is exposed or when there is or reduced uptake of nitrogen by plants.

Methane: implications of peatland dynamics and rice agriculture Natural wetlands and rice agriculture are among the most important single sources of the global C H 4 flux. From the c. 500-600 Mt C H 4 emitted annually worldwide, between 203 and 337 Mt are derived from wetland and rice soils (IPCC 2001). Estimates on the amount of C H 4 emitted from wetlands alone range from 115 to 237 Mt C H 4 a -1 (IPCC 2001, and references therein). On the other hand, aerobic soils oxidize c. 29 Mt C H 4 annually (Smith et al. 2000). The major quantitative process of atmospheric CH4

SOILS AND GREENHOUSE GASES consumption is its reaction with OH radicals in the troposphere (IPCC 2001). Methane emissions from peatlands which are not drained, mined or otherwise used agriculturally vary widely, depending on the water level and oxygen supply, nutrient availability and temperature. Bergkamp & Orlando (t999) estimated that natural wetlands in boreal and temperate regions emit 0.11-0.20 t CH4 ha -1 a q, and tropical wetlands 0.35-0.37 t CH 4 ha -~ a-1. Similar CH 4 emission rates of 0.1-0.28 t CH4 ha -1 a -1 were reported by Kasimir-Klemedtsson et al. (1997) for Finland and Sweden. Augustin et al. (1996) reviewed studies from bogs and fens in Sweden, Finland, Germany and the US (temperate zone) and reported CH 4 emissions to range from 0.01 to 2.4 t CH 4 ha -1 a -1. In their study, CH 4 emissions from bogs (i.e. ombrotrophic peatlands) were always higher than those from fens (minerotrophic peatlands). In Finland, Martikainen et al. (1996) measured fluxes of 0.15-0.45 t CH4 ha -1 a q for a series of natural bogs and fens, and Nyk~inen et al. (1996) reported fluxes of 0.15-0.26 t CH 4 ha q a q for two different years at yet another site. Values for wetlands, including other types than peatlands, were modelled by Cao et al. (1998). They estimated the mean CH 4 emissions to be 0.15, 0.55 and 0.71 t CH 4 ha -I a q for northern, temperate and tropical wetlands, respectively, considering a total wetland area of 504 Mha. The value for this area lies below the range for freshwater wetlands given by Spiers (1999), so it is assumed that freshwater wetlands, including peatlands, comprise the area modelled by Cao et al. (1998). The total emission was estimated to be 92.3 Mt CH 4 a -1. If the emission from peatlands is assumed to be proportional to that of all wetlands considered by Cao et al. (1998), the global contribution by peatlands would become 46.5-73.3 Mt CH 4 a -1 for the above-mentioned peatland area of 254-400 Mha. Smith et al. (2004) suggest a CH4 emission attributable to the n o r t h e r n hemisphere peatlands of c. 60 Mt CH 4 a -1, based on interpolar (i.e. Greenland/Antarctic) methane gradient data from ice cores and assumptions on global distributions of CH4 sources and sinks. This value lies well within the estimate given above, keeping in mind that more than 90% of the world's peatland area is located in the northern hemisphere. Based on the data of Cao et al. (1998), a mean calculated emission rate of 0.18 t CH 4 ha -1 a -1 can be derived, which is within the range of the studies from field trials cited above. It must be noted, however, that

35

measured ranges of CH 4 emissions from peatlands, even at the same site but under different conditions, can vary by two orders of magnitude. Similar to aerobic soils, natural peatlands may also oxidize some atmospheric CH4 during drier periods; however, this flux is small relative to the annual release (Nyk~inen et al. 1996). Along with the drainage and cultivation of natural peatlands for agricultural purposes or peat mining, CH4 fluxes, in general, are drastically declining, while CO2 fluxes change from a small sink to a strong source. From those studies cited above that use paired fields to distinguish between fluxes of natural v. drained conditions, it can be concluded that lowering the water table by only a few decimetres reduces CH 4 emissions in most cases by more than 90%, or even converts the peatland from a net source to a net sink of c n 4. On the other hand, CO2 release resulting from peat oxidation can reach levels exceeding those from mineral soils by one to two orders of magnitude. An emission range of 3.7-70 t CO2 ha -1 a q is given by Bergkamp & Orlando (1999) for the agricultural use of wetlands in boreal/temperate regions. Freibauer & Kaltschmitt (2001) estimated CO2 fluxes from drained organic soils to be 10 +_5 t CO2 ha q a -1 for ieys and grasslands, and 15 +_ 5 t for arable crops. The authors stress that their estimate is applicable to farming on organic soils, and not for the conversion of pristine peat soils to farmed soils. For temperate regions in Switzerland, Leifeld et al. (2005) estimated an annual loss of 35 __ 8 t CO2 ha q by peat oxidation from organic soils used for agriculture. Compared to the CO2 emission from drained peatlands, C accumulation by peat growth in intact peatlands is small. G o r h a m (1991) attributed a mean annual accumulation of c. 0.23 t C ha -1 to northern peatlands, or 0.076 Gt C globally for an approximated peatland area of 346 Mha. Global CH 4 emissions from rice agriculture amount to approximately 60 Mt CH4 (FAO 2003). This value lies within the range of estimates compiled by IPCC (2001), but is higher than the range of 25-54 Mt CH4 given by Sass et al. (1999). Rice agriculture increased almost steadily from 115 Mha in 1961 to 153 Mha in 2004 (it was 86 Mha in 1935) and increasingly contributed to the world's total agricultural area (2.56% in 1961, 2.94% in 2002). It is expected to increase to 164 Mha by 2030 (FAO 2003). Most of this area is wetland rice (around 85 %), which is considered to be the main source of CH 4 from rice agriculture (IPCC 1996). In contrast to the pronounced uncertainty with regard to global

36

J. LEIFELD

C H 4 emission from rice agriculture, various case studies address the drivers of C H 4 emission from wetland rice, and possible reduction strategies. Among the manageable factors known to affect C H 4 emission from rice fields, two are frequently cited to be of major importance:

(1) Water regime~irrigation status. For continuously flooded fields, IPCC (1996) gives a seasonally integrated emission factor (i.e. cumulative emission over the growing season) of 12-28 g C H 4 m -2. Because C H 4 production is strongly related to the redox conditions (Yu & Patrick 2004), the maximum default emission factor under continuous flooding is reduced by intermittently flooding, or when water depths are below 0.5 m (IPCC 1996). In a study evaluating the effect of rice field management on the composite effect of N 2 0 and CH4 emissions, Yu et al. (2004) noted the highest C H 4 emissions reductions, of more than 70%, for non-flooded but wet soil conditions; however, they observed that some of this reduction was offset by increased N 2 0 emissions. Park & Yun (2002) discussed measured C H 4 emissions at various sites in Korea. They showed that the mean emission factors were 0.24 g C H 4 m -2 d -1 for continuous flooding and 0.15 g C H 4 m -2 d -1 for intermittent flooding in situations where no organic matter was added. Emissions under direct seeding on dry paddies were 40-50% lower than those in transplanted rice. Integrated C H 4 emissions factors in transplanting cultures were 27 g C H 4 m -2 for early-maturing rice, but 35 g C H 4 m -2 for late varieties. During the Methane Asia Campaign, Mitra et al. (2002) revealed higher integrated emission factors of 5-29 g C H 4 m -2 and 22-57 g CH4 m -2 for fields with continuous flooding, but with different SOC contents, while the corresponding ranges for intermittently flooded fields were 0.06-3 g C H 4 m -2 and 0.6-24 g CH4 m -2, respectively. Emission reductions of >70% for non-flooding but wet conditions relative to flooding were reported by Yu et al. (2004). In a modelling study for India, Pathak et al. (2005) derived annual net emissions of 1.07-1.10 Mt C H 4 - C for continuous flooding of 42.25 Mha rice fields, and 0.12-0.13 Mt for the same area but under intermittent flooding. Under conditions of continuous irrigation, the depth of the water table also has a pronounced effect on CH4 emissions (Liu & Wu 2004).

(2) Organic matter status~organic matter addition. In most of the above-mentioned studies, SOM content and organic matter application were included as additional factors. Yu et al. (2004) reported an emission reduction of 57% for flooded fields without amendment, relative to the application of 30 t ha -1 a -1 organic manure. Liu & Wu (2004) proposed the removal of the rice straw (i.e. reducing the addition of organic matter) from the first harvest, before planting the second crop, thus reducing the CH4-emission by one order of magnitude. Significantly higher emissions were reported by Kimura et al. (2004), when senescent rice leaves at a quantity naturally occurring in the field were left in the field, in comparison to their removal. Mitra et al. (2002) showed that soils with lower SOC contents (<0.7%) emitted 5-29 g CH4 m -2 and 0.06-3 g C H 4 m -2 under continuous and intermittent flooding, respectively, and soils with a higher SOC content (>0.7%) emitted 22-57 g C H 4 m -2 and 0.6-24 g C H 4 m -2 under continuous and intermittent flooding during the growing season, respectively. In the study of Park & Yun (2002), the amendment of 5 t rice straw ha -1 in autumn increased the emission from 0.24 to 0.33 g CH4 m -2 d -1 (continuous flooding), and from 0.15 to 0.24 g CH4 m -2 d -a (intermittent flooding).

Integrated assessment of greenhouse gas fluxes from soil System analysis

An integrated assessment of any anthropogenic use of soil should account for all of the soil functions, and this equally applies to soils as sources and sinks of GHGs. Measures implemented to sequester carbon in soil often have benefits for other soil functions as well; however, their gross effect may be counterbalanced by changes in the exchange rate of N20 and CH4. Integrated assessments of G H G balances have been discussed by Robertson et al. (2000) and Robertson & Grace (2004), for example, for a maize/soybean/wheat cropping system in the US Midwest, and for a rice/wheat/cowpea cropping system in India. For the US system, annual crop rotations with three practices (conventional tillage, no tillage, organic without mineral N but with legumes in the rotation) were compared together with perennial systems (alfalfa and

SOILS AND GREENHOUSE GASES poplar) and forest successional communities (on a time-scale of 10-50 years). Net globalwarming potentials (GWP) were positive (i.e. net GHG source) for any of the annual systems, and neutral-to-highly negative for perennial crops and successional communities. They showed C sequestration to be of major importance for no tillage, perennial crops and early successions, and N 2 0 emission to be highest in the annual crops and for alfalfa. Pronounced CH 4 oxidation occurred mainly in late successional forests. For the Indian rice system, the net GWP was one order of magnitude greater, mainly due to CH 4 emissions from rice. Both studies indicated marked differences in net GWP, which must be known before management strategies aimed at reducing G H G emissions are implemented. Six et al. (2004) compiled data from soil-derived G H G emission comparisons between conventional and notillage strategies for humid- and dry-temperate climates, mainly from the northern hemisphere. They found the net GWP to be highly timedependent and net GHG mitigation to occur under no tillage in humid regions only when practised for more than 10 years. During the first years, no tillage led to high N20 emissions, and to a net loss of soil C in dry climates. A net C sequestration could only be ascertained in the long term (20 years). Cumulative soil-related G H G effects over 20 years indicated that no tillage outperformed conventional tillage only

37

in humid climates. They concluded that N 2 0 emissions drive much of the trend in net GWP, suggesting that improved nitrogen management is essential for realizing the full benefit of C sequestration. Further illuminative comments on the net CO2 balance of some mitigation measures can be found in Schlesinger (1999). Net GWP has been compared only rarely for peatlands under natural v. cultivated conditions. It has been shown that G H G fluxes in peatlands are typically much higher than those of mineral soils; that pristine peatlands are important emitters of CH4; and that they respond dynamically to drainage and cultivation measures. To assess the atmospheric impact of natural v. drained peatlands, a rough estimate of the integrated G H G fluxes of natural v. drained peatlands is provided. This calculation, based on the data discussed above, illustrates the directional flux under two different conditions: natural and managed. For CO2 fluxes, I assume a sink for intact peatlands of 0.84 t CO2 ha -1 a -1, and a mean source of 3 7 t C O z h a q a q under drainage, with an exponential rise of the emission to maximum over time. Methane emissions from the intact peatland are estimated to be 0.18 t CH 4 ha -1 a q, and are estimated to be zero after cultivation. Nitrous oxide emissions are assumed to be 0.5 kg N 2 0 - N ha -1 a -1 for intact peatland, and 5 kg N z O - N ha -1 a -1 for cultivated peatland at the beginning of drainage. Dinitrogen monoxide emission during drainage

9000 8000

Net emissions from cultivated peatland

~

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

CO2 plus N20 CO2 alone

7000 6000 5000

2

4000 3000

G u 2000 ,/"

1000

0

Net emissions from intact peatland . ,.

50

100

150

CH4 plus C-accumulatior --'--"--'peat growth

200

250

300

Years

Fig. 3. Cumulative greenhouse gas emissions of natural and cultivated peatlands (CO2-equivalents haq aq), estimated by means of typical flux rates. For details, see the text.

38

J. LEIFELD

is considered to behave proportionally to C O 2 (i.e. it declines along with the decline in annual CO2 emission rates) because C and N mineralization are tightly coupled. It is also assumed that the intact p e a t l a n d stores 2000 t C ha q (peat thickness 2 m; Batjes 1996), so that oxidative peat consumption ceases after 200 years. The results of this calculation are given in Figure 3, adopting global-warming potentials for a 100year horizon of 23 for CH4 and 296 for N20 (IPCC 2001). The curves clearly indicate that, over the selected time horizon, which is determined by the duration of peat decomposition, cultivation is highly unfavourable compared to leaving the peatland in its natural state. Methane emissions would need to accumulate for 1700 years to cause similar levels of global warming to those of the drained peatland over 200 years. Within the selected time horizon, annual CH 4 emission rates of approximately 1.59 t ha -1 would have a similar global-warming effect to that of the drained peatland. Such rates have b e e n r e p o r t e d sporadically (Augustin et al. 1996), but are considered to lie at the upper limit. Global estimates

Global estimates of the annual fluxes of C 0 2 , CH 4 and N20 from soils can be addressed tentatively with the data discussed in the previous paragraphs. This summary adopts the 40-60 Gt CO2-C lost by cultivation, 11 Gt of which are delivered from peatland drainage (Paustian et al. 1997), assuming that these changes occurred exclusively and steadily since 1850 (i.e. over a period of 150 years). Estimates of CH 4 emis-

sions from rice agriculture (currently 60 Mt), wetlands (115-237 Mt), from peatlands as a fraction of wetlands (46.5-73.3 Mt CH4), of N 2 0 emissions from agricultural soils (2.8 Mt N 2 0 - N ) and from natural soils (6.8 Mt NzO-N ) are taken from the data compiled above. Emission savings by peatland cultivation (39 Mha cultivated • 0.18 t CH4 ha q a q ) are calculated based on the data published by Paustian et al. (1997) and on the emission rate used to calculate the net flux of peatlands above. A value of 29 Mt CH 4 for global soil CH 4 oxidation is taken from Smith et al. (2000), and C sequestration by peat growth in intact peatlands from G o t h a m (1991). The annual fluxes are summarized in Table 3. Methane emissions from intact peatlands and other wetlands and N20 from natural soils are the most i m p o r t a n t single sources, followed by CH4 from rice agriculture; N 2 0 from agricultural soils; and CO2 from land-use change. The latter category is an estimated mean for the period since 1850, and fluxes may have been distinctly above or below the calculated range at any one point in time. If, for example, the soil CO2-flux is considered to be proportional to the total CO2-flux from land-use changes and land m a n a g e m e n t , annual rates may have been higher since the 1950s than before (see Houghton 2003 for comparison). The data imply that, on a global scale, soils are net G H G sources, and that this source function has been accelerated by human perturbation. Under natural conditions, soils emitted 4.86 to 7.67 Gt CO2-equivalents (2.65 to 5.45 Gt from wetlands minus 0.67 Gt from CH 4 oxidation minus 0.28 Gt by peat growth, plus 3.16 Gt

Table 3. Estimated global annual greenhouse gas fluxes for important soil-related sources (Mt C02 equivalents) CO 2 from land-use

change Mineral soils

Peatland drainage

710-1200

269

CO2 uptake CH 4 CH 4 rice CH4 intact CH 4 by peat emission agriculture peatlands oxidation4 growth savings 1 (all wetlands)

N20 from soils

N20 Natural agricultural soils6 soils5 -279

-162

1380

1070-16902 (2645-5451) 3

-667

1302

3160

1 Originating from peatland drainage. 2 Range only for intact peatlands. 3 Range for all wetlands. 4 By aerobic soils. 5 Fertilizer-induced. 6 Including background emission from agricultural soils. Based on Gorham (1991); Paustian et al. (1997); Cao et al. (1998); WBGU (1998); Spiers (1999); L. C. Smith et al. (2000); IPCC (2001); Bouwman et al. (2002); Moore (2002); FAO (2003). For the method of calculation, see the text. Negative values indicate sinks.

SOILS AND GREENHOUSE GASES as N20. Human-induced emissions are 0.71-1.20 Gt CO2 from land-use change, plus 0.27 Gt CO2 from peatland drainage, plus 1.38 Gt as methane from rice agriculture, plus 1.30 Gt as nitrous oxide from agricultural fertilization, minus 0.16 Gt methane emission savings deriving from peatland drainage. According to this rough estimate, emissions, including wetlands other than peatlands, have increased from 4.86-7.67 Gt to 8.36-11.65 Gt CO2-equivalents, namely by 51 to 84%. Total anthropogenic G H G emissions are approximately 33 Gt CO2-equivalents per year (including CH 4 and N20 from agriculture; WRI 2005). Annual soil fluxes due to land-use change and agriculture account for 3.50-3.99 Gt CO2 equivalents of the total, and are thus a significant contribution to the anthropogenic source. Not included in this examination are the naturally low sequestration rates that occur in soil along with profile development. Schlesinger (1990) indicated long-term rates of carbon accumulation of c. 2.4 g C m -2 a q, resulting in a global annual accumulation of 0.32 Gt C, or 0.40 Gt C if peat growth is also included. He points out that these rates probably overestimate actual ones, because the data include initial phases of soil developments after glacier retreat or volcanic eruptions, which typically have higher rates than older soils. He also stresses that such a global accumulation is consistent with global estimates on organic carbon transported by rivers. Hedges et al. (1997) give similar estimates on the global amount of organic carbon discharge to the sea (0.25 Gt dissolved C and 0.15 Gt particulate C). They conclude that much of this material is from forests and is soil-borne, and that there is growing evidence for rapid and remarkable extensive mineralization of terrestrial organic matter in the ocean. Together, these findings imply that, once the soil is in a steady state, much of the organic matter entering it will be discharged and subsequently mineralized in the ocean unless not oxidized in situ. Climate change is expected to affect the element balance of soils. Decreasing C stocks of temperate soils have been discussed already in the context of climate change (Bellamy et al. 2005), referring to potential CO2 sources and atmosphere-plant-soil interactions beyond the scope of this chapter. U n d e r conditions of warming, arctic and boreal regions are vulnerable to thawing. While thawing of permafrost soils seems possible (Waelbroeck et al. 1997), there is quite a lot of uncertainty about the direction and magnitude of potential warming impacts and feedback loops because of the

39

many interacting forces (Stokstad 2004). For example, thawing the permafrost potentially increases decomposition of stored C and thus may turn a system from a net sink to a net source, while on the other hand it may increase net primary production and thus enhance C inputs. Effects on both, CO2 and CH 4 emissions are also difficult to predict because they depend strongly on the hydrological situation.

Mitigation potentials A tentative attempt is made to discuss potentials for the reduction of global greenhouse-gas emissions from soils on the basis of three of the major strategies cited above. This approach does not claim to be complete, but is intended to stimulate discussions in terms of quantity and longevity of abatement. The three strategies are: (1) Carbon sequestration in soil of 1 0 - 3 0 G t C O 2 - C , according to Batjes (1999), for a timeframe of 50 years until a new steady state is reached (i.e. 0.73 to 2.2 Gt CO2 a-l), with an exponential rise to that point over time; (2) Annual reductions of N20 emissions from managed soils by 10% and 20% relative to the current level of c. 2 . 8 M t N 2 0 - N (Bouwman et al. 2002) taking into consideration an underlying annual increase in N fertilization of 1%, according to the FAO (2003), and assuming that global N20 emissions are proportional to N fertilizer application. A n annual reduction of 10-20% seems feasible in view of the much higher reduction reported in the literature; (3) Annual reduction of CH4 emissions from rice agriculture by 20 and 30% from current values of 60 Mt CH 4 and considering an increase in area to 164 Mha by 2030 (FAO 2003). The attainable annual reduction is small relative to that reported in most of the studies for measures of water and organic matter management, but it was chosen to account for the higher N 2 0 emissions reported in certain cases. The results of these strategies are summarized in Figure 4. Over the next 50 years, C sequestration will be the most effective measure, fixing between 37 and l l 0 G t C O 2 . Cumulative emission savings are 8.4-16.8 Gt CO2equivalents for improved management of N20, and 13-20 Gt CO2-equivalents for the reduction of CH4 from rice agriculture. For a scenario with small C sequestration rates, combined N20 plus CH 4 savings are in the same order of magnitude as C sequestration after 50 years. Mean annual

40

J. LEIFELD

Fig. 4. Cumulative global CO 2 mitigation potentials of soils. (la), (2b) C sequestration; (2a) reduction in N20 emissions by 10%; (2b) reduction in N20 emissions by 20%; (3a) reduction in CH 4 emissions 10%; (3b) reduction in CH 4 emissions by 30% (Gt CO2 equivalents). For further explanations, see the text.

emission savings (i.e. CO 2 plus N 2 0 plus CH4) amount to 1.2-2.9 Gt CO2-equivalents, which is a significant share of the anthropogenic emission of 3.5-4.0 Gt CO2-equivalents attributed to soils. Considering that conservation tillage is one of the measures contributing to the C sequestration capacity, it must be noted that the potential for N 2 0 emission reduction will probably be smaller because of higher N 2 0 emissions under no tillage (Six et al. 2004). While the pattern for these three strategies seems obvious for the next few decades, the relative contribution of emission savings from N 2 0 and CH4 will increase over time, because of the finite potential of soils to store C. Long-term investigations should thus specify the interaction between Csequestration and fluxes of the other GHGs, and should focus on emission savings of the nonCO2-GHGs. Compared to soil C sequestration, their cumulative potential is infinite. Peatland management has not been included in the discussions of mitigation potentials, because data on the area of cultivated peatlands and their future projection are highly uncertain. However, the integrated assessment of G H G fluxes in Figure 3 has shown that it would be reasonable to protect intact peatlands, not only in view of hydrological and ecological functions, but also with respect to global-warming mitigation. Cultivation of 1% of the actual peatland area would imply additional mean annual emissions (emissions at the beginning of cultivation are supposed to be higher) of 96-147 Mt CO2,

and a reduction in CH 4 emissions of only 11-17 Mt CO2-equivalents.

Conclusions On the global scale, soils are net sources of G H G s under natural conditions as well as under conditions of land-use change and agricultural intensification, particularly with respect to the large amounts of N 2 0 and CH 4 that they emit. Greenhouse-gas fluxes from soils are important budget components in the terrestrial elemental cycles, and are also significant if the overall anthropogenic emissions are considered. H u m a n activity has increased soil-derived fluxes, and it is likely that, despite mitigation potentials yet to be implemented, fluxes will remain above the pre-agricultural level. This is mainly induced by the elevated CH4 and N 2 0 emissions, while p r o p e r agricultural land management is qualified to maintain and restore S O M levels in many cases. There are enormous uncertainties in estimates related to (1) fluxes, particularly for CH 4 from intact and for CO2 from drained peatlands, and (2) areas, particularly of cultivated peatlands. There are also gaps in our understanding of the fate of SOC transferred by erosion, sealing and desertification, as well as in the role of black carbon formation. Beyond what has been addressed in this paper, the responses of soil-related G H G fluxes to climatic changes are crucial for the future role of soils in the global element cycles.

SOILS A N D G R E E N H O U S E GASES I thank R. Rees and Changming Fang for helpful comments on an earlier draft of the manuscript. This work is dedicated to the memory of my father.

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Soil as an important interface between agricultural activities and groundwater: leaching of nutrients and pesticides in the vadose zone L. E B E R G S T R O M

1 & E DJODJIC 2

1Department o f Soil Science, Swedish University o f Agricultural Sciences, P.O. Box 7014, SE-750 07, Uppsala, Sweden (e-maik lars. [email protected]) 2Swedish Environmental Research Institute (IVL), P O Box 21060, SE-IO0 31, Stockholm, Sweden Abstract: Agricultural non-point source pollution by plant nutrients and pesticides can

cause severe environmental disturbances, such as deterioration in the quality of surface water and groundwater. In order to prevent this, the development and implementation of appropriate countermeasures are necessary, which requires knowledge of critical soil functions in the vadose zone. The nutrient source, the transport pathway and the availability of solutes in soil are some important conditions that affect leaching. Our results show that organic nitrogen sources are often more susceptible to leaching than inorganic N fertilizers, due to poor synchronization between the N demand of the crop and the release of inorganic N from the organic N source. Large amounts of leachable N are left in the soil after the growing season. Preferential flow, in combination with where soil solutes occur, is critical for establishing safe loading rates. In some cases, the solute is located in smaller pores of the soil matrix, and is thereby protected against preferential flow and leaching. In other cases, especially soon after application of a fertilizer or pesticide, transient flow-peaks rapidly displace the solutes through macropores in the vadose zone, which can cause large leaching loads and associated water-quality problems.

The potential negative effects of various h u m a n activities on soils and natural waters has been a topic of concern for several decades. Due to problems of o v e r e x p l o i t a t i o n of agricultural land, soils suitable for production of healthy and nutritious food are disappearing rapidly in many parts of the world, especially in developing countries (Buresh et al. 1997). Adverse effects on soils and waters include: deterioration in drinking-water quality, surface waters and groundwater caused by plant nutrients (Kirchmann et al. 2002); accumulation of pesticides in the soil to toxic levels (Torstensson & Stenstr6m 1990), and redistribution of e r o d e d surface material (Lal 1990). As a result of the increasing c o n c e r n over food and e n v i r o n m e n t a l quality, a transition towards more sustainable soil use has been established in many different documents (e.g. Bergstr6m & Goulding 2005) as a common goal for world society. Within the agricultural sector, this transition is urgently needed, since agricultural food production is fundamental to the survival of a rapidly growing world population. Most of the diffuse contamination of surface waters and groundwater by nutrients and pesticides originates from h u m a n activities on agricultural soils. In the transition towards more sustainable use of agricultural soils, management practices

to reduce emissions that have a p o t e n t i a l negative impact on the environment have been developed (Bergstr6m & Goulding 2005). In this work, several guiding principles are proposed, for example, environmental indexing of fields and consideration of spatial variability within fields in relation to their contribution of contaminating soils, surface waters and groundwater within a drainage basin (Lemunyon & Gilbert 1993). A n o t h e r r e c o m m e n d a t i o n is reduction of nutrient inputs to the soil to levels slightly below those expected to give the m a x i m u m yields, by applying less fertilizer (Lord & Mitchell 1998) and by a further reduction in animal density (Sims et al. 2005), with the goal of obtaining equilibrated nutrient balances at the field level. With these guiding principles, existing solutions that limit contamination of soils and groundwater, such as the use of cover crops to reduce losses of N (Bergstr6m & Jokela 2001), buffer strips, and conservation tillage to reduce P losses (Withers & Jarvis 1998), can be improved further. Nevertheless, for successful application of countermeasures, we need a thorough understanding of soil functions in the vadose zone. The following topics are discussed in this paper to describe various soil functions and how they affect transport of solutes in the vadose

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. R (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 45-52. 0305-8719/06/$15 9 The Geological Society of London 2006.

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zone of agricultural soils: (1) the importance of the type of source, here exemplified by the difference between inorganic and organic N sources; (2) how transport pathways affect leaching of solutes (plant nutrients and pesticides); and finally (3) the need to consider the location of chemicals in the soil, in order to assess their environmental fate and potential for degradation and leaching. The causes of pollution are explained briefly in a practical context, and some existing measures for the reduction of agricultural non-point source pollution are described and evaluated. The majority of the examples presented in this overview are selected from our own research, which explains why we focus only on some of the many important processes determining the fate of solutes in soil.

Leaching and plant uptake of N from different nitrogen sources Converting conventional agricultural production to organic farming is considered by many to be an effective way of minimizing agricultural non-point source, or diffuse, pollution by plant nutrients, especially nitrogen. In organic farming systems, the nutrient requirements of plants are satisfied by using organic manures or mineral fertilizers with very low solubility (e.g. by apatite to supply phosphorus). However, for such nutrient sources to be efficient, they have to deliver soluble inorganic nutrients at a time when the crop needs them, which is usually in early summer in cold, humid regions such as Sweden. If the nutrients are released too late in the growing season, or after the crop has been harvested, there is a risk that they will leach through the vadose zone and thereby contaminate the groundwater. Mineralization of organic N is a biological process that is controlled to a large extent by the C/N ratio of the organic substrate, but also by temperature and soil moisture content. During autumn in cold and humid regions, the soil temperature and moisture content are high enough to trigger the release of inorganic N from soil organic N fractions, and this N is exposed to leaching from an often large surplus amount of precipitation. It has been shown in several studies that leaching of N from organic manures can be greater than N leaching from inorganic N fertilizers, when both are applied at similar rates (e.g. Thomsen et al. 1997; BergstrOm & Kirchmann 1999, 2004). In the study by BergstrOm & Kirchmann (1999), in which leaching of N from poultry manure and

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Fig. 1. Leaching and plant uptake of nitrogen derived from inorganic N fertilizer (ammonium nitrate; AmNi) and poultry manure stored in different ways (fresh, anaerobic or aerobic), measured in 0.3-m-diameter and 1-m-long lysimeters (modified from Bergstr6m & Kirchmann 1999).

NH4NO3 was measured in field lysimeters over a three-year period, it was found that the N losses derived from poultry manure were almost one order of magnitude larger than those from NHaNO 3 (Fig. 1). This was attributed to poor synchronization between the demand for N by the crop and the release of N from the manure. In other words, considerable amounts of manure N were released during mild and wet periods when there was no crop uptake of N. The importance of climatic conditions during the non-cropped (autumn/winter/spring) period on N leaching was also shown in a nine-year lysimeter study carried out by Dressel et al. (1992). Years with mild winters resulted in greater N leaching from manure than from inorganic fertilizers. However, even though the above-mentioned examples are quite convincing, other studies have shown that leaching of N does not increase when organic manures are used, and in some cases it has been shown that N leaching is lower when manures are used instead of inorganic N-fertilizers (e.g. Eltun & Fugleberg 1996). This is often attributed to the fact that not only the N sources are different, but also the cropping systems. In the study by Eltun & Fugleberg (1996), the crop rotation with inorganic fertilizer included two years when potatoes were grown, whereas one year's worth of potatoes were included when manure was used. During years when potatoes were grown, N leaching loads are typically large (Modramootoo et al. 1992). How then can we improve the N-use efficiency of organic N sources, so that crop needs are satisfied and leaching is avoided? Anaerobic digestion of organic manures is a

LEACHING IN THE VADOSE ZONE method by which about 50% of the total carbon in raw manures is transformed into a valuable byproduct - biogas (methane c. 60-65%, carbon dioxide c. 35%), leaving a more mineralized, nitrogen-rich liquid fertilizer that can be applied to crops, similar to synthetic inorganic fertilizers (Tafdrup 1995). Another way in which it is possible to efficiently use N and other nutrients in manure is to recover the organically bound nutrients in inorganic form by ultramicrofiltration (Cicek 2003), followed by various concentration processes (Kirchmann et al. 2005). The product obtained can then also be used as an inorganic fertilizer and be transported long distances to farms that are dependent on purchased nutrient inputs. A third approach could be to modify the organic materials. Current research is focusing on the possibility of manipulating manure quality and quantity through additions of various carbonrich and nitrogen-poor waste products (e.g. primary fibre sludge from paper-mills; Ziebilske 1987; Vinten et al. 1998; Kirchmann & BergstrOm 2003). For green manures, another possibility would be to manipulate their C/N ratio by varying their phenol content, or the types and concentrations of carbohydrates (Gunnarsson & Marstorp 2002).

The importance of flow pathways for transport of solutes in soil A condition that complicates the flow of solutes through the vadose zone is that water, and nutrients and pesticides dissolved in the water phase, often move through large pores in soil (e.g. earthworm and root channels, cracks, etc.), a process commonly referred to as preferential flow. Under such conditions, an equilibrium solute concentration throughout the soil profile cannot be obtained. This phenomenon occurs primarily in fine-textured soils with high clay contents, especially those that have the potential to swell, shrink and crack. Preferential flow can also occur in sandy soils from wetting-front instability, often referred to as finger flow (Ritsema et al. 1998). Fingering is associated with layered soils (fine- over coarse-textured layers), hydrophobic soils, air compression, and redistribution effects caused by ponding (Hillel 1987). In addition to the vertical flow component, preferential flow is also characterized by being non-equilibrium transverse to the main flow direction (Flt~hler et al. 1996). It is important to note that, irrespective of the reason for preferential flow, solute concentrations are hardly ever in equilibrium throughout

47

a soil profile; there is generally non-equilibrium along the depth gradient in most soils. Through preferential flow, solutes such as pesticides can be transported rapidly through large vertical portions of the vadose zone and bypass biologically and chemically active layers in which they would otherwise be degraded or sorbed. Exposure to preferential flow is most pronounced soon after application of a fertilizer or pesticide, when high concentrations occur in the soil solution in upper soil layers, in combination with intensive rainfall (Bergstr/3m & Stenstr6m 1998). Losses of P from agricultural soils are considered to occur mainly through surface runoff, with P bound to soil particles (Sharpley & Rekolainen 1997). However, during the last couple of decades, leaching has also been recognized as a possible transport pathway (Heckrath et al. 1995), primarily of dissolved R It then becomes a critical issue whether preferential flow or flow through the soil matrix dominates in displacing P through the vadose zone. The importance of flow behaviour for shortterm P losses in a clay and a sandy soil was investigated in a study by Djodjic et al. (1999). Leaching was measured for 29 days after application of 100 kg P ha q at the surface of 1-mlong lysimeters filled with undisturbed soil columns. Water equivalent to 100 m m was added on five occasions with seven days between each watering event, giving a total water input of 500 mm. It is important to stress that this water input is similar to the total annual precipitation for large parts of Sweden and many other countries in the temperate region. In reality, natural rainfall in those countries is not typically produced by the large storms that would cause ponded conditions. Therefore, under natural conditions the likelihood of preferential transport would be reduced considerably in comparison to this study. The added P fertilizer was labelled with the radioactive isotope 33p, so it was possible to distinguish whether the P in drainage-water originated from added P or soil R The results of the study showed that leaching was considerably higher in the clay soil (4 kg P ha q) than in the sandy soil (56 g P haq), which was largely attributed to preferential flow in the well-structured clay soil (Fig. 2). Of the 4 kg ha -1 leached from the clay soil, 72% was derived from the added P fertilizer. Another interesting observation was that, over time, P leaching decreased after each watering event in the clay soil, whereas it increased in the sandy soil, although P leaching was overall much less in the sandy soil, primarily due to high levels of iron in lower

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Fig. 2. Average cumulative P leaching from a clay and a sandy soil during a four-week period when a total of 500 mm of water was applied. The study was performed in 0.3-m-diameter and 1-m-long lysimeters (modified from Djodjic et al. 1999).

Fig. 3. Concentrations of a pesticide in water leaching from 1-m-long undisturbed soil columns of a clay and a sandy soil (modified from Bergstr6m & Stenstr6m 1998).

layers of the sand profile (Ghorayshi & Bergstr6m 1991). This flow pattern illustrates the importance of incidence of preferential flow in relation to P application. Once the solute (P) is mixed in with the soil matrix, water moving through preferential flow paths does not interact with soil, and leaching is therefore reduced. In other words, preferential flow can both increase and decrease leaching of solutes, depending on the time when it occurs in relation to application of the solute (fertilizer or pesticide). The final result is often that, although initial P losses are large due to elevated solute concentrations in water leaching through preferential flow paths in clay soils, the leaching loads over extended periods are often quite small in such soils. Indeed, leaching loads are typically larger in sandy soils, in which water movement occurs between individual soil particles within the main soil matrix (Bergstr6m & Stenstr6m 1998). The results from the pesticide leaching study performed in undisturbed lysimeters shown in Figure 3 illustrate this quite well; the peak concentration (24 pg L -1) was 5.6 times higher in the clay soil, whereas the accumulated leaching load was 2.2 times larger from the sand lysimeters. A similar explanation to that described above is given by Larsson & Jarvis (1999a) in a model evaluation of results from a leaching study with bentazone (3-isopropyl-lH-2,1,3benzothiadiazine-4(3H)-one 2,2-dioxide). That study was performed in tile-drained plots at the field site in south-west Sweden where the clay lysimeters described above for measurements of P leaching were collected. Their conclusion was that macropore flow reduced leaching by c. 50% for this weakly sorbed herbicide, because with

long intervals, the compound was stored in micropores and was thereby protected against rapid flow in macropores and was instead exposed to a slow convective transport. In a similar way, Larsson & Jarvis (1999b) showed that nitrate leaching during an eight-year period at the same site was reduced by c. 28% due to macropore flow. Occasional increases in leaching caused by macropore flow following fertilization on the soil surface were outweighed by the reduction in leaching of mineralized N during autumn and winter. This conclusion is also supported by the results of a lysimeter experiment on the same clay soil presented by BergstrOm (1995), which showed that transient peaks of fertilizer N leaching were overshadowed by general convective-dispersive movement. A dye tracer study was carried out (Bergstr6m & Shirmohammadi 1999) to evaluate the areal extent of pores participating in displacement of water and solutes with profile depth in the clay and sandy soils used in the leaching study with P presented above. Water (corresponding to 100 mm) stained with acid red (azophloxine) was applied to the same type of 1-m soil columns as those used for P leaching measurements. Subsequently, they were cut into 0.1-m increments and each transect was examined for the areal extent of stained soil. The results revealed that a considerably larger area was stained at each depth in the sandy soil as compared with the clay soil (Fig. 4). In the clay subsoil (below 30-cm depth), the mean stained area was less than 10% under transient flow conditions, whereas in the sand it was around 70%. Despite this, the arrival time of stained solution was considerably faster in the

LEACHING IN THE VADOSE ZONE

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Availability of chemicals in soil determines their environmental fate

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Fig. 4. Cross-sectional representation (0.3-m diameter) of the areal extent and location of flow paths at different depths in a sandy soil (left) and a clay soil (right) (modified from Bergstr6m & Shirmohammadi 1999). clay soil; in fact, only clear water eluted in the sand monoliths throughout the experiment. All the studies referred to above were conducted in 1-m soil profiles, in which some of the preferential flow paths may be continuous. However, in deeper profiles this is uncommon. In such cases, preferential flow paths likely end somewhere in the soil matrix, and readily sorbed solutes like P and most pesticides would be retained in the subsoil when moving vertically in soil. However, it has been shown that earthworm channels can extend down to 3-m depth in some soils (Edwards et al. 1989), and cracks have been observed as deep down as 4 m in a Danish soil (JCrgensen & Fredericia 1992). It has also been shown that preferential flow paths are more continuous in the subsoil than near the surface (Grevers & de Jong 1994), which is also an indication that preferential flow of solutes may occur to considerable depth in some soils.

The influence of preferential flow in relation to application of fertilizer or pesticide is primarily determined by its availability for leaching, in other words its location in the soil in relation to when and where water is flowing through the vadose zone. For pesticides, which are primarily degradable organic compounds, availability to microbes also determines persistence in soil. Soil is a complex and heterogeneous environment and its variability has a large impact on soil liquids, inorganic and organic solutes, gases, and also on micro-organisms. The smallest pores in soil (<0.1 pm) have a restricted exchange of water and solutes with the larger pores, and are mostly water-filled. Nevertheless, passage and transport of water and pesticide molecules are possible into these micropores, whereas entrance of micro-organisms is not; the microbes are simply too big. It is notable that those pores which are inaccessible to microorganisms can constitute a large proportion of the pore volume in fine-textured soils. For example, Adu & Oades (1978) showed that at least 90% of the pores in the surface layers of a clay-loam soil are not accessible to micro-organisms or enzymes. In other words, despite the fact that hundreds of millions of micro-organisms are present in a gram of soil, most of its pores and surfaces are sterile. Pesticides located in such pores are therefore physically protected from microbial degradation. The rate-limiting process regulating degradation is therefore diffusion of the pesticide molecules out of small pores. An experiment showing the effect of poresize distribution on degradation of the herbicide 2,4-D ((2,4-dichlorophenoxy)acetic acid) was carried out by Stenstr6m (1989). An inflection point in the degradation curve was obtained after about one week, when the rate of degradation drastically decreased, leading to very long persistence time for the remaining 2,4-D. The very slow degradation kinetics of this remaining 2,4-D indicated that diffusion of the herbicide stored within the micropores to active degradation sites in larger pores could be ratelimiting.

Minimizing agricultural non-point-source pollution In addition to the suggestions on how to improve the N-use efficiency of organic N sources mentioned above, there are several

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management practices that can be used at the farm level today to control non-point source pollution from nutrients and pesticides that are leaching from agricultural soils. Ongoing research within this sector will provide future agriculture with even better countermeasures to further reduce the pollution pressure. In such efforts, it is important to keep in mind that proposed new practices must be adaptable and adopted by farmers. As indicated above, the most critical period for large leaching loads in cold and humid areas is during the non-cropping season. The use of various catch crops or cover crops sown simultaneously with the main crop in spring or after harvest in autumn, has been the subject of much interest as a countermeasure against nitrate leaching. Many studies of catch crops have shown that they can take up large amounts of N, and thereby reduce N leaching considerably (e.g. Martinez & Guiraud 1990; Bergstr6m & Jokela 2001). Indeed, reductions in leaching loads exceeding 80% have been shown when comparing cereal cropping systems with and without a catch crop (Lewan 1994). However, the long-term effects of using catch crops repeatedly are not fully known. Much of this uncertainty is related to the large amounts of organic material that are incorporated into soil under such conditions, and how this affects release of inorganic N. Another uncertainty is related to P losses when large quantities of catch-crop residues are incorporated into soil. Increased leaching of P has been observed in many studies as a result of freezing fresh plant material with the consequent disruption of cell membranes (e.g. Miller et al. 1994). Nevertheless, there is reason to believe that catch crops will be used increasingly in the future as an effective measure to reduce N leaching. Other management options to reduce N leaching include 'controlled-release' forms of N fertilizers, which have been shown to improve N-use efficiency and reduce N leaching, but also to cut labour and application costs and to increase yields (Giller et al. 2004). This is also applicable to other plant nutrients. Minimum tillage practices have been discussed in the context of reducing N leaching (Goss et al. 1993). They also reduce erosion, and thereby P losses, and consumption of fossil fuels. Conventional mouldboard ploughing in autumn disturbs the soil, and new aggregate and particle surfaces are exposed to microbes with increased mineralization of organic matter. Minimum tillage leaves the soil less disturbed, which decreases mineralization and hence the production of leachable nitrate. A possible negative

effect of minimum or no-tillage practices on leaching of nutrients and pesticides is the fact that macropores in the topsoil remain intact. Such large pores can act as preferential flowpaths through which solutes may bypass the root zone, going to deeper soil layers and possibly also to groundwater. In addition to the P-leaching study by Djodjic et al. (1999) presented above, a follow-up experiment was carried out to examine whether a reduction in P leaching could be obtained by disrupting the continuity of macropores by tillage (Djodjic et al. 2002). However, it was clear that tillage was not enough. The added P fertilizer had to be incorporated and mixed with the topsoil to significantly reduce leaching, rather than being applied on the soil surface. These few examples show that much can be done to reduce agricultural non-point-source pollution. However, irrespective of the solute (nutrients or pesticides), the most important condition to keep in mind to reduce the risks of large leaching losses is to focus on the amounts applied to soils. In many cases, this would require a reduction of the nutrient inputs to levels slightly below those giving maximum yields, in order to obtain nutrient balances that are either equilibrated or slightly negative, and for pesticides, application only when absolutely necessary.

Conclusions Understanding the interactions between different processes in soil is the key to developing appropriate countermeasures to reduce agricultural non-point-source pollution. I n this overview we showed that the nutrient source, transport pathway and availability of solutes in soil have a strong influence on the loading rates of nutrients and pesticides that have to be considered when attempting to minimize leaching losses. Management practices that can be used at the farm level include whole farming systems, such as organic farming. However, reduced leaching is not a question of organic or conventional agriculture, but rather of modifications to existing management methods, and the introduction and use of appropriate countermeasures such as catch crops, minimum tillage, etc. Despite all our efforts, we have to keep in mind that it is virtually impossible to produce food and other agricultural products without some losses of nutrients and pesticides to groundwater and surface waters. These losses are often larger from agricultural systems than from natural terrestrial ecosystems. The major

LEACHING IN THE VADOSE Z O N E

challenge for the future is to minimize such losses while maintaining high crop productivity and quality. References ADU, J.K. & OADES, J.M. 1978. Physical factors influencing decomposition of organic materials in soil aggregates. Soil Biology and Biochemistry, 10, 109-115. BERGSTROM, L. 1995. Leaching of dichlorprop and nitrate in structured soil. Environmental Pollution,

87,189-195. BERGSTROM, L. & GOULDING, K. 2005. Perspectives and challenges in the future use of plant nutrients in tilled and mixed agricultural systems. Ambio, 34, 283-287. BERGSTROM, L.E & JOKELA, W.E. 2001. Ryegrass cover crop effects on nitrate leaching in spring barley fertilized with 15NH415NO3. Journal of Environmental Quality, 30, 1659-1667. BERGSTROM, L.E & KIRCHMANN,H. 1999. Leaching of total nitrogen from nitrogen-15-1abeled poultry manure and inorganic nitrogen fertilizer. Journal of Environmental Quality, 28, 1283-1290. BERGSTROM, L. & KIRCHMANN,H. 2004. Leaching and crop uptake of nitrogen from nitrogen-15-labeled green manures and ammonium nitrate. Journal of Environmental Quality, 33, 1786-1792. BERGSTROM, L.E & SHIRMOHAMMADI,A. 1999. Areal extent of preferential flow with profile depth in sand and clay monoliths. Journal of Soil Contamination, 8, 637-651. BEROSTROM, L. & STENSTROM,J. 1998. Environmental fate of chemicals in soil. Ambio, 27, 16-23. BURESH, R.J., SANCHEZ, RA. & CALHOUN, F. (eds) 1997. Replenishing Soil Fertility in Africa. Soil Science Society of America, Madison, WI, Special Publications, 51, 251 pp. CICEK, N. 2003. A review of membrane bioreactors and their potential application in the treatment of agricultural wastewaters. Canadian Biosystems Engineering, 45, 637-649. DJODJIC, E, BERGSTROM, L., ULFrN, B. & SHIRMOHAMMADI,A. 1999. Mode of transport of surface-applied phosphorus-33 through a clay and a sandy soil. Journal of Environmental Quality, 28,1273-1282. DJODJIC, E, BERGSTROM, L. & ULI~N, B. 2002. Phosphorus losses from structured clay soils in relation to tillage practices. Soil Use and Management, 18, 79-83. DRESSEL, J., WEIGELT, W. & MOCKEL, D. 1992. Langjfihrige Untersuchungen tiber die Wirkung von Stickstoff aus Mineraldtingung und Stallmist (Lysimeterversuche). Agribiological Research, 45, 177-185. EDWARDS, W.M., SHIPITALO, M.J., OWENS, L.B. & NORTON, L.D. 1989. Water and nitrate movement in earthworm burrows within long-term no-till cornfields. Journal of Soil Water Conservation, 44, 240-243. ELTUN, R. & FUGLEBERG, O. 1996. The Apelsvoll cropping system experiment. VI. Runoff and

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nitrogen losses. Journal of Agricultural Science (Norwegian), 10, 229-248. FL~3HLER, H., DURNER, W. & FLURV, M. 1996. Later solute mixing processes - a key for understanding field-scale transport of water and solutes. Geoderma, 70, 165-183. GHORAYSHI, M. & BERGSTROM, L. 1991. Equilibrium studies of the adsorption of dichlorprop on three Swedish soil profiles. Swedish Journal of Agricultural Research, 21, 157-163. GILLER, K., CHALK,P., DOBERMANN,A., HAMMOND,L., HEFFER, P. LADHA, J.K., NYAMUDEZA,P., MAENE, L., SSALI, L. & FRENEY, J. 2004. Emerging technologies that will increase the efficiency of use of fertilizer nitrogen. In: MOSIER, A.R., SYERS,J.K. & FRENEY, J.R. (eds) Agriculture and the Nitrogen Cycle: Assessing the Impacts of Fertilizer Use on Food Production and the Environment. Island Press, Washington, 35-51. Goss, M.J., HOWSE, K.R., LANE, P.W., CHRISTIAN,D.G. & HARRIS, G.L. 1993. Losses of nitrate-nitrogen in water draining from under autumn-sown crops established by direct drilling or mouldboard ploughing. Journal of Soil Science, 44, 35-48. GREVERS, M.C.J. & DE JONG, E. 1994. Evaluation of soil pore continuity using geostatistical analysis on macroporosity in serial sections obtained by computed tomography scanning. In: ANDERSON, S.H. & HOPMANS, J.W. (eds) Tomography of Soil-Plant-Root Processes. Soil Science Society of America, Madison, WI, Special Publications, 36, 73-86. GUNNARSSON,S. & MARSTORP,H. 2002. Carbohydrate composition of plant materials determines the pattern of N mineralization during decomposition in soil. Nutrient Cycling in Agroecosystems, 62, 1983-1991. HECKRATH, G., BROOKES, RC., POULTON, ER. & GOULDING, K.W.T. 1995. Phosphorus leaching from soils containing different phosphorus concentrations in the Broadbalk experiment. Journal of Environmental Quality, 24, 904-910. HILLEL, D. 1987. Unstable flow in layered soils: A review. Hydrological Processes, 1, 143-147. JORGENSEN, RR. & FREDERICIA,J. 1992. Migration of nutrients, pesticides and heavy metals in fractured clayey till. Geotechnique, 42, 67-77. KIRCHMANN,H. & BERGSTROM,L. 2003. Use of papermill wastes on agricultural soils: is this a way to reduce nitrate leaching? Acta Agriculturae Scandinavica (Sect. B), 53, 56-63. KIRCHMANN, H., JOHNSTON,A.E.J. & BERGSTROM, L. F. 2002. Possibilities for reducing nitrate leaching from agricultural soils. Ambio, 31, 404-408. KIRCHMANN, H., NYAMANGARA,J. & COHEN,Y. 2005. Recycling of municipal wastes in the future - from organic to inorganic forms? Soil Use and Management, 21, 152-159. LAL, R. 1990. Soil erosion and land degradation: the global risk. Advances in Soil Science, 11, 129-172. LARSSON, M.H. & JARVIS,N.J. 1999a. Evaluation of a dual-porosity model to predict field-scale solute transport in a macroporous soil. Journal of Hydrology, 215, 153-171.

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LARSSON, M.H. & JARVIS,N.J. 1999b. A dual-porosity model to quantify macropore flow effects on nitrate leaching. Journal of Environmental Quality, 28, 1298-1307. LEMUNYON, L.E. & GILBERT, R.G. 1993. The concept and need for a phosphorus assessment tool. Journal of Production Agriculture, 6, 483-496. LEWAN,E. 1994. Effects of a catch crop on leaching of nitrogen from a sandy soil: simulations and measurements. Plant and Soil, 166, 137-152. LORD, E.I. & MITCHELL, R.D.J. 1998. Effect of nitrogen inputs to cereals on nitrogen leaching from sandy soils. Soil Use and Management, 14, 78~83. MARTINEZ, J. & GUIRAUD, G. 1990. A lysimeter study of the effects of a ryegrass catch crop, during a winter wheat/maize rotation, on nitrate leaching and on the following crop. Journal of Soil Science, 41, 5-16. MILLER, M.H., BEAUCHAMP, E.G. & LAUZON, J.D. 1994. Leaching of nitrogen and phosphorus from the biomass of three cover crop species. Journal of Environmental Quality, 23, 267-272. MODRAMOOTOO, c.m., WAYO, K.A. & ENDRIGHT, P. 1992. Nutrient losses through tile drains from potato fields. Applied Engineering and Agriculture, 8, 639~546. RITSEMA, C.J., DEKKER, L.W., NIEBER, J.L. & STEENHUIS, T.S. 1998. Modeling and field evidence for finger formation and finger recurrence in a water repellent sandy soil. Water Resources Research, 34, 555-567. SHARPLEY,A.N. & REKOLAINEN,S. 1997. Phosphorus in agriculture and its environmental implications.

In: TUNNEY, H., CARTON,O.T., BROOKES, EC. & JOHNSTON, A.E. (eds) Phosphorus Loss from Soil to Water. CAB International, Wallingford, UK, 1-53. SIMS, J.T., BERGSTROM, L., BOWMAN,B.T & OENEMA, O. 2005. Nutrient management for intensive animal agriculture: policies and practices for sustainability. Soil Use and Management, 21, 141-151. STENSTROM,J. 1989. Kinetics of decomposition of 2,4,dichlorophenoxyacetic acid by Alcaligenes eutrophus JMPI134 in soil. Toxicity Assessment, 4, 405--424. TAFDRUe, S. 1995. Viable energy production and waste recycling from animal manure digestion of manure and other biomass materials. Biomass and Bioenergy, 9, 303-314. THOMSEN, I.K., KJELLERUP,W. & JENSEN,B. 1997. Crop uptake and leaching of 15N applied in ruminant slurry with selectively labelled faeces and urine fractions. Plant and Soil, 197, 233-239. TORSTENSSON,L. & STENSTROM,J. 1990. Persistence of herbicides in forest nursery soils. Scandinavian Journal of Forest Research, 5, 457-469. VINTEN, A.J.A., DAVIS,R., CASTLE,K. ~1; BAGGS,E.M. 1998. Control of nitrate leaching from a nitrate vulnerable zone using paper mill waste. Soil Use and Mangement, 14, 44-51. WITHERS, P.J.A. & JARVIS, S.C. 1998. Mitigation options for diffuse phosphorus loss to water. Soil Use and Management, 14, 186-192. ZIEBmSKE, L.M. 1987. Dynamics of nitrogen and carbon in soil during papermill sludge decomposition. Soil Science, 143, 26-33.

Understanding of a soil system derived from a single bed-rock, for improved vineyard management in Southern France M. D O S S O 1., O. P H I L I P P O N 2 & A . R U E L L A N 3

1CNEARC, Centre National d'Etudes Agronomiques des ROgions Chaudes, 1101 Avenue Agropolis, B P 5098, 34033, Montpellier, Cedex 01, France (e-mail: [email protected]) 2Conseil Environnement, Terroir et Viticulture, 140 Rue P. Larousse - Clos de Belvezet, 34090, Montpellier, France 32 Boulevard Berthelot, 34000, Montpellier, France Abstract: The 'Massif de La Clape' in Southern France is a calcareous hill with a semi-arid Mediterranean climate. All the agricultural lands are under vineyards managed by different chateaux. The objective of this work was to understand the soil system of Chateau X to solve a problem of vineyard mortality and to show how the exceptional variability of the soil conditions could be taken into account for optimal vineyard management. The soils of Chateau X are derived from sandstone that fills a basin of about 40 ha, surrounded by a calcareous formation. The study of the soil system showed the presence of four soil domains, two located at the top of the basin and mainly derived from eluviation, and two located at the bottom of the basin, which are mainly derived from illuviation. The domain located at the very bottom of the basin showed an accumulation of swelling clays in its BtM horizon. The zones of highest vine mortality were located at the bottom of the basin, on the soil with the BtM horizon. This mortality was explained by the strong discontinuity between the E and the BtM horizons, with the accumulation of Cu at toxic concentrations as a consequence of lateral water flows. In this soil, root growth was hindered by the high bulk density of the BtM horizon. In order to solve the problem of vine mortality and manage the vineyard according to the different soil conditions observed in the toposequence, several proposals are made.

The Massif de la Clape (an area 2 0 k m • 25 km), is located near N a r b o n n e in Southern France, facing the M e d i t e r r a n e a n Sea. O n average, the area receives less than 500 m m rainfall per year and has warm, dry summers. This aridity is reinforced both by winds that blow throughout the year and by the calcareous nature of the bedrocks. No runoff is observed on this massif. The entire agricultural surface (less than 20% of the total area) is dedicated to the production of high-quality wines. The spatial distribution of vineyards is linked to the geological structure, as they are mainly located in combs on marls. E a c h c h a t e a u is o r g a n i z e d around a single piece of land, which means that each grower has to cope all of the organization of a single pedological system, as defined by Boulet et al. (1982a, b, c). Chfiteau X is located in the Massif de la Clape, in a sandstone basin surrounded by a calcareous e n v i r o n m e n t (Figs 1 & 2). The vineyard covers about 40 ha, and is planted with different vine varieties of different ages (Fig. 3). Soil cultivation practices are the same all over the Chateau. The current aim of the grower is to

restructure his vineyard to obtain the Appellation d'Origine Contr616e label ( A O C ) which guarantees the high quality of his wines. We became interested in Chateau X because the grower had b e e n unable to solve the long-term problems with vine mortality in his vineyard. The objective of this work was to show how an understanding of the soil system of Chateau X could help in determining the reason for the vine mortality and optimizing soil cultivation practices for the h i g h e s t - q u a l i t y v i n e y a r d management.

Materials and m e t h o d s We used a morphological approach to study the o r g a n i z a t i o n of the soil. Field studies w e r e carried out on t h r e e m a i n t o p o s e q u e n c e s , following the m e t h o d of structural analysis described by Boulet et al. (1982a, b, c), and summarized by Ruellan & Dosso (1993, 1998). These first o b s e r v a t i o n s w e r e s u b s e q u e n t l y c o m p l e m e n t e d by additional toposequences to describe the organisation of the soil at the scale of the slope, and to m a p the different soil

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 53-61. 0305-8719/06/$15 9 The Geological Society of London 2006.

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Fig. 1. Geology of Chateau X in the Massif de La Clape, Southern France. Geomorphological transsect (from geological map, BRGM 2545,1:50 000)

Fig. 2. Chateau X seen from the east. A newly planted vineyard is in the foreground. Chateau X is in the background. volumes using the approach proposed by Ruellan & Dosso (1993, 1998). In parallel, we mapped the areas of vineyard mortality (Fig. 4). Then, we studied the relations between the vineyard and the areas affected by vine mortality: (1) at the scale of the Chateau as a whole, by assessing the relationships between the surface affected by mortality and the different soil volumes; (2) at the scale of the toposequence, by studying the root distribution; and (3) at the scale of the soil volume by studying soil exploration by roots and the soil's physical, chemical and biological fertility.

The temporal dimension of the soil cover was then taken into account by studying the seasonal functioning of the soil/plant system, as well as the impact of human interventions on the soil and vineyard.

Results and discussion This section first describes hypotheses that could not be substantiated to explain vineyard mortality. We then give a description of the soil system of the studied domain, where we show the importance of taking the soil system into account when considering plant growth, and, finally, we give recommendations for improving vineyard management.

SOIL SYSTEM AND QUALITY WINE PRODUCTION

55

Fig. 3. Map showing the ages of the vines at Chateau X.

Fig. 4. Map showing the zones of vine mortality at Chateau X. The newly planted vines that did not survive, shown in the foreground of Figure 2, were located on the right side of this map, in the rectangular plot.

Mortality problems: hypotheses tested without success

occurred irrespective of the age of the vine. E v e n the newly planted vines, which are visible on the foreground of Figure 2, died.

The age of the vines By linking the spatial distribution of the areas affected by mortality with the age of vines we clearly d e m o n s t r a t e d (Fig. 5) that mortality

Excess water during winter E v e n though less than 500 mm of rain falls each year in this area, each winter, excess w a t e r

56

M. DOSSO ETAL.

accumulates in the centre of the basin, and tractors can no longer be used as they get stuck in the wet soil. This led the grower 10 years ago to dig a ditch to drain the centre of the basin. This allowed the passage of tractors, but had no effect on mortality, as even the newly planted vines died (Fig. 5). The excess of water in winter was consequently not the main reason for vine mortality. We also know (Van Zy11983; Murisier 1996) that the development of vines in spring is linked with the growth of new roots in the upper part of the soil profile, so that an excess of water deeper in the profile in winter would normally not cause problems. The problem thus appeared to be related to the upper part of the soil.

soils of the Massif de La Clape, which are calcareous. However they assumed that all the soils of the C h a t e a u were h o m o g e n e o u s t h r o u g h o u t the soil profile. Consequently, analyses were only made of the top 20 cm of the soil. As the analysis of the soil system will show (see below), this view was incomplete. The soil system

Fungi were observed on dead plants. They were then assumed to have caused the death of the vines and appropriate treatments were applied to bring them under control. However, this did not stop vine mortality.

The surface of the soil, which is visible when the plants have lost their leaves, shows variations in colour. Walking downhill towards the centre of the basin, one can easily see changes in the colour, texture and structure of the soil. Despite there being a single bedrock, there is great soil diversity. Our purpose was to observe, describe and understand this lateral variation in the soil (Fig. 6). Figure 7 is a schematic diagram of the soil system, showing the vertical and lateral distribution of the different volumes of the soil cover throughout the basin. Four main domains can be distinguished.

The soil as viewed by the experts

Domain 1

Experts were aware that the soils of Chateau X were acid and sandy, unlike the majority of the

In the first part of the toposequence, the soils are shallow. They are either slightly differentiated -

Pathogenic fungi

Fig. 5. Map indicating the zones of mortality, superimposed on the age map. As can be seen in this figure, even the youngest vines died (see the rectangular plot on the right).

SOIL SYSTEM AND QUALITY WINE PRODUCTION

57

Fig. 6. The different toposequences and profiles (for example, P5 means five soil profiles) studied to understand the soil system at the Chfiteau X. The three most important toposequences are the sourthern ones: P9/P3, P5 and P4.

Fig. 7. Model of the organization of the soil system at Chateau X. with an A horizon on a C horizon derived from sandstone, or m o r e differentiated - with an S horizon between the A and C. The S horizon is a s a n d y - c l a y e y h o r i z o n resulting from the weathering of the bedrock. It is n a m e d 'S', as the pedological structure is fully developed and the

lithological structure can no longer be observed (which is the difference b e t w e e n the S and the C horizons). A and S are sandy-clayey horizons displaying the red colour of the w e a t h e r e d sandstone. This first part of the t o p o s e q u e n c e is sometimes very shallow (a few metres thick).

58

M. DOSSO ETAL.

Domain 2 After this first domain, a leached E horizon appears very rapidly down the slope. The red colour becomes paler, the texture of the A horizon sandier, and a real E horizon appears, which very rapidly becomes massive. The E horizon deepens and a yellow Bt clayey horizon appears. The roots of the vines develop mainly at the transition between the E and Bt horizons. At the surface, the soil is no longer reddish, and thus the disappearance of the reddish colour is an indicator of the beginning of the leaching process. Continuing downwards, the A and E horizons become deeper, paler in colour, and sandier, while the limit between the E and Bt horizons (E/Bt) is more contrasted, and the depth of the Bt increases. In domains 1 and 2, the soil profiles show no CaCO 3.

Domain 3 This is characterized by the presence of a new soil volume deep in the profile: BtCa, which is enriched with CaCO3.

Domain 4 This is located downhill, at the bottom of the toposequence. It is characterized by the presence of a Bt horizon rich in swelling clay minerals, and the colour and structure change. It no longer displays a yellow polyhedral structure, but a dark, cubic, sometimes even vertic structure. At the centre of the basin, hydromorphic features can be observed. Here, the morphological contrast between leached horizons and the Bt(M -- swelling clay mineral accumulation) horizon is the greatest. The A and E horizons are completely sandy. The E horizon contains very little organic matter; displays a massive structure, and is 50 cm deep. The BtM horizon is clayey, dark, with a very welldeveloped cubic structure, and displays compact and dense aggregates. Under the BtM horizon, the Btca horizon is not as dark in colour, and has a polyhedral structure with lighter, friable aggregates.

D o m a i n s 1, 2, 3 and 4: one soil system The vertical and lateral distribution of the different soil volumes in domains 1, 2, 3 and 4 is representative of the morphology of a single soil system. The different soil volumes depend on the same dynamic, i.e. leaching/accumulation in relation to the scale of the slope. An eluviated

domain is found uphill, while an illuviated domain is found downhill, with neoformation of swelling clay minerals. The dynamic is revealed by the morphology. This type of soil system has been studied in Chad by Bocquier (1971). A real dry season and a closed system are required for the accumulation of material that originates from the upper part of the watershed. Here the bedrock is sandstone, but as the whole basin is surrounded by a calcareous environment, lateral migration of calcium carbonates also occurs. These appear deep in the profile of domains 3 and 4, and increase towards the lower parts of the basin. Figure 8 shows the spatial extension of the different soil volumes with the presence of BtM in the lowest parts of the basin.

The soil system in relation to plant growth One of the main morphological characteristics of this particular soil system is the limit between the E and Bt horizons, which becomes gradually sharper towards the bottom of the basin. Such a vertical discontinuity and its lateral extension strongly affect the rooting pattern and the movement of water and solutes in the soil, and therefore can strongly influence plant growth and element uptake (Koundouras et al. 1999). The E/Bt limit creates a vertical discontinuity in the soil profile, preventing water from infiltrating vertically and inducing a lateral flow of water within the soil cover. The application of CuSO4 to control pathogenic fungi in vineyards is a very common practice. In soil rich in organic matter, Cu is not toxic as it is strongly complexed on organic compounds (Chaignon 2001). In the acid, sandy soils of Chateaux X, which are low in organic matter, it was assumed that Cu was leached out of the soil profile and reached the groundwater. However, our study of the soil morphology showed that this was not the case. In fact, the vertical flow is rapidly transformed into a lateral flow at the top of the Bt, suggesting that Cu would accumulate in the upper horizons, mainly at the E/Bt limit. Analyses were made to check this hypothesis. All the upper horizons were indeed enriched in DTPA-extractable Cu (from 20 to 120 mg kg -1 in the A and E horizons; and less than 1.5 mg kg -1 in the B and C horizons). At the bottom of the basin, the Cu concentration at the E/Bt limit was 40 times the background concentration, reaching a lethal level for vines (Delas 1963; Rousseau 1995). Therefore, understanding the soil system was the key to understanding the origin of the mortality problem. Copper had become an indicator of

SOIL SYSTEM AND QUALITY WINE PRODUCTION

59

Fig. 8. Map of the soil system at Chateau X, showing the extent of the different soil volumes (the E horizon being associated with Bt) corresponding to the soil domains 1, 2, 3 and 4. water flow. Besides the high Cu concentration, the high density of the BtM horizon also strongly hinders root growth. In the Bt horizon it becomes more difficult for roots to develop - from the top to the bottom of the basin - and a similar trend is observed for the E horizon, with fewer roots observed at the bottom of the basin. So it is clear that, beyond solving the problem of mortality, the soils need to be treated according to their specific position in the toposequence, i.e. the currently uniform soil-cultivation practices need to be changed.

Proposals for the care of the vineyards

Urgency The most urgent problem is the remediation of the upper horizons of domain 4 where mortality is concentrated and increasing (Fig. 9). This soil is degraded (Fig. 10), it is sandy, acid, with low organic matter, without structure, very compact when dry, and almost liquid when wet in the upper 50 cm. The initial effort should aim at improving these soils with subsoiling and with the addition of compost, to decrease toxicity due to Cu and to improve the soil structure and increase soil organic matter content. Then a cover crop should be planted to further improve

soil structure and to improve soil biological activity.

Evaluation of the Bt(M) horizon and facing risks The unique nature of Chateau X lies in the exceptional diversity of its soils, which has no equivalent in Massif de La Clape. There is no other closed basin in which sandstone occurs and where accumulation of swelling-clay minerals can be observed. We assume that this situation should allow the production of great wines as long as the cropping system is suited to the different parts of the toposequence. However the drainage ditch dug 10 years ago represents a risk for the Chfiteau X. As the system now functions as an open system, we suspect that the swelling mineral clays, and therefore the BtM horizon, will progressively disappear with the drainage water.

What is the future for this vineyard? We believe that Chfiteau X has reached the point where it has the right tools to change the vineyard in the desired direction. It is possible to rehabilitate the degraded soils and then to progressively transform and adapt the vineyard

60

M. DOSSO E T A L .

Fig. 9. The zones of mortality superimposed on the map of the soil system. As can be seen from the map, most of mortality zones are related to domain 4.

Fig. 10. In the foreground, a soil profile typical of the bottom of the basin can be seen (domain 4) (detail on the right: a 50 cm sandy horizon (A and E horizons with massive structure); then a dark-coloured BtM horizon with cubic structure; then a Btca horizon with polyhedral structure, slightly lighter in colour). An extensive area of vine mortality is visible in the background. m a n a g e m e n t to the u n i q u e soil diversity in the basin. A d e e p e r u n d e r s t a n d i n g of the soil system p r o v i d e s t h e basis for d e v e l o p i n g w h a t t h e

F r e n c h call a terroir, w h e r e g r o w e r s oversee coe v o l u t i o n b e t w e e n t h e soil a n d the vines ( A c a d 6 m i e d ' A g r i c u l t u r e 1998; M o r l a t 2001).

SOIL SYSTEM A N D Q U A L I T Y WINE P R O D U C T I O N

References ACADt~MIE D'AGRICULTUREDE FRANCE 1998. Le lien du terroir au produit, s6ance sp6cialisde du 04/02/98. Comptes Rendus de l'AcadOmie d'Agriculture de France, 84, 19-33. BOCQUIER G. 1971. Gen~se et Evolution de Deux Topos~quences de Sols Tropicaux du Tchad. Interprdtation Biog~odynamique. ORSTOM, Paris. BOULET, R., CttAUVEL, A., HUMBEL, EX. & LUCAS,Y. 1982a. Analyse structurale et cartographie en p6dologie: I - prise en compte de l'organisation bidimensionnelle de la couverture p6dologique: les 6tudes de topos6quences et leurs principaux apports h la connaissance des sols. Cahiers ORSTOM, SOrie POdologie, 19, 309-321. BOULET, R., HUMBEL, EX. & LUCAS,Y. 1982b. Analyse structurale et cartographie en p6dologie: II - une m6thode d'analyse prenant en compte l'organisation tridimensionnelle des couvertures p6dologiques. Cahiers ORSTOM, S&ie P~dologie, 19 (4), 323-339. BOULET, R., HUMBEL,EX. & LUCAS,Y. 1982c. Analyse structurale et cartographie en p6dologie: III passage de la phase analytique ~ une cartographie g6n6rale synth6tique. Cahiers ORSTOM, S&ie Pddologie, 19 (4), 341-351. CHAIGNON,V. 2001. Biodisponibiltd du Cuivre Dans la RhizosphOre de Diffdrentes Plantes Cultivdes. Cas des Sols Viticoles Contamin& par des Fongicides. ThEse 3bme cycle. Universit6 Aix-Marseille III, I N R A Montpellier, U M R Sol et Environnement, 183 pp.

61

DELAS, J. (1963) La toxicit6 du cuivre accumul6 dans les sols. Agrochimica, 7, 258-288. KOUNDOURAS, S., VAN LEEUWEN, C., SEGUIN, G. & GLORIES, Y. 1999. Influence de l'alimentation en eau sur la croissance de la vigne, la maturit6 des raisins et les caract6ristiques des vins en zone m6diterran6enne. Journal International des Sciences de la Vigne et Vin, 33/4, 149-160. MORLAT, R. 2001. Terroirs Viticoles: Etudes et Valorisation, Coll. 'Avenir Oenologie', Oeno Plurimedia, 120 pp. MURISIER, EM. 1996. Optimalisation du rapport feuille/fruit de la vigne pour favoriser la qualit~ du raisin et l'accumulation des glucides de r&erve. Relation entre le rendement et la chlorose. ThEse Doct. Ecole Polytechnique F6d6rale de Ztirich, 118 pp. ROUSSEAU, J. (1995) Utilisation du Cuivre en Agriculture Biologique - Impact sur l'Environnement et Perspectives de Diminution des Doses Employees. ITAB (Institut Technique de l'Agriculture Biologique), Paris. RUELLAN,A. & DOSSO, M. 1993. Regards sur le sol. Les Editions Foucher, A U P E L E 192 pp. RUELLAN, A. & DOSSO, M. 1993, 1998. Compldments, Index et Glossaire. Les Editions Foucher, Paris, 63 pp. VAN ZYL, J.L. 1983. Influence de l'irrigation sur la croissance et la qualit6 des vignes et raisins de Colombar. XVIIIbme Congrbs OIV Le Cap. Afrique du sud. 24/28-10-1983, 223-247.

Heavy metals in Swiss forest soils: modification of lithogenic and anthropogenic contents by pedogenetic processes, and implications for ecological risk assessment J O R G L U S T E R 1, S T E P H A N Z I M M E R M A N N 1, C H R I S T O P H N. Z W I C K Y 2, PETER LIENEMANN 2 & PETER BLASER 1

1Swiss Federal Institute for Forest, Snow, and Landscape Research WSL, Ziircherstrasse 111, CH-8903 Birmensdorf Switzerland (e-mail." [email protected]) 2Swiss Federal Laboratories for Materials Testing and Research (EMPA), Uberlandstrasse 129, CH-8600 Diibendorf Switzerland Abstract: We investigated the occurrence and effective mobility of heavy metals (HM) in

a representative collective of Swiss forest soils. The total HM contents of pedogenetic horizons were analysed, and the enrichment or depletion of a given HM relative to the original lithogenic content was assessed. The latter was calculated using the contents of Zr as an immobile reference. Chromium, Ni, Cu and Zn were mainly lithogenic, with a wide range of contents reflecting the diverse geology of Switzerland, while anthropogenic input of Pb was detected in most topsoils. Pedogenetic processes exerted a strong influence on the translocation or leaching of the HM. In acidic soils Ni, Cu and Pb were more mobile than Cr and Zn. We relate this behaviour to the strong affinity of the former HM to dissolved or colloidal organic matter. On the other hand, nutrient cycling by the vegetation probably led to an apparent reduction of Zn downward mobility. In many soils, guide levels as specified by Swiss legislation, or threshold values for effects on micro-organisms, were exceeded, indicating a potential risk for long-term soil fertility. As a consequence of translocation, guide levels for Cr, Ni and Cu were exceeded more often in subsoils than in topsoils.

The risk of heavy-metal (HM) leaching to the groundwater is linked to the mobility of a given H M in a given soil. Much has been learned in the l a b o r a t o r y about the p o t e n t i a l mobility under different soil chemical conditions (Selim & A m a c h e r 1996; K a b a t a - P e n d i a s 2001). However, considering the large variability of soils, regional assessment of the risk for groundwater contamination depends on knowledge of the contents and the effective mobility of H M in the characteristic soil types of a given area. One approach is to look at the depth distributions of H M in selected soil profiles, which gives clues about the influence of pedogenetic processes on metal translocation or leaching (Berrow & U r e 1986; R~iis~inen et al. 1997; Palumbo et al. 2000; Protasova & Shcherbakov 2004). Furthermore, input from the atmosphere, as well as the influence of nutrient recycling through plant uptake and litter deposition, can be detected (Nowack et al. 2001). In Switzerland, 42% of groundwater protection zones are situated under forests ( S A E F L & WSL 2005), where H M in soils are mainly lithogenic or are derived from diffuse atmospheric long-distance transport. In Switzerland, the

diverse geology ( L a b h a r t 1993), climate (Schiiepp 1978) and relief make it a difficult task to assess the regional status of H M contamination and the risks involved. As a consequence of the large spatial variability in bedrock composition, and the wide range of microclimatic conditions, a large variety of soil types has developed, including Leptosols, Regosols, Cambisols, Luvisols, podsolized soils and gleyic soils. The objective of the present study was to obtain an overview of the range of Cr, Ni, Cu, Zn and Pb contents in Swiss forest soils and their effective mobility in different soil types. The basis of this was the depth distribution of total H M contents in a group of soil profiles representing the most typical site conditions. As a means of coping with the high diversity and to better generalize the results from the individual soil profiles, H M contents were normalized to their lithogenic b a c k g r o u n d contents, which were calculated using the m e t h o d involving an immobile reference element (Blaser et al. 2000). In addition, the p o t e n t i a l impact of H M contents on long-term soil fertility is evaluated briefly using guide levels of the Swiss ordinance relating to impacts on the soil ( S A E F L 2001),

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 63-78. 0305-8719/06/$15 9 The Geological Society of London 2006.

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and possible effects on micro-organisms were evaluated in terms of threshold values suggested by VanMechelen et al. (1997).

overview of the site factors for each soil profile is given by Walthert et al. (2004).

Soil sampling and analysis Materials and methods Sampling sites A total of 95 forest sites, distributed among the five forest production regions of Switzerland, were investigated (Fig. 1). They represent a large variety of bedrocks, including marine sediments, molasse and glacial deposits, as well as magmatic and metamorphic rocks (Table 1). Macroclimatic boundary conditions include yearly precipitation between 650 and 2200 mm, as well as average annual temperatures between 1.4 and 10.3 ~ With elevations from 340 to 2000 m a.s.l., all exposures, inclinations from 0 to 45 ~ and locations from basins to top of slopes, a wide range of microclimatic conditions in terms of seasonal variations in soil temperature and soil moisture, are covered. The vegetation includes most of the typical and widespread forest communities, according to Ellenberg & K16tzli (1972). A detailed

Wherever possible, soil profiles were dug to the BC or C horizon, i.e. to bedrock material that is only weakly altered by pedogenetic processes. Samples were collected from pedogenetic horizons that were distinguished by morphological criteria, including colour, texture and the occurrence of redoximorphic features. Samples were dried at 60 ~ passed through a 2-mm sieve and stored at room temperature in an archive where atmospheric humidity is kept between 40 and 50%. The total contents of Cr, Ni, Cu, Zn, Zr and Pb were determined by X-ray fluorescence spectrometry (Philips PW 2400) of fine powder samples (grain size 10 to 20 pm) obtained by grinding with a rotating disc-mill (Fritsch Pulverisette 9; disc coated with tungsten carbide). The detection limit for all metals was 10 mg kg -1. The reproducibility of the measurements, determined for a reference soil, was between 5 and 10%. The accuracy of the method

Fig. 1. Map of Switzerland, showing the five forest production regions (Brassel & Br~indli 1999) and the location of the soil profiles investigated in the present study.

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was confirmed by analysing certified reference materials. For comparison with guide levels specified by Swiss legislation relating to impacts on the soil ( S A E F L 2001), extractable contents of Cr, Ni, Cu, Zn and Pb were determined. Sieved soil samples were extracted for two hours with 2 M nitric acid at 95 ~ and a soil to extractant ratio of 1 : 10 (mass : volume). The pH value was measured in a stirred 1 : 2 suspension of mineral soil in 0.01 M CaC12 solution after 30 minutes of equilibration. The translocation of iron in the soils was assessed using dithionite-citrate-bicarbonate extraction (Mehra & Jackson 1960; modified by Holmgren 1967) and ammonium oxalate extraction (Schwertmann 1964) of sieved soil samples. Soil organic-matter contents were estimated from organic C contents of ground soil samples (total carbon, determined by an automatic CN analyser, minus carbonate C analysed by gravimetric determination of CO2 driven from the soil sample by sulphuric acid).

Enrichment factors To assess the enrichment or depletion of a trace-element in a given soil horizon, the trace element concentrations can be normalized to an element which is conservative with respect to chemical weathering and which has no significant anthropogenic source. Zirconium which is hosted mainly in the mineral zircon satisfies both of these conditions, and Zr is used widely in geochemical studies of mineral weathering as a 'conservative' lithogenic element, against which relative enrichment or depletion of more reactive elements can be compared (Faure 1991). The lithogenic element content in a given soil horizon y was calculated from the Zr content in this horizon and the ratio of HM/Zr in the sample collected from the lowest horizon: (HM)lithogenic,y = (Zr)y X (HM/Zr)lowes t horizon

The truly lithogenic content of a soil horizon, however, can be calculated only if the assumptions hold that the bedrock of the entire soil is of a similar mineral and chemical composition, and that the respective element content at the lowest soil depth has not been altered during soil development. Both assumptions do not necessarily hold for all soils. In cases in which there are clear indications that the bedrock material changes within a soil profile, the lithogenic contents for a given soil horizon were calculated relative to the lowest horizon with

the same mineralogical composition. An enrichment factor (EF) can then be calculated as: EF = HMmeasured,y/HMlithogenic' y with EF <1 indicating depletion and EF >1 enrichment. The enrichment factors reflect the changes in trace-element contents of the individual soil horizons relative to the contents in the sample of the lowest soil horizon. If the possible alteration of the lowest horizon during soil development is properly taken into consideration, the enrichment depletion pattern in a soil profile can be interpreted with respect to atmospheric input, translocation and leaching by pedogenetic processes, as well as nutrient cycling of the vegetation (Blaser et al. 2000). On the other hand, enrichment in weathering residues or dilution by the build-up of soil organic matter (SOM) are masked, because these processes affect the reference element Zr as well. The former process is most relevant in soils developed from bedrock containing carbonates, and can lead to large HM contents in such soils.

Results and discussion

Heavy-metal contents o f the lowest soil horizons Since lithogenic background contents can contribute significantly to the HM contents in forest-soil profiles, we compared the range of HM contents in the lowest horizons of a group of soils that are assigned to a given geochemical unit, with the typical range found for true lithogenic contents in this unit, as determined by Tuchschmid (1995) (Table 1). Considering the small number of soil profiles in most geochemical units, the large inherent variability and the fact that the typical range comprises only 50% of Tuchschmid's data, the ranges generally compare well. Relatively high upper values for a given group of soil profiles can often be explained. Since many of the lowest horizons of soil profiles in units 12, 18 and 19 are partly decarbonated, high HM contents can be caused by enrichment in the weathering residue. In unit 6 (a paragneiss), one profile is located in an area with known Zn and Pb ores. However, the completely different ranges for Cr, Ni, and Cu contents in geochemical unit 1 (granitoid plutonics) are difficult to explain. Anthropogenic input is unlikely, since none of these soil profiles reveals topsoil enrichment in these HM. On the other hand, the soil contents of Ni and

HEAVY METALS IN SWISS FOREST SOILS Cu compare reasonably well with the range of 5 to 15 mg kg -1 and 10 to 30 mg kg -1, respectively, given by Kabata-Pendias (2001) for granites and gneisses.

Enrichment depletion patterns in different soil types In the following paragraphs, enrichment depletion patterns (EDP) for the five investigated HM in the 95 soil profiles grouped into different soil types are discussed. In order to reveal patterns that are characteristic for a given pedogenic process, profiles that exhibit the influence of several processes were not included. For example, several soils characterized by clay translocation also showed gleyic properties. Box plots were chosen to represent the distribution of EF for a given HM in a given horizon (or group of horizons) of a given soil type. The assignment of horizons was mainly based on morphological features, but, in some cases, modifications were made based on soil pH, the content of iron oxides, or the content of soil organic matter (SOM). The number of soils considered for a given soil type can vary for different HM and different horizons, because of HM contents below the detection limit or missing horizons. Although the database is quite small in some cases, a qualitative picture of the mobility of the HM in different soil types is obtained. Soil types were classified according to the World Reference Base for Soil Resources (FAO 1998). Common alternative pedogenetic names are given in some cases. As weakly developed soils, we considered Leptosols (Rendzinas) characteristic of the Jura region and Regosols (Pararendzinas) found in the central plateau, the Prealps and the Alps (a panels in Figs 2 to 6). For the EDP of weakly developed soils, only calcareous horizons (Ah, AC, C) were included. Cambisols (Brown Earths), mainly found in the Alps and in Southern Switzerland, show no signs of pedogenetic alteration other than browning (b panels in Figs 2 to 6). Only profiles that are moderately to strongly acidic throughout were considered. The pH values for the different groups of horizons are as follows: A: 3.9 _+ 0.3; B: 4.4 _+0.5; C: 4.8 + 0.6. Luvisols are soils characterized by clay translocation (c panels in Figs 2 to 6). Those included in our study are strongly acidic, with pH of c. 4 in most parts of the profile, and about half of them are calcareous in the lowest part. The pH values are: A: 3.9 + 0.5; E1 (horizons with clay depletion): 4.1 + 0.4; Bt (horizons with clay enrichment): 4.4 _+0.4; C: 6.2

67

_+1.7. Podsolized soils are characteristic for cool and humid forest in the Alps, and for many sites in southern Switzerland (d panels in Figs 2 to 6). Soil types include Podsols characterized by translocation of iron and/or SOM, Cryptopodsolic soils in which morphological characteristics of translocation are masked by deep incorporation of SOM (Blaser et al. 1997), as well as Dystric and Humic Cambisols (podsolized Brown Earths) with weak translocation of iron. All profiles are strongly to moderately acidic throughout, with the following pH values: A: 3.3 + 0.4; E (horizons with depletion in Fe and/or SOM): 3.5 _+ 0.5; Bh/s (horizons with enrichment of Fe and/or SOM): 4.1 _ 0.3; B_C (horizons below the Bh/s horizons): 4.6 _ 0.3. Gleysols and Gleyic Cambisols are characterized by temporary or constantly waterlogged conditions in parts of the profile, caused either by a fluctuating groundwater table or by poor drainage of percolation water due to a weakly permeable layer. Such soils are found mainly in the central plateau and the Prealps. We considered two groups, differentiated by their pH. Soils in the first group are strongly acidic throughout the profile, and are characterized by the translocation of iron - morphologically recognizable by a bleached layer (e panels in Figs 2 to 6). The pH values are: A_B (A or B horizons without redoximorphic features): 3.8 _ 0.5; eluvial (horizons depleted in Fe, bleached in some cases): 3.8 _+ 0.4; illuvial (horizons enriched in Fe, redoximorphic features for mainly oxidized conditions): 4.3 +_ 0.3. The second group of gleyic soils is weakly to moderately acidic in the upper part of the profile and mostly calcareous in the lowest horizons (f panels in Figs 2 to 6) with the following pH values: A_B: 5.4 +_ 1.2; ox_a (horizons with redoximorphic features of mainly oxidized conditions, without the horizon directly above the reduced horizons): 5.6 _ 1.0; ox_b (as ox_a, but directly above the reduced horizons): 6.0 _ 1.0; red (horizons with redoximorphic features of reduced conditions): 7.0 _ 0.6. The strong alteration of the lowest horizons in the gleyic soils needs to be taken into consideration when interpreting the EDP.

Chromium. In this section, we assume that Cr is present mainly in the most abundant chemical form: Cr (III) (McBride 1994; Kabata-Pendias 2001). Although Cr is not an essential micronutrient for plants, one possible way of explaining its slight enrichment in the Ah- and AC horizons of Leptosols and Regosols (Fig. 2a) is nutrient cycling. On the other hand, as the Ahand AC horizons are decarbonated to a larger

68

J. LUSTER E T A L .

(a)

(b)

14] ............................... F-t

Cr Ah

(6)

i

Cr A

(6)

i

Cr AC (12)

CTH

CrB

(6)

t

t I

t

0

Cr C (11)

Cr C

(6)

0.5 1 1.5 Enrichment factor

(c)

0.5 1 t .5 Enrichment factor

i i

(d)

Cr A

(6) _

H

Cr El

(6) _

l

4

Cr Bt (7) _

I F-4

i

t |

t

Cr A (14)_

| !

Cr E (17)_

t i

Cr Bh/s (21) _

t ! i

i

F-4

CrBC

(7)

CrBC (21)

i

'

''

'1

' ' ' '

'

'

'

'

1

'

'

'

I r

'

0.5 1 1.5 Enrichment factor

0

(f)

m

(2)

,

,

[

r

,

CrA

(4) ''

't

'

'

''

I'

''

r

,

i

i:i

i=i

i:

2

i

i i

Cr red

| i

'

r

I i i

_

Cr ox b (15) _

@

Cr illuv

i

B

Cr ox a (14) _

(4)

f

=

(10)

Cr eluv

f

0.5 1 1,5 Enrichment factor

0

(e) CrAB

f

'1

0.5 1 1.5 Enrichment factor

''

(15)

''

2

|

" ' ' ' '

0

I . . . .

I

. . . .

I '''''

~'''''

0.5 1 1.5 Enrichment factor

2

Fig. 2. Enrichment depletion patterns for Cr in (a) weakly developed soils (Leptosols, Regosols), (b) acidic Cambisols, (c) Luvisols, (d) podsolized soils, (e) strongly acidic gleyic soils, (f) moderately acidic to calcareous gleyic soils; for explanation of the horizons, see the text; the number of soil profiles considered are given in parentheses. extent than the C horizons, enrichment in the weathering residue may be another explanation. However, this would m e a n that Cr is less mobile

than the reference e l e m e n t Zr under these conditions. The accumulation of weak long-term and long-distance atmospheric input is another

HEAVY METALS IN SWISS FOREST SOILS

(a)

69

(b)

i i i

Ni Ah

Ni A

(3) i

Ni A C

(81

Ni B (6)

t i i

Ni C (7)

i i I

Ni C (6)

t t

i

' ' ' '

0.5

1

1.5

2

0

Enrichment factor

(c)

(d)

Hi

(4)

(111_

I

1-

Ni_Bh/s (t4) _

i !

t

. . . . .

1' ' ' ' I . 0.5 1 1.5 Enrichment factor t

. . . .

2

NiAB

i 1

H .......

t ....

IN i | !

O,5 1 1.5 Enrichment factor

t

! i i

Ni ox a (14) _

Ni eluv

(41

Ni ox b

(151 _ Ni iltuv

2

NiAB

(lO) _

(2)

i

! i f

Hi

~H t t

Ni red (151 =

t !

0.5 1 1.5 Enrichment factor

. . . .

!

0

(f)

0

!

1.5

. . .

(e)

(4)

'

, ,t............. I :

NiBC (14)

NiBC

(4)

1

Ni A (9) _

Ni E 1 i !

Ni Bt

0.5

I ' ' '

Enrichment factor

Ni A (3) _ Ni El (4) _

I . . . .

2

0

0.5 1 1,5 Enrichment factor

2

Fig. 3. Enrichment depletion patterns for Ni in (a) weakly developed soils (Leptosols, Regosols), (b) acidic Cambisols, (e) Luvisols, (d) podsolized soils, (e) strongly acidic gleyic soils, (0 moderately acidic to calcareous gleyic soils; for explanation of the horizons, see the text; the number of soil profiles considered are given in parentheses. explanation. The relatively small depletion of Cr in A- and B-horizons of the acidic Cambisols (Fig. 2b) demonstrates the rather weak mobility

of Cr in the pH range 4 to 5 (Kabata-Pendias 2001). In the Luvisols, the E D P of Cr follows the depletion and enrichment of clay (Fig. 2c). In

70

J. LUSTER E T A L .

(a)

(b)

@

Cu_A

Cu Ah

(4) Cu AC

Cu B

(4)

(4)

Cu_C

CuC

(4)

(4)

i

1

. . . .

0.5 1 1.5 Enrichment factor

(c)

Cu_A (3) _

I

Cu_EI

D

(4)

(4)

2

4

CuA

HI

I

,i

''

2

F----t

!

i

I

i

""

1 !

Cu Bh/s (12) _ CuBC (12)

I'"

'1"

"1'

"'!

....

1"'

'1"

CuAB

CuAB

(8)

(1)

i

i

r

|

i

i

~

1

I

i

i

1

1

I

s

i

I

i'

0.5 1 1.5 Enrichment factor

0

(f)

(e)

i t !

i

"

0 0.5 1 1.5 2 2.5 3 3.5 4 Enrichment factor

HI

"

k--I

_

Cu ox a (12)

ffl

(2)

',

,

(9)

i

Cu eluv

. . . .

Cu_E

CuBC ' "'1

I

t

(8)

i

(4)

. . . .

0.5 1 1.5 Enrichment factor

(d)

HI,

Cu Bt

I

Cu ox b (13)

FT-

Cu illuv

(2)

H I

....

H

_

Cu red

t

=

0

i

~

i

i

i

i

,

.

"

i

J

,

,

i

=

0.5 1 1.5 Enrichment factor

~

~

'

I

2

(13)

|

,,,,'i",' i',,'l',,, 0 0.5 1 1.5 Enrichment factor

2

Fig. 4. Enrichment depletion patterns for Cu in (a) weakly developed soils (Leptosols, Regosols), (b) acidic Cambisols, (e) Luvisols, (d) podsolized soils, (e) strongly acidic gleyic soils, (f) moderately acidic to calcareous gleyic soils; for explanation of the horizons, see the text; the number of soil profiles considered are given in parentheses. the podsolized soils, Cr is translocated to the same extent as Fe, with depletion in the eluvial horizons and enrichment in the Bh/s horizons

(Fig. 2d). Similarly, in the strongly acidic gleyic soils, Cr is d e p l e t e d in the A horizons and the horizons that are characterized by wet

H E A V Y M E T A L S IN SWISS F O R E S T SOILS

(a)

(b)

t i

Zn Ah

(7)

Zn B

3

t--

Zn C

(10)

#)

! !

1 ....

1 ....

0 0.5

illi

i|li

ilil

iii

I iiii

Iiil

i|ii

~ =

1 1.5 2 2.5 3 3.5 4 Enrichment factor

= I

*

0.5

..

I

Zn A

t l t

H

_

-H i

|

0

i,

i

i

i

i

i

!

[

i

i

i

i

0.5

(f) ZnAB

(2)

ZnAB (10)

_

Zn ox a (14) _

Zn eluv

(4)

Zn ox b (15) _

Zn illuv

1

= I

i

~

I

=

1.5

[.~

1

1.5

H

1: I

HI

I

H I

Znred

t

0.5

=

Enrichment factor

(e)

0

i

I !

0.5 1 1.5 Enrichment factor

(4)

i

ZnBC (21)

Zn_B..C ii

I

t

Zn_Bh/s (21)

i I

l i l l

=

t 1 t

Zn E (17)_

Zn El ZnBt

t

1

(14)_

(6) _

(g) _

i

Enrichment factor

(d)

Zn A

(7).

i 1 =

ii

i t

(7)

H 1

(7)

i | t !

Zn C

(C)

Zn A

(7)

i i i i

Zn AC (13)

71

(15) 1.5

2

Enrichment factor

0

0.5 1 1,5 Enrichment factor

Fig. 5. Enrichment depletion patterns for Zn in (a) weakly developed soils (Leptosols, Regosols), (b) acidic Cambisols, (e) Luvisols, (d) podsolized soils, (e) strongly acidic gleyic soils, (f) moderately acidic to calcareous gleyic softs; for explanation of the horizons, see the text; the number of soil profiles considered are given in parentheses. bleaching (Fig. 2e). Since the illuvial horizons with a very slight enrichment in Cr are the lowest horizons, it cannot be judged, however,

to what extent Cr is leached to the underground or merely translocated from the upper part of the soil into the illuvial horizons. The upper part

72

J. LUSTER E T A L .

(a)

(b) Pb Ah

Pb AC

(3)

(3)

I

Pb C

(1)

;

b

(3)

:

,

0

(c)

1 2 3 Enrichment factor

(d)

Pb A (1)

_

I-r-i

Pb Et (2) _

Pb A (10)_

, ,

i~

~

i~

~,

i,

i'~

10

i i i

EH i i

Pb Bh/s (14)

PbBC

_

i i !

i

PbBC (14)

(2) 2 3 Enrichment factor

0

l,

2 4 6 8 Enrichment factor

,H....i........................................................... V---

Pb E (11)_

M

Pb Bt (2) _

I~

0

4

5 10 15 Enrichment factor

0

(f)

(e) PbAB

(1)

PbAB (3)

(4)

(2)

I

H

H,' 1 H

Pb ox a Pb eluv

H

_

20

i i J

_

Pb ox b

(5)

_

| !

Pb illuv

(2)

Pb_red

! ,,

0

,~ f

,

i

,, c

,,,,~,,1,~,,~

,T ,,, ,~

1 2 3 Enrichment factor

~,

(5)

,'~

4

r

0

,

r,,,l,~,~,~

,i,~

~ ~, 1,, ,, ~ , ~ , , [ - r - - n - w -

2 4 6 8 Enrichment factor

Fig. 6. Enrichment depletion patterns for Pb in (a) weakly developed soils (Leptosols, Regosols), (b) acidic Cambisols, (c) Luvisols, (d) podsolized soils, (e) strongly acidic gleyic soils, (f) moderately acidic to calcareous gleyic soils; for explanation of the horizons, see the text; the number of soil profiles considered are given in parentheses. of the moderately acidic to calcareous gleyic soils is slightly depleted in Cr relative to the reduced horizons, which suggests some leaching

of Cr associated with Fe minerals from the profile because of water flux (Fig. 2f). Considering the E D P in the latter five soil types, Cr is

HEAVY METALS IN SWISS FOREST SOILS depleted in horizons that are strongly acidic and are depleted in Fe or clay, as shown for eluvial horizons in Luvisols, podsolized soils, and gleyic soils. This behaviour demonstrates the strong association of Cr with clay minerals and sesquioxides (McBride 1994). Low pH alone, as in acidic Cambisols, or only Fe depletion, as in reduced calcareous horizons of gleyic soils, does not lead to strong mobilization of Cr. The observed EDP are qualitatively consistent with depth profiles of Cr contents observed by others in podsolized soils (McKeague & Wolynetz 1980; Berrow & Ure 1986; R~iis~inen et al. 1997; Vaichis et al. 1998), as well as in Luvisols and acidic gleyic soils (McKeague & Wolynetz 1980). Nickel. In Leptosols, Ni is slightly enriched in the Ah and AC horizons, and thus shows a behaviour similar to Cr (Fig. 3a). The strong depletion of A and B horizons of the acidic Cambisols reflects, on one hand, the medium mobility of Ni in the pH range 4 to 5, and, on the other hand, the influence of transport bound to dissolved or colloidal organic matter (DOM; Kabata-Pendias 2001). Since the profiles are acidic throughout, Ni is very likely leached to the underground (Fig. 3b). The Luvisols are depleted in Ni in the A and E1 horizons, and enriched in the Bt horizons (Fig. 3c). Considering the similar pH of the B horizons in the acidic Cambisols and of the Bt horizons in the Luvisols, the comparison of the EDP in these two soil types suggests a strong retentive effect exerted by the translocated clay particles, an effect that cannot be overcome by the combined action of low pH and DOM. The EDP for Ni in the podsolized soils with depletion of A, E and Bh/s horizons are similar to the EDP in the acidic Cambisols (Fig. 3d). In this soil type, the illuvial horizons do not act as sink for Ni, and thus this HM is translocated very strongly and probably leached to the underground by the combined effect of low pH and transport with DOM. Nickel translocation or leaching is even stronger in the acidic gleyic soils (Fig. 3e). It is very likely that during water-logged phases the dissolution of Ni, which is associated with iron oxides, complements the effects of low pH and DOM. The stronger mobility of Ni as compared with Cr, also under moderately acidic conditions, combined with complexation by DOM, is probably responsible for Ni being leached more strongly than Cr, from the upper part of the second group of gleyic soils (Fig. 3f). As for Cr, the observed EDP are qualitatively consistent with depth profiles of Ni contents observed by other researchers in podsolized

73

soils, Luvisols and acidic gleyic soils (McKeague & Wolynetz 1980; Berrow & Ure 1986; Vaichis et al. 1998). On the other hand, R~iis~inen et al. (1997) studied a larger number of podsolized soils, but did not find stronger translocation of Ni compared to Cr in these soils. Copper. In contrast to Cr and Ni, AC and C horizons of Leptosols are slightly depleted in Cu (Fig. 4a). This reflects the higher solubility of Cu in calcareous soils, due to the formation of soluble hydroxy and carbonate complexes (McBride 1994). The A- and B-horizons of the acidic Cambisols are more strongly depleted in Cu (Fig. 4b) than in Ni, despite the weak mobility of Cu in the pH range 4 to 5 (Kabata-Pendias 2001). This behaviour suggests a very strong mobilizing effect of DOM. Indeed, Cu is known to form more stable complexes with D O M than any other HM (McBride 1994). It is very likely that transport with D O M also led to a strong depletion in Cu of the E and Bh/s horizons of podsolized soils, and leaching to the subsurface cannot be excluded (Fig. 4d). In the Luvisols, A and E1 horizons are strongly depleted in Cu (Fig. 4c). In contrast to Cr and Ni, however, the Bt horizon is also depleted in most of the investigated Luvisols. This indicates that translocated clay particles do not act as a sink for Cu. The EDP for Cu in strongly acidic gleyic soils equals the one for Ni, and suggests strong leaching to the subsurface by the combined action of transport with D O M and dissolution of Cu associated with iron oxides during waterlogged phases (Fig. 4e). Finally, transport while bound to D O M is very likely to be an important reason for the strong removal of Cu from the ox_a and ox_b horizons in the moderately acidic to calcareous gleyic soils. Further leaching of Cu from the reduced horizons is unlikely, considering the generally weak mobility of Cu under reduced conditions (McBride 1994). The smaller Cu depletion in the A horizons of the Luvisols, podsolized soils and moderately acidic to calcareous gleyic soils - when compared to the lower horizons - can be explained either by competition by SOM; by nutrient cycling by the vegetation (see discussion for Zn below); or by accumulation of anthropogenic input. Accumulation of anthropogenic Cu in organic topsoils has been widely observed (Kabata-Pendias 2001). The observed EDP are qualitatively consistent with the depth profiles of Cu contents observed by other researchers in podsolized and acidic gleyic soils (McKeague & Wolynetz 1980; Berrow & Ure 1986; Vaichis et al. 1998). In

74

J. LUSTER ETAL.

contrast to our observations, however, McKeague & Wolynetz (1980) found significant retention of Cu in the Bt horizons of Luvisols. Zinc. The most prominent feature of the EDP for Zn in Leptosols and Regosols is the strong enrichment in the Ah horizons (Fig. 5a). Since Zn is a micronutrient for plants, nutrient cycling of the vegetation rather than atmospheric input is given as the dominant cause for Zn enrichment in forest and Alpine topsoils (Blaser et al. 2000; Nowack et al. 2001). This is supported by the fact that atmospheric input of Zn could not be detected in the topsoils of industrialized areas of Switzerland (SAEFL 1993). Nutrient cycling refers to the uptake of nutrients from the rooted part of the soil profile and their reintroduction to the topsoil by litter-fall and degradation. In all other soil types, the EDP for Zn are similar to the ones for Cr (Fig. 5b to f). As for Cr, the extent of translocation or leaching in acidic Cambisols, Luvisols, podsolized and gleyic soils is smaller than for Ni and Cu. At first sight, this is surprising, considering the rather high mobility of Zn under strongly and moderately acidic conditions (Kabata-Pendias 2001). One reason might be that constant return of Zn to the topsoil by nutrient cycling leads to an apparent reduction of the effective Zn mobility within the profile. The observed EDP are qualitatively consistent with depth profiles of Zn contents observed by other researchers in podsolized soils, Luvisols and acidic gleyic soils (McKeague & Wolynetz 1980; Berrow & Ure 1986; R~iis~inen et al. 1997; Vaichis et al. 1998). Lead. Lead is strongly enriched in the Ah or A horizons of all soil types (Fig. 6a to f). This enrichment can be attributed to accumulation of anthropogenic Pb in mainly SOM (KabataPendias 2001). Based on isotopic signatures, Nowack et al. (2001) estimated the anthropogenic contribution after 1900 to the Pb content in Alpine topsoils to be between 20 and 75%, depending on the lithogenic background. Vogel et al. (1992) found Pb enrichment in all investigated topsoils of intensively used areas of Switzerland independent of the land use. In the subsoil, the E D P for Pb in Luvisols and podsolized soils follow the characteristic patterns for clay and iron translocation, respectively (Fig. 6c and d). At the same time, Pb in the illuvial horizons is enriched relative to the lowest horizon. This indicates that, despite the rather short period of Pb input, compared to the timescale of soil formation, this HM has already been incorporated deeply into these soils. Considering that Pb is even less mobile than Cr

and Cu in the pH range 4 to 5 (Kabata-Pendias 2001), transport of Pb bound to DOM is very likely to be mainly responsible for this strong translocation. The Pb depletion in the B horizon of some acidic Cambisols (Fig. 6b) and in the eluvial horizons of the strongly acidic gleyic soils (Fig. 6e) indicates that Pb is also mobile in the lower part of these soil types. Generally, Pb is considered the least mobile of the HM in acidic soils (Kabata-Pendias 2001). However, our data suggest a high mobility of Pb in our soils, which is probably mainly due to the strong competitiveness of DOM with sites on the soil solid phase for binding of Pb. Actually, the data for podsolized soils presented by McKeague & Wolynetz (1980) and Berrow & Ure (1986) support our observations. On the other hand, R~iis~inen et al. (1997) observed no translocation of Pb below the E horizon in a number of podsolized soils in Scandinavia, and Reaves & Berrow (1984) found no evidence that leaching and gleying had an influence on the depth distribution of Pb in Scottish soils. Summary for different soil types. In this section the results of the EDP for the mainly lithogenic HM Cr, Ni, Cu and Zn are summarized with respect to the characteristics of the different soil types. Lead is not discussed, because atmospheric impact is superimposed strongly on the translocation processes inherent to a given soil type. With the exception of a strong Zn enrichment in the topsoil, Leptosols and Regosols are characterized by only slight depletion or enrichments of all HM, which is in accordance with the alkaline conditions throughout the profiles. The A and B horizons of acid Cambisols are depleted in all HM, with the strength of depletion increasing in the order Cr = Zn < Ni < Cu. In acid Luvisols, the EDP for Cr, Ni and Zn follow the depletion and enrichment of clay, while Cu is depleted in the entire profiles. In podsolized soils, the translocation of Cr and Zn is as strong as that of iron. Nickel and Cu, on the other hand, are also depleted in the illuvial horizons. Also in Gleysols, Cr and Zn exhibit similar behaviour, with weak depletion of the A_B and the eluvial or ox horizons. Nickel and Cu are strongly depleted in these horizons.

Evaluation o f the H M contents in Swiss forest soils, in terms o f ecological risk Ground- and surface-water quality. The results for the EDP of HM in different soil types have implications with respect to ecological risk

HEAVY METALS IN SWISS FOREST SOILS assessment. In particular, our data suggest that the mobility of Ni, Cu and Pb in acidic soils is enhanced by complexation and transport with DOM. Leaching of Ni and Cu from the profiles is indicated by the EDP, and Pb from anthropogenic input has been deeply translocated within a short period compared to the timescale of soil formation. We therefore conclude that large contents of Ni, Cu or Pb in acidic soils pose a potential risk to groundwater quality. On the other hand, translocation with clay in acid Luvisols has to be considered to be relic. Since clay translocation does not happen in soils with pH <5 (Schachtschabel et al. 1992), this process is very likely not operational at present. However, the illuvial horizon in these soils may still act as a barrier for Cr, Ni and Zn, and reduce the risk of leaching of these metals. The downward mobility of Zn is surprisingly low in all soils. One reason for this may be the redistribution of Zn on the soil surface, through plant uptake and litter deposition. The fact that similar E D P were found for soils developed from different bedrock materials suggests that the translocation pattern of a given H M in a given soil type is not changed completely by its speciation in the source rock. However, the mobility of metals in terms of soilsolution concentration may depend very much on their chemical form (Kabata-Pendias 2001). Therefore, while the potential risk for leaching of H M to ground- or surface water in a given soil type can be evaluated by our data, nothing can be said about effective concentrations in soil solution and the potential exceedance of drinking-water threshold values. The same considerations apply if we want to extend our findings to the risks involved in local contamination from anthropogenic point sources. In addition, changes in the chemistry of the topsoil have to be considered in such a case (e.g. alkalinization by deposition of filter dust). Soil fertility. High HM contents in soils may not only pose a risk to ground- and surface-water quality, but may also affect the soils themselves. Exceedance of guide levels defined by the Swiss ordinance relating to impacts on the soil (OIS; S A E F L 2001) indicates that the long-term fertility of a soil may be at risk. These guide levels are based on nitric-acid-extractable metal contents, and are given as 40 mg kg -1 for Cu; 50 mg kg -1 for Cr, Ni and Pb; and 150 mg kg -1 for Zn. For soil horizons with a SOM content >15%, volume-based concentrations in mg L -1 are considered. A critical discussion of the concept is presented by Blaser (2003). Microorganisms in soils are particularly sensitive to

75

high H M contents. To evaluate a potential risk to the community structure or function of micro-organisms, lowest observation effect concentrations can be used as threshold values. These are defined as total HM contents at which the first negative effects on micro-organisms were observed in the field or in the laboratory. For forest soils, VanMechelen et al. (1997) recommend values from Tyler (1992) for organic litter layers (Cr: 30 mg kg-1; Cu: 20 mg kg-1; Zn: 300 mg kg-1; Pb: 500 mg kg-1). For mineral topsoil horizons, values from Kabata-Pendias & Pendias (1984) and Witter (1992) are recommended (Cr: 75 mg kg-1; Ni: 95 mg kg-1; Cu: 60 mg kg-1; Zn: 170 mg kg-1; Pb: 100 mg kg-1). In Table 2, the exceedance of guide levels of the OIS in topsoils (uppermost 20 cm of the mineral soil) and subsoils, as well as the exceedance of threshold values for micro-organisms in the uppermost 10 cm, are listed and grouped according to geochemical units (Tuchschmid 1995), soil type and region. Chromium and Ni contents exceed critical values very often Cr both in topsoils and subsoils, and Ni mainly in subsoils. Copper contents exceed guide levels of the OIS in the subsoil of nine profiles but only in one topsoil. Copper threshold values for micro-organisms are not exceeded for any of the cases. Zinc is a potential risk for micro-organisms in the topsoil of eight of the profiles. Enrichment of Zn in the topsoil by nutrient cycling thus can lead to conditions that might affect microbial communities. On the other hand, guide levels for Zn are exceeded in none of the profiles. Potentially hazardous contents or concentrations of Pb, which can be considered mainly anthropogenic in most soils (for exceptions, see Table 1), occur in 10 topsoils and six subsoils. Vogel et al. (1992) found a similar proportion of topsoils showing excessive Ni, Zn, and Pb contents in the Central Plateau, the region of Switzerland with the highest anthropogenic influence. The proportion of topsoils showing an excessive Cu content, however, was larger than in our study. Potentially hazardous contents of Cr, Ni and Cu in terms of guide levels of the OIS are mainly restricted to the group of pelitic units 6, 11, 21, 24 and 25, as well as unit 16 (serpentinites). Extremely large lithogenic Cr and Ni contents are characteristic for the latter unit (Table 1). The exceedance of guide levels for Pb is also of concern in the granitoids (unit 3) and carbonates (units 12 and 18). The grouping according to the soil types reveals the influence of the pedogenetic processes discussed by means of EDP. Translocation in acidic soils has led to a larger number

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76

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HEAVY METALS IN SWISS FOREST SOILS of exceedances of guide levels in subsoils as compared to topsoils, especially for Ni and Cu. In some podsolized soils, the translocation of anthropogenic Pb has led to the exceedance of guide levels in the subsoil. O n the other hand, the large H M contents in topsoils and subsoils of Leptosols and Regosols are most likely caused by the relative enrichment of the residue derived from carbonate weathering, which is not detected by EDP. W h e n taking a final look at the different regions of Switzerland, one realizes that the guide levels of the OIS for all metals are exceeded more frequently in soils from southern Switzerland than in soils from other regions. In the Jura, only guide levels for Pb are exceeded.

Condusions (1) Assessing the depth distribution of H M in soils by using e n r i c h m e n t d e p l e t i o n patterns ( E D P ) that are based on ratios between measured H M contents and calculated lithogenic H M contents m a k e s it possible to easily group the results for different soils also if total contents vary strongly; (2) Our study, which is based on a qualitative comparison of E D P for H M in Swiss forest soils, emphasizes the strong influence of pedogenetic processes on the translocation and leaching of the metals; (3) Nickel, Cu and Pb appear to be more mobile in acidic forest soils of Switzerland than Cr and Zn, which is very likely to be related to the strong affinity of the former group of H M to D O M . On the other hand, nutrient cycling by the vegetation seems to lead to an apparent reduction of Z n mobility; (4) In m a n y soils, guide levels of heavy metals, as specified by Swiss legislation or threshold values for effects on micro-organisms were exceeded, indicating a potential risk to long-term soil fertility. As a consequence of translocation, guide levels for Cr, Ni, and Cu were exceeded more often in subsoils than in topsoils. The authors thank R. Bucheli, D. Christen, M. Oswald, B. Peter (WSL), U. Gfeller (EMPA), and the team of the central analytical laboratory (WSL) for sample preparation and analyses. The preparation of soil profiles, their morphological description, and the sampling were carried out mainly by the team of R Lt~scher, in particular R. K6chli and M. Walser, but also by L. Walthert (all from WSL). The financial support of WSL and the Swiss Federal Agency for the Environment, Forest and Landscape is gratefully acknowledged.

77

References BERROW, M.L. & URE, A.M. 1986. Trace element distribution and mobilization in Scottish soils with particular reference to cobalt, copper and molybdenum. Environmental Geochemistry and Health, 8,19-24. BLASER, E, KERNEBEEK, E, TEBBENS, L., VANBREEMEN, N. & LUSTER,J. 1997. Cryptopodzolic soils in Switzerland. European Journal of Soil Science, 48, 411-423. BLASER, P., ZIMMERMANN,S., LUSTER,J. & SItOTYK,W. 2000. Critical examination of trace element enrichments and depletions in soils: As, Cr, Cu, Ni, Pb, and Zn in Swiss forest soils. Science of the Total Environment, 249, 257-280. BLASER, P. 2003. Wann ist ein Boden schwermetallbelastet? Eine bodenkundliche Sicht auf gesetzliche Richtwerte. Gaia, 12/1, 38-44. BRASSEL, P. & BR.~NDLI,U.-B. (eds) 1999. Schweizerisches Landesforstinventar. Ergebnisse der Zweitaufnahme 1993-1995. Haupt Verlag, Berne. ELLEN~ERG,H. & KL~3TZLI,E 1972. Waldgesellschaften und Waldstandorte der Schweiz. Mitteilungen der EidgenOssischen Forschungsanstalt fiir Wald, Schnee und Landschafi, 48, 587-930. FAURE, G. 1991. Principles and Applications of Inorganic Geochemistry. Macmillan, New York. FAO 1998. World Reference Base for Soil Resources. World Soil Resources Reports, 84, FAO, Rome. HOLMGREN, G.G.S. 1967. A rapid citrate-dithionite extractable iron procedure. Soil Science Society of America Proceedings, 31, 210-211. KABATA-PENDIAS,A. 2001. Trace Elements in Soils and Plants, 3rd edition. CRC Press, Boca Raton. KABATA-PENDIAS, A. & PENDIAS, H. 1984. Trace Elements in Soils and Plants. CRC Press, Boca Raton. LABHART, T.P. 1993. Geologie der Schweiz, 2nd edn. Ott Verlag, Thun. MCBRIDE, M.B. 1994. Environmental Chemistry of Soils. Oxford University Press, New York. MCKEAGUE, J.A. & WOLYNETZ, M.S. 1980. Background levels of minor elements in some Canadian soils. Geoderma, 24, 299-307. MEHRA,O.P. & JACKSON,M.L. 1960. Iron oxide removal from soils and clays by a dithionite-citrate system buffered with sodium bicarbonate. In: SWINEFORD, A. (ed.) Clays and Clay Minerals, Proceedings of the 7th National Conference, Washington, D.C, 1958. Pergamon Press, New York, 317-327. NOWACK, B., OBRECHT,J.-M., SCHLUEP,M., SCHULIN, R., HANSMANN,W. & KOPPEL, V. 2001. Elevated lead and zinc contents in remote alpine soils of the Swiss national park. Journal of Environmental Quality, 30, 919-926. PALUMBO, B., ANGELONE, M., BELLANCA,A., DAZZI, C., HAUSER, S., NERI, R. & WILSON,J. 2000. Influence of inheritance and pedogenesis on heavy metal distribution in soils of Sicily, Italy. Geoderma, 95, 247-266. PROTASOVA, N.A. & SHCHE~BAKOV, A.P. 2004. Microelemental composition of zonal soils in the central chernozemic region. Eurasian Soil Science, 37, 40-48.

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RAISANEN, M.L., KASHULINA, G. & BOGATYREV, I. 1997. Mobility and retention of heavy metals, arsenic and sulphur in podzols at eight locations in northern Finland and Norway and the western half of the Russian Kola peninsula. Journal of Geochemical Exploration, 59, 175-195. REAVES, G.A. & BERROW, M.L. 1984. Total lead concentrations in Scottish soils. Geoderma, 32,1-8. SAEFL (eds) 1993. Nationales Bodenbeobachtungsnetz: Messresultate 1985-1991. Schrifienreihe Umwelt Nr. 200. Eidgen6ssische Drucksachenund Materialzentrale, Berne. SAEFL (eds) 2001. Commentary on the Ordinance of 1 July 1998 Relating to Impacts on the Soil (OIS). Swiss Federal Agency for the Environment, Forest and Landscape, Berne. SAEFL & WSL (eds) 2005. Forest Report 2005 - Facts and Figures About the Condition of Swiss Forests. Swiss Federal Agency for the Environment, Forest and Landscape, Berne/Swiss Federal Research Institute WSL, Birmensdorf. SCHACHTSCHABEL, P., BLUME, H.-P., BROMMER, G., HARTGE, K.-H. & SCHWERTMANN, U. 1992. Scheffer/Schachtschabel Lehrbuch der Bodenkunde, 13th edn. Enke, Stuttgart. SCHI3EPP, M. 1978. Die Stellung der Schweiz im Europ~iischen Klimaraum und Gesamtiaberblick. In" SCHUEPP, M. et al. (eds) Klimatologie der Schweiz, volume II: Regionale Klimabeschreibungen, part 1: Gesamtiibersicht, Westschweiz, Wallis, Jura und Juranordfuss sowie Mittelland. Schweizerische Meteorologische Anstalt, Zurich, 1-42. SCHWERTMANN,U. 1964. Differenzierung der Eisenoxide des Bodens durch Extraktion mit Ammonium-

oxalat-L6sung. Zeitschrift fiir Pflanzenerniihrung und Bodenkunde, 105, 194-202. SELIM, H.M. & AMACHER,M.C. 1996. Reactivity and Transport of Heavy Metals in Soils. Lewis Publishers, Boca Raton, FL. TUCHSCHMrD,M.P. 1995. Quantifizierung und Regionalisierung yon Schwermetall- und Fluorgehalten bodenbildender Gesteine der Schweiz. Umweltmaterialien 32, SAEFL, Bern. TYLER, G. 1992. Critical Concentrations of Heavy Metals in the Mor Horizon of Swedish Soils. Swedish Environmental Protection Agency, Reports, 4078. VAICHIS,M., RAGUOTIS,A., ARMOLAITIS,K. & KUBERTAVICHENE,L. 1998. Total content of heavy metals in forest soils of Lithuania. Eurasian Soil Science, 31, 1489-1494. VANMECHELEN,L., GROENEMANS,R. & VANRANST,E. 1997. Forest Soil Condition in Europe; Results of a Large-scale Soil Survey. Technical Report, EC, UN-ECE, Ministry of the Flemish community, Brussels, Geneva. VOGEL, H., DESAULES,A. & H,~NI, H. 1992. Heavy metal contents in the soils of Switzerland. International Journal of Environmental Analytical Chemistry, 46, 3-11. WALTHERT, L., ZIMMERMANN,S., BLASER,P., LUSTER, J. & LI2SCHER, P. 2004. Waldb6den der Schweiz, Vol. 1: Grundlagen und Region Jura. Hep-Verlag, Berne. WITTER, E. 1992. Heavy Metal Concentrations in Agricultural Soils Critical to Microorganisms. Swedish Environmental Protection Agency, Report, 4079.

Reuse of agricultural drainage water in central California: phytosustainability in soil with high levels of salinity and toxic trace elements G. S. B A i q U E L O S 1 & Z . - Q . L I N 2

1USDA_ARS, Water Management Research Unit, 9611 S. Riverbend Ave., Parlier, CA 93648, USA (e-maik [email protected]) 2Department o f Biological Sciences and Environmental Sciences Program, Southern Illinois University Edwardsville, Edwardsville, I L 62026-1650, USA Abstract: Agricultural drainage waters in the western San Joaquin Valley of Central California contained high levels of salts, boron (B) and selenium (Se). Discharge of the drainage directly into the Kesterson Reservoir was hazardous to plants and wildlife. To investigate the plausibility of using plants as recipients for disposing of poor-quality drainage-waters, multi-year field studies were conducted to reuse drainage water on plants that are salt and B tolerant, and accumulate or stabilize soluble Se from the drainage. Installation of a subsurface tile drainage system in cropping fields allowed drainage waters to be collected and subsequently reused on tolerant plants. The tested plant species included Lactuca sativa, Lycopersicon esculentum, Gossypium hirsutum L., Medicago sativa L., Brassica napus var. Hyola, Helianthus annuus, Distichlis spicata L., and Salicornia bigelovii Torr. The quality of drainage water decreased along the water reuse path, with a sulphate-dominated salinity change from 4.5 to 15.2 dS mq; soluble B from 3.4 to 14.5 mg Lq; and soluble Se from 0.08 to 1.18 mg L-1. Soil salinity and concentrations of B and Se increased with decreasing quality of reused drainage waters. This study demonstrated the importance of monitoring changes in soil quality for sustaining such a plantbased system.

From the geological evidence, it is proposed that trace elements result from the dissolved mineral load ultimately draining from Cretaceous marine sedimentary strata underlying and surrounding basins such as the San Joaquin Valley in central California. Extensive volcanic eruptions during Cretaceous times are thought to be the primary source of selenium (Se) and other trace elements through deposition in the Cretaceous seas that had submerged the western US. Over time, Se was incorporated in sediments that were then uplifted and exposed to weathering and erosion. Alluvial deposits of the west central San Joaquin Valley were derived from the eastern side of the Coast Ranges (Presser 1994). The Coast Ranges evolved from complex folding and faulting of sedimentary and igneous rocks of Mesozoic and Tertiary age. Cretaceous and Tertiary marine sediments dominate the Coast Ranges, and the mountains drain into the western San Joaquin Valley. These marine shales and sandstones weathered under a sulphate regime, as evidenced by extensive salt efflorescences and evaporites at water and shale surfaces (Irwin & Barnes 1975). Weathering of

reduced shale (oxidation of pyrite, FeS2) was largely a reversal of the chemistry of the early diageneses of the shale (i.e. reduction of sulphate). Because of its similar chemical and physical properties, Se substituted for sulphur in pyrite in sedimentary rock at high concentrations (Berner 1984). Agricultural practices under the semi-arid climatic conditions in central California required intensive irrigation and subsurface drainage to prevent salt accumulation in the surface soils. Within the San Joaquin Valley, agricultural irrigation changed the groundwater recharge and discharge mechanisms and the distribution of chemical solids (including Se) (McNeal & Balisteri 1989). U n d e r natural conditions, groundwater recharge was primarily from the infiltration of water from intermittent streams in the upper parts of the alluvial fans. The pumping of groundwater for irrigation substantially altered the groundwater flow system by increasing the depth of the water table, while irrigation with pumped groundwater caused downward displacement of soil salts. The application of irrigation water caused leaching of soil salts, and increased Se and B

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. R (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 79-88. 0305--8719/06/$15 9 The Geological Society of London 2006.

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G.S. BAiqUELOS & Z.-Q. LIN

concentrations in the groundwater, which also facilitated the need for drainage systems in the San Joaquin Valley. The installation of subsurface tiled-drainage systems hindered further salinization of the root zone in the agricultural regions, and captured Se-laden groundwater that was present at elevated levels within the soil (Deverel et al. 1994). Management of large volumes of agricultural drainage water in the San Joaquin Valley is a major environmental challenge. Because of Se toxicity to wildlife attested at the Kesterson Reservoir, agricultural drainage water with high levels of Se, B, and salts cannot be discharged directly into surface waterbodies. Currently, drainage waters are collected and stored in evaporation ponds where water evaporates and B, Se and salts accumulate. Recent studies showed that the build-up of toxic elements in evaporation ponds has become a potential threat to wildlife (Ong et al. 1997). Therefore, it is important to substantially reduce the volume of drainage water resulting from agricultural production in the western San Joaquin Valley. There is an increasing trend towards reusing poor quality waters for irrigated agricultural production systems (San Joaquin Valley Drainage Implementation Program 2000). Sufficient evidence exists to support the concept of reusing saline water originating from agricultural drainage or from shallow groundwaters (Grattan & Rhoades 1990; Ayars et al. 1993; Oster 1994; Shalhevet 1994; Shennan et al. 1995; Rhoades 1999; Shannon et al. 1997). Reusing poor-quality water for irrigation can serve two purposes - one is to dispose of drainage water that would otherwise be costly to be treated, and the other is to utilize poor-quality drainage water as a water resource for growing crops that have economic value. Consequently, the successful adoption of a practical water-reuse strategy in central California requires integrated management practices related to irrigation schemes with poor-quality waters, crop selection, and monitoring soil and groundwater for the fate of toxic elements in agricultural production systems. The concept of reusing and disposing of saltand Se-laden drainage water in agricultural systems was expanded upon by Cervinka et al. (1999) as the Integrated on-Farm Drainage Management (IFDM) system (formerly termed the 'agroforestry system', see Cervinka et al. 1999). The IFDM system involved the use of freshwater (i.e. with low salinity and Se concentration) to grow salt-sensitive crops, and the use of the resulting drainage water to irrigate salt-

tolerant crops (Lin et al. 2000). Drainage water produced from the irrigation of salt-tolerant trees or grasses was, in turn, used to irrigate highly salt-tolerant, halophytic plants. By this means, the volume of drainage water was substantially reduced by evapo-transpiration, and the remaining drainage water was eventually disposed of via sprinklers into a lined solar evaporator (Cervinka et al. 1999). The sustainability and success of a drainagewater reuse strategy is dependent on managing the ever-increasing accumulation of salts and using the appropriate plant species for the varied quality waters and soils. In central California, suitable plants must be salt- and B-tolerant, and be fairly low-maintenance to grow (see Benes et al. 2004). Additional favourable characteristics would include: the plant's ability to attenuate Se, and to have economic value. Maas and Grattan (1999) have reviewed the effects of salinity on the yields of different crops, and clearly indicated that crop yields are a function of interactions between salinity and various soil, water and climatic conditions. Moreover, plant parameters such as stage of growth, varieties, irrigation methods, bed arrangements and plant population density, also influence the ability of the selected plant species to survive in a drainage-water reuse system. If quality of irrigation waters becomes increasingly poor with subsequent use in a drainagewater reuse strategy, it is imperative to select even more salt- and B-tolerant plant species for the latter stages of the drainage reuse program. When possible, the economic viability of selected crops and a low field maintenance requirement should be considered as two important criteria for the selection. The objective of this review was to illustrate a plant-based agricultural drainage-water reuse system in the San Joaquin Valley, with an emphasis on the role of soil and vegetation in operating a water reuse system under high-salt, high-B and high-Se conditions.

Materials and methods The drainage-water reuse field project was initially established in 1994 on a commercial farm (Red Rock Ranch) in the township of Five Points, California, and was comprised of reuse components A through E (see Fig. 1 for field layout). Based upon reported salt and B tolerances of different plant species, as described by Maas and Grattan (1999), the plant species exhibited in Figure 1 were specifically selected for the respective drainage reuse components. During the experimental periods (1994-2002),

REUSE OF AGRICULTURAL DRAINAGE WATER

Freshwater Irrigation

A 195 ha Crops: lettuce tomatoes

B 52 ha Irrigation

Crops: canola sunflower safflower cotton alfalfa

C 5ha Irrigation

Crops*: eucalyptus saltgrass wheatgrass

Irrigation

b 2 ha Crops: pickleweed saltbush cordgrass

81 E 0.8 ha Solar evaporator

* 14 salt-tolerant grass species were planted in this section

The volume of drainage water decreases while passing through the water reuse system

Concentrations of salts, El and Se increase while passing through the water reuse system

Fig. 1. Schematic illustration of a plant-based drainage-water reuse system, and its different components.

data were collected from sampling sites located on the following components: (A) 195 ha of saltsensitive crops, such as lettuce (Lactuca sativa) and tomatoes (Lycopersicon esculentum); (B) 52 ha of salt-tolerant crops, including cotton ( Gossypium hirsutum), alfalfa (Medicago sativa L.), canola (Brassica napus), sunflower (Helianthus anuus L. ), and safflower (Carthamus tinctorius); (C) 5 ha of salt-tolerant eucalyptus (data not shown; eucalyptus was replaced with 14 different grasses after 1999); (D) 2 ha of halophytic plants: pickleweed (Salicornia bigelovii Torr.), saltgrass (Distichlis spicata L.), saltbush (Atriplex lentiformis L.), and cordgrass (Spartina gracilis Trin.), and (E) 0.73 ha of a lined solar evaporator (which was replaced with a salt-distillation facility after 1999). The total ground surface area of the last three components C (salt-tolerant grasses), D (halophytes) and E (solar evaporator) comprised only a small portion (i.e. -3 %) of the whole water reuse system. A subsurface drainage system was installed to a depth of 1.65-2.13 m below the soil surface. All drains consisted of perforated polyethylene pipes placed on a gravel filter at least 7-8 cm thick surrounding the pipe (Lin et al. 2002). The tiles were spaced along the length of the site at about 25 m intervals, and also ran perpendicular at either end of the respective field site. Observation tubes were installed randomly to a depth of 3 m (when possible), to monitor the depth and

quality of groundwater (data were not reported). The collected drainage water was pumped from a drainage sump and routed through a central distribution manifold. The grower followed typical furrow irrigation practices when using available drainage water (drainage water availability tends to decrease during the winter growing season). Irrigation scheduling was primarily based on the weather data provided by the local California Irrigation Management Information System. Losses of soluble salts via surface water runoff, percolation beyond 90 cm, and reduction of soluble Se to elemental Se were not presented in this study, due to technical limitations. These variables, in addition to spatial and drainage line variability under field conditions, are acknowledged as essential for creating any type of mass-balance expressions. The soil on the field site was predominantly classified as ciervo clay (fine, Semitic, thermic, vertic haplocambid). A hydraulic auger (Giddings Rig) was used to collect soil cores at 0 to 30, 30 to 60, and 60 to 90 cm depths, during the course of the field trials. Collected soil samples were dried at 65 ~ thoroughly mixed and sieved with a 2-mm screen. Water-soluble Se and B, and EC were determined in a soil water extract of 1:1. The different harvested plant organs (leaves, stems and roots) were washed with de-ionized water, dried at 50 ~ for

82

G.S. BAiqUELOS & Z.-Q. LIN

seven days, weighed, and ground in a stainlesssteel Wiley mill equipped with a 0.83-ram screen. Plant samples were acid digested with concentrated HNO3-HzO2-HC1. Selenium in soil and plant tissues was analysed by an atomic absorption spectrophotometer with an automatic vapour accessory, and B concentrations were determined by an inductively coupled plasma spectrometer. Volatile Se was also collected at limited sites by a sampling chamber system that was described in detail by Lin et al. (2000, 2002). Briefly, the dimensions of the Plexiglas chamber were 0.71 m long, 0.71 m wide and 0.76 m high; the chamber provided an internal volume of 0.38 m 3 and enclosed a ground area of 0.5 m 2. Under field conditions, volatile Se was collected continuously during a 24-h sampling period at an airflow rate of 0.36m3h -1 through 20 activated-carbon filters.

Results

Irrigation with different quality waters Irrigation of salt-sensitive species with goodquality water (designated as component A) produced the first drainage water that would subsequently be reused on selected vegetation in component B (Fig. 1). The quality of water used for irrigation in the other components - B, C and D, decreased with each subsequent use (Table 1). For example, qualitative differences were clearly observed in water used for irrigation in component A as compared to component D, as follows: water salinity (EC) increased from 0.7 dS m -1 to 15.1 dS m-l; soluble B concentrations increased from 0.7 to 21.2 mg L-l; and soluble Se concentrations increased from ND (not detectable) to 1.3 mg L -1. Due to the increased acreage planted in components A and B, the greatest amount of drainage water was produced in these two components (data not shown). The quality of the drainage waters produced from these two components decreased to the greatest extent among all the components: EC increased from 4.5 to 15.2 dS m q, soluble B increased from 3.4 to 14.5 mg L -a, and soluble Se increased from 0.08 to 0.12 mg L q (Table 1). There was no substantial increase in EC and B concentrations in drainage waters produced from components C to D (Table 1). In general, while volumes of drainage water decreased due to evapo-transpiration along the path of drainage-water reuse, concentrations of salts, B and Se increased in the drainage waters (Table 1).

Accumulation o f salts, B and Se in irrigated soil The reuse of drainage water resulted in an increased accumulation of salts, B, and Se in the soil from 0 to 90 cm at harvest of each year for all reuse components of the drainage reuse system as follows: D > C > B > A (Table 1). Significant increases and downward movement of soluble salts, B and Se were also observed at the deepest depth (60-90 cm) at post-harvest for each of the water reuse components (A through D). Typical annual examples of these changes throughout the soil profile are presented for the pre-planting and post-harvest stages, in Table 1, for each of the reuse components. These changes are clearly illustrated in the mean comparison from 0 to 90 cm between component A (pre-planting) and component D (post-harvesting) as follows; soil EC levels increased from 1.6 to 30 dS m -1, soilextractable Se concentrations increased from <14 lag L -a to 1.3 mg L q, and extractable B concentrations increased from 1.5 to 31 mg L q (Table 1).

Plant tolerance in water reuse systems Stand establishment for the selected crops for each respective component was generally good during the time period of study. Typical plant yields are shown in Table 2 for each drainage reuse component. Careful monitoring and salt management practices of some sort will, however, be absolutely necessary for the soil to sustain the observed plant growth over longterm use of components C and D. Symptoms of leaf burn/necrosis from excessive salt or B build-up in the soil were only observed in canola growing near the ends of furrows, where there was an obvious excessive accumulation of salts on the soil surface. Mean tissue B concentrations were greatest in canola at 226 mg kg q DM, and generally under 100 mg kg -1 D M (except for cotton and eucalyptus at 118 and 165 mg kg q, respectively). The low accumulation of plant B indicates that B was not causing plant damage, despite the high concentrations available within the reuse system.

Attenuation o f excessive Se in the drainage reuse system Selenium can be removed through phytoextraction and volatilization. The amount of Se remaining after irrigation with drainage was determined at selected sites under field

R E U S E OF A G R I C U L T U R A L D R A I N A G E W A T E R

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REUSE OF AGRICULTURAL DRAINAGE WATER conditions, although this drainage-water reuse system was not designed and established with a primary focus on Se removal from soil. In the different components, the plant Se concentrations ranged from a low of 0.1 mg kg -a in lettuce to a high of 13 mg Se kg -~ DM in pickleweed shoots (Table 2). The estimated amounts of Se mass removed by uptake are shown in Table 2, based upon plant concentrations of Se and yields (or biomass) per ha per growing season for annual crops in components A and B, and on an annual basis for perennial crops grown in components C and D. Rates of Se volatilization for some plant species are also compiled in Table 2. The mass of Se removed by volatilization was estimated for the plant species, with the respective growing season (in days) for each of the species. The greatest mass of Se removed by volatilization was 620 g ha -1 year -1 in the pickleweed field, and the lowest was by cotton at 20 g ha -1 per year. Selenium removed by volatilization in unvegetated (or bare soil) was c. 167 g Se ha -a year -1.

Final disposal of drainage in the solar evaporator The drainage water collected from component D was discharged by sprinklers into the solar evaporator, and its discharge was programmed according to daily evaporation rates, to prevent excessive saline water ponding in the area. Salinity levels, total Se and extractable B concentrations are presented in Table 1. Because the solar evaporator was lined with a plastic sheet, there were no output pathways of salts, B and Se from the system, except for Se volatilization. The annual rate of Se volatilization was generally below 10 pg Se m -2 d -a. Low levels of Se volatilization probably resulted from low levels of organic matter, and therefore, low biological activities in a solar evaporator. Volatilization of Se accounted for only 1% of the annual total Se input in the drainage-water reuse system. The chemical form of Se in the solar evaporator was dominated by selenate (data not shown).

Discussion In this study, the grower selected a part of the farm to use 'clean water' on salt-sensitive crops and use another part of the farm for growing more-tolerant crops with poor-quality drainage waters. Segregating the farm into areas growing salt-sensitive and salt-tolerant crops entailed less operational complications. Reusing drainage

85

water produced from salt-sensitive crops, e.g. from component A, grown in these Westside soils, will eventually increase salinity and traceelement levels on the same soil site in component B if without having a salt management strategy. In order to address this reality as early as possible in the drainage-water reuse system, we annually moved our field site within 'component B' (consisting of moderately salt tolerant plants, e.g. canola, sunflowers), to different locations on the farm. Canola and sunflowers were our preferred choices for 'component B', because of both canola's ability to accumulate and volatilize Se (Bafiuelos 2002), and sunflowers' ability to both tolerate high-B/saline conditions and to be used in rotation with canola. Both crops produce viable economical products, e.g. biofuel, Se-enriched animal forage. Moreover, canola can be grown as a rain-fed cover crop, and hence requires fewer applications of drainage water containing salt, B and Se. Vegetation selected and grown in component B will be of great importance in the illustrated reuse program, because viable economic products are required by growers who accept or attempt this water reuse concept. The latter stages of the drainage-water reuse strategy, e.g. reuse components C and D, are very much dependent on the ability of selected long-term crops, e.g. saltgrass, cordgrass, pickleweed, to tolerate the highest levels or concentrations of sulphate-salinity and B from the irrigation water, and especially in the soil. To some extent, winter rains and application of clean water at the beginning of each growing season in early spring may help reduce salinity and B levels in the soil surface at the beginning of each growing season in components C and D. These rains may be helpful to growers who are generally unfamiliar with irrigation practices with such poor-quality waters used in components C and D. In addition, if plant transpiration rates are eventually reduced by salineor B-induced stress, less water will be taken up by the plant species in components C and D and waterlogging could occur and result in a low redox potential in the root zone. Benes et aL (2004) are presently investigating water relationships with salt-tolerant plant species in component D. Clearly, growing crops successfully in components C and D requires more information on both irrigation production practices and on salt-management strategies that will be essential for growers using a drainagewater reuse system. Decreases in the infiltration rate (IR) of a soil will likely occur over time, with continual application of poor-quality water, because of the

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gradual deterioration of the soil's structure and the formation of a surface seal. The IR is more sensitive to exchangeable Na, EC and pH than hydraulic conductivity. Drainage from soils where water is reused must be adequate, so that the salts, including exchangeable Na and B, are removed from the root zone. Boron is more difficult to leach than other salts, because it is adsorbed on to clay materials - hydroxyl oxides of A1, Fe and Mg (Keren & Bingham 1985). Other constituents of drainage water, particularly arsenic, chromium, molybdenum and dissolved solids, can also create problems associated with disposing/reusing such waters (Ong et al. 1997). Soil clay content and type will be important in influencing the long-term stability of soil structure and hydraulic properties in a waterreuse system, because of the large surface area of clay particles and their thin, platy shapes and negative lattice charges, which are balanced by exchangeable cations. In addition, some of the soils of the San Joaquin Valley are of a gypsiferous nature. When excess water is applied for leaching purposes, gypsum may dissolve, Ca is released, and it exchanges with adsorbed Na. Consequently, the soil water becomes a Na2 SO4-type water of high salinity, with a much greater solubility than gypsum. Gypsum's dissolution will act as a salt source until it is dissolved. Hence, the sodium adsorption ratio (SAR) of the poor-quality water is a critical parameter to monitor, because it tends to cause swelling and soil dispersion. In addition to salinity, the potential phytotoxicity of B; its relative immobility in the root zone; and boron's interaction with salinity, may be major limitations to long-term reuse of drainage water containing B concentrations of >5 mg L -1 (Letey et al. 2001). Although B did not appear to exert any noticeable effects on the selected plant species used in this drainagereuse system, its increasing concentrations over time will need to be managed. If leaching is used as a management tool for extractable soil-B concentrations in the root zone, at least three times more water may be required to leach B than that amount of water necessary to leach non-B salts (Oster et al. 1999). In this regard, two fundamental questions need to be addressed: is adequate low-salinity- or freshwater available for this management practice? Where does the leachate go? Readers may find more detailed information on drainage water by perusing the following research articles: Rhoades (1999), Oster et al. (1999) and Ayars (2003). Unique to the drainage-water reuse strategy

in central California is the presence of Se in the drainage waters used for irrigation. Because Se is a toxicant of concern, extended research is needed for the removal of Se from the drainage reuse system. Under field conditions, Se ecotoxicity was minimized with various management approaches, including the use of Zon Gun Propane Canon (a loud sound-making device) and bird-scare flagging tape to keep birds away from the drainage-reuse sites, particularly in the sections of D (halophytic plant section) and E (solar evaporator). In this study, removing Se via biological volatilization was observed only at some sites with selected plant species. Among the species tested, pickleweed, a salt-tolerant vascular plant species ( S a l i c o r n i a bigelovi), volatilized the greatest amount of Se, followed by canola and saltgrass (Lin et al. 2002). Identifying key plants for volatilization may represent a contributing technology for the remediation of Se contaminated waters and soils, because Se is volatilized from the ground surface and passes harmlessly into the atmosphere (Terry et al. 1992; Frankenberger & Karlson 1994; Terry et al. 2000). Vegetation is generally important to Se volatilization, not only because plants volatilize Se directly, but also because plants may create rhizosphere environments that support specific soil micro-organisms that may contribute significantly to Se volatilization (Azaizeh et al. 1997). Plant utilization will always be the most important consideration for the growers when adapting such a drainage-water reuse system. Currently, field studies are under way to test the feasibility of identifying useful salt-tolerant forage crops that could grow well with saline drainage water (Grattan et al. 2002). Bafiuelos and Mayland (2000) have harvested Seenriched plants grown under saline conditions, and have used them as part of a feed ration for sheep and dairy cows. Selenium is an essential trace element for animals, and Se deficiencies are generally a far greater problem than Se toxicities in animals in many other regions in the world (Mayland 1994). While excess Se caused ecologists to be concerned about the safety of wildlife in the San Joaquin Valley, Se deficiency in the diets of cattle is more of a problem there and in other regions. Harvesting Se-enriched crops (such as canola produces products, including Se-enriched feed and oil for biofuel) is of potential economical importance for growers who carefully reuse drainage water as an additional source of water for irrigation. Cattle and sheep may consume seleniferous plant tissues up to 5 mg kg -1 DM for a portion of their feed ration without suffering from Se toxicity (Mayland et al. 1989). Growers and animals will,

REUSE OF AGRICULTURAL DRAINAGE WATER however, require an adjustment period in order to familiarize themselves with utilizing or feeding upon crops that may not typically be in regular use in livestock production; palatability can be increased by blending with other forages.

Condusions Irrigation management is essentially the most important strategy for reducing the volume of fresh water applied and drainage water produced in any agricultural region worldwide. Since salts are imported from the central California soils with irrigation water, a means of ultim a t e l y isolating salts from productive agricultural soils is required for sustainability. Otherwise, salts will accumulate. When, however, drainage water is produced, using drainage water for irrigation can serve two purposes: one is to dispose of water that would otherwise be costly to remove; the other is to use drainage water as a valuable irrigation resource and reduce the requirement for goodquality irrigation water. The long-term concern regarding reusing drainage water of the qualities already discussed is that it may lead to an excessive salinity buildup, as well as contribute to the deterioration of soil quality, namely soil permeability and tilth. The likelihood of these problems increases as the S A R (sodium adsorption rate) increases and as EC increases. The consequence can be impermeable and crusted soils with poor stand establishment. This p r o b l e m can be managed by the use of soil amendments (e.g. gypsum) and a p p r o p r i a t e tillage practices. Furthermore, leaching salts from topsoils by subsurface tile-drainage systems will be essential when the soil-plant system becomes saltsaturated. M o n i t o r i n g p r o g r a m m e s for soil salinity should be implemented for understanding the complexity of salinity's spatial variability and its dynamic nature. Selecting plant species that are suitable for the different components of the drainage-water reuse systems will always be complicated by drainage-water compositions; trace elements, e.g. B and Se; field variability; specific cultivation and irrigation practices; and p r o d u c t utilization for the selected species. All these factors become even more complicated in the reuse components of C and D. If plants or plant products are to be considered as part of a feed ration, other trace elements, e.g. Mo, Cu, and even the macronutrient-S, need to be monitored for their potential antagonistic relations within animals. Lastly, a major issue in the present water reuse strategy that needs to be addressed

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is the presence of residual Se at toxic levels in the solar evaporator. Hence, it is essential that one apply water to the pond at the rate at which it evaporates, thus not ponding any water that would attract waterfowl. The authors gratefully acknowledge the funding support by California State University Agricultural Research Initiative (to Banuelos) and Office of Science (BER), U.S. Department of Energy (DEFG02-03ER6321 to Lin) during the preparation of this publication, and to J. Diener, owner of Red Rock Ranch.

References AYARS,J.E. 2003. Field crop production in areas with saline soils and shallow saline groundwater in the San Joaquin Valley of California. Journal of Crop Production, 7, 353-386. AYARS, J.E., HUTMACHER, R.B., SHONEMAN, R.N., VAIL, S.S. & PFLAUM,T. 1993. Long term use of saline water for irrigation. Irrigation Science, 14, 27-34. AZAIZEH, H.A., GOWTHAMAN,S. & TERRY, N. 1997. Microbial selenium volatilization in rhizosphere and bulk soils from a constructed wetland. Journal of Environmental Quality, 26, 666--672. BAlqUELOS,G.S. 2002. Irrigation of broccoli and canola with boron and selenium-laden effluent. Journal of Environmental Quality, 31,1802-1808. BAIqUELOS,G.S. & MAYLAND,H.E 2000. Absorption and distribution of selenium in animals consuming canola grown for selenium phytoremediation. Ecotoxicology and Environmental Safety, 46, 322-328. BENES,S., GRATrAN,S., FENCH, C. & BASINAL,L. 2004. Plant selection for IFDM. In: JACOBSEN, T. & BASINAL,L. (eds) Managing Agricultural Irrigation Drainage Water: A Landowner's Manual A Guide from California State Water Resources Control Board. Hudson Orth Communications, 5, 1-6. BERNER,R.A. 1984. Sedimentary pyrite formation: an update. Geochimica et Cosmochimica Acta, 48, 605-615. CERVINKA, V., DIENER, J. ET AL. 1999. Integrated System for Agricultural Drainage Management on Irrigated Farmland. Bureau of Reclamation, US Department of the Interior, Final Research Report, 4-FG-20-11920, 41. DEVEREL, N.J., FIO, J.L. & DUBROVSKY, N.M. 1994. Distribution and mobility of selenium in groundwater in the western San Joaquin Valley of California. In: FRANKENBERGER,W.T., JR & BENSON,S. (eds) Selenium in the Environment. Marcel Dekker, New York, 157-184. FRANKENBERGER, W.T., JR & KARLSON, U. 1994. Microbial volatilization of selenium from soils and sediments. In: FRANKENBERGER, W.T., JR & BENSON, S. (eds) Selenium in the Environment. Marcel Dekker, New York, 369-389. GOOROHOO,D., BENES,S., ADHIKARI,D. & SENATORE, K. 2005. Characterization of Soil Irrigated with

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Saline-Sodic Drainage Water: Chemical Composition. Presented at International Salinity Forum, Riverside, CA, 24-27 April, 2005. GRATFAN, S.R. & RHOADES,J.D. 1990. Irrigation with saline round water and drainage water. In: TANJI, K.K. (ed.) Agricultural Salinity Assessment and Management Manual. American Society of Civil Engineers, Manuals and Reports on Engineering Practice. ASCE, New York, 71, 432-449. GRATTAN, S., GRIEVE, C., POSS, J., ROBINSON, P., SUAREZ, D. & BENES, S. 2002. Evaluation of Salttolerant Forages for Sequential Reuse Systems. UC Salinity/Drainage Research Program, 2001-2002 Annual Report, UC Center for Water Research, 97-118. IRWIN,W.E & BARNES,I. 1975. Effect of geologic structure and metamorphic fluids on seismic behavior of the San Andreas fault system in Central and Northern California. Geology, 3, 713-716. KEREN, R. & BINGHAM,ET. 1985. Boron in water, soil, and plants. In: STEWART, B.A. (ed.) Advances in Soil Science. Vol. 1, Springer Verlag, New York, 229-276. LETEY, J., GRATTAN, S., OSTER, J.D. & B1RKLE, D.E. 2001. Findings and Recommendations to Develop the Six-year Activity Plan for the Department's Drainage Reduction and Reuse Program. California Department of Water Resources Final Report #98-7200-B80933. Sacramento, CA. LIN, Z.-Q., CERV1NKA,V., PICKERING,I.J., ZAYED,A. t~ TERRY, N. 2002. Managing selenium-contaminated agricultural drainage water by the integrated onfarm drainage management system: role of selenium volatilization. Water Research, 12, 3149-3159. LIN, Z.-Q., SCHEMENAUER,R.S., CERVINKA,V., ZAYED, A. & TERRY, N. 2000. Selenium volatilization from the soil - Salicornia bigelovii Torr. treatment system for the remediation of contaminated water and soil in the San Joaquin Valley. Journal of Environmental Quality, 29, 1048-1056. MAAS, E.V. & GRATTAN, S.R. 1999. Crop yields as affected by salinity. In: SKAGGS, R.W. & VAN SCHILFGAARDE, J. (eds) Agricultural Drainage. Agronomy Monogram, 38, American Society of Agriculture, Crop Science Society of America, Soil Science Society of America, Madison, WI, 55-110. MCNEAL J.M. & BALISTERI, L.S. 1989. Geochemistry and occurrence of selenium: an overview. In: JACOBS,L.W. (ed.) Selenium in Agriculture and the Environment. Soil Science Society of America, Special Publications, 23. American Society of Agriculture and Soil Science Society of America, Madison, WI, 1-13. MAYLAND, H.F. 1994. Selenium in plant and animal nutrition. In: FRANKENBERGER, W.T., JR & BENSON, S. (eds) Selenium in the Environment. Marcel Dekker, New York, 29-46. MAYLAND, H.E, JAMES, L.J., PANTER, K.E. & SONDEREGGER,J.L. 1989. Selenium in seleniferous

environments. In: JACOBS, L.W. (ed.) Selenium in Agriculture and the Environment. Soil Science Society of America, Special Publications, 23. American Society of Agriculture and Soil Science Society of America, Madison, WI, 15-50. ONG, C.G., HERBEL, M.J., DAHLGREN, R.A. & TANJI, K.K. 1997. Trace element (Se, As, B) contamination of evaporates in hypersaline agricultural evaporation ponds. Environmental Science and Technology, 31, 831-836. OSTER, J.D. 1994. Irrigation with poor quality water. Agricultural Water Management, 25,271-297. OSTER, J.D., SHA1NBERG,J. ~; ABROL, J.P. 1999. Reclamation of salt-affected soils. In: SrCAGGS,R.W. & VAN SCmLFGAARDE, J. (eds) Agricultural Drainage. Agronomy Monogram, 38, American Society of Agriculture, Crop Science Society of America, Soil Science Society of America, Madison, WI, 659-691. PRESSER, T.S. 1994. Geologic origin and pathways of selenium from the California Coast Ranges to the west-central San Joaquin Valley. In: FRANKENBERGER, W.T., JR & BENSON, S. (eds) Selenium in the Environment. Marcel Dekker, New York, 139-155. RHOADES, J.D. 1999. Use of saline drainage water for irrigation. In: SKAGGS, R.W. & VAN SCHILEGAARDE,J. (eds) Agricultural Drainage. Agronomy Monogram, American Society of Agriculture, Crop Science Society of America, Soil Science Society of America, Madison, WI, 615-658. SAN JOAQUIN VALLEY DRAINAGE IMPLEMENTATION PROGRAM 2000. Evaluation of 1990 Drainage Management Plan for the Westside San Joaquin Valley, California. Final Report submitted to the Management Group of the San Joaquin Valley Drainage Implementation Program, January 2000, San Joaquin Valley Drainage Implementation Program and University of California Ad Hoc Coordination Committee, 87 pp. SHALHEVE%J. 1994. Using water of marginal quality for crop production. Major issues. Agricultural Water Management, 25,233-269. SHANNON,M.C., SUHAYDA,C.G. ETAL. 1997. Water Use of Eucalyptus camaldulensis, Clone 4544, in Saline Drainage Reuse Systems. California Department of Water Resources Report 2. SHENNAN, C., GRATTAN, S.R. ET AL. 1995. Long-term feasibility of irrigating processing tomato with saline drainage water in a three-year rotation with cotton. Journal of Environmental Quality, 24, 476-486. TERRY, N., CARLSON, C., RAAB, T.K. & ZAYED,A.M. 1992. Rates of selenium volatilization among crop species. Journal of Environmental Quality, 21, 341-344. TERRY, N., ZAYED, A.M., DE SOUZA, M.P. & TARUN, A.S. 2000. Selenium in higher plants. Annual Review of Plant Physiology and Plant Molecular Biology, 51, 401-432.

The effects of agricultural practices on arbuscular mycorrhizal fungi J A N J A N S A a, A N D R E S W I E M K E N 2 & E M M A N U E L

FROSSARD 1

1 E T H Zurich, Institute o f Plant Sciences, Eschikon 33, CH-8315 Lindau (ZH), Switzerland (e-maik [email protected]) 2University o f Basel, Institute o f Botany, Hebelstrasse 1, CH-4056 Basel, Switzerland

Abstract: Arbuscular mycorrhizal fungi (AMF) form symbiotic associations with the majority of land plants, including many important agricultural crops. These fungi facilitate plant nutrient uptake, promote soil aggregation and use a significant portion of reduced carbon from the plants. AMF functional traits differ considerably among and within species, meaning that functional properties of a mycorrhizal community depend on its composition. Here we review studies exploring the effects of agricultural practices such as tillage, crop rotation, fertilization, pesticide application, irrigation and grazing on AMF communities. Although it is difficult to generalize the results of studies performed under different soil and climatic conditions, some universal patterns emerge. For example, soil tillage reduces the abundance of Scutellospora spp.; phosphorus fertilization lowers the extent of AMF root colonization; and diversification of crops results in more diverse AMF communities. We now need to design simple and reliable field tests for quantifying the effects of AMF communities on crop growth, yields and sustainability of the agro-ecosystems.

Soil is a reservoir of biodiversity, particularly with respect to the diversity of microbial communities. Though micro-organisms are essential for the functioning and sustainability of all natural ecosystems, they are frequently ignored, due to their small size and methodical difficulties in studying them (Prosser 2002). It is estimated that only about 1% of the bacterial species present in soils can be cultivated. Identification methods independent of culturing, such as molecular and chemical fingerprinting, have recently shed more light on the enormous taxonomic diversity and composite function of microbial communities in the soil (Nannipieri et al. 2003). However, the link between functional and taxonomic diversities of soil microbial communities is still a poorly resolved issue.

Arbuscular mycorrhizal fungi In this review, we focus on one group of soil micro-organisms, namely the arbuscular mycorrhizal fungi (AMF), belonging to the phylum Glomeromycota (Schtil31er et al. 2001). All known members of this very ancient phylum (except one, Geosiphon pyriform&) are associated with plant roots, forming the so-called mycorrhizal symbiosis (SchgBler 2002). The roots of the majority of land-plant species are colonized by these fungi, whose hyphae extend into the surrounding soil, where the spores are also formed. The range of A M F hosts includes

agricultural crops such as maize, wheat, potatoes, sunflower, millet, rice, bananas, cassava, yams, flax, coffee and soybeans, to name but a few. Only a minority of plant species do not associate with AMF. These so called non-mycorrhizal plant species occur in several families, such as Brassicaceae, Chenopodiaceae, Cyperaceae, Caryophyllaceae, Juncaceae and A m a r a n thaceae (Hirsch & Kapulnik 1998), of which rapeseed, mustard, cabbages, cauliflower, sugar beet and spinach are important agricultural crops. All A M F depend on the association with living plant roots to fulfil their life cycle. Despite numerous attempts, no method has yet been established for A M F cultivation without plant roots for an extended period of time. Worldwide, only about 150 A M F species have been described, based mainly on morphological differences between spores (Smith & R e a d 1997). For a long time, spore morphotyping was virtually the only way to identify A M F species, but recent advances in molecular approaches have allowed identification of these fungi in roots and in soil, by using signature D N A sequences (Redecker et al. 2003). However, greater diversity has been observed on the molecular level than at the morphological level, supposedly as a consequence of the unusual, multigenomic cellular organization of the A M F (Koch et al. 2004). It has been proposed that these fungi have reproduced solely asexually since their evolutionary origin about 500 million years ago (Sanders et al. 2003). These two facts

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 89-115. 0305-8719/06/$15 9 The Geological Society of London 2006.

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(absence of sexual reproduction and multigenomic nature) are likely reasons for the apparent DNA sequence polymorphism within a single AMF spore (Sanders 2004a).

Global and local distribution of AMF AMF are present worldwide in almost every terrestrial ecosystem, and their distribution is affected by environmental factors (Rillig 2004a). Probably, their distribution is affected only to a lesser extent by isolated evolution on different continents, as it is currently supposed that these fungi had already undergone their major evolutionary radiation during Gondwanaland times (Simon et aL 1993). Locally, A M F disperse rapidly through wind, animal vectors, and growth through the soil (Allen & Allen 1992). Temperature, the plants associated with them, and soil type, all appear to be important factors determining occurrence of AMF species and composition of AMF communities. However, our knowledge of those interactions is still far from complete (Fitter et al. 2004). For example, there is a growing body of evidence, contradicting earlier beliefs, on preferences in associations between particular AMF and plant species (Bever et aL 1996; Vandenkoornhuyse et al. 2002).

AMF functions The fungi forming an arbuscular mycorrhizal symbiosis are conferring a multitude of benefits to the plants. They facilitate plant uptake of some mineral nutrients from soil, thus alleviating nutrient deficiency. This is particularly important for nutrients with little mobility in the soil, such as phosphorus (P), zinc (Zn) and copper (Cu) (Marschner & Dell 1994), but also for nitrogen (N) (Govindarajulu et al. 2005). Under acidic soil conditions, mycorrhizal cereal plants were also shown to take up more potassium (K) than non-mycorrhizal plants (Clark 2002). Contradictory evidence is available for the role of AMF in plant uptake of elements such as iron, manganese (Mn) and aluminium; both enhancements and reductions of their uptake by plants due to mycorrhizal symbiosis have been reported. This variation appears to be determined by the identity of both fungal and plant genotypes, as well as by environmental context (Clark & Zeto 2000). The beneficial influence of mycorrhizal association on plant nutrition and growth is usually greater in soils with low fertility (Douds & Millner 1999). Besides their effects on plant nutrition, AMF also play an important role in modulation of

plant resistance to pathogens and tolerance to environmental stresses (Newsham et aL 1995), and in soil stabilization by promoting soil aggregation (R. M. Miller & Jastrow 2002). This is accomplished by mechanical binding of soil particles by AMF hyphae, and through hyphal exudation of compounds such as glomalin (Rillig 2004b). A M F receive plant-reduced carbon (C), and redistribute some of it throughout the soil - thus having a major impact on soil microbial activity (Jakobsen et al. 2002). AMF also affect both inter- and intraspecific competition of plants (van der Heijden et al. 2003). For example, inoculation of maize and sorghum with G l o m u s c l a r u m and Gigaspora margarita in northern Cameroon resulted in reduction of Striga hermonthica infestation by 30% and 50%, respectively (Lendzemo et al. 2005). A remarkable functional diversity exists within and among AMF species (Jansa et al. 2005). Therefore, it is difficult to link taxonomic and functional diversities of the AMF, as it is unclear what functions could be assigned to a species and what to a genotype. As different species and genotypes of AMF have different functional properties, it is probable that the composition of A M F community strongly affects ecosystem functioning. A change in composition of AMF may have functional consequences for the plants and the soil, and may affect the productivity and/or sustainability of an ecosystem and the extent of delivery of ecosystem services (Rillig 2004a).

Agricultural practices and AMF Farmers implement agricultural practices for optimizing seedbed conditions, weed and pest control, soil fertility, water supply, etc. It has been suggested previously that some of these practices may affect some of the benefits conferred to plants by symbionts such as mycorrhizal fungi and nodule-forming bacteria, but it has been noted that the available data were not sufficient to test this hypothesis thoroughly (Kiers et al. 2002). In this review we bring together some of the studies on the effects of different agricultural practices on AMF spore and hyphal densities in soil and colonization of roots. We also examine whether and how the different agricultural practices affect diversity and composition of AMF communities (Fig. 1). Further, we will discuss the mechanisms underlying those changes, as well as the potential impacts of such changes for plant growth and performance. On purpose, here we will only briefly cover some specific practices applied in horticulture (such as soil sterilization), and we

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[ Consequences I

Agricultural management Low

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Size of AMF community AMF infectivity AMF diversity, species abundance Plant nutrition and growth Soil properties

Tillage croprotation Fertilization Pesticide use Irrigation Burning, grazing

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

> High

Intensity, direction of changes Management practice application, cessation Fig. 1. Conceptual framework of this review, dealing with the effects of various agricultural management practices on the size, diversity and composition of AMF communities, as well as on the AMF-mediated effects on plants and soil. cover neither contained systems such as glasshouses nor the systems relevant to forestry. Specifically, we will concentrate on: soil tillage, crop rotation, inputs of fertilizers, use of pesticides and soil sterilization, irrigation, and burning and grazing of vegetation. Additionally, we will briefly mention three phenomena, which closely accompany human activities in agroecosystems, namely inputs of pollutants, soil compaction and topsoil movement. A short section will then discuss the possibility of reversing negative human impacts on A M E

Soil tillage Soil tillage is used to control weeds, prepare seedbeds, incorporate crop residues and fertilizers, and increase water infiltration and soil temperature (Warkentin 2001). Soil tillage may also increase the rate of soil erosion. Currently experienced expansion of agricultural land under reduced or no-tillage management is primarily driven by the need to reduce soil erosion (P. L. O. D. Machado & Silva 2001). Upon reducing soil tillage, alternative ways to control weeds, such as application of herbicides, must be established (A. E L. Machado et al. 2005).

The proportion of fungal to bacterial biomass is usually higher in non-tilled than in conventionally tilled soils (Spedding et al. 2004). This relative increase in fungal biomass could be attributed to pronounced development of A M F under non-tillage, as concluded from analyses of fatty acid profiles in differently tilled soil (Drijber et al. 2000). Tillage reportedly reduced A M F spore and hyphal length densities, as well as decreased glomalin concentrations in both temperate and tropical soils (Wright et al. 1999; Boddington & Dodd 2000a). The composition and diversity of A M F spore communities were affected by tillage in a number of studies (Kabir 2005). For example, spores of Glomus etunicaturn and G. caledonium were more abundant in tilled temperate soils, whereas other species (e.g. Glomus occultum, Scutellospora pellucida, Acaulospora paulinae and Entrophospora infrequens) were more abundant in non-tilled soils (Galvez et aL 2001; Jansa et al. 2002). These results based on the observation of spores were later corroborated by analysing A M F communities within maize roots by means of D N A markers. This survey confirmed that Scutellospora spp. was absent in maize roots growing on a Swiss tilled Luvisol, but was relatively

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abundant in the maize roots growing in the nearby non-tilled soil (Jansa et al. 2003). The extents of root colonization by G l o m u s c l a r o i d e u m and G. i n t r a r a d i c e s were, on the other hand, lower under no-tillage than under conventional tillage (Jansa et al. 2003). Given the differences in both the spore and the hyphal length densities, it is not surprising that the rates of AMF colonization of roots are usually different in differently tilled soils. Faster A M F colonization is commonly observed in the nontilled as compared to the tilled soils, resulting in a greater proportion of the roots being colonized by A M F in the non-tilled soil early in the season (Mozafar et al. 2000). However, the different rates of root colonization in differently tilled soils could sometimes be masked by other factors such as crop rotation. For example, root colonization of wheat following rapeseed was faster under non-tillage than under conventional tillage, whereas the rates of root colonization were not significantly different in differently tilled soils in maize following wheat, or in wheat following maize in a Swiss sandy soil (Anken et al. 2004). Due to faster establishment of A M F colonization in roots under no-tillage, early uptake of P and Zn by plants is usually higher under notillage than under conventional tillage (M. H. Miller 2000), and this may eventually translate to yield-increases by plants such as maize (Grant et al. 2001). However, some researchers report maize being colonized faster by AMF under no-tillage, resulting in higher P uptake, but not translating to a yield increase compared to conventional tillage (Galvez et al. 2001). This indicates that other factors may sometimes limit plant growth and yield under no-tillage management, such as low soil temperature or weed development (Galvez et al. 2001). Besides the nutritional effects of A M F on plants grown under no-tillage, higher length densities of AMF hyphae and higher concentrations of glomalin in no-tilled soils are likely to contribute to greater stability of soil aggregates in no-tillage as compared to tilled soils (Wright et al. 1999). Indirectly, this notion was also supported by application of captan (a fungicide) to differently tilled soil. Application of the fungicide caused a more pronounced decrease in the aggregate stability of the non-tilled than the tilled soils, whereas application of oxytetracycline (an antibiotic impairing bacterial activity) did not show a differential effect in differently tilled soils (Bossuyt et al. 2001). It is, however, likely that some other factors, such as soil type and texture, turnover of soil organic matter, and time of exposure to no-tillage management may

modulate the effects of A M F on physical properties of the soil, such as aggregate stability (Rhoton 2000; Kabir 2005). Since the densities of both the spores and AMF hyphae in soil decline with depth, it has first been proposed that tillage reduced the spore and hyphal densities in the topsoil simply by mixing the top soil with lower soil layers (T. E Smith 1978). More recently, the direct effects of soil disturbance imposed by tillage on the AMF have been recognized. Cutting or sieving of soil was shown to reduce its mycorrhizal infectivity (i.e. its capacity to cause colonization of plant roots) and, subsequently, to reduce mycorrhizal development and P uptake benefits to plants such as maize and soybean (Jasper et al. 1989; Goss & de Varennes 2002). Differential tolerance of AMF species to soil disturbance, fertilizer accumulation, and preferential associations with some weed species are likely reasons for the observed shifts in AMF community composition due to soil tillage. For example, cutting of a preestablished mycelium network reduced root colonization by G i g a s p o r a rosea and increased that by G l o m u s m a n i h o t i s in D e s m o d i u m ovalif o l i u m plants (Boddington & Dodd 2000b). This mechanism probably explains the low abundance of S c u t e l l o s p o r a spp. and the dominance of G l o m u s spp. in tilled soils (Jansa et al. 2002, 2003). Increase in P availability in the uppermost layer of non-tilled soil, due to the absence of fertilizer incorporation (Salinas-Garcia et al. 2002), may also create less-conductive conditions for some A M F species. Additionally, under no-tillage, other weed species are usually encountered than in the tilled soils (Streit et al. 2002), and this is probably another important determinant of composition and dynamics of AMF communities.

Crop rotation Crop rotation is important to maintain and improve soil quality, nutrient and water availability; enhance N inputs through biological N fixation; prevent soil erosion, break down the life cycles of plant pathogens; and to control weeds (Watson et al. 2002; Haramoto & Gallandt 2004). The densities of A M F spores and hyphae, as well as AMF infectivity, are usually lower in bare soil or in soils planted with nonmycorrhizal crops such as rapeseed, than in the soils cropped with mycorrhizal host species (Oliveira & Sanders 1999; Allen et al. 2001; Jansa et al. 2002). Spore densities in the soil, mycorrhizal infectivity and root colonization may also vary depending on the identity of the

AMF IN AGRICULTURAL SOILS crop species (Hendrix et al. 1995). For example, A M F spore densities were higher under maize than under soybeans in Iowa (Troeh & Loynachan 2003) and A M F infectivity several times higher was recorded under tall fescue (Festuca arundinacea) than under soybeans in Kentucky (An et aL 1990). More than 30 AMF spores per gram of soil were found in Swiss grassland soil, whereas less than 10 spores per gram were usually found in soils under continuous maize (Oehl et al. 2003). A legume (Crotalaria brevifolia) intercropped with coffee in Brazil caused an increase in A M F spore density in the soil (Colozzi & Cardoso 2000). The colonization of sorghum roots in Niger and Burkina Faso was 10-15% higher when rotated with either cowpeas or groundnuts than if grown continuously (Alvey et al. 2001), and colonization of roots of irrigated cotton in Australia was higher when intercropped with peas or wheat than when grown continuously (Hulugalle et aL 1999). Crop rotation strongly affects both the diversity and composition of A M F spore communities in the soil, with higher A M F diversity usually encountered under rotated crops than under monocultures (Vestberg et al. 1999; Oehl et al. 2003). Soils under continuous soybeans in Kentucky were dominated by Gigaspora spp., whereas A M F communities under other crops (maize, sorghum, tall fescue) rotated with soybeans were dominated by Glomus spp. The rescue markedly reduced the spore density of Glomus macrocarpum, whereas sorghum increased the spore density of the same species (An et al. 1993). In other studies carried out in the USA, Gigaspora gigantea, Glomus albidum, G. mosseae and G. etunicatum dominated A M F communities under maize; G. caledonium and G. microcarpum were abundant under soybeans; and Glomus occultum was most abundant under wheat and barley (Johnson et al. 1991; Douds & Millner 1999; Troeh & Loynachan 2003). Crop rotation of maize with soybeans also changed A M F community composition as compared to maize monocropping in Nigeria (Sanginga et al. 1999). Coffee and Crotalaria breviflora stimulated populations of different A M F in their rhizospheres in Brazilian soil. Scutellospora and Gigaspora spp. spores were more abundant under Crotalaria, whereas Acaulospora spp. occurred more often under coffee (Colozzi & Cardoso 2000). By means of molecular markers, Mathimaran (2005) showed that Scutellospora heterogama was more abundant in roots of continuous maize, as compared to the maize rotated with Crotalaria grahamiana in western Kenya.

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Changes in infectivity and composition of A M F communities due to crop rotation may affect the magnitude of symbiotic benefits conferred to the crops. For example, bare-soil fallow reduced A M F infectivity and, subsequently, P acquisition of sunflowers in Australia (Thompson 1987). In Pennsylvania, the winter wheat cover crop increased A M F infectivity in soils as compared to bare soil, and this resulted in a better yield of maize following wheat in the subsequent season (Boswell et al. 1998). Early dry-matter production and P uptake of maize were usually higher upon rotation with soybeans or sunflower than under maizebare-soil fallow or maize-rapeseed rotation (Arihara & Karasawa 2000). Affinities of different A M F species towards certain host plants have been recognized since a few years (Bever et al. 1996). This species preferentiality (which is distinct from species specificity implicating full incompatibility of certain plant-fungus combinations) could explain why A M F composition and diversity is affected by different crop-plant species. Intercropping and crop rotation considering more plant species may thus result in more diverse A M F communities in the soil, and may also prevent selection for specific mycorrhizal 'cheaters'. Mycorrhizal 'cheaters' are AMF species which consume a lot of C from the plant without improving the acquisition of mineral nutrients (N. C. Johnson et al. 1992). Analogously, this also applies for less-intensive cropping systems with higher abundance and diversity of weeds, as compared to high input cropping systems with lots of herbicide inputs. On the other hand, higher diversities of A M F communities in unplanted soil or in soil planted with non-mycorrhizal species such as sedge (Carex sp.) compared to soil planted with mycorrhizal host plants (D. Johnson et al. 2004) is due to the higher A M F community evenness in soil where symbiosis is not established. The mycorrhizal plant narrows the diversity of the A M F communities by favouring some members of the indigenous community. On the other hand, the mycorrhizal host plant obviously increases the infectivity, spore and hyphal length densities of the A M F (as well as biomass and activity of soil bacteria) compared to the unplanted soil or soil under non-host-plant species (D. Johnson et al. 2004). In contrast to the rather unequivocal evidence of crop rotation with different mycorrhizal host plants on A M F community composition (= abundance of particular species), the effects of different mycorrhizal hosts on A M F spore densities in soil are less apparent. This is because different A M F species may (but must

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not) be sporulating at different rates (Sanders 2004b). For example, in a tropical soil in Cameroon, the clearance of rainforest resulted in a reduction of A M F spore density in the soil within three months and in a subsequent steep rise in the spore numbers associated with the planted tree T e r m i n a l i a i v o r e n s i s (Mason et al. 1992). The composition of the A M F spore communities in the rainforest and under Terminalia was, however, quite different. It is thus not possible to conclude that lowering plant diversity necessarily results in changes in A M F spore densities. Although different plant species are preferentially associating with different AMF species, some overlaps exist. This may be important for intercropping, where different plant species are grown simultaneously in the same soil and may potentially share the same mycorrhizal hyphal networks. The significance of such sharing for the redistribution of mineral nutrients and the partitioning of C costs among different plant species is still little explored and deserves further attention (Simard & Durall 2004).

Inputs of fertilizers Fertilizers are applied to optimize crop nutrition and to replenish nutrients removed by harvested crops. The densities of A M F spores and hyphae in soils may slightly increase upon moderate application (5-15 kg P ha -1 a -1) of mineral P fertilizers to P-deficient soils such as shown in a lateritic loam in Western Australia (Thomson et aL 1992). The AMF are usually unaffected by moderate P inputs in more fertile soils such as Swedish Cambisols (Martensson & Carlgren 1994). Increasing inputs of mineral P fertilizers usually decrease the densities of A M F spores and hyphae. This has frequently been observed under different soil and climate conditions upon application of _>50kg P ha -I a -1 (Douds & Millner 1999; Kahiluoto et al. 2001). The spore and hyphal densities may increase, be unaffected, or decrease in response to mineral N fertilization (Bethlenfalvay et al. 1999; Eom et al. 1999). For example, several years of N fertilization at levels between 100 and 170 kg N ha -1 a -1 in the form of N H 4 N O 3 decreased the spore densities, particularly those of Gigasporaceae, in four Northern American grassland soils, but increased spore densities in a limestone-derived soil on Konza Prairie (N. C. Johnson et al. 2003). In contrast to mineral P and N fertilization, the application of organic fertilizers such as cattle manure and green manure mostly increase both the spores and hyphal

densities in soils (Baby & Manibhushanrao 1996; Gryndler et aL 2001). Like A M F spores and hyphal densities, the application of moderate amounts of mineral P fertilizers into P-deficient soils may increase plant-root colonization by AMF (Picone 2002). A further increase in fertilization levels, resulting in a situation where the soil P availability threshold is exceeded for the crops (this depends on the plant species and environmental conditions: see Picone 2002 and references therein), commonly decreases the AMFs colonization of roots of various plant species (Gryndler & Lipavsk 5, 1995; Kahiluoto et al. 2001). The negative responses of root colonization to mineral P fertilizers are usually fast and long-lasting. For example, arbuscular colonization of A g r o p y r o n d e s s e r t o r u m was reduced three days after P fertilizer application (Duke et al. 1994), and the colonization of cereal roots was still lower in fields previously overfertilized with P, even 10 years after stopping P fertilizer inputs (Dekkers & van der Werff 2001). On the other hand, reasonable degrees of root colonization (>70% of the root length being colonized) were observed in maize and soybeans in North American soils with available P well above those required for maximum yield (Lu & Miller 1989; Khalil et al. 1992). Moreover, AMF were still colonizing plants growing in a P-polluted soil with available P (calcium acetate-calcium lactate extraction) content reaching 12 g kg -1 soil (Renker et aL 2005). The responses of A M F root colonization to N fertilization are less consistent than for E Inputs of N into soil, through mineral fertilizers or aerial depositions, usually decrease root colonization (EgertonWarburton & Allen 2000; Baum et al. 2002), but moderate N inputs may have positive effects on the colonization of roots of tall-grass prairie plants (Eom et al. 1999). Both omitting and adding an excess (350 kg N ha -1 a -1) of mineral N inputs resulted in reduction of AMF root colonization in maize, compared to moderate (175 kg N ha -1 a-1) N inputs (Gryndler et al. 1990). Nutrients such as P and N supplied in organic fertilizers such as farmyard manure do usually have much less inhibitory effects on the A M F colonization of roots than mineral fertilizers (Gryndler et al. 1990; Joner 2000). In contrast to P and N, inputs of K and sulphur (in form of sulphate) seem to have only small effects on the levels of A M F colonization of plant roots (Heijne et al. 1992; Berreck & Haselwandter 2001). Increased supply of magnesium (6 to 12 mM in hydroponic solution) may, on the other hand, stimulate the colonization of maize roots by G l o m u s spp. (Gryndler et aL 1992).

AMF IN AGRICULTURAL SOILS Both P and N fertilization may affect the composition of AMF spore communities in soil (Ezawa et aL 2000; Treseder & Allen 2000), but the diversity of the communities is usually not affected (Hamel et al. 1994; Mathimaran et aL 2005). For example, in Cedar Creek, Minnesota, spore abundances of Gigaspora gigantea, G. margarita, Scutellospora calospora and Glomus occultum decreased, and that of G. intraradices increased, after eight years of P and N fertilization (Johnson 1993). In archived soil samples from the period 1937 to 1999, A M F spore communities in Californian shrublands markedly changed in response to anthropogenic N enrichment, with Acaulospora, Scutellospora and Gigaspora spp. completely disappearing, and Glomus aggregatum and G. leptotichum dominating A M F communities at later sampling dates (Egerton-Warburton et al. 2001). The abundance of Scutellospora sp. in Hawaiian soil was lower under N than under P fertilization; the occurrence of Glomus sp. was higher in the fertile than in the N-limited soil; and spore abundances of Gigaspora and Acaulospora spp. were not significantly affected by soil fertility (Treseder & Allen 2002). In a Swiss soil fertilized either with mineral or organic fertilizers, Glomus spp. were similarly abundant, but spores of Acaulospora and Scutellospora spp. were more abundant in soil that only received organic fertilizers (Oehl et al. 2004b). Also, different forms of organic fertilizers differentially affected A M F communities in other studies. For example, addition of leaf compost combined with either chicken litter or cow manure, enhanced spore populations of some A M F species (Glomus etunicatum and G. mosseae) relative to those found in soils fertilized with raw dairy-cow manure or with mineral fertilizer (Douds et aI. 1997). Much less is known about the responses of AMF communities in roots to fertilization, as compared to our knowledge about the responses of spore communities (reviewed above). One earlier study from Western Australia was based on the different morphologies of A M F root colonization structures (hyphae, arbuscules and vesicles). This study reported a change in the composition of A M F communities inside the roots, due to P fertilization; the colonization by Glomus spp. and Acaulospora laevis was increasing, and the colonization by Scutellospora calospora decreasing with increasing P inputs (Thomson et al. 1992). Another study employing molecular identification of AMF reported change in the composition of A M F communities in maize roots, due to P fertilization, in a Kenyan Ferral-

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sol; the extent of root colonization by Acaulospora mellea decreased in response to P fertilization in maize rotated with Crotalaria grahamiana, whereas the colonization by several other A M F species was not affected (Mathimaran 2005). In soils showing high concentrations of available P, the ways in which plants benefit from mycorrhizal symbiosis in terms of improving P uptake are lower than in soil containing little available P. This has been documented many times both in glasshouse experiments and in the field for various plant species (Miller et al. 1995; Treseder 2004). On the other hand, application of moderate amounts of P fertilizer into Pdeficient soils may increase mycorrhizal benefits, whose maximum is encountered under low, but not under extremely low, P availability (Bolan et al. 1984). For example, upon cultivation of leeks in silty-clay soil in Jordan, the colonization of roots decreased after application of >20 kg P ha -1 a -1 (supplied as triple superphosphate), and the benefits due to mycorrhizal inoculation (either with Glomus mosseae or with G. fasciculatum) were greatest upon addition of 20 kg P ha -1 a-1 (A1-Karaki 2002). Both P and N fertilization applied to soils for several years were shown to favour communities of AMF, which were generally less beneficial for the plants than the communities from unfertilized soils (Johnson 1993; Corkidi et al. 2002). General decrease in A M F hyphal-length density in soil in response to mineral P and N fertilization may result in lower concentrations of glomalin, which may in turn contribute to lower stability of soil aggregates upon mineral fertilization of the soils (Lovelock et al. 2004). However, evidence for this happening in agricultural soils is still scarce (Ryan & Graham 2002). Soil P availability is likely to directly influence development of the A M E A slight increase in P availability stimulated spore germination and hyphal growth of Glomus etunicatum, whereas a further increase in P availability reduced the hyphal growth dramatically (de Miranda & Harris 1994). Different A M F species respond differently to increases in P availability, and such differences may underlie shifts in A M F community composition upon soil P fertilization. For example, Scutellospora heterogama spore abundance decreased more rapidly with increasing P availability than that of Gtomus spp. (de Miranda & Harris 1994). The other mechanism possibly playing a role in reduction of A M F development under conditions of high soil P availability is mediated by the host plant. Under conditions of higher soil P availability, plants accumulate less-soluble carbohydrates in

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their root, thus possibly reducing carbon supply to their symbionts (Amijee et al. 1993). It has been shown that some AMF species such as S c u t e l l o s p o r a c a l o s p o r a are particularly sensitive to such decreases in carbohydrate concentration in the roots, meaning that these species will probably be eliminated first by P fertilization (Thomson et at. 1986). The effects of N fertilization on AMF are also both direct and indirect. Like R moderate levels of N fertilization, especially in an N-limiting system, may positively affect the development of both root colonization and sporulation. For example, low P combined with ambient N promotes the sporulation of different A M F species, but excess N can be inhibitory, e.g. for G i g a s p o r a spp. under hydroponic conditions (Douds & Schenck 1990). The results from field studies generally confirm this notion. For example, root colonization by AMF was increased by N and P additions into N- and P-limited sites in Hawaii, respectively (Treseder & Allen 2002). Nitrogen fertilization caused a decrease in A M F root colonization in grasses grown in soils containing sufficient P for plant growth (18.4 mg bicarbonate-extractable P kg-1), but not in grasses grown in soils with lower P availability (6.6 mg kg -1) (Corkidi et al. 2002). Therefore, the inconsistencies of experimental results quoting differential responses of AMF to N application could be attributed to the initial N status of the systems, as well as to the N:P ratios of available elements in the soils (Miller et aL 2002; Treseder & Allen 2002). It is also becoming evident that not only the nutritional content, but also other materials applied together with fertilizers, can influence the effects of fertilization on the AME This is particularly true for organic fertilizers (Khalil et al. 1992). The nutrients in organic fertilizers are usually less available and being released over a longer period of time that those applied as mineral fertilizers (Oehl et al. 2004a). It has been shown previously that slowly degradable polysaccharides such as cellulose and chitin are strongly promoting A M F sporulation (Gryndler et al. 2002; Gryndler et al. 2003). Application of low quantities of fertilizers, preferably in organic forms, may thus prove to be the way to allow optimal utilization of mycorrhizal symbiosis in low-fertility soils (Treseder & Vitousek 2001), whereas long-term application of high quantities of mineral N and P fertilizers may select for A M F communities that are less beneficial for the host plants (Kiers et al. 2002).

Use of pesticides and soil sterilization Pesticides refer to all substances used to control pests and diseases. Nowadays, approximately

60% of commercial pesticides are herbicides, 25% insecticides, and 15% fungicides (Anaya 1999). The effects of pesticides on AMF were unknown at the times of their introduction, but they were explored soon after that (Trappe et al. 1984; Schiiepp et al. 1987). In order to control pests or diseases in intensive production systems such as the vegetable-production or flower industries, the soil may be sterilized by steaming, methylbromide/chloropicrin fumigation, or by solarization. Sterilization unspecifically kills all living organisms in the treated soil layer, unlike the pesticides that mostly target only a certain range of organisms. The effects of pesticides and soil sterilization on the AMF and their consequences for plants will be reviewed in this section. The effects of pesticides on AMF are highly variable depending on the pesticide type, crop and A M F species, timing, application rate, and environmental conditions (Menge 1982; Sukarno et al. 1993; Schreiner & Bethlenfalvay 1997). It is not surprising that soil fumigants such as methylbromide and formaldehyde are most harmful to the AMF, temporarily eliminating them from the soil (Haas et al. 1987; Udaiyan et al. 1995). Soil solarization, a sterilization approach that is comparatively less destructive for AMF than soil fumigation with methamsodium or methylbromide (but similarly effective for the control of weeds) is likely to exert only small direct effects on AMF. However, it may indirectly reduce the infectivity of A M F communities by reducing perennial weed populations, with which the AMF are occasionally associated (Schreiner et al. 2001). The effects of fungicides on the AMF are more variable than those of soil fumigants. For example, systemic fungicides and benzimidazole derivatives such as benomyl and carbendazim can significantly reduce root colonization, viability of soil hyphae, and spore development of many AMF species (Dodd & Jeffries 1989a; Sugaranam et al. 1994; Kling & Jakobsen 1997). Other fungicides such as captan do not affect and can even enhance - AMF colonization, hyphal development in soil and spore production (Schtiepp & Bodmer 1991; Venedikian et al. 1999; Kj011er & Rosendahl 2000). Herbicides may negatively affect colonization of plant roots, hyphal and spore densities in soil if they impair the vitality of the host plant (e.g. a weed species), otherwise their effects on the A M F remain limited (Dodd & Jeffries 1989b; Allen & West 1993; Mujica et al. 1999). Insecticides and nematicides cause much less damage to the A M F than benomyl and other fungicides (Wan & Rahe 1998). Some insecticides may cause transient decrease in root colonization, hyphal

AMF IN AGRICULTURAL SOILS and spore densities in soil, whereas others such as monocrotophos usually promote root colonization of plants such as sorghum (Vijayalakshmi & Rao 1993), probably through decreasing the activity of hyphae-feeding nematodes and/or by improving the health of the host plants. Pesticides, particularly the fungicides, may affect community composition and diversity of AMF spore communities in soils (comparable data for A M F communities in the roots are not available yet). This is because pesticides may differentially suppress different A M F species. For example, germination of spores of G. m o s s e a e was inhibited by carbendazim and propiconazole, but not by tridemorph, whereas germination of G. g e o s p o r u m was inhibited by neither of those (Dodd & Jeffries 1989a). Additionally, captan was shown to reduce spore population of G. rosea, but had positive or no effects on spore populations of G l o m u s etunicaturn and G. m o s s e a e (Schreiner & Bethlenfalvay 1996). In contrast to pesticide application, soil sterilization is not likely to directly influence AMF community composition and/or diversity, because it is by its nature rather unspecific. However, since soil sterilization is only effective in the soil layer which is in direct contact with the fumigant or which is physically treated (e.g. heated), it affects mostly the AMF present in the uppermost soil layers. For example, it has been noted that application of soil fumigants such as methamsodium or methylbromide only eliminated A M F from the top 20 cm of the soil layer (Jawson et al. 1993; Kapulnik et al. 1994). This may be important for the recovery rates and composition of secondary A M F communities, as AMF spores are also present in soil well below 1 m depth, and as the composition of the spore communities changes with soil depth (Oehl et aL 2005). Soil sterilization and application of fungicides reducing A M F colonization of roots and development of mycelium in soil were shown to reduce P, Zn and N uptake by, and the growth of, different plants (Kapulnik et al. 1994; Schweiger & Jakobsen 1998; Dhillion & Gardsjord 2004) as well as soil aggregate stability (R. M. Miller & Jastrow 2002). The reduction of nutrient uptake by plants may, however, be insignificant for plants growing under conditions where nutrients are sufficient (Newsham et al. 1994). On the other hand, growth of citrus plants increased following benomyl application, as compared to untreated plants, probably by reducing the C costs of the symbiosis under P-sufficient conditions (Graham & Eissenstat 1998). Finally, the A M F can also significantly contribute to uptake of

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herbicides atrazine and trifluralin by roots of maize and soybeans and thus may potentially play a role in the toxicity of herbicides to some plants (Nedumpara et al. 1999). The mechanisms of action of pesticides on AMF are in most cases not precisely known. Pesticides may affect A M F both directly and indirectly, through their effects on the host plants, or on the soil biota (Bethlenfalvay 1993). Fungicides such as benomyl directly inhibit many A M F species. Benomyl has thus been frequently used for experimental killing of native A M F in soils (see, for example, O'Connor et al. 2002). The effects of other pesticides are more likely to be modulated by their effects on other microbes and/or the plants and, therefore, many conflicting reports about the effects of pesticides on A M F are available (Sugaranam et al. 1994). For example, the increase in A M F colonization of plants after the application of metalaxyl (a fungicide), was explained as a secondary effect following primary suppression of P y t h i u m sp. infestation (Seymour et al. 1994). Similarly, an increase in the colonization of clover roots by both G l o m u s c o r o n a t u m and Gigaspora m a r g a r i t a after application of fenamiphos (a nematicide), was due to reduction of grazing on the A M F mycelium by fungivorous nematodes (Bakhtiar et al. 2001). All in all, fungicides (and especially systemic fungicides) appear to cause the greatest damage to AMF, especially if applied as soil drenches rather than foliar application (Diedhiou et al. 2004). Timing of pesticide application appears to be important, because some fungicides are fungistatic rather than fungicidal. For example, spores of G l o m u s m o s s e a e were able to germinate normally following dressing with carbendazim or chlorothalonil and transfer on to media without the fungicides (Venedikian et al. 1999). The effects of pesticides on the A M F are also modulated by environmental conditions such as soil chemical and physical properties governing sorption and leaching of the pesticides, and by soil biological activities governing their degradation (Zhao et al. 2005).

Irrigation Water has always been a vital part of agriculture. Nowadays, about 60% of the world's available fresh water is used for irrigation of agricultural fields (Matson et al. 1997). Here we will deal with the effects of the water status of soil on the AMF, with particular reference to irrigation. Whereas considerable work has already been done on the influence of water deficiency on A M F development and on the mycorrhiza-mediated effects on their host

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plants, less research has been conditions on conditions of excess water, although some aquatic and wetland plants regularly associate with the AMF (Clayton & Bagyaraj 1984). Even though irrigation may closely relate to soil salinity, we will not specifically discuss the possible links between soil salinity and A M F here. The reader is kindly referred to other reviews on this topic, for example by Juniper and Abbott (1993). The densities of AMF spores and hyphae in soil, and the colonization of roots, usually correlate with the redox potential of the soils. Spores are usually abundant in upland soils, they are less abundant to rare in wetlands, and are rare or completely absent underwater (Khan & Belik 1995). For example, spore densities and root colonization were higher in a dry rice nursery than in a wet nursery (Solaiman & Hirata 1998). On the other hand, A M F spore densities of G l o m u s and Acaulospora spp. under sorghum and maize exposed to water deficit were lower compared to that when the water supply was sufficient (Simpson & Daft 1990). The root colonization of various plants in arid and semi-arid conditions was increased by moderate levels of irrigation (White et al. 1992; Caravaca et al. 2005), whereas intensive irrigation of citrus plants in Israel, resulting in waterlogging and low oxygen (02) availability in soil, caused a reduction in root colonization compared to moderate irrigation levels (Levy et al. 1983a). Thus, we conclude that either high or low water availability can reduce A M F colonization of the roots (Entry et al. 2002). This is also supported by observations from natural ecosystems, where the greatest root colonization levels were usually found in moist soils compared to very dry or flooded soils (S. R Miller 2000). The composition of AMF communities can potentially be affected by irrigation. For example, spore communities with different compositions were associated with wetland grass, Panicum hemitomon, along a hydrological gradient in South Carolina. Acaulospora trappei and Glomus clarum were more abundant in wetter soil, whereas A. laevis, G. etunicatum and Scutellospora heterogama preferred drier soils, and Glomus leptotichum showed no consistent preferences either for wet or dry soil conditions. These results suggested that different A M F species were not physiologically equivalent in their tolerance to wetland conditions (Miller & Bever 1999). It has been documented many times that A M F may improve the drought tolerance of plants, and that the mycorrhizal benefits in terms of biomass and P uptake are usually

higher under reduced water availability (RuizLozano 2003). For example, under dry conditions Glomus mosseae and G. fasciculatum improved N uptake by lettuce more than under conditions of ample water supply (Tobar et aL 1994), and the proportional growth response of field-grown maize to inoculation with G. etunicatum increased with increasing drought stress (Sylvia et al. 1993). Inoculation of maize with Glomus intraradices under glasshouse conditions resulted in higher P content and plant biomass after moderate drought stress. The mycorrhizal plants sustained higher sugar concentrations in shoots during the drought period, and also recovered more rapidly when the irrigation was restored (Subramanian et al. 1997). Mycorrhizal citrus plants in Israel were, however, more susceptible to drought stress than the non-mycorrhizal ones. This was probably because the plants inoculated with Glomus intraradices grew bigger and depleted the soil water faster and more efficiently than the non-inoculated ones, leading to more pronounced water stress during drought (Levy et al. 1983b). The effects of different A M F species on the response of plants such as lettuce to drought are variable. For example, leaf area was only slightly reduced in plants colonized by G. deserticola, whereas there was a major decline in the leaf area of plants colonized by G. occultum in response to experimentally applied drought (Ruiz-Lozano et al. 1995). As shown above, extreme drought will be detrimental for both plants and the AMF. Under such conditions, moderate irrigation may increase root colonization by AMF and sustain the greatest nutritional and growth benefits from the symbiosis. The symbiosis with A M F may alleviate plant stress caused by moderate moisture deficits, using several mechanisms such as uptake of water via hyphae; altered plant hormonal levels having an impact on stomatal conductance; lowering leaf osmotic potential; improved nutrition of the plants; and improved recovery after drought periods (Aug6 2004; Sanchez-Blanco et aL 2004). The AMF can also directly affect soil wettability and water infiltration properties through their effects on soil aggregation and by hyphal deposits of highmolecular-weight compounds such as glomalin (Bearden 2001). Excess water will generally decrease the colonization of plant roots and also the mycelium length and spore densities in the soil. This seems to be related to the availability of O2 and carbon dioxide (CO2) in the soil. Oxygen is apparently indispensable for A M F development and functioning, whereas CO2 reportedly does not cause a lot of damage even

AMF IN AGRICULTURAL SOILS at very high concentrations (up to 16%), provided that 0 2 supply is not short (Saif 1984). Oxygen concentrations below 16% in the soil atmosphere decreased the AMF colonization of roots of different plants, with the decline being very steep under 8% O2 in the atmosphere (Saif 1984). Mycelium growth of G l o m u s mosseae virtually stopped if O2 concentration fell below 3 % (Le Tacon et al. 1983). On the other hand, systematically more A M F spores were sometimes found in poorly drained (frequently flooded) soils than in well-drained soils, whereas root colonization of plants was not dependent on soil drainage. This probably indicated a shift in the life history traits of native A M F communities - with the A M F in poorly drained soils forming spores capable of surviving anoxic events (Khalil & Loynachan 1994). Anoxic conditions may specifically promote the development of bacterial plant pathogens (Sturz et al. 1997) that may, in turn - directly or indirectly - impair mycorrhizal symbiosis. The effects of soil irrigation can further be confounded by effects caused by other substances delivered to the soil with the irrigation water, either intentionally or not. If Ploaded or heavy-metal-contaminated or saline water is used for irrigation, or if the water is supplied in a way resulting in soil erosion, the irrigation may eventually result in reduction of activity of AMF (Ortega-Larrocea et al. 2001). For example, colonization of roots was generally lower in salt-stressed tomato and peanut plants than in their non-stressed counterparts (Gupta & Krishnamurthy 1996; A1-Karaki & Hammad 2001).

Grazing and burning of vegetation Grazing and burning are the two most important management practices shaping the diversity and productivity of grasslands (Watkinson & Ormerod 2001). Fire is adopted in many parts of the world to control the composition of vegetation cover, improve the quality and quantity of forage, and to control pests (Fuhlendorf & Engle 2004). Grazing goes hand in hand with domestication of ruminants such as goats and sheep, and is widely adopted in natural or seminatural grassland ecosystems around the world (Diamond 2002). In this section we will refer to the effects of burning and grazing of aboveground biomass on AMF, and we leave out the effects of underground grazing (e.g. by collembolans), because it does not fall within the range of management practices deliberately imposed on to agro-ecosystems. Fire is a major environmental disturbance

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affecting many living organisms, including the AMF (Hartnett et al. 2004). For example, A M F spore density and infectivity were lower in Australian soil immediately after fire, but recovered within six months to the levels found before the fire (Bellgard et al. 1994). Spore densities were also lower in burned than in unburned sand prairies in Illinois (Dhillion & Anderson 1993), but the recovery took longer than a year. In the long run, burning of vegetation may lead to an increase of AMF biomass and activity in soil. For example, spring burning of native North American prairie resulted in an increase in A M F spore and hyphal densities in the soil during the subsequent season, as well as in higher root colonization as compared to unburned sites (Bentivenga & Hendrick 1991; Eom et al. 1999). Experimental burning of Trifolium pratense associated with A c a u l o s p o r a scrobiculata under glasshouse conditions also resulted in increasing spore densities in the pots (Vilarifio & Arines 1993). Grazing may have variable effects on the A M E Whereas it usually leads to a decrease or has no effect on the spore densities in soil (Klopatek & Klopatek 1997; Lugo & Cabello 2002), it may both decrease and increase hyphal densities in soil and root colonization of plants (Gange et al. 2002; Lugo et al. 2003; Kula et al. 2005). For example, hyphal-length density in soil decreased upon experimental defoliation of grazing-intolerant T h e m e d a triandra, whereas it increased upon defoliation of grazing-tolerant L o l i u m perenne, Digitaria eriantha, or Trifolium pratense (Vilarifio & Arines 1991; Allsopp 1998). Likewise, root colonization of grazingintolerant Miscanthus sinensis in the fields was impaired by experimental defoliation (shearing of leaves), whereas the root colonization of grazing-tolerant Z o y s i a j a p o n i c a remained unaffected (Saito et al. 2004). The overall colonization levels of roots are usually increased in grasslands exposed to moderate grazing intensities for an extended period of time (Eom et al. 2001; Eriksson 2001), whereas overgrazing was shown to reduce root colonization. This was probably because overgrazing resulted in compositional shifts in plant communities towards species relying less on the symbiosis with AMF. For example, plant species that decreased their abundance in response to overgrazing in Florida had higher average percentage root colonization than species increasing in their relative abundance in response to overgrazing (Mullahey & Speed 1991). Grazing and burning of plants may affect the composition of underground A M F communities, as well as their diversity (Allen et al. 2003).

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For example, spring burning of native North American prairie resulted in a decrease in AMF spore diversity (Eom et al. 1999). Over a twoyear period, burning of prairies in Illinois caused a decrease in the spore abundances of G l o m u s spp. and Scutellospora heterogama, and, over a three-year period, it caused a decrease in spore abundance of Sclerocystis rubiformis (Dhillion & Anderson 1993). Like the burning, decreases in AMF spore diversity and shifts in A M F community composition due to grazing have also been reported previously (Bethlenfalvay & Dakessian 1984; Eom et al. 2001). For example, grazing of a mountain grassland in Argentina reduced the abundance of Scutellospora and promoted that of A c a u l o s p o r a spp. (Lugo & Cabello 2002). The diversity of AMF in roots (assessed by amplification and sequencing of 18S rDNA) of grazing-intolerant Miscanthus was lowered by experimental defoliation, but it was unaffected in roots of grazing-tolerant Zoysia (Saito et al. 2004). Obviously, survival, regrowth, and re-establishment of plants on areas affected by fire depend on specific morphological and physiological adaptations to withstand heat or to disperse seeds for a longer distance - with all this having little to do with mycorrhizal symbiosis. However, the growth of a plant on a previously burned soil can be affected by mycorrhizal infectivity and by the composition of A M F communities in such soils. Spring burning of tallgrass prairie in Kansas increased mycorrhizal colonization of roots, and eventually resulted in a greater plant growth compared to unburned prairie (Bentivenga & Hendrick 1991). It has also been demonstrated that grazing changed the composition of AMF communities in soil in the Yellowstone National Park, and that the AMF communities in grazed soil were more beneficial for the growth of Poa pratensis compared to soil from a fenced (i.e. ungrazed) area (Frank et al. 2003). Mycorrhizal symbiosis usually increased regrowth of plants following defoliation, and some strongly mycotrophic plants such as A n d r o p o g o n gerardii and Sorghastrum nutans showed overcompensation in response to herbivory when they were mycorrhizal (Kula et al. 2005). However, the capacity of plants to compensate for grazing damage usually decreases with an increase of grazing intensity (Hetrick et aL 1990). Both burning and grazing remove aboveground biomass, including meristems and photosynthetically active tissue. Nutrients contained in the biomass are either lost (N upon burning), or returned to the soil in mineral (P after burning) or organic forms (e.g. as N and P

in urine or manure). In addition to the biomass removal, heat generated by fire is likely to impair the viability of the AMF (Rashid et al. 1997), resulting in lower infectivity and root colonization as compared to unburned sites. It also affects the physico-chemical properties of soils (Goforth et al. 2005). Nutrient (such as P) are flushed following the burning, and changes in root architecture (roots are usually more fibrous and faster growing) in burned sites may also contribute to the generally seen lower root colonization as compared to unburned sites (Hartnett et al. 2004). Increased colonization of plants after grazing may be driven by increased demand for mineral nutrients by the plants (Eom et al. 2001) and increased C allocation to roots and root exudates by moderately grazed plants (Holland et al. 1996). This increased C availability for the AMF may stimulate build-up of storage organs such as vesicles and spores (Titus & Lep~ 2000). On the other hand, plant biomass removal by intensive grazing may impair the photosynthetic capacity of the plants, and may eventually result in lower C supply to the AMF (Nakano et al. 2001). Thus, it appears that the intensity of both burning and grazing is of central importance for their impact on mycorrhizal symbiosis, because it determines the direct damage to the AMF and/or the capacity of plants to compensate for the damage (Gehring & Whitham 1994; Kula et aL 2005). Additionally, the different tolerances of different plant and AMF species to the damage imposed by either burning or grazing are likely to contribute to the magnitude of the observed effects (Allsopp 1998; Klironomos et al. 2004).

Inputs of pollutants and heavy metals Inorganic and organic pollutants are deposited on the soil together with fertilizers, waste or contamination produced by human activities such as mining, ore processing, industry, burning fossil fuels, nuclear weapons testing, nuclear accidents, etc. (Meharg 2003). These pollutants include, for example, Zn, cadmium (Cd), lead (Pb), radioactive isotopes of caesium (Cs), strontium (Sr), uranium (U), polyaromatic hydrocarbons (PAH), and polychlorinated biphenyls. All these inputs not only affect the crops, but also exert a multitude of side-effects on inhabitants of soil. Here we review the effects of organic pollutants and heavy metals on AME The development of AMF can be hampered by elevated concentrations of heavy metals and organic pollutants in soil. For example, spore densities and mycorrhizal infectivity were

AMF IN AGRICULTURAL SOILS reduced by heavy-metal (Pb, Zn, Cd, Cu and chromium) pollution of soils (Khan 2001; Mozafar et aL 2002). The sporulation of different A M F was also lowered by experimental addition of heavy metals such as Cu, Zn, Cd and Pb in the form of water-soluble salts either into soil (Liao et al. 2003; Andrade et al. 2004) or into axenic growth media (Pawlowska & Charvat 2004). Roots of Agrostis capillaris in Dutch soil contaminated with Cd, Pb and Zn were less colonized by A M F than those from noncontaminated soil (Ietswaart et al. 1992). In another study, the colonization in Agrostis capillaris was relatively more depressed by Cu than by Zn or Cd pollution (Griffioen et al. 1994). On the other hand, 238U added in concentrations of up to 87 mg kg -1 did not affect colonization of clover by G. intraradices (Rufyikiri et al. 2004a). The presence of organic pollutants such as hydrocarbons and phenanthrene in the soil may also reduce mycorrhizal infectivity and root colonization of plants (Cabello 1997; Gaspar et al. 2002). The effects of pollutants on the development of A M F seem to be modulated by the type of plant species and the availability of the pollutant. For example, the extent of root colonization of maize and ryegrass by indigenous AMF in non-polluted soil was not significantly affected by addition of PAHs, whereas the colonization of clover and leek decreased after spiking of the soil with the PAHs (Joner & Leyval 2001). Soil pollution can also impact the diversity and composition of A M F communities in soil and in plant roots. For example, species richness and diversity of A M F spore communities were highest in German soil receiving intermediate levels of sewage sludge, but decreased in soils receiving the highest amounts of heavy-metal contaminated sludge (del Valet al. 1999a). Similarly, the densities of spores of some A M F species decreased upon irrigation with sewage water containing heavy metals, but spore densities of sporocarp-forming species such as G. mosseae and Sclerocystis spp. remained unaffected (Ortega-Larrocea et al. 2001). High chromium content in the soil reduced the species richness and diversity of A M F spore communities in Pakistan, with Gigaspora spp. dominating in the contaminated soil, whereas different A M F such as Glomus, Scutellospora and A c a u l o s p o r a spp. were present in the nearby non-polluted soil (Khan 2001). Strawberry plants in heavy-metal-polluted soils in Poland were predominantly colonized by Glomus gerdemannii, whereas otherwise common A M F such as G. intraradices and G. mosseae were much less abundant (Turnau et al.

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2001). In the UK, the abundance of vesicles in T h y m u s polytrichus roots were positively correlated with the levels of heavy-metal contamination of soils (Whitfield et al. 2004), indicating either a functional switch of the A M F towards forming more vesicles, or (more likely) the presence of different A M F species colonizing the roots in contaminated and noncontaminated soils. Differences in tolerance to heavy-metal pollution were noted among different A M F species and/or genotypes, with the fungi from heavy-metal-contaminated sites generally being more tolerant to higher concentrations of heavy metals than A M F from uncontaminated sites (Leyval et al. 1994; del V a l e t al. 1999b). A specific G l o m u s Brl isolate obtained from the roots of the yellow zinc violet (Viola calaminaria) growing on a heavy-metal-contaminated site in Germany was able to establish the symbiosis and enable growth of maize, barley and alfalfa under heavy-metal levels that were lethal for the non-mycorrhizal plants (Hildebrandt et al. 1999). Such functional protection of plants from heavy-metal stress has been repeatedly described (Chert et al. 2004). For example, in a Zn-contaminated soil, A n d r o p o g o n gerardii plants were growing better with A M F than without them, with the protection being greatest with A M F isolated from polluted sites (Shetty et aL 1995). Similarly, the A M F can mitigate the adverse effects of organic pollutants. For example, mycorrhizal symbiosis was only functional (in terms of enhancing plant P and Zn uptake) after hydrocarbon pollution if plants were colonized with AMF previously isolated from hydrocarbon polluted sites (Cabello 1999). It has also been suggested that the dissipation of condensed PAHs may be enhanced in the presence of arbuscular mycorrhiza, probably due to the presence of mycorrhizae-associated microflora (Joner et al. 2001). The tolerance of some A M F species/genotypes to heavy metals could occlude the potential adverse effects of the heavy metals on the A M F spore and hyphal densities and root colonization levels (Selvaraj et aL 2005). For example, maize root colonization was not affected by very high heavy-metal (Zn, Cd and nickel) contamination in a field site in France, most probably because the AMF inhabiting the polluted soils were tolerant to that pollution (Weissenhorn et al. 1995a). Similarly, A M F spore density in a Norwegian soil with naturally high heavy metal (Cd, Zn, Cu and Mn) contents was not different from non-polluted soils (Leyval et al. 1995). The breadth of A M F tolerance to heavy-metal pollution is variable, but

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tolerance of AMF to one pollutant does not necessarily predispose the A M F for tolerance against other stresses (Weissenhorn et aL 1995b). The stability of the tolerance is also not well understood, and needs to be studied. It has been, for example, documented that an isolate of G l o m u s sp. from a Mn-contaminated soil showed higher tolerance to Mn when kept in the original soil than the same A M F kept in nonpolluted soil for two years (Malcovfi et al. 2003). Unlike nutritional elements such as P or Zn, whose uptake by the roots is generally enhanced by AMF, the fungi probably do not transfer significant amounts of nickel, Cs and U through their hyphae towards the roots (Guo et al. 1996; Joner et al. 2004). The studies reporting some Cs transport capacity towards the roots by A M F hyphae under in vitro conditions (de Boulois et al. 2005) are probably overestimating the magnitude of fluxes, as the bioavailability of Cs in a solidified liquid medium is unrealistically high compared to any soil. In soils, the AMF may actually prevent the uptake of those elements by the plants (Berreck & Haselwandter 2001; Rufyikiri et al. 2004b). In the case of nutrients that may become toxic in high concentrations (such as Zn), A M F usually enhance uptake by plants under deficient conditions and suppress uptake under toxic conditions. For example, the contribution of G l o m u s m o s s e a e to Zn uptake by red clover reached its maximum at a Zn addition level of 50 mg kg -1 in calcareous soil, above which it decreased gradually, reducing uptake under high Zn concentrations as compared to nonmycorrhizal plants (Chen et al. 2003). It is likely that A M F can immobilize elements such as Cd, Cs, Zn and Pb in their hyphae under excessive concentrations of these elements, and at the same time the AMF can provide plants with sufficient amounts of nutrients such as P and N, but the mechanisms and regulation remain largely unknown (Joner et al. 2000; Joner et aL 2004).

Soil compaction Soil is compacted by pressure exerted by agricultural machinery. It leads to degradation of soil structure, lower porosity, reduced water and air movement, increased soil erosion, impaired root growth and ultimately to reduced plant growth (Entry et al. 2002). Here we briefly review the knowledge about the effects of soil compaction on the AMF. Under field conditions in Alabama, soil compaction reduced root development, but not

the proportion of maize roots colonized by A M F (Entry et aL 1996). The colonization of bean roots in Michigan, however, decreased upon soil compaction caused by excessive tillage and traffic (Mulligan et al. 1985), but the density of AMF spores in the soil remained unaffected. The proportion of the root length of different plant species colonized by A M F in pots was either unaffected (Nadian et al. 1996; Li et al. 1997) or lower in compacted soil (Nadian et al. 1997; Yano et al. 1998). At the same time, the growth of roots was usually severely impaired in compacted soils. The growth of A M F hyphae in compacted soil was, however, less affected than that of roots. For example, an increase in soil bulk density from 1.1 to 1.6 g cm -3 did not affect the development of G. intraradices hyphae in the soil (Nadian et al. 1996). In another study, the hyphae of G. m o s s e a e were still able to extend 30 mm from clover roots in compacted (1.8 g cm -3) soil, whereas the hyphae expanded up to 50 mm from the roots in a non-compacted (1.3 g cm-3) soil (Li et al. 1997). Recently, it has been shown that AMF may adapt their hyphal morphology to available pore size in the substrate by decreasing mean diameter of hyphae in substrate with finer pores (Drew et al. 2003). The decrease in the mean diameter of hyphae in response to lower substrate porosity was significant for G. intraradices, but not for G. m o s s e a e , indicating the possible variability of such traits between different A M F species (Drew et al. 2003). The root colonization of subterranean clover by different AMF species was also differentially susceptible to soil compaction. The colonization by G. m o s s e a e and G. e t u n i c a t u m was already reduced by medium soil compaction (bulk density 1.4 g cm-3), but the colonization by G l o m u s intraradices and G l o m u s spp. was only lower in soils with a density of 1.6 g cm -3 as compared to soils with a density of 1.2 g cm-3 (Nadian et al. 1998). The differential response of different G l o m u s spp. colonization to increases in soil bulk density was attributed to differential tolerance to the relative lack of 02 in soil, and to reduction in the available pore space in compacted soil that could be colonized by AMF hyphae (Nadian et al. 1998). Increasing soil compaction may severely restrict root growth, and this may lead to lower shoot biomass and P uptake (Yano et al. 1998). A M F can improve P uptake and growth of plants in compacted soils, but the magnitude of mycorrhizal benefits (relative increase in plant growth or P uptake due to AMF) usually decreases with increasing soil compaction

AMF IN AGRICULTURAL SOILS (Nadian et al. 1998). Additionally, A M F can also directly affect root growth and morphology in compacted soil. For example, root growth of C y m b o p o g o n w i n t e r i a n u s was promoted by inoculation with G. intraradices at low soil densities, but it was reduced by the same at high soil densities (Kothari & Singh 1996). All in all, the effects of soil compaction on A M F communities and mycorrhizal functioning have not yet been adequately studied under field conditions, and results from pot experiments are still very fragmentary (Entry et al. 2002). This subject, therefore, deserves further attention.

Topsoil movement The densities of A M F spores and hyphae in the soil, as well as the proportions of root length colonized by AMF, are usually highest in the top 15-25 cm and decrease with soil depth (Kabir 2005). The diversity and composition of the A M F communities also change with soil depth (Oehl et al. 2005). The topsoil can thus be regarded as a reservoir of AMF inoculum, and removing this layer either through mining, landscaping or erosion may result in a significant reduction of activity, as well as in changes in the composition of indigenous AMF (Carpenter et al. 2001; da Silva et al. 2005). For example, removing topsoil during road-construction activities resulted in less-diverse A M F spore communities under secondary vegetation in Venezuela, with Scutellospora and Gigaspora spp. being absent in the previously disturbed soils (Cuenca et al. 1998). These changes may have important consequences for establishment, nutrient uptake and growth of plants on the displaced soils, as well as for soil stability. On the other hand, importing topsoil and/or A M F inoculum into places affected by landscaping or erosion, may greatly benefit plant nutrient uptake and/or growth (Saxerud & Funke 1991; Mohammad et al. 1995). Therefore, there is currently a lot of interest in the reclamation of disturbed habitats such as former industrial sites using A M F inoculation for improved plant establishment and growth (Dodd et al. 2002; Gianinazzi & Vosfitka 2004).

Reversal of human impacts After reviewing the effects of various agricultural practices on the AMF, it appears important to consider briefly the reversibility of human impacts on the AMF communities. This involves processes which take place after cessation of the management activity, during system recovery

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and return to the original situation - be it extensive crop production, grassland, or forest. For example, more than 40 years of cereal cropping and tillage reduced the diversity of A M F communities in Argentina, but the A M F communities became similar to their original pattern before the cultivation after just three seasons of red clover cultivation (Menendez et al. 2001). Likewise, compositional differences in A M F communities under continuous and cereal-rotated soybeans were disappearing after one season of soybeans on rotated plots (Hendrix et al. 1995). The negative effects of tillage on the spore and hyphaMength densities in Chilean Alfisol could, however, only be reversed by prolonged (c. 20 years) cessation of tillage (Borie et al. 2000). Similarly, restoration of A M F communities affected by P fertilization may take a rather long time. For example, A M F infectivity in Dutch soil previously fertilized for 23 years with high rates of P application (52.5 kg P ha -1 a -1) was lower than in soil fertilized with either zero or 17.5 kg P ha q a -1 ten years after the cessation of fertilization (Dekkers & van der Werff 2001). It has been suggested that the recovery of AMF infectivity in Spanish soils taken out of production agriculture in the semi-arid Mediterranean zone may take up to 45 years to reach levels encountered in soil that has never been cultivated (Roldan et al. 1997). On the other hand, three years of vegetative regeneration after cattle were excluded, had not improved the physicochemical properties of severely degraded soil in Costa Rica, but had significantly improved AMF infectivity and diversity in the soil (Carpenter et al. 2001). It is possible that inoculation with AMF, coupled with management of vegetation cover, may speed up the rehabilitation of degraded soils, like the recuperation of industrial land or soil affected by landscaping (see above). Obviously, the capacity of the system to recover depends on several factors - one of the important ones being the availability of the inoculum of different A M F species. As suggested recently, some A M F species may escape soil-management practices such as tillage by preferentially inhabiting deeper soil horizons (Oehl et al. 2005). This indicates that it is possible to revert changes in A M F communities caused by tillage in reasonably short time. This is also confirmed by the study of A M F communities in Swiss soil, indicating the ample presence of Scutellospora pellucida in soil previously intensively managed (fertilized and tilled), where tillage had been prevented for the last 13 years (Jansa 2002; Jansa et al. 2003).

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The above examples indicate that systemchanges such as shifts from high- to low-input management schemes may be potentially important for improving activity and diversity of A M F communities and for maximizing mycorrhizal benefits in agro-ecosystems (Kurle & Pfleger 1994; Munyanziza et al. 1997). For example, a Swiss loess soil managed under organic or bio-organic practices for 15 years had higher A M F infectivity than the same soil managed conventionally (M~ider et al. 2000). The colonization of different plants in rotation under an organic system was on average 40% higher than under conventional management (M~ider et al. 2002). Similarly, AMF colonization of wheat under organic management was two to three times higher than that under conventional management (Ryan et al. 1994), and the colonization of clover and ryegrass was lower in conventionally managed than in biodynamic pastures in Australia (Ryan et al. 2000). Also, the colonization of rye in Germany was higher under biodynamic management than under conventional cultivation (Sattelmacher et al. 1991). It was suggested that these differences were due to the combination of several factors, such as the use of different forms of fertilizers and the application of pesticides, in addition to the different crop rotations (Sattelmacher et al. 1991). Whether the AMF under low-input and organic farming systems also contribute to a greater extent to plant growth and nutrition is, however, not yet clear. This will require field studies employing either selective fungicides that specifically kill the AMF; plant genotypes incapable of forming the symbiosis; and/or radioactive tracers and root-exclusion compartments coupled with mathematical modelling.

Conclusions Several management practices have been identified that affect mycorrhizal development, diversity and composition of AMF communities in agricultural soils (see Table 1 for a summary). Among those, the most important appear to be (in the order given): crop rotation, tillage, pesticide (particularly fungicide) application, and fertilization (Gavito & Miller 1998; Kiers et al. 2002). However, it is difficult to generalize across a wide range of studies, because the effects of the practices on the AMF composition, and especially on mycorrhizal functioning, also depend on the environmental context (soil type, fertility, pollution, climate and microbial communities). Let us mention once again soil tillage as an example: less disturbed A M F under no-tillage conditions may (and sometimes do) improve the

P nutrition of crops such as maize, but the development of the plants may be at the same time slower due to lower soil temperatures in non-tilled soil. The efficiency of use of fertilizer applied in a no-tillage system is also usually lower than under conventional tillage, since the fertilizers are not incorporated into the soil; the increases in soil density being observed under no-tillage may aggravate the development of certain diseases and eventually impair mycorrhizal symbiosis (McGonigle & Miller 1996; Sturz et al. 1997). Similarly, decreases in the mycorrhizal infectivity in soil due to fertilization appear to be modulated by other properties of the agro-ecosystem such as tillage and the type of plant species being considered, and thus a causal link to the fertilization is difficult to establish (Ezawa et al. 2000). These examples illustrate the need to consider all sides of the triangle plant-fungus-soil (environment) when thinking about AMF in soil as affected by agricultural practices. Additionally, because most of the current knowledge about the identity of A M F in agricultural soils is based on spore surveys, which may be only of limited relevance for AMF communities inhabiting the roots, or for hyphal communities in soil (Sanders 2004b), there is an urgent need to use alternative tools for identifying A M F under field conditions. Such tools are being developed and improved and their usefulness is being demonstrated under field conditions (Friese & Allen 1991; Redecker et al. 2003). Our knowledge of the human-induced changes in AMF communities in the soils is, in spite of all efforts, still quite scattered and incomplete, above all with respect to the quantification of the changes. It is thus highly desirable to collect more information about human impacts on the soil biotic component in general and on the AMF in particular, because there is an urgent need to design sustainable agricultural systems for the future (Harrier & Watson 2003; Atkinson et al. 2005). It is also important to deal with new and so far neglected issues such as the effects of transgenic crops on AMF communities (Picone 2002). Although it has been argued that A M F may not play a critical role in the nutrition of plants in fertile soils, A M F may well need to be considered in lowinput systems or for non-nutritional benefits such as soil structuring and disease resistance (Ryan & Graham 2002). We would like to acknowledge discussions with F. Oehl, as well as useful comments by three anonymous referees that helped to improve the quality of this contribution.

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Soil sealing and soil properties related to sealing WOLFGANG

BURGHARDT

Department o f Soil Technology, University o f Duisburg-Essen, Germany (e-maik wolfgang, burghardt@uni-essen, de) Abstract: Sealing implies separation of soils from the atmosphere and above-ground biosphere by impermeable layers. Sealing has a strong impact on soils. The degree of sealing is related to the type of land use and to the population density. Sealed areas are still increasing, and it is often the most fertile soils which are sealed. The negative effects of sealing are partial or total loss of soils, e.g. for plant production and habitats, and an increase in floods, as well as an increase in health and social costs. On the other hand, sealed areas contribute hugely to the gross national product of a country. The most effective contributions for reducing soil sealing are expected to come from the engineering sciences. Other contributions could be made by the economic and social sciences. It is also important to address the problem at a political level, e.g. by passing an international convention to prevent soil consumption. Despite all these problems, soils can still exist underneath sealed areas. Their properties can be used to mitigate the negative effects of sealing.

Our m o d e r n way of life has a major impact on soils. Construction of buildings, streets, airports, landfills and greenhouses, all seal large soil surfaces. At first sight, this seems to be mainly a quantitative problem of soil being eliminated by various kinds of land uses. But, u n d e r n e a t h sealed areas, soils can still exist and retain some of their functions. T h e r e f o r e we have to develop ways to decrease the growth of sealed areas and to mitigate the negative effects of sealing by using the soils located u n d e r n e a t h sealed areas.

Definition and extent of sealing The term 'sealing' is internationally used in soil science for describing the effect of soil crusting ( F A O / I S R I C 2000). In E u r o p e a n land-use planning, the term 'soil sealing is also used to indicate the separation of soils from the atmosphere and from the above-ground biosphere by constructions ( B e r l e k a m p & Pranzas 1992; B6cker 1985; Bunzel 1992). The Task Group on Soil Sealing, Soils in Urban Areas, Land Use and L a n d Use Planning of the E u r o p e a n Commission (Burghardt et aL 2004a) identified three different ways of defining sealing: (1) Definition following a systems approach: Soil sealing is the separation of soils by layers and o t h e r bodies (consisting of totally or partly i m p e r m e a b l e material) from other compartments of the ecosystem, such as the biosphere, atmosphere, hydrosphere, anthroposphere and other parts of the pedosphere.

(2) A definition following a purpose-related

approach: Soil sealing is the covering of the soil surface with an impervious material, or the changing of its n a t u r e so that the soil becomes impermeable, so that the soil is no longer able to perform the range of functions associated with it. (3) Soil sealing, including natural characteristics: Changing the nature of the soil so that it behaves as an impermeable medium. This definition includes c o m p a c t i o n of soils. Compaction of soils or subsoils may affect larger areas than the sealing as defined in definition (2). Types and degree of sealing can be different. Four types of sealing can be distinguished (Burghardt 1993): (1) total surface sealing by r o a d - m e t a l s and buildings; (2) partial sealing of the soil surface by pavements, e.g. cobbles and concrete paving slabs; (3) subsurface sealing of u n d e r g r o u n d car parks covered by a soil layer, and, to complete the definition of sealing, interruption of material and energy flow between the environmental compartments; (4) vertical sealing by walls exposed by construction in the soil. The degree of sealing is defined as the p e r c e n t a g e of an area which is covered by impermeable material, as defined above in the paragraphs (1), (2) and (3). It is not identical

FROM:FP,OSSARD,E., BLUM,W. E. H. & WARKENTIN,B. E (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266,117-124. 0305-8719/06/$15 9 The Geological Society of London 2006.

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Table 1. Urban land-use type and degree of sealing, using the example of the city of Witten, Germany (Clever & KorndOrfer 1991; Burghardt 1993) Urban land-use type

Sealed area (%) 80-100 70-90 60-80 50-70 40-60 20-40 20-40 70-100 80-100 80-100 0-20

Closed-block built-up areas Open-block built-up area and block border built-up area Terraced houses Large-block built-up area Detached houses Stately home Allotments Industrial and commercial areas Streets, large parking areas, including the accompanying green areas Strongly sealed sports grounds Public greens, public gardens

with an area used for settlements and roads. The areas of these land-use types also include open spaces. A wide range of degrees of sealing can be found in urban areas. The degree of sealing in urban areas is loosely related to the type of land use (Table 1). B6cker (1985) has classified the degrees of sealing (Table 2). The proportion of soils affected by sealing can be quite high, as shown by examples from Germany. Statistical data do not give the direct values of sealing, however. They describe surfaces used for settlements, traffic and other types of sealed land use (e.g. areas, recreation areas, cemeteries). These areas covered 12.8% of the area of G e r m a n y in 2004 (Statistisches Bundesamt Deutschland 2005a). The buildings and surrounding plots contributed 6.7% and traffic areas 4.9%. A m o n g the federal states of Germany, the surfaces used for settlements, traffic and other sealed soil-use types varied from 6.7 % in Mecklenburg-Western Pomerania, with a population density of 75 inhabitants/m 2, to 21.0% in North Rhine Westphalia, with 530 inhabitants/m 2 (Fig. 1). The surfaces used for settlements and traffic are not totally sealed, and are therefore not identical with the sealed areas. The degree of sealing is about 49% in Germany. The sealed area of Germany is about 6.3%, or 270 m2/person. The percentage of surfaces used for settlements is also high in other Central E u r o p e a n countries: for Belgium it is 18%, for the U K it is 15% and it is 10% for the Czech Republic (unpublished EUROSTAT, cited in Dosch 2001). For the years 1997-2001 the mean daily rate of growth of asettlements and traffic areas in G e r m a n y was 1.29 km 2 (Vistaverde 2006). Given a degree of sealing of 49%, this would imply an increase of the sealed areas of about 0.60 km2/day or 2.8 m2/person per year or

Classification of degree of sealing (BOcker 1985;AKS 1997)

Table 2.

Designation

Sealed area

Class

0-15 10-50 45-75 70-90 85-100

Very low Low Mean High Very high

(%)

$1 $2 $3 $4 $5

25 A

20

== .m m

lO in o

5

y=3.1" 10-2X+5.36 I r= 0.96***; n = 13

< 0

,

0

i

i

i

,

100 200 300 400 500 Population density (persons/km 2)

i

600

Fig. 1. Relationship between population density, and surfaces dedicated to settlement, traffic and other types of sealing land-uses in the states of the Germany (data from Degau 2002).

200 m2/person, within a h u m a n life-span of 75 years. The second problem is that the wrong type of soils are covered by sealing. Settlements started in ancient times on fertile alluvial plains, deltas and coastal areas, mountain valleys and other areas of soils with high agricultural productivity.

SOIL SEALING AND SOIL PROPERTIES Now w h e n populations are c o n c e n t r a t e d in urban areas, the surrounding fertile soils are also covered by construction and are then sealed. New areas of settlements and roads are established at the expense of farmland (Liischer 2003; B B R 2003a). In the U S A about 4000 km2/year of farmland are lost due to urbanization. China lost 50 000 km 2 of farmland from 1987 to 1992 ( U N E P 2002). Settlements and roads are concentrated on flat land (De Vries & Burghardt 1989). The limited fertile-soil areas in the valleys of the Alps are used for settlements, with 9.1% being used in Switzerland and 11.9% in Austria (Dosch 2001). On the other hand, more and more industrial and settlement areas have been a b a n d o n e d in recent times. The reasons for this are changes in the economy from industry to commerce and services, and the m o v e m e n t of low-income populations into the centre of cities. As an example, e m i g r a t i o n from eastern G e r m a n y will result in the demolition of 350 000 flats in the time period 2001-2010 (BMVBS 2006). This process has h a p p e n e d already in many other E u r o p e a n areas, and will perhaps go on in the countries of Eastern E u r o p e that joined the E u r o p e a n U n i o n in 2004. U r b a n sprawl, suburbs and satellite towns c a n n o t offer a d e q u a t e infrastructure for elderly people. With an ageing population, one can expect a return of elderly people into city centres in the n e a r future (see Quast, in B u r g h a r d t et al.

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2004a). It is expected that the city of Berlin will e x p e r i e n c e a d e m a n d for up to 100000 additional apartments, preferably in the inner city, from elderly people in the next 15 years (FAZ, 2 February 2006, p. 3).

What are the effects of sealing? M a n y investigations have focused on the negative effects of sealing. The impacts of soil sealing on urban climate, on soil properties and on water transfer have been studied (Lassen 1979; Paluska 1985; Pietsch 1985; Wessolek 1988; Kuttler 1998). Table 3 gives some examples of these negative effects. The soil is often excavated before sealing, resulting in 'total erosion'. Health costs are incurred from heating of the urban environment and fine dust (PM10 , PMzs) blown from fast-drying surfaces in urban areas. Often populations with low incomes are concentrated in strongly sealed areas, as these are the only places where they can afford to rent a flat. Therefore in sealed areas, social problems, and the costs related to them, tend to accumulate. But there are also numerous advantageous effects to sealing (Table 4). These touch nearly all spheres of our lives. The close relationship to gross n a t i o n a l p r o d u c t m a k e s it difficult to control the increase in soil sealing and a c o m m i t t e d c o h e r e n t policy is required. However, public security, the emergency services (fire, health) and communication are all improved by sealing areas.

Table 3. The negative effects of sealing Loss of soil by excavation Loss of fertile land Loss of flora and fauna Habitat degeneration by dissection of areas Increase in storm-water discharge, causing floods, with their associated safety costs

Increase in health costs, caused by creating the urban heat-island effect; production of fine dust (PM10); loss of CO2 sinks Generation of low-quality living areas, with their associated social costs

Table 4. Dependence of gross national product on sealing in Germany (calculated from the data of Statistisches Bundesamt Deutschland 2005b) Creation of income, gross national product Total gross national product Directly dependent on soil sealing (constructions, cars, transport, education) Indirectly dependent on soil sealing (mining, industry, engineering, services) Having no strong effects on sealing (agriculture, forestry, power and water supply)

Billions of euros

Percentage

2130 850

100 40

1160

54

120

6

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W. BURGHARDT

H o w can sealing p r o b l e m s be s o l v e d or mitigated? Soils are limited resources which are not renewable on a h u m a n time-scale. On the other hand, t h e y can be d e g r a d e d or even d e s t r o y e d ( A m u n d s o n et al. 2003). Therefore policies must be i m p l e m e n t e d to preserve soils (see Montanarella's paper in this volume). There are several options for reducing soil sealing and mitigating soil-sealing effects (Table 5). International conventions restricting soil consumption must be established and implemented as soon as possible. Sealing often consumes valuable soils, including high-productivity fertile soils, soils within valuable watersheds, and soils of special interest as habitats for wildlife. P r o t e c t i o n of soils against sealing must be taken into account in land-use planning. To prevent the misuse of soils, the soils in urban areas and their surroundings should be surveyed (Liischer 2003). The descriptions of soil properties in soil maps and in soil information systems should be made accessible for p l a n n i n g purposes. Specific instructions for surveying u r b a n soils are already available (AKS 1997) and have been proposed for sealed soils (Burghardt et al. 2004b). Efforts to avoid increased coverage of soil by settlements has already been partly successful. In the UK, 61% of new buildings were constructed on a b a n d o n e d industrial and commercial ground (brownfields), and in L o n d o n it was 90% ( B B R 2003b). Financial instruments to decrease soil sealing could include p a y m e n t for lost eco-services from soils - i.e. services such as food and wood, clean water, fresh air, carbon sequestration, heat storage, and flood prevention by water storage.

The health and social costs caused by sealing also need to be compensated for. The largest c o n t r i b u t i o n to effectively reducing soil sealing will come from engineering. Most important are technologies to change production lines and storage facilities from horizontal to vertical arrangements. The advance of e-commerce might also be said to be obviating the need for sealing. Table 5 lists some additional options for mitigating sealing, and more ideas are presented in Burghardt et al. (2004a).

Soils o f sealed areas and their properties Sealing is not always related to the total removal of soils by excavation. Soils can still survive u n d e r n e a t h sealed areas. The example of Figure 2 shows a large tree within a totally sealed car park in the centre of the city of Essen, Germany. One would assume that this would be a hostile environment for tree growth. However, the soil underneath the sealed area is rooted and supplies the tree with water and nutrients. The above example shows that sealed soils can still retain some soil functions. Trees and bushes can grow on sealed areas. Storm water infiltrates underneath sealed areas. Sealing can also protect soils from the deposition of pollutants. The possibility exists for mitigating the negative sealing effects by using the remaining soil properties. Building concepts should be adapted so as to make use of these remaining soil functions. A first step would be to accept that soils and substrates covered by a sealing layer still exist, and that soils underneath sealing covers should be investigated (Komossa et aL 2002; Burghardt et al. 2004b). In u r b a n areas, soils often develop on recently deposited substrates, and therefore soil

Table 5. Options for solving or mitigating sealing problems 9 Policy: Produce a convention on the restriction of the consumption of soils 9 Planning and decision-making: - Controlled land use and city growth - Use only soils of lowest fertility and with the lowest habitat value - Need for proof and public documentation of the need for sealing, and its effects - Reuse of urban derelict land (brownfield land) for settlement areas * Fiscal measures: - Payment for lost natural resources - Payment for social and health costs related to sealed areas - Funds for re-establishing soils when the reason for sealing no longer exists 9 Technical measures: - Sealing without excavation and soil transport (technical erosion) - Change production lines and storage from horizontal to vertical arrangements - Subsurface construction - Subsurface transport (pipelines) - Movable lightweight constructions: removable prefabs - Advancing e-conamerce

SOIL SEALING AND SOIL PROPERTIES

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Fig. 2. Examples of sealed soils. Large tree on a totally sealed site (left) and rooted soil underneath streets (below) in Essen, Germany. (Photos by Burghardt.) properties are strongly related to substrate properties (Burghardt 1996). A dual soil-classification system, based on the classification of substrates and on soil development is recommended for surveying urban soils (Burghardt 1995, 1997, 2000). A proposal for the inclusion of sealed areas in soil classification already exists (Burghardt 2001, 2002). The soils are called Ekranosols (Figure 3, left-hand photo) and in the case of partial sealing by cobbles or concrete paving slabs, they are known as Dialeimmasols (Wenikajtys & Burghardt 2002). The profile of an Ekranosol shows a solid horizon on the surface made from the sealing material, underneath a horizon made up of a base layer of gravel, followed either by a complete natural soil; a soil from a technological substrate (e.g. rubble, ash, cinder, slag, sewage and industrial sludge) (Blume et al. 1989); by a truncated soil; or by deep excavation of natural rock layer (Komossa et al. 2002). If they show no signs of soil-development processes, Ekranosols are raw soils with domi-

nantly lithic properties, and can be classified as the subtype Ekranoliths. The presence of permeable gravel layers over compacted soils underneath streets will favour soils under stagnant conditions. Such soils could be known as Ekrano-pseudogleys. Sealed soils showing reduction horizons would be Ekranoreductosols. The photo in Figure 3 shows an example of an Ekrano-reductosol. The soil profile of a c. 100-year-old Ekranosol presented in Figure 3 had a high slag content and a low fine-earth content in the basal layer of the street construction. Underneath this basal layer, the former topsoil, which was originally a loamy silt, was slightly compacted, but not more than would be found on arable land. The basal layer had a high carbonate content, but the natural soil directly under the basal layer had a pH of around neutral. A thin residual fossil A h horizon with a high organic carbon (Corg) content is observed just below the base layer. The profile had a low to medium content of plant-available potassium, and a medium

122

W. B U R G H A R D T

Fig. 3. Example of the soil properties of an Ekranosol on a main street, city of Essen (Komossa et aL 2002; photo by Burghardt). Horizontal dashed line: border between the technological substrate of the base layer and the natural soil.

SOIL SEALING AND SOIL PROPERTIES content of plant-available phosphorus, b o t h determined in an acid calcium lactate extract (Hoffmann 1991). High heavy-metal concentrations, determined in aqua regia extract, occurred in the former topsoil, as shown in Figure 3 for lead. The soil u n d e r n e a t h the sealing cover was still able to fulfil several functions, e.g. supporting the growth of trees in the street and allowing storm-water infiltration. Preserving the soil properties u n d e r n e a t h the sealing cover would allow the soil to be restored after desealing following a change of land use, as can be expected for m a n y suburban residential areas or former industrial sites.

Conclusions Sealing is a big threat to soils, as it covers soils with i m p e r m e a b l e layers or leads to soil consumption. However, there are options for limiting the sealing of high-quality soils. To achieve this, soils must be surveyed, and information about soils in and around urban areas must be m a d e available for u r b a n land-use planning. The most effective contributions for reducing soil sealing are expected to come from the engineering sciences. Other contributions could be made by economic and social sciences. Investigations on soil sealing are not only of interest to the amount of sealed areas and their annual growth. Soils in sealed areas can contribute to the solution of urban problems. Therefore, it is of interest to survey and monitor the soils of sealed areas, and to investigate the quality of these soils. This will help to develop measures to mitigate sealing effects, by using the properties of residual soils located u n d e r n e a t h the sealing cover. I thank the following organizations and individuals for their initiative, support and contributions to my research: (1) European Commission, Directorate-General, Environment: Soil Thematic Strategy: (a) Working Group on Research, Sealing and Cross-cutting Issues; (b) Task Group 5 on Soil Sealing, Soils in Urban Areas, Land Use and Land Use Planning; (2) AKS Arbeitskreis Stadtb6den (Working Group, Urban Soils) of the German Soil Science Society; (3) IUSS WG SUITMA - Soils of Urban, Industrial, Traffic and Mining Areas; (4) ICLEI (International Council of Local Environmental Initiatives); (5) Professor Dr Ewin Amann, Department of Economics, University of Duisburg, Essen;

123

(6) Prof. Johannes G. Quast, Department of Landscape Planning, University of Duisburg - Essen. For more information on Soil Thematic Strategy, please consult the Soil CIRCA e-library: http://forum.europa.eu.int/Public/irc/env/Home/main.

References AMUNDSON, R., GUO, Y. & GONG, P 2003. Soil diversity and land use in the United States. Ecosystems, 6, 470-482. AKS (Arbeitskreis Stadtb6den der Deutschen Bodenkundlichen Gesellschaft) 1997. Empfehlungen des Arbeitskreises StadtbOden der Deutschen Bodenkundlichen Gesellschafi far die bodenkundliche Kartierung urban, gewerblich, industriell und montan iiberformter Fliichen (StadtbOden). 2. Auflage. Teil 1: Feldftihrer. Sekretariat Btiro far Bodenbewertung, Kiel. BBR (Bundesanstalt ftir Bauen und Raumordnung) 2003a. Siedlungsfl~ichenentwicklung 2001. Ver~inderungen 1997-2001. World Wide Web Address: http://www.bbr.bund.de/raumordnung/siedlung/ umwelt2001.htm#1. BBR (Bundesanstalt ftir Bauen und Raumordnung) 2003b. England - Siedlungsfl~ichenentwicklung. World Wide Web Address: http://www.bbr.bund.de/ raumordnung/siedlung/england.htm. BERLEKAMP,L.-R. & PRANZAS,N. 1992. Erfassung und Bewertung yon Bodenversiegelung unter hydrologisch-stadtplanerischen Aspekten am Beispiel eines Teilraums yon Hamburg. (Ascertainment and assessment of soil sealing, considering hydrological city-planning aspects, using the example of a part of Hamburg). Ph.D. thesis, University Hamburg. BLUME, H.-R & BVRGHARDT, W. ET AL. 1989. Empfehlungen des Arbeitskreises StadtbOden der Deutschen Bodenkundlichen Gesellschafi far die bodenkundliche Kartieranleitung urban, gewerblich und industriell iiberformter Fliichen (StadtbOden). Umweltbundesamt, Texte 18/89, 171 S. BMVBS 2001. Programm Stadtumbau Ost. http://www. bmvbs.de/Staedtebau-undWohnungswesen/ Staedtebau-1553/Stadtumbau-Ost.htm. BOOKER, R. 1985. Bodenversiegelung - Verlust vegetationsbedeckter Fl~ichen in Ballungsr~iumen Landschaft und Stadt, 17 (2), 57-61. BUNZEL,A. 1992. Begrenzung der Bodenversiegelung. Planungsziele und Instrumente. Deutsches Institut far Urbanistik (Difu), Beitr~ige zur Stadtforschung, Vol. 8, Berlin. BURGHARDT, W. 1993. Formen und Wirkung der Versiegelung. Symposium Bodenschutz 29.30.6.1992, Zentrum far Umweltforschung der Westfiilischen Wilhelms Universitiit, 111-125. BURGHARDT,W. 1995. Zur Gliederung yon Stadtb6den und ihrer Substrate. Mitteilungen der Deutschen Bodenkundlichen Gesellschafi, 76, 997-1000. BURGHARDT,W. 1996. Substrate der Bodenbildung in urban, gewerblich und industriell tiberformten B6den. In: ARBEITSKREISSTADTBODENDER DBG (eds) Urbaner Bodenschutz, Springer-Verlag, Berlin, Heidelberg, New York, 25-44.

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BURGHARDT, W. 1997. Soil mapping instruction for urban and industrial sites - characterization of substrates by layers and mixtures. Proc., International Conference, Problems of Anthropogenic Soil Formation, June 16-21, V. V. Dokuchaev Soil Institute, Moscow, 41-47. BURGHARDT, W. 2000. The German double track concept of classifying soils by their substrate and their anthropo-natural genesis: the adaptation to urban areas. Proceedings of the First International Conference on Soils of Urban, Industrial, Traffic and Mining Areas, Essen, 12-18 July 2000, Vol. I, 217-222. BURGHARDT, W. 2001. Soils of low age as specific features of urban ecosystem. Soil Anthropization VI Proceedings, International Workshop, Bratislava 20-22 June 2001, J. Sobocka (ed.). Soil Science and Conservation Research Institute, Bratislava, Slovakia, 11-17. BURGHARDT, W. 2002. Diskussionspapier bisher bekannter Stadtb6den. Mitteilungen der Deutschen Bodenkundlichen Gesellschaft, Bd. 99, 3-4. BURGHARDT,W., BANKO,G. ETAL. 2004a. Sealing soils, soils in urban areas, land use and land use planning. In: VAN-CAMPEN, L., BUJARRABAL, g., GENTILE, A.R., JONES, R.J.A., MONTANARELLLA, L., OLAZABAL, C. • SELVARADJOU, S.-K. (eds) Reports of the Technical Working Groups; Established Under the Thematic Strategy for Soil Protection. Volume VI, Research, Sealing and Cross-Cutting Issues, CUR 21319 EN/6. Office for Official Publications of the European Communities, Luxemburg, 771-817. BURGHARDT, W., SCHNEIDER, J. & KUEHN, K. 2004b. Soil monitoring instruction on sealed areas in the European Union. In: VAN-CAMPEN,L., BUJARRABAL, B., GENTILE, A.R., JONES, R.J.A., MONTANARELLLA,L., OLAZABAL,C. & SELVARADJOU, S.-K. (eds) Reports of the Technical Working Groups, Established Under the Thematic Strategy for Soil Protection. Volume VI, Research, Sealing and Cross-Cutting Issues, CUR 21319 EN/6. Office for Official Publications of the European Communities, Luxemburg, 787-793. CLEVER, M. & KORNDOERFER,K. 1991. Beitr~ige zum kommunalen Bodenschutz: natur-und kulturraumliche Analyse der Stadt Witten (Contributions to the municipal soil protection: natural and cultural space analysis of the city of Witten). Diplomathesis, Department of Soil Technology of the University Essen. DEGAU, M. 2002. Nutzung der Bodenfl~iche: Fl~ichenerhebung 2001 nach Art der tats~ichlichen Nutzung. In: Statistisches Bundesamt: Wirtschaft und Statistik, 6/2002, Verlag Metzler-Poeschel, Stuttgart, 480-487. DE VRIES, H. & BURGHARDT,W. 1989. Durch StraBen beanspruchte B6den, dargestellt am Beispiel GroBraum Bonn. Mitteilungen der Deutschen Bodenkundlichen Gesellschafl, 59/11, 1149-1154. DOSCH, E 2001. Fliichenverbrauch in Deutschland und Mitteleuropa - Struktur, Trends und Steuerungsoptionen durch das Bodenbandnis. Report 1. Internationale Jahrestagung des Boden-Btindnis europ~iischer St~idte und Gemeinden. 12-13 November 2001, Osnabrtick.

FAZ 2006. Zurtick in die Stadt. Berlin w~ichst in Zentrum. In: Frankfurter Allgemeine Zeitung, February 2, 2006, Frankfurt/Main, p.19. FAO/ISRIC 1990. Guidelines for soil profile descriptions, 3rd end. Food and Agriculture Organization of the United Nations, Rome. HOFFMANN, G. 1991. Die Untersuchung von BOden. Handbuch der landwirtschaftlichen Versuchs- und Untersuchungsmethodik (Methodenbuch), Vol. 1. VDLUFA, Darmstadt. KOMOSSA,M., HAGEDORN,A. & BURGHARDT,W. 2002. Ekranolithe - BOden unter Straflen. Mitteilungen der Deutschen Bodenkundlichen Gesellschaft, Bd. 99, 13-14. KUTrLER, W. 1998. Stadtklima. In: SUKOPP, H. & WTTTI6, R., StadtOkologie. G. Fischer Verlag, Stuttgart, Jena, Ltibeck, Ulm. LASSEN, D. 1979. Unzerschnittene verkehrsarme R~iume in der Bundesrepublik Deutschland. Natur und Landschafi, 54 (10), 333-334. LI)SCHER, C. 2003. 10 Jahre Fahrplan Fruchtfolgefliichen (FFF). Erfahrungen der Kantone, Erwartungen an den Bund, Eidgen6ssisches Department ftir Umwelt, Verkehr, Energie und Kommunikation UVEK, Bern, Schweiz. PALUSKA, A. 1985. Urbane Bodenversiegelung und ihre Auswirkungen auf die Grundwasserneubildung. Ein Beitrag zum Landschaftsprogramm aus dem Hamburger Raum. In: STRUBELT,W. (ed.) Boden das dritte Umweltmedium, Beitriige zum Bodenschutz. Forschungen zur Raumentwicklung, Band 14. Bundesforschungsanstalt fiir Landeskunde und Raumordnung, Bonn. PIETSCH, J. 1985. Versiegelung des Bodens in der Stadt und ihre Auswirkungen. In: STRUBELT,W. (ed.) Boden das dritte Umweltmedium, Beitriige zum Bodenschutz. Forschungen zur Raumentwicklung, Band 14. Bundesforschungsanstalt for Landeskunde und Raumordnung, Bonn. STATISTISCHES BUNDESAMT DEUTSCHLAND 2005a. Bodenfl~ichen nach Art der tats~ichlichen Nutzung in Deutschland. World Wide Web Address: http:t/www.destatis.de/basis/dllrvgrOlad.htm. STATISTISCHES BUNDESAMT DEUTSCHLAND 2005b. Bruttoinlandsprodukt und BruttowertschOpfung. World Wide Web Address: http://www.destatis.de/ indicators/d/ugrtab7.php. UNEP (United Nations Environmental Programme) 2002. Global environmental outlook 3. World Wide Web Address: http://www.gridunep.ch/geo/ geo3/english/index.htm. VISTAVERDE 2006. Statistische Bundesamt: Fl~ichenverbrauch hat sich beschleunigt. World Wide Web Address: http://www.vistaverde.de/news/ Wirtschaft/0204/29-flaechenverbrauch.htm. WENIKAJTYS,M. & BURGHARDT,W. 2002. Dialeimmasol Ritzenboden von Gehwegpflastern. Mitteilungen der Deutschen Bodenkundlichen Gesellschaft, Bd. 99, 23-24. WESSOLEK, G. 1988: Auswirkungen der Bodenversiegelung auf Boden und Wasser. In: Bodenversiegelung im Siedlungsbereich. Informationen zur Raumentwicklung, Band 8/9. Bundesforschungsanstalt ftir Landeskunde und Raumordnung, Bonn.

Cultural soilscapes E. C. W E L L S

Department o f Anthropology, University o f South Florida, Tampa, Florida 33620, USA (e-mail: cwells@cas, usf.edu) Abstract: While soil is normally studied as the outcome of natural geological and chemical processes, soil research by archaeologists, geographers and other social scientists focuses on the human behavioural dimensions of soil formation. As a result, 'cultural soilscape' is an analytical concept common to both Earth and social sciences that encourages a more holistic, transdisciplinary approach to studying soil formation processes. This paper introduces the concept of cultural soilscape and reviews important archaeological work on this theme over the past decade.

Soil is usually considered to be the product of numerous intersecting natural processes, including those associated with the erosion of geological materials, topography, climate, living organisms and time. Recently, social scientists but especially archaeologists and geographers have proposed that 'culture' (learned and shared knowledge and beliefs that produce, and are produced by, human behaviour) should be added to this list, and that soil should be understood and studied as a product of social forces as much as natural ones (see Wagstaff 1987). Indeed, soil surveyors and pedologists have long recognized that physical, biological and chemical properties of soil may be altered significantly as a direct result of human activity. It is not surprising, then, that recent approaches to soil morphology and related research consider the cultural soilscape, which can be defined as a given area of the Earth's surface that is the result of spatially and temporally variable geomorphic, pedogenic and cultural processes. As the physical embodiment of human/environment relationships, the cultural soilscape is an important analytical domain, because it reveals the consequences of the complex and multilayered dialectic between human behaviour and soil bodies over long periods. In this paper, I review the varied ways in which geo-archaeologists and soil scientists have collaborated to study cultural soilscapes over the past few decades. In doing so, I hope to convince the reader that transdisciplinary studies of ancient, recent and contemporary cultural soilscapes provide unique, but complementary, datasets with which to model soil formation processes.

The cultural soilscape as an analytical domain Since its inception as a scientific discipline nearly a century ago, archaeology has concerned itself with the study of human impacts on the manufacture, use and discard of material culture, that is, artifacts and the technologies developed to adapt to societies' changing circumstances over time (Binford 1962; Clarke 1968). The fundamental analytical unit was considered to be the prehistoric site - an area of human settlement or activity composed of artifacts and features as well as other evidence for human behaviour. Very recently, however, the goal of archaeology has shifted to understanding landscapes and entire regions in the past, rather than a single site, and to read the history of human activity all the way up to the global scale (Anschuetz et al. 2001). As a result, archaeologists have drawn theoretical and methodological insights from other social and natural sciences to study the use, m a n a g e m e n t and meaning of landscapes (Butzer 1982), The convergence of different disciplinary approaches for landscape research is not unique to archaeology. Many social and natural scientific disciplines have independently come to the conclusion that landscapes have cultural characteristics in addition to natural ones (e.g. Cosgrove 1984; Jackson 1994). The concept of landscape, then, can be understood as a composite of all factors and features that constitute the visual and perceived impact of an anthropogenic environment upon the human senses (e.g. Bender 1992; Bradley 1998). In other words, landscapes are the physical and spatial manifestations of the relationship between humans and their environment (Marquardt & Crumley 1987). For studying landscapes as

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. E (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266,125-132. 0305-8719/06/$15 9 The Geological Society of London 2006.

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social and natural phenomena, it is clear that geology, pedology and other Earth sciences must play a central role. Indeed, these varied strands have combined to form new, transdisciplinary approaches to landscape research, such as geo-archaeology and historical ecology, which focus on the physical evidence for human/environment relations over very long time spans, on the order of hundreds or thousands of years (e.g. Davidson & Shackley 1976; Stein & Farrand 1985; Collins et al. 1990; Holliday 1992; Crumley 1994; Kirch & Hunt 1997; Ashmore & Knapp 1999). Central to these new approaches is an appreciation of cultural soilscapes. While functioning soil is formed through the physical disintegration and chemical decomposition of rocks through weathering and subsequent arrangement by biological and chemical agents, the cultural soilscape is formed through these processes plus human behavioural variables acting to disturb and displace soil. In this way, the concept of cultural soilscape is more inclusive than the generic term, anthrosol (soil modified by human activity), because it encompasses not only the materials that can be perceived and used as resources by humans (for growing food, raising buildings and earthworks, and so on) but also other humans and the social and historical frameworks imposed by people upon their physical surroundings (e.g. Waters & Ravesloot 2001). Hence, the concept of cultural soilscape is appropriate for understanding the ways in which humans modify their physical environment at the landscape scale (versus the pedon scale for anthrosols), as well as the ways in which the physical environment shapes human behaviour. To be sure, cultural soilscapes are rarely shaped by a single person. Instead, they materialize a palimpsest of many individual ideas, beliefs and practices. In this way, cultural soilscapes are historically contingent on local patterns and processes, and act as reservoirs of shared ecological knowledge and its manifestation in the soil record.

Archaeology and soil memory in the study of cultural soilscapes Since culture and soil 'coevolve' alongside one another, studies of cultural soilscapes necessarily employ diachronic, integrated, and multiperspective approaches, which juxtapose archaeological, historical and ethnological studies with research in the Earth sciences. By providing unique and compelling arguments about long-term patterns of land use, archaeol-

ogy occupies an important niche within the intellectual and scientific world, because it provides information at a scale and resolution that makes it suitable for studies of human/soil dynamics (van der Leeuw & Redman 2002). For archaeologists, the extent to which soils trap and preserve human impacts is critical for studying cultural soilscapes. Over the past decade, soil science has come to play an essential role in archaeologists' toolkits, since many human actions result in physical changes to soils as well as in the deposition of a range of chemical compounds (see Scudder et al. 1996). Recent technological advancements in analytical methods and instrumentation have made detecting and studying these impacts relatively quick and inexpensive; thus, this line of research is becoming standard practice in many archaeological investigations. Indeed, when combined with archaeology, cultural soilscapes often can be studied with the most basic tools of pedology: soil morphology, particle-size distribution analysis, pH, clay mineralogy, and patterns of chemical element accumulation (see Holliday 2004). In archaeological studies of cultural soilscapes, the core concept is 'soil memory', that is, how soils encode the physical, biological and chemical effects of different human activities. These effects can include modification to soil structure (often leading to compaction), soil reaction (pH), aeration and water drainage, nutrient cycling and soil organism activity, soil temperature regimes, as well as the addition of anthropogenic materials and other contaminants. These effects, in turn, may influence other soil properties, for example, water infiltration and permeability, water-holding capacity, and root penetrability. Studying changes to the physical structure and composition of soils has long been an integral part of archaeological research (e.g. Cornwall 1958: Limbrey 1975; Courty et al. 1989; Feller et al. 2003). Much less is known about the chemistry of anthrosols composing cultural soilscapes that developed prior to the Industrial Revolution, even though the chemical effects of modern human activities on soils have been well studied (see McBride 1994; Sparks 2003). A good deal of information on the chemistry of ancient anthrosols has come from geochemical applications in archaeology. For example, early work by European geographers (e.g. Arrhenius 1929, 1931) revealed high concentrations of soil phosphate in areas of ancient human occupation, which can be explained by the observation that human activities involving the deposition of refuse and organic waste

CULTURAL SOILSCAPES increase the amount of phosphorus and other elements (mainly calcium, carbon, and nitrogen) in soils. Once deposited, phosphorus ions attach to iron, aluminum or calcium ions to form relatively stable chemical compounds that can be detected and quantified with appropriate techniques (see Eidt 1985; Wild 1986). The basic idea is that the surfaces of certain soil particles, particularly clays, hold ions carrying a negative charge (anions) that act like magnets to attract positive ions (cations). Cations - including those from calcium, magnesium and potassium generated by human activities become attached to the soil particles in a process known as cation exchange. Since these compounds are rapidly fixed to the mineral surfaces of sediments, they tend to remain stable and immobile (resistant to horizontal and vertical migration) for very long periods. To establish the quantity and distribution of chemicals that have accumulated on the surfaces of sediment grains as a result of human activity, one compares the concentrations of elements in anthrosols to those of soils unaffected by human settlement and land use. The work of European geographers and others in the early twentieth century (reviewed in Proudfoot 1976; Bakkevig 1980; Bethell & Mfit6 1989; Craddock et al. 1986) developed an impressive array of techniques for measuring available soil phosphate (using mild-acid extraction procedures) and total soil phosphate (using strong-acid digestion procedures) as a means to prospect for archaeological sites. More recent studies (e.g. Coultas et al. 1993; Dunning 1993; Schuldenrein 1990; Sullivan 2000) have improved some of these methods, but employ them using more human-centred perspectives, by viewing patterns of phosphate deposition as a measure of the ways in which humans interfere with phosphorus cycling in ecosystems. In addition, researchers have begun to examine the implications of elements other than phosphorus for site formation processes (e.g. Linderholm & Lundberg 1994; Middleton & Price 1996; Entwistle et al. 2000b; Wells 2004).

Recent exemplary research at macro-, meso-, and micro-scales The manner by which humans mould cultural soilscapes is studied in archaeology a t three nested spatial scales: macro-, meso-, and microscales. The first concerns large portions of the soilscape, including resource catchment zones and agricultural field systems. The second deals with mid-level spatial domains, such as garden plots and landscapes between settlements. The

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third level relates to localized patterns of land use among households and communities, for instance, food preparation or consumption areas and formal spaces for public gatherings. Macro- and meso-scale approaches to cultural soilscapes tend to focus on studying human ecodynamics, plant-subsistence strategies and habitat variation, by pooling information from settlement and demography along with climate reconstructions (oxygen isotope studies of lake sediment cores) and land-use history (phosphate changes across soil strata). For example, work by Cowgill (1961; Cowgill & Hutchinson 1966) in the mid-1960s showed how soil chemical analysis of lake sediment cores in the Guatemalan Pet6n rainforest can be used to investigate the effects of intensive cultivation of the region's soils prior to the sixteenth century. Cowgill and her colleagues found that widespread cultivation accelerated soil erosion, which resulted in the accumulation of thick, clay-rich, gleyed deposits in Pet6n lakes. Not long after this work, another study, by Provan (1971,1973) in the early 1970s, applied chemical analyses to anthrosols from Bjellands0yn~e, an Early Iron Age farm site in Norway. By investigating exchangeable sodium, potassium, calcium, magnesium, organic carbon, total phosphorus and total nitrogen, Provan observed that the distribution of brown podsols correlated with cultivated parcels of land, while ironhumus podsols signalled undisturbed soils. Aside from these isolated cases, however, only since the mid-1970s has soil chemistry played a major role in archaeological research concerned with reconstructing past agricultural practices. At first, these studies were largely limited to research on enrichment and depletion of certain plant macronutrients, namely phosphorus, nitrogen, and potassium, over time. More recently, however, soil chemistry has been incorporated into large, multidisciplinary projects that combine archaeological survey and excavation with physical and chemical analysis of sediments and palynology. For example, Dunning (1993, 1996; Dunning & Beach 1994, 2000; Dunning et al. 1997, 1998), Fedick (1995, 1996; Fedick et al. 2000; Fedick & Morrison 2004), and others (e.g. Pope et al. 1996; Wingard 1996; Beach 1998; Rosenmeier et al. 2002) examine anthrosols produced by the preHispanic Maya peoples of Central America to answer questions about the nature and variability in agricultural practices and how these practices allowed the Maya to adapt to a complex mosaic of microenvironmental variations in soil since at least 1000 BC. Previous studies in this region show that the Maya developed creative

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agrotechnologies to deal with soil variation in their karstic environment, farming on drained fields to stabilize surface hydrology and improve the soil characteristics of the plant-root zone (Pohl et al. 1990; Pope & Dahlin 1989; see also Turner & Harrison 1983). By combining archaeology, pedology and geomorphology to investigate the ways in which highly fertile and effectively irrigated agricultural fields were created and maintained, this research provides key information on the environmental parameters of cultural adaptation to tropical soilscapes. In addition to elemental studies, analysis of stable carbon isotope ratios (i.e. ~3C/12C ratios) has become particularly important to prehistoric agricultural studies because of the characteristic isotopic signatures of C3 and C4 plants (Nordt 2001). Corn, for instance, is a C4 plant enriched in 13C relative to most other cultivated and wild plants. As a result, distinctive carbon isotope signatures produced by ancient corn crop residues can be preserved in the humic component of soil organic matter: humus extracted from buried A-horizons of soils that were likely used for the cultivation of corn is enriched in a3C. These kinds of studies have been used to trace the spread of maize agriculture at Caracol in the tropical forests of Belize (e.g. Webb et al. 2004) and to reconstruct manuring and other soil fertility enhancing activities at Orkney in northern Scotland (e.g. Simpson et al. 1997, 1998; Bull et al. 1999). Work on micro-scale cultural soilscapes is a relatively recent development, and is largely concerned with reconstructing the types and locations of certain domestic activities. For example, phosphate deposition is associated with the preparation and consumption of foods and beverages; sodium and potassium compounds are generated by the production of wood ash in hearths and kilns; iron oxide and mercuric sulphide are accumulated in soils through the use of hematite and cinnabar used as pigments in burials and caches; and iron and titanium oxides from microphenocrysts embedded in volcanic glass are deposited in soils as a result of obsidian tool manufacture and use (e.g. Middleton & Price 1996; James 1999; Vizcalno & Cafiabate 1999; Wells et al. 2000; Scudder 2001; Knudson et al. 2004; Middleton 2004; Wells 2004; Cook et al. 2005; Marwick 2005; Sampietro & Vattuone 2005). This work builds on earlier studies (e.g. Cook & Heizer 1965; Heidenreich & Konrad 1973) that explored the possible geochemical pathways of a range of different elements and compounds in anthrosols at archaeological sites in North

America. By mapping the distribution of certain combinations of elements across archaeological sites, these studies were able to determine the precise locations of human settlement within broader landscapes, as well as human activities within archaeological sites. Two sets of micro-scale studies exemplify current research on cultural soilscapes. First, Entwistle and colleagues (Entwistle & Abrahams 1997; Entwistle et al. 1998, 2000a, b) study the impacts of arable cultivation and animal husbandry on soil development and characteristics at historical farm sites on the islands of northwestern Scotland. They combine historical archaeological research (in which written records can be correlated with archaeological finds to confirm the locations and functions of certain features) with physical and geochemical studies of soils to investigate land-use patterns, with the greater goal of distinguishing habitation and cultivation areas. They find that the spatial covariance of certain major elements or rare-earth elements and trace metals corresponds to activity loci. For example, calcium and strontium signatures are found in cultivated soils that were enriched with shell sand and bone or fish remains to enhance soil fertility, while agricultural plots that exhausted soil fertility and nutrient capacity contain soils with significantly depleted concentrations of zinc, nickel, magnesium and copper. In contrast, habitation zones are enriched in potassium, thorium, rubidium and caesium. While it is unclear which specific human activities resulted in the deposition of these chemical elements, the signatures are nonetheless useful for studying spatial patterns of activity loci in the past, which may help model site form and function. The second example involves reconstructing daily practices in pre-Hispanic residential sectors of Piedras Negras, Guatemala, and Cer6n, E1 Salvador. Here, Terry and colleagues (Terry et al. 2000, 2004; Wells et al. 2000; Parnell et al. 2001, 2002a, 2002b) focus on phosphorus and heavy metals (namely copper, iron, mercury, manganese, lead and zinc) to infer specific activities for different kinds of architecture and features. For example, they find that interior spaces in residential buildings contain soils enriched in phosphorus, which they interpret as evidence for the preparation and consumption of meals. On the other hand, they interpret low phosphate signatures surrounding building exteriors as possible evidence for ancient roof drip-lines in which rainwater presumably would have washed away organic debris enriched in phosphors. Areas of craft

CULTURAL SOILSCAPES production involving the use or application of pigments are marked by combinations of heavy metals, with broad, linear patterns of copper and manganese concentrations possibly indicating directions of sweeping and related cleaning activities. Heavy metals also are found in association with plaster fragments at the bases of walls along the exterior faqades of buildings, suggesting that some walls may have been painted in antiquity. Finally, they use soil chemistry to create typologies for different kinds of refuse deposits, including those that contain high amounts of organic matter versus those that contain craft production debris. Knowing the spatial distribution of contrasting refuse types across a site, in addition to the types and locations of activities that produced this refuse, may help in reconstructing ancient domestic activity patterns.

Future directions Future d e v e l o p m e n t s in the archaeological study of cultural soilscapes inevitably will depend on ethno-archaeological research aimed at linking soil chemical signatures of m o d e r n (observable) h u m a n activities with those of prehistoric peoples in archaeological sites, to u n d e r s t a n d the cultural pathways by which elements and compounds are deposited in soils. This work has largely just begun (e.g. Barba & Ortiz Butr6n 1992; Barba et al. 1995; Middleton & Price 1996; Fernfindez et al. 2002; Wells & Urban 2002; Terry et al. 2004), but the results show great promise. W h e n these kinds of h u m a n - c e n t r e d investigations are c o m b i n e d with broader landscape approaches, it will be possible to study and u n d e r s t a n d cultural soilscapes in a far more detailed m a n n e r than has been the case. For example, more sensitive and flexible frameworks for classifying, mapping, and studying soils can be created. Soil taxonomy was developed in the mid-1970s in the US to characterize soil variation by measuring certain quantifiable properties, including diagnostic horizons and soil moisture and temperature. However, it has been pointed out that this taxonomy creates classes that are only partially related to l a n d f o r m (Young & H a m m e r 2000), possibly because cultural impacts are not considered. Sandor (1992), who considers the cultural soilscape, has w o r k e d closely with indigenous communities to develop folk typologies for soil classes, some of which crosscut those described using traditional soil taxonomy. While these kinds of classification approaches are not widely applicable, they strongly suggest that human use and perception

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of cultural soilscapes have i m p o r t a n t consequences for u n d e r s t a n d i n g soil f o r m a t i o n processes.

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SCUDDER, S.J., FOSS, J.E. & COLLINS, M.E. 1996. Soil science and archaeology. In: SPARKS, D.L. (ed.) Advances in Agronomy. Academic Press, New York, 1-76. SIMVSON, I.A., BOL, R., DOCKR1LL, S.L., PETZKE, K.-J. & EVERSHED, R.E 1997. Compound-specific 815N amino acid signals in palaeosols as indicators of early land use: a preliminary study. Archaeological Prospection, 4, 147-152. SIMPSON,I.A., DOCKRILL,S.J., BULL, I.D. & EVERSHED, R.E 1998. Early anthropogenic soil formation at Tofts Ness, Sanday, Orkney. Journal of Archaeological Science, 25, 727-746. SPARKS, D.L. 2003. Environmental Soil Chemistry. Academic Press, New York, Second Edition. STEIN, J.K. & FARRAND, W.R. (eds) 1985. Archaeological Sediments in Context. University of Maine, Orono. SULLIVAN,A.E 2000. Effects of small-scale prehistoric runoff agriculture on soil fertility: the developing picture from upland terraces in the American Southwest. Geoarchaeology: an International Journal, 15, 291-313. TERRY, R.E., FERNANDEZ, EG., PARNELL, J.J. & INOMATA,T. 2004. The story in the floors: chemical signatures of ancient and modern Maya activities at Aguateca, Guatemala. Journal of Archaeological Science, 31, 1237-1250. TERRY, R.E., HARDIN, EJ., HOUSTON, S.D., NELSON, S.D., JACKSON,M.W., CARR, J. & PARNELL,J.J. 2000. Quantitative phosphorus measurement: a field test procedure for archaeological site analysis at Piedras Negras, Guatemala. Geoarchaeology: an International Journal, 15, 151-166. TURNER II, B.L. & HARRISON, ED. (eds) 1983. Pulltrouser Swamp: Ancient Maya Habitat, Agriculture, and Settlement in Northern Belize. University of Texas Press, Austin. VAN DER LEEUW, S. & REDMAN, C.L. 2002. Placing archaeology at the center of socio-natural studies. American Antiquity, 67, 597-605. VIZCMNo, A.S. & CAIqABATE,M.L. 1999. Identification of activity areas by soil phosphorus and organic matter analysis in two rooms of the Iberian sanc-

tuary 'Cerro E1 Pajarillo.' Geoarchaeology: an International Journal, 14, 47-62. WAGSTAFF, J.M. 1987. Landscape and Culture: Geographical and Archaeological Perspectives. Basil Blackwell, London, UK. WATERS, M.R. & RAVESLOOT, J.C. 2001. Landscape changes and the cultural evolution of the Hohokam along the Middle Gila River and other river valleys in south-central Arizona. American Antiquity, 66, 285-299. WEBB, E.A., SCHWARCZ, H.E & HEALY, RE 2004. Detection of ancient maize agriculture in the Maya lowlands using the stable carbon isotope compositions of soil organic matter: evidence from Caracol, Belize. Journal of Archaeological Science, 31, 1039-1052. WELLS, E.C. 2004. Investigating activity patterns in prehispanic plazas: weak acid-extraction ICP/AES analysis of anthrosols at Classic period El Coyote, northwest Honduras. Archaeometry, 46, 67-84. WELLS, E.C. • URBAN, EA. 2002. An ethnoarchaeological perspective on the material and chemical residues of communal feasting at El Coyote, northwest Honduras. In: VANDIVER,P., GOODWAY,M. & MASS, J. (eds) Materials Issues in Art and Archaeology VI. Materials Research Society, Warrendale, Pennsylvania, 193-198. WELLS, E.C., TERRY, R.E., PARNELL,J.J., HARDIN,EJ., JACKSON, M.W. & HOUSTON, S.D. 2000. Chemical analyses of ancient anthrosols in residential areas at Piedras Negras, Guatemala. Journal of Archaeological Science, 27, 449-462. WILD,A. 1986. The retention of phosphate by soil: a review. Journal of Soil Science, 1, 221-238. WINGARD, J.D. 1996. Interactions between demographic processes and soil resources in the Copfin Valley, Honduras. In: FEDICK, S.L. (ed.) The Managed Mosaic: Ancient Maya Agriculture and Resource Use. University of Utah Press, Salt Lake City, 207-235. YOUNg, EJ. & HAMMER, R.D. 2000. Defining geographic soil bodies by landscape position, soil taxonomy, and cluster analysis. Soil Science Society of America Journal, 64, 989-998.

From agricultural geology to hydropedology: forging links within the twenty-first-century geoscience community EDWARD

R. L A N D A

US Geological Survey, 430 National Center, Reston, Virginia 20192, USA (e-maik [email protected])

Abstract: Despite historical linkages, the fields of geology and soil science have developed along largely divergent paths in the United States during much of the mid- to latetwentieth century. The shift in recent decades within both disciplines, towards greater emphasis on environmental-quality issues and a systems approach, has created new opportunities for collaboration and cross-training. Because of the importance of the soil as a dynamic interface between the hydrosphere, biosphere, atmosphere and lithosphere, introductory and advanced soil-science classes are now taught in a number of Earth and environmental science departments. The National Research Council's recent report, Basic Research Opportunities in Earth Science, highlights the soil zone as part of the land surface to groundwater 'critical zone' requiring additional investigation. To better prepare geology undergraduates to deal with complex environmental problems, their training should include a fundamental understanding of the nature and properties of soils. Those undergraduate geology students with an interest in this area should be encouraged to view soil science as a viable Earth-science specialty area for graduate study.

Geology's traditional subspecialties have looked at the E a r t h divided by major rock types (e.g. igneous, s e d i m e n t a r y and m e t a m o r p h i c petrology); by links to engineering, chemistry, physics, and biology (i.e. engineering geology, geochemistry, geophysics and palaeontology/ geobiology); by a focus on a specific environment or resources (i.e. oceanography, hydrogeology and economic geology); or by a focus on the static and dynamic properties of E a r t h materials (i.e. mineralogy, geomorphology and structural geology). Soil science looks at the outer skin of the Earth. By definition, its place as a geology subspecialty is clear. But the reality is that soil science in the United States has developed in a remarkably separate and distinct way from geology, and it is typically regarded as an agricultural science rather than an E a r t h science. In the past two decades, however, both fields have shown major shifts towards a focus on environmental-quality issues. Soils play an essential role in supporting life on Earth; are a scarce resource of great economic value; are easily disturbed and respond quickly to environmental changes; and indeed, as Haft (2002) has noted, bear the brunt of human impact on the land surface. These realities have no national boundaries. While the examples in this paper are, based on the author's experience, largely from the United States, it is h o p e d that the discussion will encourage others to compare and contrast the status of soil science within the larger Earth-science community in their nations.

The need for interdisciplinary approaches to address complex issues of resource- and wastem a n a g e m e n t in the surficial e n v i r o n m e n t has created greater opportunities for geologists and soil scientists to work together. With this shift has come the need for greater cross-training of students in each field. To encourage such interaction, it is of value to look at areas of connection in the past and present, and to examine ways to bridge the gaps that divide the disciplines.

Historical setting Soil science emerged as an integrated discipline from work in geology on soil formation, and work in agricultural chemistry on plant nutrition. Agricultural geology was a recognized specialty for geologists in the nineteenth and early twentieth centuries (Tandarich 1998). At the beginning of the twentieth century, there were geological institutes in Hungary and elsewhere in E u r o p e with departments of agrogeology (Szabolcs 1997). In the United States, many early reports from state geological surveys contained sections on the distribution of soils and crop production (e.g. Jackson 1840). While geology d e p a r t m e n t s exist in the colleges of arts and sciences at many institutions of higher learning in the United States, soilscience d e p a r t m e n t s typically have been confined to the land-grant universities in each state. The Morrill Acts of 1862 and 1890

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266, 133-140. 0305-8719/06/$15 9 The Geological Society of London 2006.

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established the land-grant institutions to promote education in agriculture, and it was within this academic setting and its associated system of agricultural experiment stations that the field of soil science traditionally has been taught and practised. In some universities outside of the United States (e.g. Massey University in New Zealand, the Czech University of Agriculture in the Czech Republic, Ghent University in Belgium and Wageningen University in the Netherlands), there currently are joint soil-science and geology departments. Indeed, this was the case at some American land-grant colleges at the turn of the twentieth century. Soils occupy a niche that gives them a dynamic character and a subtle memory. The organic content of surface soil can change rapidly (in as little as 10 to 100 years) in response to climatic and ecological changes or land management practices. In contrast, extensive weathering of a soil's mineral content requires much more time. Soils thus acquire their basic attributes at very different rates. They reflect both the present and the past, recording how they have changed in response to recent events while they document changes (like weathering) that have occurred over tens of thousands of years.

This elegant quote from University of Wisconsin soil scientists Kevin McSweeney and John Norman (1996) sums up the central role of soils in biogeochemistry. Indeed, the field has its roots in the 1926 treatise The Biosphere by Russian soil scientist Vladimir I. Vernadsky. Research on cycling of nutrients such as carbon, nitrogen, sulphur and phosphorus, and of contaminant elements such as arsenic, is under way in many soil science departments today. Soil scientists have made major contributions in the fields of geology and hydrology. For example, over the past five decades, soil scientists at the US Geological Survey (USGS: the largest geoscience research agency in the United States government) have been leaders in studies on: (1) the isolation and characterization of the nature and properties of soil and aquatic humic substances; (2) the retention of trace elements by iron and manganese oxides, and development of selective extraction techniques for use in geochemical prospecting; (3) the effects of industrial activities on the distribution of trace elements in soils and plants; (4) carbon cycling and sequestration in watersheds; (5) the movement of water, soil gases and contaminants in the unsaturated zone at nuclear-waste burial and other sites;

(6) the geochemical forms of radionuclides in uranium mill-tailings, and their mobilization by microbial processes; (7) the relation of groundwater quality to land use; and (8) the modelling of the transport of reactive solutes in porous media. At present, the most requested dataset in the USGS National Geochemical Database consists of analyses of 1300 surface soils collected from non-cultivated fields with native vegetation in the conterminous United States during the 1960s and 1970s; these data are widely used to establish baseline concentrations of metals and other elements (Smith et al. 2003). With the goal of expanding the database and extending its coverage to all of North America, the USGS, and partners in Canada (Geological Survey of Canada; Agriculture and Agri-Food Canada) and Mexico (Consejo de Recursos Minerales; Instituto Nacional de Estadistica Geografia e Informatica) have recently established a Geochemical Landscapes Project to sample about 10 000 sites over the span of about a decade. Total elemental analyses, as well as partial extractions (deionized water extraction; simulated human gastric-fluid extraction) to assess bioaccessibility, will be performed on O-, A-, B- and C-horizon materials. The microbial communities in the A-horizon will be documented by techniques such as phospholipid fatty-acid analysis (PLFA), enzyme assays, and B I O L O G carbon source metabolic profiling. Selected samples will be analysed for organic contaminants. Other soils-related research uses artificial intelligence systems to describe habitat suitability of soils for hosting the fungus Coccidiodes. Airborne spores of this organism can cause coccidioidomycosis ('valley fever') in humans (Bultman et al. 2004). Living up to its new motto science for a changing world', the USGS is now different from what many in the geological and soil science communities may remember or recognize. In the USGS and other geosciences agencies today, where topics such as pesticide-, nutrient- and pathogen-transport, carbon storage, medical geology and rangeland quality are now commonplace subjects for investigations, research opportunities for soil scientists and geologists with training in soil science are growing rapidly on multiple fronts.

'The Critical Zone' In its most recent assessment of Earth science research in the United States, the National

AGRICULTURAL GEOLOGY TO HYDROPEDOLOGY Academy of Sciences/National Research Council (2001) noted the need for more multidisciplinary, integrative studies of the Earth's heterogeneous surface and near-surface environment 'where complex interactions involving rock, soil, water, air, and living organisms regulate natural habitats and determine the availability of life-sustaining resources.' The focus in the NAS National Research Council (2001) report on the 'Critical Zone' [defined as including the soil zone (pedosphere), unsaturated vadose zone, and the saturated groundwater zone] is striking, and far more emphatic than in the previous major assessment by the group (National Research Council 1993). In order to better understand biogeochemical processes active in the Critical Zone, to assess human impacts there, and to help society adapt to the consequences of these impacts, soil science and the study of coastal zone processes were singled out as disciplines requiring greater attention and funding by the National Science Foundation's (NSF) Earth Science Division. The research direction and training priorities established in the 2001 report can be expected to guide funding decisions in the US Earthscience community during the coming decade. Following up on the NAS National Research Council (2001) report, an NSF-sponsored workshop 'Frontiers in Exploration of the Critical Zone' (October 24-26 2005, University of Delaware) addressed four major questions (Brantley et al. 2006): (1) How are the rates of physical and chemical weathering perturbed by environmental forcing? (2) How do important biogeochemical processes occurring at Critical Zone interfaces govern long-term sustainability of soil and water resources? (3) How do processes in the Critical Zone nourish ecosystems and how do they respond to changes in external forcing? (4) What processes in the Critical Zone control biosphere atmosphere exchanges of atmospherically important gases and particulates? A proposal from a team of soil scientists and geologists (the Weathering System Science Consortium; http://www.wssc.psu.edu/) is already in place to study the Earth's weathering engine (Anderson et al. 2004). The plan calls for establishment of three highly instrumented 'node' field sites to study weathering at the soil profile- and catchment-scales, and a broader network of 'backbone' soil sites across a range

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of lithologies, ecosystems and topographies; data from these sites will be used in development of models to interpret weathering systems. A recognition of the importance of soils is not limited to the basic science community. Geologists and engineers are increasingly dealing with soil-contamination clean-ups and wetland delineation. Demand for housing and commercial space in the United States has led to the rapid expansion of suburban communities in areas beyond the reach of existing sewer systems. More than 37% of new development in the United States uses on-site wastewater disposal systems (Dix 2001). This has created an increasing need for expertise in soil suitability for the safe disposal of wastewater using septic fields and alternative technologies. Environmental health specialists, engineers and geologists who are called upon to make these siting decisions look to soil scientists for guidance on issues such as soil structure and redox status as indicators of suitable soils. In assessing the long-term fate of contaminants in mine and mill tailings, soilforming processes provide a scientific framework. For example, a knowledge of the similarities between acid sulphate soils and pyritic tailings can aid in the management of each. As the mission of geoscience agencies shifts from classical geological projects (mineral resource assessments, bedrock mapping, palaeontology) to environmental issues focused on the surficial environment (water quality, land-use planning, remedial action at contaminated sites), hiring of soil scientists and providing mid-career training in soil science to the existing workforce can add a necessary component of expertise in multidisciplinary environmental investigations.

K-12 education Although the thrust of this paper is on the need for integrated soil science and geology education in colleges and universities, the underpinning of these undergraduate and graduate programmes is Earth-science education at the kindergarten to high-school level (grades K-12). At present (2001 statistics), less than 7% of the high-school students (grades 9-12; ages 14-17 yr) in the United States will take a class in Earth and space science (Ridky 2002). This is a steep decline from the 26% figure in 1968 (Earth Science Curriculum Project 1969). The decline mirrors a general drop in science literacy in American society, and presents a challenge for recruiting the next generation of Earth scientists. New national education standards for K-12 do, however, call

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for training in E a r t h and space science. In Virginia, for example, fourth-grade students are now taught soil formation, soil horizons and soil properties (colour, texture, structure and moisture-holding capacity). Geochemical cycles are taught in grades 9-12 (National Research Council 1996). There are other bright spots in the K-12 picture. Recent efforts to improve the teaching of soils at the middle-school level (grades 6-8; ages 11-13 yr) have been undertaken by the American Geological Institute as part of its enquiry-based Investigating Earth Systems curriculum (http://www.agiweb.org/education/ ies/soils/index.html. N A S A soil scientist Elissa Levine hosts a comprehensive K-12 soilscience education home page (part of NASA's G L O B E Program) at http://www.globe.gov/fsl/ welcome.html. Among the topics included are ion exchange, thin sections, soil monoliths and soil characterization. Interest in stimulating K12 interest in soils is also apparent in other nations. For example, G. V. Dobrovol'skii and colleagues at Moscow State University have chronicled the teaching of soil science in Russian primary and secondary schools, and the availability of texts from the early twentieth century to the present (Dobrovol'skii et al. 2002). With the support of the Soil Science Society of America, a major exhibit on soils at the Smithsonian Institution's National Museum of Natural History in Washington, DC is in preparation (http://www.soils.org/smithsonian). This is the most-visited museum in the world, with six to nine million visitors annually. The soils exhibit will complement existing Earth-science halls devoted to fossils and minerals (including the very popular display of the Hope Diamond), and an insect zoo with termite and ant mounds. Tentative topics for the soils exhibit include the role of soil in the environment; food and medicine from soils; soils in cultural history; and careers in soil science. This exhibit and its potentially huge audience represent an unprecedented opportunity to reach a large number of students with the message: Soils Sustain Life. Smaller outreach and educational efforts, prepared by individuals and teams of soil scientists in various states, include a 'living soil tunnel' - a trailer unit taken to schools, and soilpit demonstrations for grade-school students (grades K-5; age 5-10 yr).

College and university level Introductory geology classes and textbooks must cover a wide range of topics. Soils and

weathering generally receive one-chapter treatment. The focus in most introductory texts is on the soil profile. The modern system of soil classification used in the United States is often ignored in favor of the old 'podsol-pedalferpedocal-laterite' system. The new system, with its 11 soil orders, is an elegant scheme, developed over several decades, based on observable and measurable soil features, and using nomenclature with information-filled prefixes and suffixes. Its inclusion in introductory geology classes will fill an obvious gap. For supplementary topics in upper division courses (years 3 and 4 of 4 yr undergraduate curriculum), instructors should consider including material on the dynamic properties of soils - e.g. discussions in geochemistry classes of ion exchange and nutrient cycling, and discussions in hydrogeology classes of shrink-swell behaviour and water movement. Suggested topics for inclusion in undergraduate geology courses are given in Table 1. Introductory and, in some cases, advanced soil-science classes are presently taught in a rapidly expanding number of geography and Earth- and environmental-science departments in the United States. Cross-training across traditional academic boundaries is growing. Among the geology faculty at a variety of institutions, one now sees soil scientists, or geologists with graduate or post-doctoral training in soil science, in a variety of new faculty positions targeting surficial processes. These trends are also apparent in graduate student-, postdoctoral- and research-support staff recruitment and hiring. Cross-disciplinary training and hiring is a reality for soil scientists, as well as for geologists. For example, the undergraduate soil-science programme at Pennsylvania State University now requires a course in hydrogeology. Students from soil science with interests in palaeosols and saturated zone hydrology may opt for graduate training in geology departments. Tackling groundwater vulnerability, soil contamination, land-use and other issues dealt with by regulatory agencies and environmental consulting companies requires a breadth of Earth-science expertise. As a result, employment opportunities for students with cross-disciplinary training in soils and geology can be expected to grow in coming years. In the United States, advanced undergraduates in geology and other fields wishing to explore career options in soil science can get research experience as summer interns at Washington State University, Pennsylvania State University and the University of California at

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Table 1. Possible soils topics for inclusion in undergraduate geology courses

Ion exchange, and origin of permanent and pH-dependent charges on minerals Concept of bioavailable nutrients. Use of plant-available soil-testing procedures and other selective extraction procedures (in contrast to total-element measurements) Bulk density, soil structure and their effects on water flow Soil (and rock) colour description using the Munsell charts. Use of soil colour to identify redox conditions in soil profiles, and implications for land-use management Concepts of plant-available water, water-holding capacity and use of the soil-moisture retention curve to calculate available water Five factors of soil formation Soil taxonomy; its nomenclature and diagnostic criteria Soil micromorphology, and use of soil thin-sections (Stoops 2003; accompanying CD contains hundreds of thin-section images) Indices of soil quality Role of micro-organisms in cycling of carbon, nitrogen, sulphur and iron Examination of soil and rock materials in criminal investigations (Murray & Tedrow 1975; Murray 2004)

Davis under the NSF-funded Integrative Graduate Education, Research, and Training (IGERT) programme. The IGERT centres are also bringing geology and soil-science faculty, post-docs, and graduate students together on research projects dealing with biogeochemistry, surface and colloid chemistry, biodegradation of soil contaminants, and soil-water movement. The IGERT programme is expanding to other campuses. In Europe, similar interactions across Earthand environmental-science disciplines are made available for doctoral students by the E U Marie Curie Training Sites Programme (MCTS). Among those programmes with research opportunities in soil science are the MCTS Postgraduate Centre for Biogeochemistry at the University of Newcastle upon Tyne (http://www. ceg.ncl.ac.uk/mariecurie/mariecuriel.htm), and the Chemical Speciation, Biological Availability and Ecotoxicological Effects of Contaminants in Soil and Water MCTS of the Netherlands Research School for the Socio Economic and Natural Sciences of the Environment (SENSE), a consortium of environmental research groups at seven Dutch universities (including the soilquality group at Wageningen University and Research Centre) (http://www.sense.nl/MCTS). Maintaining enrolment of students majoring in soil science and geology is a challenge at many American colleges and universities. With declining student numbers and shrinking budgets, as well as proactive steps toward intellectual integration, have come moves to consolidate departments and to change their focus.

These transformations within academic institutions are not unique to our science, or to our era. To use the language of the business world, one sees 'morphing and reinvention' (e.g. in 2004, the D e p a r t m e n t of Wood and Paper Science at the University of Minnesota became the Department of Bio-based Products) to meet perceived changes in the scope of the field. Mergers and acquisitions represent another restructuring strategy; e.g. in 1975, the Department of Soils and Plant Nutrition at the University of California at Davis merged with the Department of Water Science and Engineering, and with the Atmospheric Science programme of the Department of Agricultural Engineering, to become the D e p a r t m e n t of Land, Air and Water Resources. At the University of Maryland, soil science, plant sciences, and landscape architecture are now all housed under the umbrella of the Department of Natural Resource Sciences and Landscape Architecture. These examples, far from being isolated anecdotes, represent the main trend. Nevertheless, counter-examples exist. For example, at the University of Maryland, the geology and agronomy faculties (now in separate departments and colleges) were in the same department until 1973 (Robert Ridky, pers. comm. 2002). Although transformations of organizational boundaries are traumatic, one side-benefit of mergers may be enhanced interdisciplinary teaching and research - including those across the soil-science-geology interface. Likewise, it should be noted that the evolution or

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restructuring of classical geology departments into departments with a broader environmental geoscience orientation is not without peril. The preservation of the core geology base should be a guiding principle in such actions, in order to yield graduates able to pursue more traditional geoscience career paths (e.g. within the petroleum and mining industries), as well as those seeking employment in government agencies, non-profit organizations, and commercial enterprises with a more integrated geoscience approach. The same can be said of classical soilscience departments and their need to preserve core strengths in soil fertility and plant nutrition, and other areas not closely allied with the environmental geoscience interface. Recruitment of new soil scientists is a continuing concern, especially now with a large wave of retirements anticipated in the next decade. While the 'big three' sciences (biology, chemistry, physics) have prominence in the high-school curriculum, subspecialty sciences such as soil science do not have such exposure. What mechanisms are most effective in providing career information to high-school and undergraduate students? Little research seems to have been done in this area. Older approaches, such as career brochures and limited outreach programmes, may not be the most effective route in the present, webdominated information age. The same holds true for other subspecialty sciences such as entomology and plant pathology. Recruitment of new talent is critical to the future problemsolving needs of society.

Professional society activities The Soil Science Society of America (SSSA) is the major soil-science professional society in the United States, with some 5700 members. In 1993, SSSA became a member of the American Geological Institute, which is a federation of 43 geoscience societies. The place of soil science in the greater Earth-science community has been greatly enhanced by this affiliation. The SSSA has an outreach programme whose goals are to spread soil-science knowledge to other fields of science, and to heighten the awareness of soil science as an Earth-science discipline. Its sponsorship of a 2001 workshop aimed at promotion of soil-science training in the undergraduate geology curriculum at non-land-grant colleges and universities was a recent step in that direction. A landmark event will occur in 2008, when SSSA and the Geological Society of America (GSA) will hold a joint annual meeting. Among the existing co-operative efforts of SSSA and

G S A is publication of the Vadose Zone Journal, and co-sponsorship of the new biogeosciences website (http://www.biogeosciences.org/). Other Earth-science professional societies in which a good mix of soil scientists and geologists exists include the Friends of the Pleistocene, the Hydrology Section of the American Geophysical Union, and the Clay Minerals Society. Informal, one-on-one interactions within these societies have had a major role in helping to integrate the disciplines. As exemplified by this book, the 32nd International Geological Congress in Florence in 2004 had a strong soil-science presence. Integration at the international level has continued at the International Union of Soil Sciencessponsored 18th World Congress of Soil Science in Philadelphia in 2006 (http://www.18wcss.org/). Symposia included: (1) soil geochemical patterns at the regional, national, and international scales; (2) arid soils: genesis, geomorphology and geoarchaeology; (3) soils on limestones: their properties genesis and role in human societies; (4) imprint of environmental change on palaeosols; (5) soil mineralogy and geophysics: environmental and soils management and mineral exploration; and (6) soils and natural hazards.

The path forward In the late 1880s and early 1890s, John Wesley Powell, the second director of the USGS and a strong supporter of soil surveys, pressed to have the agency moved from the Department of the Interior to the newly formed Department of Agriculture (Amundson & Yaalon 1995). Such were the ties between geology and soil science a century ago. We are now at another interesting crossroads. The modern societal and scientific perception of soils has been extended beyond soil as solely a medium for plant growth. Soil is now also viewed as a natural body, a structural mantle, a water-transmitting mantle and an ecosystem component (Smith & Hudson 2002). Clear links exist from these perspectives of soils to geography, geomorphology, engineering geology, hydrology and hydrogeology, and biogeochemistry. This is an exciting time to be working at the soil science/geology interface. Hydrogeologists are investigating macropore flow, a phenomenon first studied by soil physicists, as a possible aquifer-recharge mechanism in semi-arid regions (Wood et al. 1997). At the

AGRICULTURAL GEOLOGY TO HYDROPEDOLOGY same time, we have soil scientists using acoustic b a t h y m e t r y data collected by the National Oceanic and A t m o s p h e r i c A d m i n i s t r a t i o n to help map subaqueous soils (Bradley & Stolt 2002). Within the geological and geographical sciences, there has been a recent emergence of many subspecialty and hybrid fields, in which a knowledge of the properties and behaviour of soils, and an appreciation of their spatial heterogeneity and temporal dynamics is critical. These include: E c o h y d r o l o g y - the study of plant-water inter-

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continuum of their development as geoscientists. Soil-science d e p a r t m e n t s offer u n i q u e opportunities for training, with a disciplinary depth not generally available elsewhere, in pedology, soil physics, soil chemistry and mineralogy, and soil microbiology. The views expressed here are those of the author, and do not reflect official policy of the US Geological Survey. This paper is updated and adapted from an earlier journal article: Landa, E. R. 2004. Soil science and geology: connects, disconnects and new opportunities in geoscience education. Journal o f Geoscience Education, 52, 191-196. The permission of the Journal of Geoscience Education for its usage here is gratefully acknowledged. Robert Ridky, Educational Coordinator for the USGS, provided useful dialogue and much appreciated encouragement. Any use of trade, product or firm names in this report is for descriptive purposes only and does not imply endorsement by the US government.

actions; of the effects of hydrological processes on the distribution, structure, and function of ecosystems; and of the effect of biological processes on components of the hydrological cycle (Eagleson 2002; Newman et al. 2003). E t h n o p e d o l o g y - the study of the perception, classification, appraisal, use and managem e n t of soils by indigenous p e o p l e (WinklerPrins & Sandor 2003). H y d r o g e o p h y s i c s - the application of geophysical techniques for hydrogeological characterization of the shallow subsurface (Hubbard & Rubin 2002; Miiller 2003). H y d r o p e d o l o g y - the bridging of traditional pedology with hydrology, geostatistics and soil physics for application in soil-landscape modelling (Lin 2003). N a n o g e o s c i e n c e - the study of geological processes (typically near-surface) involving particles smaller than 100 n a n o m e t r e s (National Science Foundation 2002). N e o g e o m o r p h o l o g y - the study of the change of the Earth's surface as a result of human activity (Haft 2002).

AMUNDSON, R. ~ YAALON,D.H. 1995. E.W. Hilgard and John Wesley Powell: efforts for a joint agricultural and geological survey. Soil Science Society o f America Journal, 63, 1485-1493. ANDERSON,S.E, BLUM,J. ETAL. 2004. Proposed initiative would study Earth's weathering engine. Eos,

While t o d a y there are m a n y cross-cutting issues in geology and soil science - such as carbon sequestration, water quality and vadosezone hydrology and contamination - there are also differences in perception, training and institutional affiliations that tend to keep the disciplines and individual geologists and soil scientists apart. The realities of dealing with complex e n v i r o n m e n t a l problems, and the actions of scientific advisory groups and professional societies, are helping to bridge these gaps. Exposure of geology undergraduate students to an in-depth e x a m i n a t i o n of the nature and properties of soils will be of benefit to their professional d e v e l o p m e n t and later work experience. Graduate training in a soilscience department should be viewed as a viable option for undergraduate geology students with interests in the surficial environment, and as a

Expression o f Vegetation Form and Function.

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85, 265, 269. BRADLEY, M.E & STOLT, M.H. 2002. Evaluating methods to create a base map for a subaqueous soil inventory. Soil Science, 167, 222-228. BRANTLEY,S.L., WHITE,T.S. ET AL. 2006. Frontiers in Exploration of the Critical Zone: Report of a workshop sponsored by the National Science Foundation (NSF), October 24-26, 2005, Newark, DE, 30p; http://www.wssc.psu.edu/booklet.pdf accessed 31 July 2006. BULTMAN,M.W., FISHER,ES. t~z GETTINGS,M.E. 2004. Coccidioidomycosis: mitigating risk. GeoHealth News, 3 (1), 2-6. World Wide Web Address: http://energy.er.usgs.gov/medical__geology.htm. DIx, S.E 2001. Onsite wastewater treatment: a technology and management revolution: part 1. Water Engineering & Management, 148, 24-28. DOBROVOL'SKII,G.V., ORLOV,D.S. & ROZANOVA,M.S. 2002. Soil science in secondary school and popular books on soils. Eurasian Soil Science, 35, 210-215. EAGLESON, P.S. 2002. Ecohydrology: Darwinian Cambridge University Press, Cambridge, 443 pp. EARTH SCIENCE CURRICULUM PROJECT 1969. E S C P Newsletter, 20 (October 1969). American Geological Institute, Alexandria, Virginia. HAFF, RK. 2002. Neogeomorphology. Eos, Transactions o f the American Geophysical Union, 83, 310, 317. HUBBARD,S. & RUBIN,Y. 2002. Study institute assesses the state of hydrogeophysics. Eos, Transactions o f the American Geophysical Union, 83, 602, 606.

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JACKSON, C.T. 1840. Report on the Geological and Agricultural Survey of the State of Rhode-Island, Made under the Resolve of Legislature in the Year 1839. B. Cranston and Co., Providence, Rhode Island. LIN, H. 2003. Hydropedology - bridging disciplines, scales, and data. Vadose Zone Journal, 2, 1-11. MCSWEENEY, K. & NORMAN, J.M. 1996. Soil land modeling: issues of scale. Geotimes, 41, 22-24. MULLER, M. 2003. Opening doors for geophysics in soil sciences. Los, Transactions of the American Geophysical Union, 84, 243. MURRAY,R.C. 2004. Evidence from the Earth: Forensic Geology and Criminal Investigation. Mountain Press Publishing, Missoula, Montana, 240 pp. MURRAY, R.C. & TEDROW, J.C.E 1975. Forensic Geology: Earth Sciences and Criminal Investigation. Rutgers University Press; New Brunswick, New Jersey. NATIONAL ACADEMY OF SCIENCE, NATIONAL RESEARCH COUNCIL (BOARD ON EARTH SCIENCES AND RESOURCES, COMMISSION ON GEOSCIENCES, ENVIRONMENT, AND RESOURCES) 2001. Basic Research Opportunities in Earth Science. National Academy Press, Washington, DC. NATIONAL RESEARCH COUNCIL (BOARD ON EARTH SCIENCES AND RESOURCES, COMMISSION ON GEOSCIENCES, ENVIRONMENT, AND RESOURCES) 1993. Solid-Earth Sciences and Society. National Academy Press, Washington, DC. NATIONAL RESEARCH COUNCIL (NATIONAL COMMITTEE ON SCIENCE EDUCATION STANDARDS AND ASSESSMENT) 1996. National Science Education Standards. National Academy Press, Washington, DC. NATIONAL SCIENCE FOUNDATION 2002. Report of the Nanoscience Workshop (Berkeley, California, June 14-16, 2002). NEWMAN,B.D., SALA, O. & WILCOX,B.P. 2003. Conference promotes study of ecohydrology of semi-arid

landscapes. Los, Transactions of the American Geophysical Union, 84, 13, 17. RIDKY, R.D. 2002. Why we need a corps of Earth science educators. Geotimes, 47, 16-19. SMITH, D.B., GOLDHABER, M.B., WILSON, M.A. & BURT, R. 2003. A Proposal for Upgrading the National-Scale Soil Geochemical Database for the United States. United States Geological Survey Fact Sheet FS-015-03. World Wide Web Address: http://pubs.usgs.gov/fs/fs-015-03/. SMITH, H. & HUDSON, B.D. 2002. The American soil survey in the twenty-first century, In: HELMS, D., EFFLAND, A.B.W. & DURANA, RJ. (eds) Profiles in the History of the US.SoU Survey. Iowa State University Press, Ames, Iowa, 303-313. STOOPS, G. 2003. Guidelines for Analysis and Description of Soil and Regolith Thin Sections. Soil Science Society of America, Madison, Wisconsin. SZABOLCS,I. 1997. The First International Conference on Agrogeology, April 14-24, 1909, Budapest, Hungary, In: YAALON,D.H. & BERKOWICZ,S. (eds) History of Soil Science: International Perspectives, Advances in Geoecology, 29. Catena Verlag GmbH, Reiskirchen, Germany. TANDARICH, J.E 1998. Agricultural chemistry: disciplinary history; agricultural geology: disciplinary history. In: GOOD, G. (ed.) Sciences of the Earth: an Encyclopedia of Events, People, and Phenomena. Garland Publishing, New York, 19-23, 23-29. VERNADSKY, V.I. 1926. The Biosphere [translated to English and reprinted with commentary; 1998] Copernicus (Springer-Verlag), New York, 192 pp. WINKLERPRINS,A.M.G.A. & SANDOR,J.A. (eds) 2003, Ethnopedology (special issue). Geoderma, 111 (3-4), 165-538. WOOD, W.W., RAINWATER, K.A. & THOMPSON, D.B. 1997. Quantifying macropore recharge: examples from a semi-arid area. Ground Water, 35, 1097-1106.

Australian examples of the role of soils in environmental problems E A. HAZELTON

Faculty o f Engineering, University o f Technology, Sydney, Australia (e-maik Pam. Hazelton @eng. uts. edu. au) Will everyone, please at some stage during the next few days take hold of a handful of soil and show it some respect? Grab some, take a hard look at it, marvel at its workings, sing in praise of its chemistry,bless its bugs for being there and before piacing it on the ground, say thank you. If ever the environment had a poor relation, it is the soil. It has none of the romance of furry animals,birds or butterflies and yet it is fundamental to life on earth. (Heiney 1997)

Soil as the ultimate e n v i r o n m e n t a l interface Soil is a distinctive, identifiable part of the e n v i r o n m e n t . It is a u n i q u e resource, and, together with water, is fundamental for life on Earth. Soil and its properties result from the interaction of chemical, physical and biological activities. Soil is, in effect, the ultimate interface between the geosphere, the atmosphere, the hydrosphere and the biosphere (Rimmer 1998). The soil type is influenced by environmental factors, including the p a r e n t m a t e r i a l from which it weathers, vegetation, climate, topography, and availability of water. The interaction with soils is of mutual benefit for most life-forms (Yaalon & Arnold 2000). The soil serves as a habitat for macro- and microbiota within and on its surface. These biota transform living and dead organic matter, a function that forms a very close link with the biosphere. Soil's biology can influence the uptake of nutrients and water by plants (Rovira & Ridge 1983), and facilitates the life cycle of growth, s u s t e n a n c e and decay. The ensuing biodiversity provides the natural environments, such as national parks, for all to enjoy. Soil filters and modifies the surface water and groundwater; it is also the medium in which vegetation grows. Without plant cover, the soil will erode. The wind transports the fine particles, sometimes as dust storms, polluting both the air and water. Yet, despite the fact that soil is a key component of the environment, its properties and their effects are not always understood or easily articulated. Soil is used as the medium in which crops are grown, but it is also abused in m a n y ways because all types of waste are disposed of in it. In general terms, the e n v i r o n m e n t can be defined as the surroundings or conditions that influence the life and work of a p e r s o n or c o m m u n i t y (Anon., Macquarie Dictionary

1989). Ecologically, the environment is all the conditions that influence the habitat, behaviour and development of an organism. Geographically, it represents all the geographical features of an area, such as land, vegetation, water and the system connecting these. Essentially, the environment is the world in which we live.

T h e effects o f land use on soils The effects of h u m a n impact on soils are linked to the way in which the soils are used and the land managed. Traditionally, most soil scientists worked with soils to devise better methods of farm m a n a g e m e n t to increase and sustain agricultural production while maintaining environmental integrity. However, due to population growth, especially in A s i a and Africa, land clearing and m i s m a n a g e m e n t of agricultural land has resulted in a significant decline in soil quality. In more recent years, the functions of soil in the e n v i r o n m e n t h a v e b e e n r e c o g n i z e d by communities w h e n rural hinterlands have been modified and developed for urban residential areas. Soil-survey information previously used as a resource inventory for the assessment of land capability, suitability and agricultural and forest sustainability, is now also required to be i n t e r p r e t e d for u r b a n living. Soil p r o p e r t i e s such as sodicity, s h r i n k - s w e l l , salinity and acidity, which cause p r o b l e m s for rural land-users, n o w p r e s e n t the p l a n n i n g and construction constraints for engineers, environmentalists, ecologists, planners and landscape architects. This shift from rural to u r b a n soilm a n a g e m e n t and design includes building requirements such as pre- and post-construction; s e d i m e n t and erosion m a n a g e m e n t ; landfill design for waste management; specialist assessment for the t r e a t m e n t of disturbed and/or contaminated sites; and the landscaping

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. R (eds) 2006. Function of Soilsfor Human Societiesand the Environment. Geological Society, London, Special Publications, 266,141-147. 0305-8719/06/$15 9 The Geological Society of London 2006.

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of areas for the well-being of communities. Soil is also used for the foundations of buildings, as well as for the construction of dams and landfills. The change in the quality of groundwater and local waterways because of excavation and movement of the soil from construction sites alone has become a major environmental issue that has had to be addressed. Prior to any planned change in land use, soil data should be used for conservation and restoration of native woodlands (Towers et al. 2002), or to interpret the relationships existing between endangered vegetation species or communities with specific soil types (Tozer 2003). Especially in residential development, to meet the needs of the planner, architect and developer, it is not uncommon that housingestates are no longer developed on a natural landscape, but on one that is totally altered. This creates some profound soil problems, such as biologically dead stockpiled topsoil; the mixing of horizons, causing perched water tables and severe interface problems for moisture movement (infiltration); and poor revegetation response to landscaping (Handreck 1994).

Soil properties and the environment The chemistry, physics and biology of the soil are little understood by professionals in the Earth sciences or in environmental engineering. Soil scientists are specifically qualified to evaluate and interpret soils and soil-related data. Their ability to recognize or predict the behaviour of soils and the subsequent change in the environment is vital to prevent or minimize the often costly, long-term, detrimental effects of development or disturbance in any landscape. Identification and an understanding of some of the basic soil properties, such as sodicity, slaking, salinity, acidity, shrink-swell behaviour and their effects on the environment are just as important for engineers, architects, hydrologists and landscape architects, as they are for agriculturalists, foresters and horticulturalists. Soil scientists, therefore, have begun to interact with a broad range of stakeholders (Bouma 1997) and to work in a multidisciplinary team of professionals. However, because of the increasing need for land-use change, this collaboration and accessibility of soil expertise and knowledge should be more widely used in decisionmaking (Bouma 2001) if environmental problems are to be prevented when soils are disturbed.

Environmental impact of unstable soil structure Sodicity

Soil sodicity results from the level of exchangeable sodium cations in the soil. The increase in exchangeable sodium cations can result in very severe surface crusting, leading to an increase in runoff (Sumner et al. 1998), low infiltration and hydraulic conductivity. Also, these soils disperse and have very hard, dense subsoils (Hazelton & Murphy 2006). Sodicity severely decreases the structural stability of the soil to raindrop impact and wetting, resulting in cloddiness and poor workability; high soil strength; and poor aeration (Murphy 2003). The effect of these poor soil-water and soil-air relations results in restricted plant growth ( C h a n & Abbott 1995). The soil ecological systems degrade and there is a decline in soil health. This lowers crop yield and also inhibits revegetation at urban sites, increasing the risk of soil erosion from exposure to wind and rain. Dust storms are not uncommon in areas that have been degraded, for example, in the 1930s the Tennessee Valley in the USA and the Mallee country in the 1960s in Australia. Sodic soil is highly susceptible to severe gully and tunnel erosion. This soil property increases levels of stress, undermining engineering structures, which may rupture. For example, in Albury, NSW, Australia, the excavation of a drain and realignment of a creek to allow the maximum use of a site was severely affected by tunnel erosion. Prior to excavation, this area could have been readily identified as susceptible to tunnel gully erosion of the drainage channel, by conducting routine dispersion tests. Also the site should have been suspected of being sodic, as it was in a rocky area that formed part of a contact zone between shales and granite (Crouch 1977). Low wet bearing strength is another sodicity problem, because the soil is usually pliable and deforms easily under pressure when wet. It is generally unsuitable for foundations without specific engineering design, and makes site access difficult. Soil strength, settlement and swelling behaviour are properties that affect local roads and streets, because they influence the ease of excavation, grading and trafficsupporting capacity. In road-cuttings, the shear strength (the ability to support loads without shear failure) of exposed faces dictates the maximum angle achievable before road batter (side-slopes) failure. This ultimately determines the stability of slopes and the need for specific

SOILS IN THE AUSTRALIAN ENVIRONMENT revegetation techniques to avoid wind and water erosion to prevent site degradation such as rilling of road batters because of increased runoff. It decreases the sediment deposited on road surfaces, preventing road accidents and the clogging of roadside drains, which often create a flood hazard in low-lying residential areas (Hazelton 2001).

Salinity

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soil erosion potential. Slaking is the breakdown of aggregates in water to aggregates with a diameter of less than 0.25 mm. It occurs in two stages (Hillel 1980), with the final stage being a series of small explosions, each marked by the escape of a bubble of air, shattering the aggregate into fragments. One of the main reasons for slaking of dry soil aggregates on wetting with water is that the clay swells. This can only be prevented if the rate of wetting is very slow or if the clay particles are bonded together by organic matter (Emerson 1959a). Slaking decreases infiltration into the soil, and is an important property to consider for irrigation of land (Collis-George & Greene 1979). Farm dams, unless adequately treated, can fail, resulting in land inundation and loss of property.

Soil salinity differs from sodicity; it is a measure of the level of dissolved salts in the soil solution. The effects of salinity can be observed in many countries including Australia, the USA, Pakistan, Iran and Sri Lanka. There are two types of soil salinity: dryland salinity occurring on land not subject to irrigation, and irrigated salinity. Both describe areas where soils contain high levels of salts. Salinity can develop naturally or from human intervention. The human impact, for example, of clearing of trees for farmland or poor water management following overgrazing, disturbs the natural ecosystems and changes the hydrology of the landscape, so that the movement of salts into rivers and on to land is accelerated. Salinity degrades the environment. It reduces the viability of agricultural soils, restricting crop growth and resulting in loss of productivity, with severe economic implications. Salinity also poses a problem in urban areas. There is expensive, continual maintenance of roads and houses, where adverse structural impacts are due to high water tables and rising salt. Salt crystals can form and degrade bricks, resulting in wall failure. High salt levels can increase the corrosivity of soils, especially in relation to iron/steel material and concrete. This affects structures such as bridges and other infrastructure such as roads and footpaths. It is important that saline soil is identified in areas of land-use change, because of its low permeability and high water tables. These waterlogged soils, because of sparse vegetation cover, have a high erosion hazard and need specific erosion control measures (Hicks 2003). Heritage sites are also affected. For example, the ancient temples at Luxor in Egypt have been corroded by rising damp and salt. This has not only an environmental impact but also economic implications, because Egypt depends significantly on the tourist industry.

Suspension of sediment that cannot be collected in gross pollutant traps is a result of dispersion. Clay dispersion is defined by Shainberg et aL (1989) as the formation of a stable suspension of particles in water. The dispersion is the breakdown of micro-aggregates to individual sand, silt and clay particles ( C h a n & A b b o t t 1995). Suspended material - depending on its properties - may disperse, causing turbidity that can contribute to an increase of water temperature and a decline in water quality. As a result, an aquatic ecosystem can be destroyed or severely degraded. If the health of an aquatic ecosystem has been affected, the riparian (river-bank) vegetation may also degrade, and the terrestrial ecosystem will begin to fail. The soil quality will decline as the vegetation degrades, increasing the possibility of erosion because of loss of surface cover. If the suspended sediment had been excavated from a site contaminated with heavy metals, for example, the metal ions will attach themselves to the dispersed clay particles and be transported along the waterways, which could become polluted if in contact with a highly acid environment. This will also lead to a decline in water and land quality and degrade their respective aquatic and terrestrial ecosystems. In Australia, because of the number of variables that must be taken into consideration to address these environmental management problems, a catchment management approach is often used (Brock 2001).

Slaking

Shrink-swell

Another soil property, slaking, has a negative effect on the environment because of the high

Most soils shrink on drying and swell on wetting, the volume ranging from 5 to 20% (Mills et al.

Dispersion

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1980). The percentage of clay in a sample and the clay type are critical factors in determining the shrink-swell capacity of a soil. However, the valence of the adsorbed cations also has an effect. The balance of sodicity and salinity in the soil affects its shrink-swell behaviour and whether the clay fraction disperses or remains flocculated on wetting (Abbott 1991). However, soil sodicity and salinity are often confused, and it is important to clearly distinguish them, because they have the opposite effects on clay swelling and dispersion. High sodicity but low salinity, promotes swelling and dispersion (Abbott 1991). The capacity of the soil to shrink and swell can assist water entry by crack development (Blake et al. 1973); impede water movement by swelling (Loveday 1981); and restrict seed germination by crust development (Watt 1974), resulting in poor crop and vegetation response both in agriculture and at residential sites. In engineering works, the volume change that occurs in clay soils may take place at any time in the post-construction period. Due to shrink-swell in the subgrade, distortion of road pavements - as well as cracked bricks, pipes and foundation failure - may occur (Hazelton 2001). Linear shrinkage, volume expansion, and Atterberg limit tests are used to identify potential problems in urban areas, and to assess the suitability for earthworks such as earth dams and detention basins (Hicks 2003). Shrink-swell soils require special foundation design for buildings. The specific depth and volume-change relationship for a particular soil is dependent on the type of soil and level of groundwater (Cheng & Evett 2005). For dams, the wall batters (sideslopes) must be flatter and the wall must not dry out (Crouch et at. 2003). Sometimes the soil swells with changes in the subsoil moisture content. When this occurs, it can lead to problems with underground services, where changes in soil moisture can be one of the contributing factors to water loss from leaking pipes and ultimately water-main failure (Palmer & Hazelton 1994). In the field, soils with a high shrink-swell capacity are easily identified by an uneven soil surface caused by soil movement. 'Gilgai' soils are black, grey and brown cracking clays found generally in north-western NSW, Australia. These soils have various types of surface humps and hollows formed by massive shrinking and swelling. Special optic-fibre cables for telecommunications had to be manufactured and used in these areas to prevent failure.

Acidity

In many countries, including Australia, there are naturally occurring acid soils. Soil acidity can also be the result of human impact, because fertilizers often make soils more acid. Due to differences in the chemical composition of parent materials, soils will become acidic after different lengths of time. According to reports from the Commonwealth Scientific and Industrial Research Organisation (CSIRO), in Australia 33 million hectares of farming land have highly acidic soils. Acidity occurs particularly in the productive higher-rainfall areas in south-eastern and south-western parts of Australia and also in Queensland. The effects of pH on the availability of toxic elements such as manganese and aluminium, and on nutrient elements, can interfere with the normal growth of plants. Acid-rain deposition in the UK and Europe generally leads to acidification that can permanently reduce the productive capacity of agricultural and forest land because roots and special types of fungi (mycorrhizae) are affected. Water resources can become acidified from the waters draining from the acid soil. Acidification may lead to decreased plant growth or changes in plant communities that may shift toward acid-tolerant species, resulting in a change in biodiversity. Another acid soil problem is the presence of naturally occurring acid sulphate soils, which occur in low-lying coastal and estuarine areas in many countries, including Australia, South-East Asia, western Finland and the USA. There are an estimated 20 million hectares worldwide (Dent 2000). Acid sulphate soils contain large amounts of sulphides (mainly pyrite) (Hashimoto & Roy 1996). They are often known as potential acid sulphate soils (White & Melville 1993). The potential environmental hazards of acid sulphate soil materials are directly related to the net acidity and the rate that actual acidity is released into the environment (Ward et al. 2004). On the south-east coast of Australia, for example, coastal wetlands and swamps have been drained to increase the area of grazing land. Oxidation of these soils through dredging of clogged drains or excavation for construction works results in a rapid pH change from neutral to extremely acid (<4.0), with the production of sulphuric acid. Export of acidity from drained coastal floodplains to estuaries is of major ecological concern (White et al. 1997). The acid runoff water and subsequent mobilized aluminium ions from the clay are released into creeks, resulting in fish kills (White et al. 1995). Fish breeding-grounds can decline or be lost,

SOILS IN THE AUSTRALIAN ENVIRONMENT and a major change in the ecology of a stream can occur. However, the acid environment is toxic not only to marine invertebrates and vertebrates but also to vegetation. As the vegetation dies, scalding occurs (Rosicky et al. 2000), and the denuded area becomes more susceptible to soil erosion. If the erodible sediment releases toxic metal (Sundstrom et al. 2002), which is then washed into streams, it has a deleterious effect on the water quality, and aquatic ecosystems are put under further stress. Acid sulphate soil is also an environmental problem for urban planners who want to develop housing-estates on flat coastal land that has become too valuable to be used for farming. Development results in exposure and oxidation of these soils during excavation for house foundations, and for infrastructure such as roads and services. In the eastern United States, Delaware, Maryland and Virginia have active acid sulphate soils resulting from the construction of roads, canals, housing developments, shopping centres, landfills and surface mines (Fanning et al. 2004).

Soil, h u m a n health and the environment Soil contamination and disturbance are a concern for human health and the environment. Poor understanding of the chemical properties of the soil can result in environmental problems, leading to ill health and also death. Most examples of ill health associated with the soil are caused by toxic concentrations of elements such as lead and arsenic in food and water. Toxic concentrations may reflect the natural condition of the soil (weathered from, for example, arsenic-bearing rocks) or may be due to human activity where the area has become contaminated (Oliver 1997). Soil that is contaminated is not just limited to old industrial sites, and can be found throughout urban and regional areas where previously there were farms. For example, land-use practices, such as the use of copper and mercury in fungicides, may contribute to the accumulation of elements that are toxic to plant growth. It is difficult to assign an absolute value indicating toxicity from heavymetal contamination, because it varies according to site and case-specific variables and is dependent on many factors. Individual metals respond differently to changes in soil conditions, and the bioavailable fraction (the mobile fraction able to be absorbed by the human body or taken up by plants or organisms) will also be different for each metal (Bernard 1997). Many soil properties, such as soil pH, clay content, organic matter and the cation-exchange capacity, influence their retention and mobility

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(Helmke & Naidu 1996). To predict the environmental impact of metals in the soil environment, it is essential to understand the processes affecting their behaviour. Acid sulphate soil also poses a health hazard to humans, animals and the environment because mosquitoes that carry many diseases, such as malaria, and, in Australia, Ross River fever, are attracted to the acid drainage water. Mosquitoes multiply in areas of acid sulphate soil drainage, because of a breakdown in the aquatic ecosystem, especially as their natural predators, fish, have been displaced by the acid conditions (Green 1993). There is also a relationship between disease and regional types of soils (Oliver 1997). A lack of understanding of soil properties can have a significant effect on humans and their interaction with the environment. For example, the local nature of toxicity and deficiency in the epidemiological literature (Lauwerys et al. 1984), found that the average cadmium in a sample of autopsied bodies was approximately twice as great in Li6ge as in Brussels. Little is known about whether the people had suffered from the subclinical effects of this toxicity. But it is important to know why there is excess or deficiency, and whether it relates to the soil in which the food was grown. To determine the risk to human health and the environment, soils that are considered to be potentially contaminated must be assessed and classified. Determination of this risk is on the basis of exposure as well as the toxicity and availability of the contaminants present. A consistent approach using specific guidelines and legislation is required for the assessment and management of these soils.

Conclusions 'Soil helps us to understand the self-limitations of life, its cycles of death and rebirth, and the interdependence of all species' (Kirschenmann 1997). If society is to achieve a better quality of life and a sustainable environment, education at all levels is required so that the links between soil-water-air are understood. Only through education will the need be perceived for substantial legislation for the protection of soil and that of the environment. Because of the pressure of population growth on landscapes, legislation must provide a framework to control all aspects of land-use change. The process of environmental assessment and planning must take into account the impact of an activity on the total environment, to ensure its protection and to maintain biodiversity in which ecological

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sustainability is an integral part of the function and health of the soil. Soil is the common ground on which we all stand (Kirschenmann 1997). Soil is the key component in environmental management for sustainability. Therefore, 'people should marvel at its workings, its chemistry and its bugs and will be pleased to say thank you to the soil, the basic commodity on which life depends' (Heiney 1997). References ABBOTT,T.S. 1991. Swell/shrink phenomena and structural stability of clay soils. In: HAZELTON,P.A. & BANNERMAN, S. (eds) Soil Technology-Applied Soil Science. A Course of Lectures. Australian Society of Soil Science, NSW Branch and Soil Science, School of Crop Sciences, University of Sydney, Sydney, Australia, 41-53. ANON. 1989. Macquarie Dictionary, Second Edn, Jacaranda Press, Victoria, Australia. BERNARD, A.M. 1997. Effects of heavy metals in the environment on human health. In: PROSr, R. (ed.) Contaminated Soils: Third International Conference on the Biogeochemistry of Trace Elements. Paris, 15-19 May 1995, Colloque, INRA editions, Paris, France, 21-34. BLAKE, G., SCHLICHTING, E. & ZIMMERMANN, P.V. 1973. Water recharge in a soil with shrinkage cracks. Proceedings of the Soil Science Society of America, 37, 669-672. BOUMA, J. 1997. The role of quantitive approaches in soil science when interacting with stakeholders. Geoderma, 78, 1-12. BOUMA, J. 2001. The role of soil data in the land-use negotiation process. Soil Use and Management, 17, 1-6. BROCK, P.M. 2001. Catchment management as a context for soil management. In: CAttLE, S.R. & GEORGE, B.H. (eds) Describing, Analysing and Managing Our Soil. Proceedings of the DAMOS 99 Workshop held at The University of Sydney, 22-26 November 1999. Published Jointly by University of Sydney and Australian Society of Soil Science NSW Branch, 173-184. CHAN, K.Y. & ABBOTT T.S. 1995. Physico-chemical properties controlling soil structure. In: HAZELTON, P.A. & KOPPI, A.J. (eds) Soil Technology-Applied Soil Science. A Course of Lectures, 3rd edition, Published Jointly by Australian Society of Soil Science and Department of Agricultural Chemistry and Soil Science, University of Sydney, 61-75. CHENG LUI & EVITI', J.B. 2005. Soils and Foundations. Prentice Hall, Pearson Education, South Asia. COLLIS-GEORGE, N. & GREENE, R.S.B. 1979. The effect of aggregate size on the infiltration behaviour of a slaking soil and its relevance to ponded irrigation. Australian Journal of Soil Research, 17, 65-73. CROUCH, R.J. 1977. Tunnel-gully erosion and urban development: a case study. In: SHERARD, J.L. & DECKER, R.S. (eds) Dispersive Clays, Related Piping, and Erosion in Geotechnical Projects,

ASTM STP 63, American Society for Testing and Materials. CROUCH, R.J., REYNOLDS,K.C., HICKS R.W & GREENTREE, D.A. 2003. Soils and their use for earthworks. In: CHARMAN,EE.V. & MURPHY,B.W. Soils: Their Properties and Management, Sydney University Press, Sydney. DENT, D.L. 2000. An International perspective. In: AHERN, C.R., HEY, K.M., WATLING, K.M. & ELDERSHAW,V.J., "Acid Sulfate Soil: Environmental Issues, Assessment and Management' Technical Papers. Queensland Department of natural resources, Brisbane, 11-12. EMERSON, W.W. 1959a. The structure of soil crumbs. Journal o f Soil Science, 10, 235-244. FANNING, D.S., COPPOCK, C., ORNDORFF, Z.W., DANIELS,W.L. & RABENHORST,M.C. 2004. Upland active acid sulfate soils from the construction of new Stafford County, Virginia, USA Airport. Australian Journal of Soil Research, 42, 527-536. GREEN, D. 1993. Rivers of death. Fishing World, April, 38-41. HANDRECK, K. 1994. Home gardens and landscaping. In: PETERSON, D.R., WEATHERLEY,A.J. & WHITE, R.E. (eds) Extended Abstracts of the Workshop 'Soil in the City', University of Melbourne, Parkville, Victoria. HASH1MOTO,T.R. & ROY, RS. 1996. Strategic controls on the distribution of potential acid sulfate soils in the coastal lowlands of south eastern Australia. In: Proceedings, 2nd National Conference on Acid Sulfate Soils, Robert Smith and Associates, Acid Sulfate Soils management Advisory Committee (ASSMAC), Coifs Harbour, NSW, 41-42. HAZELTON,EA. 2001. Soil properties and engineering. In: CATTLE, S.R. & GEORGE, B.H. (eds) Describing, Analysing and Managing Our Soil. First edition. Proceedings of the DAMOS 99 Workshop held at The University of Sydney, 22-26 November 1999. Published Jointly by University of Sydney and Australian Soil Science Society, 277-283. HAZELTON, EA. & MURPHY, B.W. (eds) 2006. Interpreting Soil Test Results: What Do All the Numbers Mean? CSIRO Press, Melbourne. HEINEY, P. 1997. Weekend. The Times, London, 23 August 1997. HELMKE,EA. & NAIDU,R. 1996. Fate of contaminants in the soil environment: metal contaminants. In: NAIDU, R. ET AL. (eds) Contaminants and the Soil Environment in the Australasia-Pacific Region. Kluwer Academic Publishers, Dordrecht, The Netherlands, 69-93. HICKS, R. 2003. Soils and urban land use. In: CHAIRMAN, RE.V. & MURPHY, B.W. (eds) Soils: Their Properties and Management. Sydney University Press, Sydney. HILLEL, D. 1980. Fundamentals of Soil Physics. Academic Press, London. KIRSCHENMANN,E 1997. On becoming lovers of soil. World Wide Web Address: http:www.igc.org/ wsaala/lovers.html. LAUWERYS,R., BUCHET,J.P. ETAL. 1984. Potential risk of cadmium for the general population: update of the Leige study. In: Proceedings of the 4th

SOILS IN THE A U S T R A L I A N E N V I R O N M E N T International Cadmium Conference, Munich, Cadmium Association, London, 113-114. LOVEDAY,J. 1981. Soil Management and Amelioration. In: ABBOTT, T.S., HAWKINS, C.A. & SEARL, P.G., National Soils Conference 1980 Review Papers. Australian Society of Soil Science, Sydney. MmLS, J.J., MURPHY,B.W. & WICKHAM, H.G. 1980. A study of three simple laboratory tests for the prediction of soil shrink-swell behaviour. Journal of Soil Conservation NSW, 36, 77-82. MURPHY, B.W. 2003. The nature of soil. In: CHARMAN, RE.V. & MURPHY,B.W. (eds) Soils: Their Properties and Management, Sydney University Press, Sydney. OLIVER, M.A. 1997. Soil and human health: a review. European Journal of Soil Science, 48, 573-592. PALMER,B. & HAZELTON,P. 1994. Sydney Water Board St George water area main failure analysis. In: PETERSON,D.R., WEATHERLEY,A.J. & WHITE, R.E. (eds) Extended Abstracts of the Workshop "Soil in the City' The University of Melbourne, Parkville, Victoria, 9-10. RIMMER, D. 1998. New Scientist, London, 14 November 1998. 1611 (2160), SS1-SS4. ROSICKY,M.A., SULLIVAN,L.A. & SLAVICH,RG. 2000. Pyrite concentration and surface reformation in and around acid sulfate soil scalds on the NSW coast. In: MACDONALD, B.C.T., KEENE, A.F., CARLIN, G. & SULLIVAN,A. (eds) Proceedings, 5th International Acid Sulfate Soil Conference. Tweed Heads, NSW. ROVIRA, A.D. & RIDGE, E.H. 1983. Soil-borne root diseases and wheat. In: Soils: an Australian Viewpoint. CSIRO Division of Soils, Melbourne, 722-754. SHAINBERG, L., SUMNER,M.E., MILLER, W.P., FARINA, M.RW., PAVAN, M.A. & FLY, M.V. 1989. Use of gypsum on soils: a review. In: STEWART,B.A. (ed.) Advances in Soil Science, 9 Springer-Verlag, New York, 1-111. SUMNER, M.E., MILLER, W.P., KOOKANA, R.S. & HAZELTON, RA. 1998. Sodicity, dispersion and environmental quality. In: SUMNER, M. & NAIDU, R. (eds) Sodic Soils - Distribution, Properties, Management, and Environmental Consequences. Oxford University Press, New York.

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SUNDSTROM, R., ANSTROM, M. & OSTERHOLM,P. 2002. Comparison of the metal content in acid sulfate soil runoff and industrial effluents in Finland. Environmental Science and Technology, 36, 4269-4272. TOWERS,W, HESTER, A.J., STONE, M.D. & GRAY, H. 2002. The use of soil data in natural heritage planning and management. Soil Use and Management, 18, 26-33. TOZER, M. 2003. The native vegetation of the Cumberland Plain, Western Sydney: systematic classification and field identification of communities. Cunninghamia, 8, 1-76. WATT,L.A. 1974. The effect of water potential on the germination behaviour of several warm season grass species, with special reference to cracking black clay soil. Journal of Soil Conservation, NSW, 3tl, 28-41. WARD, N.J., SULLIVAN,L.A. & BUSH, R.T. 2004. Soil pH, oxygen availability, and the rate of sulphide oxidation in acid sulfate soil materials: implications for environmental hazard assessment. Australian Journal of Soil Research, 42, 509-514. WHITE, I. ~; MELVILLE, M.D. 1993. Treatment and containment of potential acid sulphate soils: formation, distribution, properties and management of potential acid sulfate soils. CSIRO Centre for Environmental Mechanics. Technical Reports 53, Canberra. WHITE, I., MELVILLE, M.D., LIN, C., VAN OPLOO, P., SAMMUT,J. & WILSON,B.R 1995. Identification and management of acid sulphate soils. In: HAZELTON, RA. & KoPPI, A.J. (eds) Soil Technology-Applied Soil Science, A Course of Lectures. 3rd edition. Published Jointly by the Australian Society of Soil Science and the Department of Agricultural Chemistry and Soil Science, University of Sydney, Australia, 463-497. WHITE, I., MELVILLE,M.D., WILSON,B.R & SAMMUT,J. 1997. Reducing acidic discharges from coastal wetlands in eastern Australia. Wetlands Ecology and Management, 5, 55-72. YAALON, D.H. & ARNOLD, R.W. 2000. Attitudes towards soils and their societal relevance: then and now. Soil Science, 165, 5-12.

P o l i c i e s for a s u s t a i n a b l e u s e o f soil r e s o u r c e s LUCA MONTANARELLA

European Commission, Joint Research Centre, Institute for Environment and Sustainability, TP 280, 1-21020 Ispra (VA), Italy (e-maiL luca. montanarella@jrc, i 0

Abstract: Soil protection policies have been developed at National level in a number of countries. The US Soil Conservation Act, the German Federal Soil Protection Act, and the Soil Action Plan for England are just some examples of such policies. International policies for a sustainable use of soil resources have been developed both at regional level and at global level. The Alpine Convention and the upcoming EU Thematic Strategy for Soil Protection are two examples of regional policies addressing soil protection. No specific binding global policy exists for soil protection. Nevertheless, a number of environmental protection policies address soil-protection aspects, like the climate-change convention, the biodiversity convention and the desertification convention. Further development of the EU Thematic Strategy for Soil Protection will make it possible to explore new methods of international cooperation in this field.

The recognition that soils are a limited resource and therefore need to be protected for the benefit of future generations was always at the top of the preoccupations of rural populations, particularly farmers. The history of agriculture from the early Neolithic to modern times has been constantly characterized by the development of improved agricultural practices in order to preserve the precious soil resources and increase their fertility (Demoule & Perles 1993; Lowdermilk 1994; Mazoyer & Roudart 1997; Cauvin 2000; Price 2000). Only during the past 100 years has a dramatic change of population distribution taken place, leading to an extensive urbanization of large portions of the rural population. In Europe, the current percentage of population active in agriculture varies from c. 5 to 10% of the total, with some countries having less then 3% farmers. Consequently, a large proporiton of today's population has completely lost the direct linkage to soils and their sustainable use. Soil protection, as a normal day-to-day task of a large part of the population, has been essentially left to a small number of farmers, often lacking the means for implementing effective soilprotection measures. Additionally, large parts of previously cultivated land have been abandoned, mainly in the more marginal areas, leading to rapid soil degradation. Typical examples of such land abandonment can be found in many areas around the Mediterranean basin, particularly in ancient terraced landscapes (Lasanta et al. 2001). The first assessment at global level of the status of soil degradation (Oldeman et al. 1991) has reported about the extensive processes of

soil degradation occurring in large areas of the world. These phenomena have triggered the first attempts to develop soil-protection policies able to reverse the negative trend. Typically, such policies have been developed at national level, following the perception by the majority of the (urban) population that soil degradation was taking place in the countryside.

National policies Several national soil-protection policies have been developed over the past few years in several countries of the world (Hannam & Boer 2002). Certainly, the most remarkable for its impact, which had extended to successive legislative initiatives in other countries, has been the US Soil Conservation Act of 1935 and its successive amendments. In the following sections we will briefly review the US Soil Conservation Act; the German federal law for soil protection; and the soil-action plan for England, as examples of national soil-protection policies.

The US Soil Conservation Act The US Soil Conservation and Domestic Allotment Act of 1935 (http://ipl.unm.edu/cwl/ fedbook/soilcons.html) has undoubtedly become one of the pieces of national soil-protection legislation which has had a deep impact on soilprotection strategies and in general on the development of soil conservation both in the US and at a global scale (Helms 1998). The act was conceived as the response of legislators to the serious threat posed by soil

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. R (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266,149-158. 0305-8719/06/$15 9 The Geological Society of London 2006.

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degradation in the Midwest of the United States of America, the renowned 'dust bowl'. Chronicles of that time report that the US Congress was finally convinced of the urgent need for such an act by the huge dust clouds visible over Washington out of the windows of the Congress buildings, during the hearings and debates around the soil-conservation act. Indeed, the history of the soil-conservation act can tell us many lessons about the factors influencing the creation of an effective soil-protection policy (Arnold 2004). There needs, first of all, to be a perceived requirement to conserve and protect soil and land resources. This need must be perceived by the majority of the population in order to be appealing to legislators. The dramatic drought and erosion events in the US Great Plains: the 'dust bowl', spreading clouds of eroded soil over large parts of the urbanized east coast, was immediately perceived as a problem. Of course, such events did not develop from one day to the other, but were the consequence of many years of land degradation combined with serious social and economic problems. The soil-conservation act was conceived in the context of the dramatic economic depression of the 1930s. It was therefore aiming primarily to achieve soil conservation, but was also highly attentive to the need for the creation of new job opportunities and economic development. One of the main pillars of the act has been the creation of the Soil Conservation Service, an agency within the Department of Agriculture, to deal with soil-erosion concerns on privately owned rural land throughout the US. This agency rapidly developed as a leading institution, both in the US and globally, in soil conservation, soil survey and soil science in general. The creation of such an agency, with enough staff and resources to have a considerable impact on the way that rural land was managed by local communities was the key to the success of the Soil Conservation Act. The Soil Conservation Service, as formulated and modified over seven decades, has served the purpose of providing technical support for the control of erosion through the strong partnership with Soil and Water Conservation Districts. The Districts are units of state-level government charged with helping rural landowners care for their land and water resources. Through cooperative agreements, the expertise and personnel of the US government were provided to assist at the local county-level. Perhaps one of the more notable achievements was the demonstration that a voluntary

participatory programme of soil conservation was possible within the diversity of American culture. It was done with numerous partnering arrangements that changed in response to new challenges and mandates. As monitoring techniques improved, it was possible to design and build reasonable baseline data from which trends of resource conditions could be measured more accurately (Nusser & Goebel 1997). The physical quantities of conservation practices and measures could be determined; the number of farms served; the acres of land treated, the tons of soil loss occurring and that saved; and so on. The statistics are impressive; however, the network of databases, the information technology, the people-to-people interactions, and an increasing stewardship of natural resources are even more impressive. Private landowners in rural areas are able to ask for and receive attention concerning use and management of their soil resources. Many practices or measures are eligible for some degree of financial assistance that is determined at the local level but financed at the federal level. Soil maps, farm plans, and followthrough activities became available to those willing to commit to a resource-management plan for their land. The range of conservation activities is large - tree planting, drainage guidance, crop rotation advice, earth construction specifications, revegetation assistance, wildlife habitat help, grazing plans and livestock watering systems, woodlot evaluation, waste-management systems, etc. Urban dwellers have benefited from a responsive agricultural system and cheaper food; a cleaner rural environment; and improved watershed management and assistance when natural stresses occur. And everyone has an opportunity to better understand and contribute to the stewardship of the environment - whether rural or urban. The citizens benefit by having technology information systems that provide up-todate accurate information about the status and trends of land conditions, mainly rural. Agricultural production was stabilized in large measure because causes of loss of production were addressed, such as erosion, crop rotations, liming, water management, and later soil fertility and crop advice services. This was always a partnership of federal, state and local governments combining to give attention to individual landowners and users. During wartime, production could be, and was, significantly increased in the short run, then reduced later where most appropriate for the care of the land. Agriculture is vital to the US economy, thus improved knowledge of the capabilities of

SOIL PROTECTION POLICIES land help to guide policies to protect, conserve and wisely use those resources. Conservation has contributed to cleaner water and air, reduced sediment damage, rural communities are maintained, cheaper food is available, and many areas of environmental health concern are receiving attention. A major monitoring system, the National Resource Inventory, is now in place to provide more information to guide future policies. The US Soil Conservation Act is certainly a good example of a success story in soil conservation policy-making. It continues to be a major component of the US agricultural policy through the subsequent amendments by the US Farm Bill. The 1996 Farm Bill repealed certain elements of the soil-conservation act and amended the act to provide financial and technical assistance through the new environmental-quality incentives programme described in the Food Security Act of 1985. The Soil Conservation Service, established by this act, was repealed by the Department of Agriculture Reorganization Act of 1994 and replaced in that statute by the Natural Resource Conservation Service (NRCS). The major limitation of the act remains its narrow view on soil conservation in relation to the single threat that it was aiming to address: soil erosion. Modern views on soil conservation are now considering soils in their multifunctionality as environmental compartments, not only delivering food and fibre production, but also delivering a number of additional environmental services, like clean drinking-water, biodiversity, cultural heritage, etc. In this sense, the US Soil Conservation Act has been somehow limited in scope Several lessons can be learned from the US experience in soil conservation (Montanarella et al. 2004). Soil conservation can best be achieved through participatory approaches involving all actual stakeholders. Sound conservation strategies require a solid scientific and technical background. The substantial achievements of the US Soil Conservation Service in reversing the negative trend and actually improving soil conditions in the US demonstrate that effective soil conservation is achievable through voluntary approaches. But goodwill is not sufficient, there needs to be also a background of scientific knowledge; an infrastructure for the transfer of that knowledge to the stakeholders; and, last but not least, substantial economic resources to sustain the soil-conservation efforts in the long term.

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The G e r m a n Federal Soil Protection A c t A more recent example of national soil protection legislation is the Federal Soil Protection Act of G e r m a n y of 1998. Within the E U member states, it is probably the most comprehensive piece of legislation dealing with soil protection, and certainly was one of the elements triggering the new developments towards a thematic strategy for soil protection of the European Union. The main difference to the previously outlined US Soil Conservation Act is the incorporation into legislation of the concept of soil multifunctionality. The purpose of this act is therefore to protect or restore the multiple functions of the soil on a permanent sustainable basis. This shall include prevention of harmful changes (to the soil), rehabilitation of the soil, of contaminated sites and of waters contaminated by such sites, and 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. The Act also attempts to define soil for the purpose of the legislation. Soil within the meaning of this Act is the upper layer of the Earth's crust, as far as this layer fulfils the soil functions, and including its liquid components (soil solution) and gaseous components (soil air), except groundwater and beds of bodies of water. For the purpose of this definition, it becomes crucial to also define the multiple functions that the soil performs. The act recognizes three groups of functions: (1) Natural functions: (a) as a basis for life and a habitat for people, animals, plants and soil organisms; (b) as part of natural systems, especially by means of its water and nutrient cycles; (c) as a medium for decomposition, balance and restoration as a result of its filtering, buffering and substanceconverting properties, and especially groundwater protection. (2) Functions as an archive of natural and cultural history; (3) Functions useful to humans as: (a) a medium that holds deposits of raw materials; (b) land for settlement and recreation; (c) land for agricultural and silvicultural use; (d) land for other economic and public uses, for transport, and for supply, provision and disposal.

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The recognition that soils perform these functions reflects the aim of the policy to be an act for the protection of soil as an environmental compartment as such, analogously with air and water protection. This is a radical shift from previous policies mostly aiming at protecting soils for single functions, like agricultural production. Another basic element of the act is the introduction into soil protection of the concept of the liability of the landowner for damage caused to the soil. This also opens the way to a number of implications in assessing off-site effects of soil degradation, that in several cases can be very substantial, like, for example, groundwater contamination, mudflows, flooding, etc. The principle of the 'polluter pays' is a key element of the act. This aspect of obligations to prevent hazards is of course radically different from the previously described US Soil Conservation Act, which was putting the major emphasis on developing, together with stakeholders, a positive loop of remediation through incentives, technical support, education, etc. In this sense, the G e r m a n act has more of a prescriptive nature, including a large part dealing with thresholds, obligations, values, requirements, orders, investigations, and, finally, value compensation and fines. This is a vocabulary that is totally absent in the previously described US Soil Conservation Act. While the act recognizes soil multifunctionality, it is much more limited in considering the various threats to these functions. Most of the act is focused on the issue of soil contamination, particularly on the problem of contaminated sites. Some provisions are included for the problem of soil sealing (Art. 5) and good agricultural practices (Art. 17), including soil erosion, compaction, soil biodiversity and soil organic matter. Nevertheless, the act remains essentially an act aiming at the remediation of contaminated sites.

The SoiI Action Plan for England A second example of European national soil protection policy is the recently presented Soil Action Plan for England ( D E F R A 2004). The First Soil Action Plan for England commits the government and partners to actions which will improve the protection and management of soils within a whole range of land uses. The Action Plan builds on the earlier Draft Soil Strategy published as a consultation paper in 2001. It is complemented by an Environment Agency report on the State of Soils in England and Wales. The plan sets out an ambitious programme of

work for the next three years, to help move towards a clearly stated vision for the nation's soils. The actions are often only the first, important step in the process. The aim of this first plan is to achieve as much as possible by properly embedding soils into ongoing work, to gather the evidence; and to build consensus and partnerships with others in government and outside, in order to provide the foundation for future action. The Action Plan contains 52 actions on issues ranging from soil management on farms to soils in the planning system; soils and biodiversity; contamination of soils; and the role of soils in conserving cultural heritage and landscape. All of the actions take steps towards more sustainable soil use and protection. The plan takes a pragmatic approach in identifying actions that cover all aspects of soil protection, ranging from erosion control, to the protection of biodiversity and cultural heritage. The actions are scheduled with well-defined milestones and expected results over the period 2004-2006, and are designed for rapid implementation and feasibility in a short timeframe. They are based on the recognition that soil protection is an issue that cuts across many policy areas where already existing legislative instruments can already be used effectively for improving soil conditions in England. By involving the many actors actively operating in the field of soil conservation, a number of synergies can be found that make the soil-action plan for England one of the first examples of streamlining the different processes and policies towards a common goal of sustainable use of soil resources.

International policies Soil degradation is not a strictly local (national) problem, but it has many aspects that make it of relevance for international policies as well. Starting from the issue of the various off-site effects of soil degradation, like erosion, contamination, floods and landslides, which can easily affect bordering countries, up to the truly global implications of soil degradation, like the interlinkages with global climate change through the effects on the levels of greenhouse gases in the atmosphere. Besides the direct linkage to trans-boundary off-site effects of soil degradation, there are also very relevant indirect aspects of soil degradation that justify the development of international soil-protection policies. The most relevant are surely the socio-economic aspects of soil degradation, particularly in developing countries, that are often linked with malnutrition and poverty

SOIL PROTECTION POLICIES and trigger major migrations of large parts of the rural population. Indirectly, soil degradation is relevant to WTO negotiations, with large areas of the world suffering from soil degradation as an indirect consequence of agricultural market distortions (Lahmar et al. 2002). International policies for a sustainable use of soil resources have been developed both at regional level and at global level. Some examples are reported below.

Trans-national (regional) policies The recognition that soil protection has a transnational dimension has initiated several attempts to develop an international legal framework for soil protection (Hannam & Boer 2002). Two European examples are reviewed briefly below, the Alpine Convention and the EU Thematic Strategy for Soil Protection.

Alpine Convention Soils play a central role in the life and development of mountainous lands. They provide a vital substratum for humans, animals, plants and micro-organisms. They are a key component of the mountainous ecosystems, in particular for water and nutrient cycling. In addition, they constitute a major gene reservoir and are a place of major regulation and transformation

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processes. As a central element of the mountainous landscape, soils have always occupied a central position in the cultural and economic life of human communities. In particular, they constitute the basis for agriculture, including forestry and livestock breeding, and a major environment for recreational and tourist activities. It is therefore essential to promote the conservation and sustainable use of soil resources, taking into account the specific sensitivity of mountainous soils to degradation and alteration processes. The Alps represent one of the most sensitive ecosystems in Europe, because of aggressive development in the recent past; huge numbers of tourists; and severe environmental damage. Based on the fact 'that the Alps are one of Europe's largest inter-related natural regions and that, due to their specific and diverse nature, culture and history, they represent an excellent location for habitation and the execution of economic, cultural and recreational activities situated in the heart of Europe', an Alpine convention was established following the results of the first Alpine Conference of the Ministers of the Environment held in Berchtesgaden in 1989. Members of the Alpine Convention are Germany (D), France (F), Italy (I), Slovenia (SLO), Liechtenstein (FL), Austria (A), Switzerland (CH), Monaco (MC) and the European Community (Fig. 1).

Fig. 1. Geographical coverage of the Alpine Convention.

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The Alpine Convention stipulates in Article 2, Paragraph 1, that all parties shall, with regard to the principle of precaution and co-operation and the 'polluter-pays' principle, guarantee a global policy designed to preserve and protect the Alps, with careful consideration being given to the interests of all Alpine states, their Alpine regions and the European Community, by making prudent and sustainable usage of the resources available. Trans-national cooperation in the Alpine region shall be increased and extended to include additional territorial regions and special areas. Article 2 also stipulates that 'in order to attain the objective specified under paragraph 1', the parties to the agreement shall take appropriate measures in particular in the following fields: population and culture, regional planning, air cleanness, soil conservation, water economy, protection of nature and maintenance of the landscape, mountain agriculture, mountain forests, tourism and recreation, traffic, energy and waste management. Furthermore, the contracting parties shall also ensure that the public is regularly kept informed in an appropriate manner about the results of research, monitoring and actions taken. This means that the dissemination of information should not be restricted to selected individuals, public organizations and research institutions, but that the media will be used to ensure that a wider section of the public, both within the Alpine areas and outside, will be appropriately informed. The adoption of the protocol on soil protection (Bled, 20 October 1998) sets new goals in the framework of the Alpine Convention. The protocol is based on an ecosystem perspective and recognizes the Alps region for its ecological diversity and highly sensitive ecological systems whose functional capacity must be preserved. The principal objective of the protocol is to reduce the quantitative and qualitative damage to soil through the use of appropriate agricultural and forestry land-use methods which do not harm the soil. It promotes minimal interference with soil, soil-erosion control, restrictions on the sealing of soil, and soil rehabilitation. The protocol sets out the functions of soil, including natural functions, cultural functions and land-use functions, emphasizing the need to be safeguarded and preserved to maintain an ecological balance in the region, and soil diversity, for future generations (Article 1). In particular, the protocol has a number of specific obligations for the parties to observe with regard to:

(1) legal and administrative measures to protect the soil and apply the precautionary principle (Article 2); (2) consider the objectives of the protocol in other policies (e.g. forestry, agriculture, nature protection) (Article 3); (3) coordination and cooperation between institutions and territorial authorities to develop synergies for soil protection (Article 4); (4) a commitment to support international cooperation among institutions on soil research (Article 5). Chapter II of the protocol contains many specific measures for contracting parties to adopt to use soil sparingly and cautiously (Article 7), preserving wetlands and moorlands (Article 9); delineating areas for rehabilitation and special land management (Article 10), and areas affected by soil erosion (Article 11); protecting soil for agriculture and forestry (Article 12), against the impacts of tourism (Article 14); protecting soil against pollutants (Article 15); and recognizing the ecological characteristics and values of soils (Article 16). Chapter III sets out specific obligations in relation to the maintenance of ecological harmony and to preserve the environment, and to make soil information available to the public.

EU Thematic Strategy for Soil Protection Soil protection has never ranked high among the priorities for environmental protection in Europe. Soils are commonly not well known by European citizens, particularly since only a small fraction of the European population is currently living in rural areas and having direct contact with soils. The majority of the urban population in Europe have only scant understanding of the features and functions of soils. The most common perception is that soils are good dumping sites for all kinds of wastes, and that soil can be quite useful as a surface for building houses and infrastructure. Only during the last two to three years has the need for a coherent approach to soil protection come on to the political agenda in Europe and was therefore introduced as one of the thematic strategies (EC 2002) to be developed within the European Community's 6th Environment Action Programme (6th EAP). The rationale behind the development of a coherent approach to soil protection is based on the recognition of the multifunctionality of soils. Soils are no

SOIL PROTECTION POLICIES longer considered only as dumping sites; construction surfaces or means for production (agriculture), but also as a fundamental environmental compartment performing vital ecological, social and economic services for E u r o p e a n citizens: filtering and buffering contaminants to allow us to have clean drinking-water, a pool of biodiversity, a source of raw materials, a sink for a t m o s p h e r i c carbon dioxide, an archive of cultural heritage, etc. These functions are now recognized as having importance equal to the traditional functions commonly attributed to soils: production of food, fibre and wood (agriculture and forestry) and surfaces for housing and infrastructure (spatial development). In order to develop a soil-protection policy, it is important to identify that soils have distinctive features that make them quite different from the other environmental compartments, like air and water. Soils are first of all highly diverse both in space and over time (Bullock et al. 1999). Prop-

155

erties can be completely different for two soils only a few metres apart. The development of a c o m m o n soil map of E u r o p e has helped in describing the very high spatial variability of soils across the E u r o p e a n continent (Fig. 2). Soils are not static, but develop over time. The time-scale for these changes is usually very long (hundreds of years). Therefore, for policymaking purposes, we will consider soils as essentially a n o n - r e n e w a b l e resource. The high variability of soils implies that any soil-protection strategy needs to have a strong local element built in. It is at the local level that we can act in specific ways that are appropriate to the features of these particular soil types. This of course brings out the important distinction that needs to be made in identifying the actors that must develop and i m p l e m e n t soil-protection measures. It should be recognized that while there are important local elements that need to be built into soil-protection strategy, there are

Fig. 2. Soil map derived from the Soil Geographical Database of Europe at a scale of 1:1 000 000.

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L. MONTANARELLA

nevertheless, clearly identified off-site effects of soil degradation that justify an European or even global approach to soil protection. Erosion, decline of organic matter, soil contamination, soil compaction, soil sealing and loss of biodiversity have very important off-site consequences, like silting of hydropower stations, increase of atmospheric carbon dioxide, contamination of drinking and bathing waters, contamination of food, increased frequency of flooding and landslides, etc. All these off-site effects seriously threaten human health and have substantial economic implications. A key feature for developing a soil-protection strategy is the recognition of the implications linked with the fact that soils in Europe commonly have property rights. The majority of soils are on private property, and this brings up a series of environmental liability implications. A coherent approach to soil protection in Europe is just beginning. The goals set out in the communication 'Towards a thematic strategy for soil protection' will take time to be achieved and will need further steps, as outlined in the final conclusions of the Council on this thematic strategy. An efficient soil-information system capable of providing answers to the questions raised by policy-makers is a key requirement before any further action can be effectively undertaken. Soil information is available in Europe but is scattered in different institutions both at the national and at the European level. The proposal for a common approach to soil monitoring that the Commission will put forward will address this problem and propose solutions that will take into account the existing soil-information systems and propose a framework allowing for the interchange of data in a harmonized way across the EU. In the longer term, the availability of policyrelevant soil information will allow efficient implementation of the necessary measures to achieve soil protection for sustainable development in Europe.

Global policies Soil degradation processes are not confined to the European Union and the United States, but constitute a major worldwide problem with significant environmental, social and economic consequences. As the world population increases, the need to protect soil as a vital resource, particularly for food production, is increasing. Growing awareness in the international community of the need for global responses has led to an increasing number of international initiatives. The 1972 Council of Europe's Soil Charter

called on states to promote a soil-conservation policy. The World Soil Charter and the World Soils Policy sought to encourage international cooperation in the rational use of soil resources. The U N E P Environmental Guidelines for the Formulation of National Soil Policies set out a procedure for preparing national policy with a built-in sustainable land-use element. At the Earth Summit in Rio de Janeiro in 1992, the international community agreed on a global partnership for sustainable development and established the Agenda 21 framework. As a result of that, several conventions were launched. The 1992 Framework Convention on Climate Change (UNFCCC) recognizes the role and importance of terrestrial ecosystems as sinks of greenhouse gases, and that land degradation problems and changes in land use can exacerbate the emission of gases to the atmosphere. The 1997 Kyoto Protocol promotes sustainable development and calls on each party to implement policies and measures to protect and enhance sinks and reservoirs of greenhouse gases. The 1992 Convention on Biological Diversity (CBD) aims to conserve biological diversity. Fundamental to the CBD is the concern that biological diversity is being significantly reduced by human activities, including soil and land management. In several Conferences of the Parties of the Convention, decisions have been taken aiming at the protection of soil biodiversity and the reduction of the negative effect on it of certain agricultural practices, including the excessive use of inputs. The BGBD programme (Below Ground Biodiversity project) (http://www.bgbd.net), co-financed by GEF, which aims to better manage and conserve B G B D in tropical agricultural landscapes, is a good example of a programme linked to CBD and aiming at the protection of soil biodiversity in tropical environments. The 1994 Convention to Combat Desertification (UNCCD) acknowledges that arid, semiarid and dry subhumid areas together account for a significant proportion of the Earth's land area and represent the habitat and source of livelihood for a large segment of its population. The objective of the U N C C D is to prevent and reduce land degradation, rehabilitate partly degraded land, and reclaim desertified land through effective actions supported by international cooperation and agreements. The U N C C D contains five regional annexes covering Africa, Asia, Latin America and the Caribbean, the Northern Mediterranean (relevant for four E U member states: Greece, Italy, Portugal and Spain) and Central and

SOIL PROTECTION POLICIES Eastern Europe (relevant for most new E U member states). The elaboration and implementation of Regional Action Programmes and National Action Programmes, form valuable policy instruments to combat desertification and soil degradation phenomena in the affected areas. Furthermore, the Committee of Science and Technology, a subsidiary body within the Convention, produces a significant amount of information and advice on scientific and technological matters relating to land degradation worldwide. Initiatives have been developed over the past few years towards evaluating the possible need for a global soil convention ( W B G U 2001). In the later part of the 1990s, much of the non-UN attention paid to soil-conservation issues has been channelled through non-governmental organizations, among which the Tutzing Initiative has been prominent. From the mid-1990s, people connected with the Protestant Academy in Tutzing, Germany, worked towards a draft text for a convention. This was finalized in July 1998 (Held et al. 1998). The Danish Ministry of Food, Agriculture and Fisheries financed a recent analysis (Wynen 2002) of the various options at hand for the achievement of soil protection at global scale. The results of this study show the minimal effectiveness of current U N Conventions (UNFCCC, UNCBD, U N C C D ) in achieving soil protection globally. It also concludes that adding a new convention focusing on soil protection would not be the solution either, since implementation of current U N Conventions is already difficult. Two alternatives are proposed. One is to extend current conventions in order to cover soil protection more explicitly and develop guidelines and codes of conduct. The second option is to design an infrastructure that would provide the tools for proactive approaches to soil protection on voluntary basis, which seems to be more feasible and the only realistic way forward. Such an infrastructure could make use of a solid scientific basis through the creation of an Intergovernmental Panel for Land and Soil (IPLS) in analogy with the IPCC of the climatechange convention. This Intergovernmental Panel on Land and Soil (IPLS) should give scientific advice to the public, like the Intergovernmental Panel on Climate Change (IPCC). An Intergovernmental Panel on Land and Soils (IPLS) would have ( W B G U 2001): (1) To serve as a clearing house for ongoing and periodic assessment of global land and

(2)

(3)

(4)

(5)

157

soil degradation, and its impact on environmentally sustainable soil and land resources, and to propose regulatory management strategies. To assess and synthesize globally the scientific, technical and socio-economic information relevant for the understanding of the risk of human-induced land and soilquality change and show the pivotal role of soil and land use in ecosystem services at all scales. To address the variety of land-use and soilmanagement issues, including desertification, as related to environmentally sustainable development, food security, poverty alleviation and multilateral environmental agreements. To stimulate and involve the scientific community to advance and develop the science of soils and sustainable land use in a multidisciplinary context. To assist actively national, regional and global decision-makers in developing policies to assess, monitor and mitigate negative impacts of land and soil use.

Conclusions This short overview of policies for sustainable soil use has demonstrated that different approaches are possible for achieving soil protection. Voluntary instruments (incentives or subsidies) and involuntary instruments can be used to promote sustainable soil management. Crucial in any case is the availability of policyrelevant soil information in order to take knowledge-based decisions. Unfortunately, there is still a lack of sufficient data about soils in many parts of the world, including Europe, hampering objective evaluation of soil conditions for policy-making purposes (Van-Camp et al. 2004). The recognition of the trans-boundary character of soil degradation processes has focused the attention of international institutions on the need for development of soil-protection policies at a global level. The further development of the E U Thematic Strategy for Soil Protection will allow the exploration of new methods of international cooperation in this field.

References ARNOLD,1~ 2004. Lessons for Europe: the Experience of the US Soil Conservation Service, Proceedings of the 2nd SCAPE Workshop, Riomaggiore, Italy, 2004. BULLOCK, P., JONES, R.J.A. & MONTANARELLA,L. (eds) 1999. Soil Resources of Europe. European

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Soil Bureau Research Reports 6, C U R 18991 EN. Office for Official Publications of the European Communities, Luxemburg. CAUVIN, J. 2000. The Beginnings of Agriculture in the Near East: a Symbolic Interpretation. Cambridge University Press, Cambridge, UK. DEMOULE, J. & PERLES, C. 1993. The Greek Neolithic: a new review. Journal of Worm Prehistory, 7, 355-416. D E F R A 2004. The First Soil Action Plan for England: 2004-2006. March 2004, DEFRA, London. EC 2002. Communication of 16 April 2002 from the Commission to the Council, the European Parliament, the Economic and Social Committee and the Committee of the Regions: Towards a Thematic Strategy for Soil Protection [COM (2002) 179 final]. World Wide Web Address: http://europa.eu.int/ scadplus/printversion/en//lvb/128122.htm. HANNAM, I. & BOER, B. 2002. Legal and Institutional Frameworks for Sustainable Soils: a Preliminary Report. IUCN, Gland, Switzerland, and Cambridge, UK. HELD, M., KUMMERER, K. & ODENDAHL, K. 1998. Preserving soils for life. The Tutzing Project 'Time Ecology'. Proposal for a 'Convention on Sustainable Use o f Soils'. Verlag, Munich. HELMS, D. 1998. A Historical Guide to the U. S. Government. Oxford University Press, New York, 434-439. LAHMAR, R., HELD, M. & MONTANARELLA,L. (eds) 2002. People Matter, Food Security and Soils, Editions Charles Leopold Mayer, Paris. LASANTA, T., ARNAEZ, J., OSER1N, M. & ORTIGOSA, L.M. 2001. Marginal lands and erosion in terraces fields in the Mediterranean mountains. Mountain Research and Development 21 (1), 69-76. LOWDERMILK, W.C. 1994. Conquest of the Land

Through Seven Thousand Years. USDA, Soil Conservation Service. US Government Printing Office, 99, 1-30. MAZOYER, M. & ROUDART,L. 1997. Histoire des Agricultures du Monde, du N~olithique d la Crise Contemporaine. Editions du Seuil, Paris. MONTANARELLA,L., MICHELI,E. ~c ARNOLD, R. 2004. Soil conservation services in the European Union and in the United States of America. Proceedings of the 4th International Conference on Land Degradation, Cartaghena, Italy. NUSSER, S.M. & GOEBEL, J.J. 1997. The National Resources Inventory: a long-term multi-resource monitoring programme. Environmental and Ecological Statistics, 4 (3), 181-204. OLDEMAN,L.R., HAKKELING,R.T. & SOMBROEK,W.G. 1991. World Map of the Status of Human-Induced Soil Degradation (GLASOD). ISRIC/UNEE Wageningen, The Netherlands. PRICE, T.D. (ed.) 2000. Europe's First Farmers. Cambridge University Press, Cambridge, UK. VAN-CAMP, L., BUJARRABAL, B., GENTILE, A.-R., JONES, R.J.A., MONTANARELLA,L., OLAZABAL, C. & SELVARADJOU,S.-K. 2004. Reports of the Technical Working Groups Established Under the Thematic Strategy for Soil Protection. CUR 21319 EN/1, Office for Official Publications of the European Communities, Luxembourg. WBGU (GERMAN ADVISORY COUNCIL ON GLOBAL CHANGE)2001. World in Transition: New Structures for Global Environmental Policy. Earthscan, London and Berlin (can be also dowloaded as full text at: http://www.wbgu.de). WYNEN, E. 2002. A UN Convention on Soil Health or What Are the Alternatives?, Proceedings of the 14th IFOAM Organic World Congress, Victoria, Canada, August 2002.

Assessing anthropogenic inputs to soils by comparing element contents and their spatial distribution in O- and A-horizons M. I N / k C I O , V. P E R E I R A & M. P I N T O

Department o f Geosciences, University o f Aveiro, P-3810-190 Aveiro, Portugal (e-maik vpereira@geo, ua.pO Abstract: This paper reports the results of a low-density geochemical survey covering the entire surface of Portugal, taking organomineral horizons (A) and organic horizons (humus, O) as the sampling media. The main purpose of the study was to compare element contents and their spatial variability in the two soil horizons in order to distinguish between natural and anthropogenic inputs. Contents are compared using an index of enrichment, IE = (content in O in mg kg-a)/(content in A in mg kg-1) and the IE maps are used to discuss the origin of the elements. It was found that element contents in the upper horizons of Portuguese soils are controlled by the nature of the parent material, pedogenetic processes and anthropogenic additions. Our approach, evidencing spatial relationships, proved useful to the confirmation and refinement of the information obtained from biomonitoring surveys. Another advantage is a significant reduction in sampling costs, because by comparing the two uppermost soil horizons there is no need for sampling to great depths.

The concept of low-density/low-cost sampling for geochemical mapping (Garrett & Nichol 1967) has gained renewed interest over the past 15 years (Darnley et al. 1995; Plant et al. 2001). The need for geochemical mapping as a method of providing multi-element databases documenting the present composition of the surface environment is now recognized as a priority in many countries throughout the world. Such data are necessary for environmental legislation, and provide a reference against which future changes can be quantified. Applications to environmental protection, agriculture, forestry, human and animal health, are well documented in the literature (Appleton & Ridgway 1993; Tarvainen 1996; Xie Xuejing et al. 1997). A low-density geochemical survey of Portugal, taking soils as sampling media was carried out at the University of Aveiro between 1994 and 1999. A statistical summary of the survey, as well as a few maps, are presented in Ferreira et al. (2001) and Inficio et al. (2002). All the data (nearly 44 000 individual pieces of data) are stored in a database. The publication of the first Soil Geochemical Atlas of Portugal is in preparation. A knowledge of anthropic v. natural origin f or potentially harmful elements in soils is important for assessing human impact before fixing guide values and quality standards. The distinction between the geogenic background and anthropogenic inputs is currently attempted by comparing element levels in the top soil layer with those from a greater depth. Many authors have compared element contents in A and C

horizons (e.g. Curl~ & SefO'k 1999; Cannon et al. 2004), in O and C horizons (Reimann et al. 1998) or in O, B and C (Reimann et al. 2001). To our knowledge, comparison between the O and A horizons has never been considered for such purpose. There is general agreement in taking the C horizon to reflect the soil geogenic composition, but it usually lies within a depth range of 50-200 cm, which renders sampling very timeconsuming and expensive. The main aim of the present study was to attempt the distinction between natural and anthropogenic origin for 20 elements in the Portuguese soils, by comparing element contents and their spatial distribution in the two uppermost soil horizons, the O and the A.

Materials and m e t h o d s

Soils and geology A simplified soil map of Portugal is presented in Figure la. The major soil groups considered for the present work and the equivalent FAO (1988) soil units, as well as the number of samples in each group of soils, are listed in Table 1. A geological sketch-map of Portugal is shown in Figure lb.

Sampling and sample preparation The whole continental area of Portugal, which is about 89 000 km 2, was covered by a low density geochemical survey. Sampling, sample

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENT1N,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society,London, Special Publications, 266,159-170. 0305-8719/06/$15 9 The Geological Society of London 2006.

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M. INACIO E T A L .

Fig. 1. (a) Major soil types in Portugal (simplified from Cardoso et al. 1973). (b) Lithological map of Portugal (simplified from Atlas Digital do Ambiente, 1992).

ELEMENT CONTENTS IN O- AND A-HORIZONS

161

Table 1. Selected information on the major soil groups considered for this study Major soil groups

Arenosols Cambisols Fluvisols Leptosols Luvisols Planosols Podsols Calcic Soils Vertic Soils Other

FAO (1988) equivalence

Area (%)

Arenosols Cambisols, except Calcaric or Vertic Cambisols Fluvisols, except Calcaric Fluvisols Leptosols Luvisols, except Calcaric or Vertic Luvisols Planosols Podsols Calcaric Cambisols, Calcic Luvisols and Calcaric Fluvisols Vertisols, Vertic Luvisols and Vertic Cambisols Rock outcrops, urban areas, other soil groups

preparation and analysis followed the recommendations of the IGCP (International Geological Correlation Programme) Project 259 'International geochemical mapping' (Darnley et al. 1995). The A-horizon was sampled from 652 sites, at a density of 1 site/135 km 2 from the 0-20 cm layer. Humus samples (O-horizon) were collected where present (195 samples), mostly in forested areas located in the north and centre of the country. On each site, a composite sample made up of five grabs was collected over an area of about 100 m 2. For both horizons, duplicates were taken every 10 sites. The samples were dried at 35-40 ~ passed through a 180 pm plastic sieve, homogeneized and quartered. They were given random numbers in order to remove any systematic relationship between order of analysis and geographical location.

Number of samples A-horizon

O-horizon

1.86 31.35

9 249

2 113

3.72 17.53 20.48 0.16 7.06 6.12

8 144 90 0 80 49

2 33 13 0 17 15

3.32

20

0

8.4

3

0

one hour shaking and then standing for 20 hours. Organic-matter content (OM) was estimated gravimetrically by loss-on-ignition; duplicate 5-g oven-dried samples were ignited for 16 hours in a furnace at 450 ~ Particle-size analysis was performed by the classical combination of sieving and the pipette method. The accuracy and precision of element analysis were checked by analysis of two 'in house' standards; two international standards; duplicates (analytical splits) of randomly selected samples; and by field duplicates, and were found acceptable. However, 15% of the analytical results for Hg were rejected because the limit of detection (1 mg kg -1) was too high for the purpose of this study. Detailed results of the extensive quality-control procedures can be found in In~cio Ferreira (2004).

Data processing and m a p p i n g Analysis a n d analytical quality control The chemical analysis was performed in the ACME Analytical Laboratories Ltd.,Vancouver, Canada. Representative 0.500-g subsamples were extracted for one hour with aqua regia (3-2-1 concentrated HC1-HNO3-H20) at 95 ~ and the extracts were analysed by I C P - A E S for A1, As, Ba, Ca, Co, Cr, Cu, Fe, K, Hg, La, Mg, Mn, Ni, P, Pb, Sr, Th, V and Zn. Some additional soil analyses were carried out in the D e p a r t m e n t of Geosciences, University of Aveiro. The pH was determined on duplicate samples in a water suspension, using a soil:water ratio of 1:2.5. Electrical conductivity (EC) was measured on duplicate samples in a soil:water suspension (1:10) after

Basic statistical parameters for each variable, correlation matrices and box-plots were calculated using Statistica | software. The distribution maps were prepared following recommendations in Darnley et al. (1995). The colour maps for all elements, as well as for pH, E C and OM in A horizons, based on 652 soil samples, were plotted by kriging using a variogram model adjusted for each variable. The software used was Surfer 7.0 (Keckler 1999) and the geostatistical analysis was performed by Variowin 2.21 (Pannatier 1996). Dot maps were preferred for representing the spatial distribution of elements in the organic horizon, because only 195 samples are available. These maps were obtained using the technique

162

M. I N A C I O E T A L .

described in Bj6rklund & Gustavsson (1987), in which the diameter of the dots is related to metal content by a continuous function curve defined by the user. For the compilation of the maps and other statistical calculations, contents below the detection limit were replaced by half that value. R e s u l t s and discussion Table 2 shows some basic statistical parameters for the 20 elements selected and the variables pH, EC and OM in O- and A-horizons. The element content of a soil depends on (1) the nature of the soil parent material, (2) the influence of pedogenic processes, and (3) anthropogenic additions either directly as herbicides, pesticides and fertilizers, or indirectly as pollutants. The spatial distribution of most elements in humus and topsoils, as assessed by visual inspection, is quite similar and generally appears to be controlled by lithology and soil type. To exem-

plify these relationships, the distributions of A1 and Ni are given in Figures 2 and 3. As one of the main constituents of the Earth's crust, total A1 in rocks commonly ranges from 0.45 to 10% (Kabata-Pendias 2001). In magmatic rocks, the highest values (7.8 to 8.8%) are found in intermediate and mafic rocks (diorites, syenites, gabbros, basalts), as well as in acid granitoids (7.2 to 8.2%), while the figures for ultramafic rocks are much lower, in the range of 0.45 to 2%. In sedimentary rocks, the A1 content ranges from 10 to 0.4%, with the highest values for argillaceous rocks and the lowest for limestones and sandstones. Soils developed on granitic and mafic or ultramafic lithologies, respectively (Cambisols and Vertic Soils) had high A1 contents. Podsols and Calcic Soils, formed from sandy parent materials and limestones, had the lowest A1 contents. Nickel contents are highest in ultramafic rocks (1400 to 2000 mg kg-1), decreasing with the increasing acidity of the rocks, down to 5 to 15 mg kg -1 in granites. Sedimentary rocks contain Ni in the range 5 to 90 mg kg -1, with the

Table 2. Basic statistics Element

O-horizon A

AI (%) As (mg kg-1) Ba (mg kg-1) Ca (%) Co (mg kg-1) Cr (mg kg-1) Cu (mg kg -1) Fe (%) Hg (~tg kg-1) K (%) La (mg kg-l) Mg (%) Mn (mg kg -1) Ni (mg kg-1) P (%) Pb (mg kg -1) Sr (rag kg-1) Th (mg kg -I) V (mg kg -1) Zn (mg kg 1) pH EC (mS/m -3) OM (%) Clay (%)

1.68 16 61 0.68 7 19 21 1.94 71 0.22 23 0.35 583 17 0.074 32 20 6 27 72 5.1 1.64 23.7 .

Median GM

1.51 11 52 0.24 5 15 17 1.88 60 0.17 20 0.28 378 11 0.067 27 15 3 21 64 4.9 1.22 20.9 .

Observed range

A-horizon Expected range

1.32 0.17-5.57 0.27-3.78 10 <2-139 2-45 49 12-316 16-151 0.24 0.01-15.28 0.03-2.04 5 <1-52 1-22 13 <1-215 3-54 17 3-114 5-55 1.59 0.24-6.1 0.34-3.86 61 15-475 25-145 0.17 0.03-0.74 0.05-0.57 18 2-88 5-54 0.27 0.03-1.92 0.06-0.87 374 19-3355 74-2231 11 2-440 3-39 0.065 0.015-0.236 0.026-0.153 28 6-158 11-70 15 3-211 5-59 4 <2-67 <2-19 21 3-174 7-68 61 8-331 20-134 5.0 3.7-6.8 4.0-6.6 1.28 0.03-8.92 0.46-3.88 21.2 9.96-80.2 10.7-48.0 . . .

A

2.07 17 65 0.80 10 26 21 2.66 55 0.23 29 0.39 586 22 0.049 26 16 7 32 59 5.2 0.22 7.8 15.1

Median GM

1.84 11 53 0.10 8 21 16 2.74 50 0.15 25 0.29 394 16 0.038 21 10 5 27 55 5.0 0.39 6.1 14.0

Observed range

Expected range

1.69 0.19-9.30 0.43-4.50 10 <2-266 <2-55 51 6-422 16-155 0.11 <0.01-23.24 0.01-4.08 7 < 1-84 1-27 18 <1-336 4-59 14 <1-245 3-53 2.26 0.23-6.49 0.56-4.66 46 <10-285 15-105 0.16 0.02-1.52 0.05-0.68 24 1-155 9-60 0.26 0.01-4.24 0.04-1.03 357 13-4965 59-1857 13 <1-880 3-56 0.036 0.004-0.610 0.010-0.135 21 2-585 8-51 11 2-217 3-40 5 <2-87 <2-23 25 3-192 7-75 46 5-738 10-113 5.1 3.6-8.1 4.0-7.5 0-4.5 0.19-1.37 6.4 1.2-41.5 2.4-19.8 13.4 2.7-39 5.0-30.7

Summary statistics of the analytical data for 195 organic horizons (167 for Hg) and 652 A-horizon samples (358 for Hg). A, arithmetic mean; GM, geometric mean; observed range, minimum-maximum; expected range, 5th-95th percentile.

ELEMENT CONTENTS IN O- AND A-HORIZONS

163

Fig. 2. Spatial distribution and cumulative frequency curves of A1 in organic and A-horizons. highest values for argillaceous rocks and the lowest for sandstones (Kabata-Pendias 2001). In Portugal, Ni was significantly higher in Vertic Soils formed from the weathering of mafic and ultramafic igneous rocks, as well as in Leptosols developed on metasedimentary rocks. The lowest contents of Ni were found in Podzols and Cambisols, i.e. in soils derived from sandy parent materials and granitic rocks.

Compar&on between O- and A-horizons The correlation between element contents in O and A horizons is presented in Table 3. The correlations (Spearman rank order and Pearson, P < 0.001) between element contents in the two media are generally high, particularly for A1, As, Ba, Co, Cr, Fe, La, Ni, Th and V. This reflects the similarity in their spatial distri-

bution, showing that the composition of both Oand A-horizons is strongly influenced by the composition of the substratum. The lowest correlations were observed for Hg and Pb, suggesting that their contents were influenced by other sources, e.g. by atmospheric input. The box-plots for ten selected elements, Ca, Mg, Mn, P, Sr, A1, As, La, Fe and Th, are given in Figure 4. The ratio of the median values obtained for the 20 elements and for organicmatter contents in the O- and the A-horizons considering the entire dataset are compared in Figure 5. The median contents in organic horizons are, at least, a factor of 1.5 higher for Ca, Mg, Mn, P and Sr and slightly higher for Ba, Cu, K, Pb and Zn. On the other hand A1, As, La, Fe and Th are significantly higher (KolmogorovSmirnov Test, P < 0.01) in topsoils which are also slightly enriched in Co, Cr, Ni and V.

164

M. INACIO E T A L .

Fig. 3. Spatial distribution and cumulative frequency curves of Ni in organic and A-horizons.

Table 3. Correlation coefficients

A1 As Ba C. Pearson C. Spearman

Ca Co

Cr

Cu Fe Hg

K

La Mg Mn Ni

P

Pb

Sr

Th

V

Zn

0.93 0.90 0.91 0.85 0.95 0.91 0.84 0.91 0.50 0.88 0.91 0.90 0.85 0.93 0.80 0.59 0.81 0.84 0.93 0.85 0.92 0.90 0.89 0.80 0.94 0.89 0.83 0.92 0.53 0.86 0.91 0.87 0.82 0.90 0.78 0.54 0.70 0.82 0.89 0.80

Spearman's rank order and Pearson correlation coefficients between element contents in O- and A-horizons (n = 195 samples; 167 for Hg). These correlation coefficients are significant at P < 0.001.

T h e e n r i c h m e n t in Ca, Mg, M n , P a n d Sr and, to a lesser d e g r e e , in Z n , C u a n d K, in the o r g a n i c h o r i z o n s is likely to be r e l a t e d to ion u p t a k e by plants a n d to the recycling of plant r e s i d u e s in the litter. H i g h m e t a l c o n t e n t s in o r g a n i c horizons, c a u s e d by t h e r e c y c l i n g of

plant residues, have been widely reported ( R e i m a n n et al. 1997; K a b a t a - P e n d i a s 2001). S t r o n t i u m is n o t a n u t r i e n t , but is very similar to C a in m a n y b i o c h e m i c a l a n d g e o c h e m i c a l properties. T h e e l e m e n t s with h i g h e r contents in topsoils,

ELEMENT CONTENTS IN O- AND A-HORIZONS

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Fig. 4. Box-plots for A1, Ca, Fe, Mg, R As, La, Mn, Sr and Th. O-horizon - white boxes; A-horizon - black boxes. A1, Fe, La,Th, and also V, As, Co and Cr, are typically lithogenic, major and minor components of many silicate minerals.

Spatial relationships Comparing medians on the entire data can be misleading, therefore it was decided to study carefully the regional distribution patterns by making one-to-one comparisons for each element at each sampling site. Contents were compared for each element at each sampling site by defining an index of enrichment as: IE = (content in O in mg kg-1)/(content in A in mg kg-1).

The mapping of this index produced three different patterns. A few examples are given in Figures 6 to 8. The maps for Ca, Mg, Mn and Sr in Figure 6 show high values of IE uniformly distributed throughout the study area. This supports the hypothesis of nutrient storage in humus - a pedogenetic process common to all soils. High values of IE confined to specific locations suggest anthropic contamination. This is the case for As, Cu, Hg, Ni, R Pb, V and Zn. Many of these locations can be related to known sources of contamination like industries, large cities or agriculture. This is exemplified in Figure 7 for As, V and R Known sources of As are a chemical plant (for pyrite roasting) located at

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Fig. 5. Ratio between median content of 20 elements and organic matter (OM) in O- and A-horizons. Barreiro, and three coal-fired power stations at Oporto, near Lisbon and at Sines; the spots in the interior are probably related to past exploration of Sn-W mines. The V pattern is very similar to the Ni (Figure 9), and the combination of these two elements can be related to fuel combustion at a Portland cement plant located near Figueira, and at several power plants located north of Lisbon and Setfibal. The axis Lisbon-Setfibal-Sines concentrates many industries, including a petrochemical plant and an oil refinery, releasing many metals, mainly As, Ni, V and Zn, to the surrounding environment. The pattern for P is characterized by higher values of IE roughly grouped in two areas, one east of Oporto and the second along the west coast near important agricultural areas where phosphate fertilizers are certainly used. The IE is generally low all over the study area for A1, La, Fe (Figure 8) and Th, indicating that the concentration of these elements in Portuguese soils is chiefly controlled by the lithology. C o m p a r i s o n with b i o m o n i t o r i n g surveys The results of the present investigation were compared with two biomonitoring surveys of Portugal: one using the epiphytic lichen Parmelia sulcata (Freitas et al. 1999) to monitor airborne inputs of As, Cr, Hg, Ni, Pb, S, Sb, Se and V, and the second using the mosses Hypnum cupressiforme and Scleropodium touretii (Figueira et al. 2002) to evaluate the atmospheric deposition of Cd, Cr, Cu, Fe, Mn, Ni, Pb and Zn. Levels in the soil O- and A-horizons were considerably higher than in mosses and lichens for all the elements compared. The main

anthropic sources for the metals measured in the biomonitoring surveys can be identified in the soil IE maps. The comparison between the maps in Figure 9 shows that the high Ni contents in mosses and lichens in the northeast of the country can be interpreted as resulting from windblown soil dust rich in Ni (Prud~ncio et al. 1999; Figueira et al. 2002). In other areas (in the regions of Figueira, Lisbon and Setfibal) the high Ni content of lichen and mosses is explained by Ni emissions from industrial activities as suggested by the IE map.

Conclusions The analysis of metal contents in the fine fraction (<180 pm) of O- and A-horizons provided enough contrast to detect enrichment between the two. The A-horizons have generally significantly higher contents of lithogenic elements. Enrichments in the O-horizons may be of natural or anthropic origin, or both. The spatial distribution of the enrichment index (metal content in O)/(metal content in A) proved helpful to differentiate, at a regional scale, natural background values in soils from areas affected by anthropogenic contamination. Results obtained with the enrichment index were complementary to those obtained from biomonitoring studies assessing the metal contents of lichen and mosses. The interest of this enrichment index remains, however, restricted to sites where both A- and a Ohorizons can be sampled from similar soil types developed on similar parent material. In this study, the enrichment index could not be determined in the south-east of Portugal. Finally, it

E L E M E N T CONTENTS IN O- A N D A - H O R I Z O N S

Ca

Mg

Mn

Sr

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A

IE 0-1 ~ o2-5 C) 5 - M a x ' Fig. 6. Spatial distribution of IE; the ratio between contents of Ca, Mg, Mn and Sn in O- and A-horizons.

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M. IN./ICIO E T A L .

AS

P

V

0

9

~

Sines~~/

Setabal

-

V

z,_ iguel

0-1 2-5 5_Max. 1-2 lE

lOOkm

Fig. 7. Spatial distribution of IE; the ratio between contents of As, P and V in O- and A-horizons.

AI

Fe

La

IE 0-1 o 1-2 2-5 5-Max. -

I

Fig. 8. Spatial distribution of IE; the ratio between contents of AI, Fe and La in O- and A-horizons.

s h o u l d b e e m p h a s i z e d t h a t this a p p r o a c h allows o n l y v e r y g e n e r a l c o n c l u s i o n s to b e d r a w n o n t h e o r i g i n of t h e e l e m e n t s , a n d o f c o u r s e m u c h is left to b e e x p l a i n e d .

The first author would like to thank the 'Fundaq~o para a Ci6ncia e a Tecnologia' for the award of a Ph.D. scholarship.

ELEMENT CONTENTS IN O- AND A-HORIZONS

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Fig. 9. Spatial patterns of (a) Ni content (mg kg-1) in lichens; (b) Ni content (mg kg -1) in mosses; (e) IE, the ratio between contents of Ni in O- and A-horizons; and (d) Ni content (mg kg -1) in A-horizons.

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References APPLETON, J.D. & RIDGWAY, J. 1993. Regional geochemical mapping in developing countries and its application to environmental studies. Applied Geochemistry, 2 (Suppl.), 103-110. ATLAS DIGITAL DO AMBIENTE 1992. Minist6rio das Cidades, do Ordenamento do Territ6rio e Ambiente. World Wide Web Address: http://www. iambiente.pt/atlas/. BJORKLUND,A. & GUSTAVSSON,N. 1987. Visualization of geochemical data on maps: new options. Journal of Geochemical Exploration, 29, 89-103. CANNON, W.E, WOODRUFF, L.G. & PIMLEY, S. 2004. Some statistical relationships between stream sediment and soil geochemistry in northwestern Wisconsin - can stream sediment compositions be used to predict compositions of soils in glaciated terranes? Journal of Geochemical Exploration, 81, 29-46. CARDOSO, J.C., BESSA, M.T. & MARADO, M.B. 1973. Carta dos solos de Portugal na escala 1:1 000 000. Agronomia Lusitana, 33, 481-602. ~URLiK, J. & SEFCiK,R 1999. Geochemical Atlas of the Slovak Republic - Soils. Ministry of the Environment of the Slovak Republic, Bratislava, Slovak Republic. DARNLEY,A.G., BJORKLUND,A. ETAL. 1995. A Global Geochemical Database for Environmental and Resource Management. Recommendations for International Geochemical Mapping. Final Report of IGCP Project 259. UNESCO Publishing. FAO 1988. FAO-UNESCO Soil Map of the World, Revised Legend. World Soil Resources Report, 60, FAO, Rome. FERREIRA, A.M., INACIO, M., MORGADO, P., BATISTA, M.J., FERREIRA, L., PEREIRA, V. (~ PINTO, M.S. 2001. Low density geochemical mapping in Portugal. Applied Geochemistry, 16, 1323-1331. FIGUEIRA, R., St~RGIO, C. (~z SOUSA, A.J. 2002. Distribution of trace metals in moss biomonitors and assessment of contamination sources in Portugal. Environmental Pollution, 118, 153-163. FREITAS, M.C., REIS, M.A, ALVES, L.C. & WOLTERBEEK, H. 1999. Distribution in Portugal of some pollutants in the lichen Parmelia sulcata. Environmental Pollution, 106, 229-235.

GARRETT, R.G. & NICHOL, I. 1967. Regional geochemical reconnaissance in eastern Sierra Leone. Transactions of the Institution of Mining and Metallurgy, B76, 97-112. INACIO FERREIRA, M. 2004. Dados geoquimicos de base de solos de Portugal Continental, utilizando amostragem de baixa densidade. Ph.D. thesis, University of Aveiro, Portugal. INACIO,M., PEREIRA,V. & PINTO,M. 2002. Correlation of geochemical data based on coarse and fine fractions of soils. 17th World Congress of Soil Science, Bangkok, Thailand, Scientific registration No. 1 2 0 3 , CD-ROM, 1-10. KABATA-PENDIAS,A. 2001. Trace Elements in Soils and Plants 3rd edn, CRC Press, Boca Raton, FL. KECKLER, D. 1999. Surfer for Windows Version 7, User's Guide. Golden Software, USA. PANNATIER, Y. 1996. Variowin - Software for Spatial Statistics. Analysis in 2D. Springer-Verlag, New York. PLANT, J.A., SMITH,D., SMITH,B. & WILLIAMS,L. 2001. Environmental geochemistry at the global scale. Applied Geochemistry, 16, 1291-1308. PRUDt~NCIO, M.A., GOUVEIA, M.A., FREITAS, M.C., CHAVES, L. t~ MARQUES, A.E 1999. Soil versus lichen analysis on elemental dispersion studies (north of Portugal). International Workshop on Biomonitoring of Atmospheric Pollution, Lisbon, 21-24 September 1997. I A E A TECDOC. REIMANN, C., CARITAT, P., NISKAVAARA,H., FINNE, T.E., KASHULINA, G. &. PAVLOV, V.A. 1998. Comparison of elemental contents in O- and Chorizon soils from the surroundings of Nikel, Kola Peninsula, using different grain size fractions and extractions. Geoderma, 84, 5-87. REIMANN, C., KASHULINA, G., CARITAT, P. ~z NISKAVAARA, H. 2001. Multi-element, multimedium regional geochemistry in the European arctic: element concentration, variation and correlation. Applied Geochemistry, 16, 759-780. TARVAINEN, T. 1996. Environmental Applications of Geochemical Databases in Finland. Synopsis. Geological Survey of Finland, Espoo. XIE XUEJING,X., XUZHAN, M. & TIANXIANG,R. 1997. Geochemical mapping in China. Journal of Geochemical Exploration, 66, 99-113.

Nutrient balances for improving the use-efficiency of non-renewable resources: experiences from Switzerland and Southeast Asia H. M E N Z I 1 & E G E R B E R 2

aSwiss College o f Agriculture (SHL), CH-3052 Zollikofen, Switzerland (e-maik harald, menzi@shl, bfh. ch) 2Food and Agriculture Organisation o f the United Nations (FAO) - Livestock Environment and Development Initiative ( L E A D ) , 00100 Rome, Italy Abstract: Simple nutrient balances can be a valuable tool to assess and visualize whether farming systems are in equilibrium or not with respect to nutrient inputs and outputs. Thus, they can raise awareness about environmental risks of agricultural production, and especially intensive livestock production. They are also a valuable tool for policy development and implementation processes. In this paper, we discuss how a nutrient-balance approach can help to improve the awareness of different stakeholders about the ecological impacts of agriculture and the efficient use of nutrients. While the example from Switzerland demonstrates the potential ecological benefits of compulsory nutrient balancing, the examples from Asia demonstrate how a combination of nutrient balancing and geographic information system (GIS) was used to identify hot-spots with respect to insufficient and excessive P input, and we will discuss a nutrient-balance method developed to improve nutrient management in the rapidly growing intensive livestock-production sector.

Potential role of the nutrient-balance approach in agriculture with respect to soil fertility and sustainable resource management

increased by a factor of eight to 10. At the same time, the pressure on the e n v i r o n m e n t (soil, water, air) increased considerably. U r g e n t concerns arising from this d e v e l o p m e n t are, for example:

In a competitive market context, commercial agriculture tends to specialize and to be located in areas where production costs are minimized. This has a strong negative effect on nutrient cycles, especially when crop and livestock production are split (specialization) and the nutrients excreted by livestock are not brought back to the feed production areas, because of prohibitive transport costs. This results in areas with net nutrient deficits, and others with strong n u t r i e n t overloads. N u t r i e n t deficits can be compensated by chemical fertilizers, but at the expense of non-renewable resources. M a r k e t - o r i e n t a t e d crop p r o d u c t i o n also c o m m o n l y aims at increasing productivity. Where land is limited, which is the d o m i n a n t situation, this translates into increasing yields. This aggravates the risk of nutrient imbalance, n u t r i e n t losses and of using n o n - r e n e w a b l e resources with a low efficiency, because the increase in inputs usually surpasses the increase in yield. For example Tilman et al. (2002) showed that while global cereal production increased by about a factor of 2.5 between the beginning of the 1960s and the end of the 1990s, N fertilizer and irrigation water use

(1) the threat to soil fertility arising from the a c c u m u l a t i o n of heavy metals, pesticide residues and excessive nutrient inputs; (2) the pollution of ground- and surface-water resources and eutrophication of susceptible terrestrial and aquatic ecosystems; (3) the depletion of easily accessible resources of phosphorus. According to H e r r i n g & Fantel (1993) and Stewart et al. (2005), the presently k n o w n reserves that can be exploited at a relatively low cost will be exhausted within the next few centuries; (4) the high fossil-fuel requirements for the p r o d u c t i o n of N chemical fertilizers, r e q u i r e m e n t s which c o n t r i b u t e to local pollution and climate change; (5) the consumption of water, which contributes to the fast depletion of water resources in many parts of the world; (6) the increasing pollution of ground- and surface-water resources through agriculture and other anthropogenic sources, which in m a n y parts of the world e n d a n g e r s the regular and safe supply of water for h u m a n and animal consumption, and seriously harms aquatic ecosystems.

FROM:FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. E (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266,171-181. 0305-8719/06/$15 9 The Geological Society of London 2006.

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In many parts of the world, especially Africa and parts of Asia (mainly western China and South and West India), Latin America and Eastern Europe, yield increase with limited fertilizer use and inadequate soil management leads to soil nutrient depletion and thus to soil degradation. In the case of excessive inputs or of nutrient depletion, the balance approach is a valuable tool to identify and visualize the problem and to help evaluating potential mitigation options. It can thus make an important contribution towards more sustainable agricultural systems with improved use-efficiency of non-renewable resources and a reduced threat to the environment (impacts on soil fertility, emissions to water and atmosphere, threat to vulnerable ecosystems). Element mass-balances are frequently used in research to assess how far agro-ecosystems are from a steady-state equilibrium with respect to inputs and outputs of nutrients (e.g. nitrogen N, phosphorus - P, potassium - K), heavy metals or other substances (Smaling & Oenema 1997; von Steiger et al. 1998). They are also used to model the potential surplus load of different pollutants in soil-monitoring programmes or the emission potential of farming systems (Keller et al. 2005). The balance approach is been increasingly used as a controlling instrument in the implementation of environmental policy at the farm level. A survey performed by Goodlass et al. (2003) across Europe and the United States of America identified 50 input-output accounting systems in which nutrients are considered. Experience from European countries and other continents has shown that the nutrientbalancing approach is promising not only for policy implementation but also as an awarenessraising and planning instrument, because it shows the equilibrium or disequilibrium of inputs and outputs in a form that is well understood by non-experts. This contribution gives details of experience from Switzerland and from Southeast Asia and discusses which aspects are important for the successful use of the balance approach at the practical level, involving farmers, agricultural extension services and policy makers.

Challenges of using element balances for the management of agricultural production systems D i f f e r e n t c o n c e p t s o f e l e m e n t balances

A nutrient balance in agriculture is the summary of the book-keeping of nutrient inputs

and outputs of a defined system (Oenema et al. 2003). A nutrient surplus or deficit is calculated as the physical difference between nutrient inputs and outputs. An overview of the different approaches used for calculating nutrient balances in crop production is given by Oenema et al. (2003). The most commonly used approaches in agricultural production systems are: (1) The farm-gate balance. This calculates the balance between what enters the farm (imports) and what leaves it (exports), ignoring farm-internal fluxes. (2) The soil surface or field balance, which compares inputs and outputs of crop production (including grassland). In this approach, internal fluxes are important, especially on livestock farms (nutrients in crops grown on the farm that are used as feedstuffs, and the nutrient content of manure produced on the farm). Two approaches can be used to calculate the outputs, i.e. the amount of nutrients taken up by crops: (a) considering the actual offtake of nutrients in harvested crops (yield x nutrient content of the harvest product), or (b) considering the nutrient requirements of crops reported by local agricultural research stations. As long as nutrient requirement recommendations are based on off-take, as it is often the case for P and K, the two methods should give similar results; (3) The nutrient excretion balance of animals, which balances elements in feed with element retention through livestock growth and production. This approach focuses on livestock production only, and gives no direct indication of the nutrient balance for a certain area. Farm-gate and soil-surface-balance methods may also take into account fluxes which can hardly be influenced by the farmer or which are difficult to quantify, such as inputs from the atmosphere through deposition or nitrogen fixation, and emissions through volatilization, denitrification or leaching. These fluxes can be particularly high for nitrogen. In general, the mass-balance methodology becomes more complicated if one deals with substances that are very mobile or with substances that can easily be transformed (e.g. organic compounds). As the necessary data are not always available for such substances, the interpretation of the results might be problematic and the results would not be reliable enough to be used for farm management.

NUTRIENT BALANCES TO IMPROVE EFFICIENCY

Validity and comparability o f element balances A direct validation of element balances at the field, farm or regional scale is often difficult. Nevertheless, it can be assumed that balances should be reliable and reflect the reality as long as the available input data are valid, complete and the assumptions of the model are appropriate. In the case of E the balance can be assumed to reliably identify potentially significant longterm changes in the soil's P status. Except for losses through runoff and erosion, there are hardly any significant fluxes that are not directly quantifiable through farming-practice records. The inputs and losses that are difficult to assess, as discussed above for nitrogen, make the results of nitrogen balance difficult to interpret. For example, Oenema et al. (2003) identified denitrification and leaching as the largest uncertainties in the nitrogen balance for The Netherlands. This is an important argument reason for the reluctant use of nitrogen balance in programmes with direct financial consequences for farmers. It should also be noted that to identify surpluses resulting from livestock production, it is sufficient to monitor the P balance, because this is usually the first element for which a surplus can be observed, and because of the strong correlation between nitrogen and phosphate content in livestock excretions. Schuepbach (2002) compared the N and P balance for different model farms, using five different balance approaches, a farm-gate and a soil-surface balance from The Netherlands; a farm-gate and a soil-surface balance from Germany; and the Swiss soil-surface balance 'Suisse-Bilanz'. The five methods gave quite diverging results. While many of the differences concerning inputs and outputs could be explained by different farming practices in the three countries or by differences in the calculation approach, the largest difference resulted from the assumed or allowed losses (Schuepbach 2002; Frossard et al. 2004), Thus, we conclude that the comparison between different balancing methods is delicate, and that a correct interpretation of element balances requires a good knowledge of the methodology used (assumptions used, aims and consequences of the balance) and a proper implementation.

Requirements f o r balancing tools The interest in nutrient balances as an indicator of sustainability in comparison with other approaches (e.g. soil or water sampling) lies in

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its robustness and in the possibility of calculating it on the basis of statistics and default values. Thus, nutrient balances have been used to estimate nutrient flows at farm, regional and national level (e.g. Brouwer et al. 1995; Saleem 1998; Scoones & Toulmin 1998; Bindraban et al. 2000; O E C D 2001). The lack of available or reliable data or biased information provided by the farmer can be major restrictions on the practical use of the balance approach at farm level. The balance should therefore be limited to input variables that can be assumed to be available on farms. Furthermore, appropriate procedures to check the reliability of the data should be established. The use of an element-balance approach on a farm is usually bound to practical consequences for the farmer. This can be (1) an awarenessraising effect which in the long term should improve the sustainability of the production system; (2) a monitoring result which shows how far the farm complies with defined standards; or (3) a planning tool to improve farm management. It is therefore essential that the general concept is understandable to the farmers, and that the results are brought into direct relation to farm-management practice. This implies that input data and results should, as far as possible, be related to information or fluxes that the farmer is aware of and that can be estimated with reliable precision. Fluxes like nitrogen fixation, atmospheric deposition or emissions through volatilization or leaching are much more difficult to communicate than fertilizer requirements and actual inputs in the form of mineral fertilizers or feed. Furthermore, the general concept of the calculation should be as simple as possible. If the balance calculation is to be performed by the farmers themselves, the methodology and the tool (PC-program or form for manual calculation) should be particularly user-friendly. Good participation of the farmers will only be achieved if they at least partly understand why this calculation is done and they do not fear negative consequences from the results. This implies that they should be at least partly aware of the environmental consequences of a nonequilibrium situation, and that the results of the balance are related to incentives rather than punishments. When the balancing approach is used in policy implementation, the requirements of simplicity, limited need for input information, and robustness, are especially important. Nutrient balances can be a valuable planning and management aid to the farmer. They can, for example, be:

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(1) simple substitute for fertilizer plans, showing whether additional mineral fertilizer would be needed to supply crops with nutrients, or whether the amounts of nutrients provided by livestock would exceed crop needs; (2) a farm-planning tool showing the livestock density that can be reached in a given farm or area without producing major manure surpluses and thus providing the basis for sustainable local recycling of the animal excreta; (3) a tool with which the implications of different farm-management strategies on the nutrient management and the environmental risk can be assessed. In case a balance approach will be used for different purposes, e.g. as monitoring tool for policy implementation and as a planning aid for farmers, it is essential that it follows the same general procedure and supplies congruent results for all applications as long as the same input data are used. If this is not the case, misunderstandings or even legal conflicts will be inevitable. It is also important that the method is designed to handle a broad variety of farm types, located in diverse contexts.

Possibilities for overcoming the diverging requirements Taking into account these partly conflicting demands can be a major challenge for an element-balancing methodology. It may therefore be necessary to: (1) Consider parameters which are familiar to farmers: (a) To omit fluxes like nitrogen fixation, deposition, leaching, etc. (such fluxes can be considered in the interpretation guidelines); (b) To consider the amount of nitrogen (N) that is available to crops in manures (Navailable) instead of the total N c o n t e n t (Ntotal) , or t o limit the balance to elements which are less volatile and not transformed; (2) Use fertilizer recommendations from local agricultural research stations rather than actual crop-uptake, because such recommendations are a commonly used planning aid on most farms. This implies that reliable recommendations are available and that they should be based on 'standard' or good practice, rather than the results of individual experiments or on a yield maximization strategy;

(3) Use guide-values for livestock nutrient excretions, rather than for nutrient contents of manure, because of the high variability of manure composition. An approach based on manure analysis should be avoided, because it is costly and is uncertain due to sampling difficulties. For non-ruminants it would also be possible to use information about diets and production parameters (e.g. feed-conversion ratio, slaughter weight and number of cycles per year) to estimate excretions; (4) Use a simple general approach, but with possibilities for more detailed inputs if these are available and if a farm-specific result is desired. This can be achieved by having a detailed calculation procedure using default values in the background of a PC-application, which could be adapted to specific conditions by advanced users. Such an approach would be well suited to meet the expectations of different user groups with variable background knowledge and input data available.

The Swiss nutrient-balance concept Manure has always been of great importance in Switzerland, contributing a major part of the nutrients used in crop production. Farmers therefore required planning aids to take into account manure nutrients in their fertilization strategies. In addition, the development of environmental policies for agriculture has also required an appropriate nutrient-balance methodology to be developed.

The Swiss ecological direct payments programme In the beginning of the 1990s, agricultural policy in Switzerland underwent a major revision. Previous market intervention subsidies were transformed to direct payments to farmers, compensating for structural disadvantages (e.g. in mountain areas) and for special ecological and animal-friendly farming practices. The importance of the ecological payments programme was continuously increased over the years. Today 95% of Swiss farms adhere to this programme. To receive the direct payments, farmers have to comply with approximately 40 criteria. Balanced nitrogen and phosphate fluxes are two of these criteria. Other criteria relevant to soil management include crop rotation and pesticide use. Further criteria deal with biodiversity and animal welfare.

NUTRIENT BALANCES TO IMPROVE EFFICIENCY

The concept of the nutrient-balance calculations The general concept of the nutrient-balance calculation 'Suisse-Bilanz' is that the nitrogen and phosphate inputs to crop production (including grassland) in the form of manure, mineral and organic fertilizers should not surpass the nitrogen and phosphate requirements of the crops (Uebersax & Schuepbach 2004). The nutrients in manure are calculated on the basis of the guide values provided by the agricultural research stations on excretions of different livestock categories. For phosphate, no losses are assumed, and crop requirements are considered to equal crop uptake. For nitrogen, losses of 15 to 30% are assumed for different animal categories, to account for ammonia losses in houses and during manure storage. Of the nitrogen applied to crops in manure, 60% was originally counted as available to crops. This number was derived from experimental experience, assuming good or even best farming practice. The nitrogen requirements of crops are considered to be independent from the yield, while the phosphate requirements of crops are linearly adjusted to yield. The yield-dependent requirements of grassland are calculated via the roughage consumption (guide values) of the livestock held on the farm. The maximum exceedance of the balance tolerated for nitrogen and phosphate is 10%. Beyond this limit, the direct payment to the farmer can be strongly decreased. The balance is calculated each year by the farmer or extension officers. It can be checked by local or national authorities. When the nutrient-balance concept was originally introduced as a policy implementation tool, it was decided to keep the calculation procedure simple, so that the approach would be well understood by farmers and so that calculations could be made manually. Implicitly, it was accepted that the balance could not take into account detailed farm-specific structural or management conditions. Over the years there was a continuous flow of suggestions to differentiate the calculations, because the simplified approach would too strongly discriminate against specific farm types with conditions diverging from the standard conditions assumed. The most important changes that were consequently adopted were: (1) lower availability of N for solid manure used on pasture and arable crops, and a lower availability of N for liquid manure used on arable crops; (2) lower availability of N excreted during grazing;

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(3) a stronger differentiation between animal categories and housing systems with respect to N losses; (4) corrections for special feeding regimes (protein-reduced feed for pigs and poultry); (5) correction for non-fertilized grassland; (6) tolerance for losses in the roughage balance. These refinements have probably helped to build up the good acceptance of the balance approach by farmers. Nevertheless, they have also made the calculations more susceptible to manipulation, and the complex tool is now difficult to understand for persons not regularly working with it. This can harm the credibility of the method and has probably decreased the value of the nutrient-balance calculation as an awareness raising tool. At least during the first few years, it was observed that the more detailed the calculations became, the more they were also questioned by farmers who feared that their specific situation was not duly considered. Furthermore, the risk that the approach is considered too lenient by the public has also increased. With all these refinements, it has also become increasingly difficult to perform the balanced calculations manually. Excel- or Access-based PC programs are mostly used today.

Guide values for nutrient excretions of livestock A major revision of the guide values relating to the amount and composition of manure was carried out in the early 1990s. The new guide values were based on two important sets of data (Menzi & Besson 1995): (1) nutrient excretion values calculated for different livestock categories; (2) experimental results on the amount of slurry and solid manure produced in different housing systems by major livestock categories. Guide values on the composition of different types of manure were derived from these two datasets. Nitrogen volatilization and losses in animal houses and during manure storage were also considered. Guide values were preferred to direct measurements of manure composition, because reliable manure analyses are difficult due to sampling problems and seasonal variations in the manure composition. Nutrient excretions (N, P, K, Mg and Ca) were determined from mass balances of inputs in feed minus retention in animal growth and products (e.g. milk, eggs). The main challenge of this approach was to get a reliable estimate

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of typical diets. Official feeding recommendations and feed composition guide values were an important basis for the estimate, although it also had to be considered that actual feeding practices may differ considerably from recommendations. For important variables like the milk yield of dairy cows, correction functions of the base values were worked out. Such balance calculations were originally done for about 30 livestock categories. Today, excretion values are available for about 50 categories. The excretion and manure guide values were first published in 1994 (Walther et al. 1994). A first revision was published in 2001 (FAL & R A C 2001), and the next is due for 2008. Periodic revisions are important to take into account major shifts in production (e.g. improved feed-conversion efficiency for pigs and poultry), especially if such guide values are used in policy implementation.

Effects o f the implementation o f equilibrated nutrient balance in the legislation Partly as a consequence of the nutrient-balance requirements in the direct payments programme, and partly due to the reduction of livestock numbers (which resulted in a decrease in the amount of nutrients in manure by about 20%), the amount of mineral fertilizer used has decreased between 1980 and 2002 by about 15% for nitrogen and 70% for phosphate. Contrary to the fears originally expressed by farmers, this reduction in nutrient inputs to crop production, which was considerable for P, did not lead to any obvious decrease in yields. The efficiency of nutrient use in Swiss agriculture has thus significantly improved. Jarvis & Menzi (2004), for example, have shown that the nitrogen-use efficiency of Swiss dairy farms is high as compared to the UK or The Netherlands. The requirements to balance nutrient inputs and outputs has certainly led to much more conscientious nutrient m a n a g e m e n t by a majority of the farmers. This is, for example, visible in the fact that farmers today are interested to know how they could reduce ammonia losses and that they try to improve their manure management. In general, the direct payments programme has considerably increased farmer's awareness of ecological issues, even though many farmers do complain that the limits imposed by the programme are too restrictive.

Identification of nutrient surpluses in Southeast Asia Background In developing countries and especially in Southeast Asia, livestock production is growing at unprecedented rates, driven by the increasing demand for animal products as a consequence of population growth, urbanization and income growth (De Haan et al. 1998). According to Delgado et al. (1999), the annual growth rate for meat production was 8.4% in China and 5.7% in Southeast Asia between 1982 and 1994. Between 1997 and 2020, the average annual growth of consumption is predicted to be 3.0% in China, 3.3% in Southeast Asia and 2.8% for the developing world. The livestock sector responds to the increasing demand for livestock products by structural changes, including: intensification and specialization of production; shift towards monogastric species; rapid growth in the size of production units; and concentration in areas with favourable transport infrastructure and good access to markets and services (De Haan et al. 1998). With this development, the environmental impacts of livestock production are becoming a more and more serious problem, especially in areas around large urban centres in Southeast Asia. As livestock and crop activities are disconnected, recycling of livestock excreta becomes more difficult. This development is reinforced by the shift from solid to liquid manure systems, as observed in pig production. While there is a market for solid manure in most Asian countries, there is no experience with recycling liquid livestock wastes. Due to the very high dilution of the slurry, resulting from cleaning and cooling activities, such a practice would also be economically uninteresting because of the large costs of storage and transport. The liquid wastes are therefore mostly released to water courses or wasteland. This results in serious water and soil pollution. Even if manures are recycled on agricultural land, doses are often excessive, and therefore lead to a strong accumulation of nutrients and heavy metals in the soil, which in the long term will harm soil fertility (Menzi 2005). Aiming at enhanced policy-making, the International Livestock, Environment and Development Initiative (LEAD), coordinated by FAO, develops decision-support tools for assessing livestock-environment interactions, for early warning, activity targeting and decision-making. In this context, balance

NUTRIENT BALANCES TO IMPROVE EFFICIENCY calculations were used to assess the contribution of livestock to nutrient fluxes for the whole of Southeast Asia, especially in the areas with nutrient surpluses. Nutrient-balance

results

Based on national and subnational statistics and spatial modelling, soil-surface phosphate balances were calculated at the pixel level. Details of the methods are given by Gerber et al. (2005). The phosphate balance is very diverse across the studied area. Areas with a negative phosphate balance (more than a 10% deficit) are mostly located in western China and South and West India and Bangladesh (Fig. 1). This is important information with regard to the poten-

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tial phosphate depletion of soils. In India, this situation is probably related to the scarce use of chemical fertilizers, and the high yields. For the whole study area, 39% of the agricultural land were estimated to be in a balanced situation (-10% to +10%). Nutrient surpluses were identified in northeast India, East China, the coast of Vietnam, Java Island, and Central and North Thailand, with especially high surpluses at the periphery of urban centres. For the whole study area, it was estimated that 24% of the agricultural land were subject to phosphate surpluses. High (more than 20 kg of P205 per ha of agricultural land) and very high (more than 40 kg of P205 per ha of agricultural land) phosphate surpluses were calculated for 15% and 4% respectively of the study area.

Fig. 1. Estimated phosphate (expressed in kg P205 ha-x) mass balance in selected Asian countries - 1998 to 2000. Source: Gerber et al. (2005).

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Nutrient-balance surpluses can result from both livestock excretions and mineral fertilizers. To identify the primary source of surplus in different parts of Asia, the relative contributions of livestock excreta and mineral fertilizers were analysed in all areas with a phosphate surplus higher than 10 kg P205 ha -1. In most of the areas with a phosphate surplus, mineral fertilizers represent the bulk of the phosphate load. According to prior studies (FAO 2001), these areas are mostly lowlands in which rice is the dominant crop: i.e. the Ganges basin, eastern and southern Thailand, the Mekong delta, and eastern China. Manure represents more than half of the phosphate surplus in northeastern and southeastern China, Taiwan, and at the periphery of urban centres such as Hanoi, Ho Chi Minh, Bangkok and Manila, which are characterized by high densities of monogastric species and limited crop production. These observations suggest that there is a high potential for better integration of crop and livestock activities. In areas receiving excessive amounts of phosphate, a proportion of the chemical fertilizers could be replaced by manure, decreasing substantially the environmental impacts on land and water. While such substitution may seem to be the logical solution in theory, its practical implementation at provincial and farm level raises a series of issues and constraints. A d e q u a t e action, including policy formulation, and development of technological packages, are required to facilitate the process.

Nutrient-flux model for intensive livestock production in Southeast Asia Drawing on the results of the regional analysis, L E A D launched a series of pilot projects to develop and implement new systems of AreaWide Integration (AWI) of specialized livestock and crop activities. The programme aimed at developing ways to 'remarry' livestock and crop production on a regional scale, in order to enable sustainable manure management without losing the economies of scale. National projects were carried out in Thailand (Rattanarajcharkul et al. 2000), China (Fang et al. 2000), Vietnam (Dan et al. 2004) and Mexico. These projects had the following aims: (1) awareness-raising about the problem of the environmental risks of livestock production, especially for specialized livestock farms without sufficient land to recycle the manure;

(2) interpretation of the present situation and evaluation of potential solutions (manure recycling and treatment); (3) development and demonstration of appropriate techniques for good practice in manure management; (4) initiate the formulation and implementation of appropriate policy to mitigate the environmental risks of livestock production (discharge standards, land-use planning, etc.); (5) capacity building for the introduction of new methods and policies for manure and nutrient management in agriculture at the regional level (in the whole of SE Asia); (6) initiate a regional network in SE Asia of policy-makers, scientists and agricultural service extension officers, working in the context of manure and nutrient management. To achieve these aims, to monitor results and to support farmers in the correct management of manure and fertilization, tools were needed to quantify nutrient fluxes and balances, manure quality and value, and environmental effects. To satisfy these needs, the nutrient-flux model NuFIux-AWI was developed.

The general concept o f the m o d e l The 'terms of reference' for the nutrient-flux calculation model were quite diverse: the model should be a user-friendly and easy-to-update tool which can cope with different levels of detail and reliability of input data (e.g. regional calculations with standard values, as well as farm-specific calculations). It should calculate nutrient balances in the context of policy formulation and implementation, but should also supply a planning aid to improve manure management on progressive farms. Finally, the model should be easy to adapt for use in different countries. The model was designed to calculate fluxes for pigs and poultry in more detail than for ruminants, because monogastric species contribute more than two-thirds of the manure in most project regions, and because detailed information on ruminant rations are difficult to obtain. The model starts the calculations with the most reliable available information: production parameters for livestock (beginning and end weight, amount and composition of feed used), nutrient requirements (recommendations) for different crops, etc. As shown schematically in Figure 2, the steps for the calculation are then: (1) nutrient excretions of different livestock categories; (2) amount and composition of fresh

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Fig. 2. Schematic structure and path of calculation of the model NuFlux-AWI (Menzi et al. 2002).

manure; (3) amount and composition of manure available for crops; (4) nutrient requirements of crops; and (5) nutrient balance. The balance is calculated for N, E K, Mg. Ca, Cu and Zn. A special module provides a tool for planning the distribution of the different types of manure to different crops. The model is equipped with default values for all the relevant variables concerning livestock and crop production. Different sets of default values were established for each project area, in conjunction with local experts. The whole set of default values can be replaced when introducing the model in a new region. These default values allow simple calculations in which only animal numbers and crop surfaces need to be known. On the other hand, if more-specific information is available, each default value can be replaced on the user-friendly input screens. The language can be switched during the calculation. More-specific information about the model is given by Menzi et al. (2002).

Reliability o f the m o d e l The accuracy of the model depends on the reliability of the available input data. While the

results on nutrient excretions are probably already within an accuracy of _+20% or better for intensive pig and poultry production, the results on manure quantities and composition are not very reliable yet, mainly because practically no information was available on the distribution of the excreta to solid and liquid manure, storage losses (e.g. lagoons) for the project regions, or comparable conditions. Nevertheless, the calculations can give a first approximation of important variables like the dilution of the pig slurry or the approximate amount of slurry and solid manure that has to be handled. Hopefully, measurement results from projects will make it possible to continuously improve the reliability of the results. The iterative validation and updating of the model should therefore be continued.

Potential use o f the m o d e l So far, the model has been used to assess the nutrient balance and manure management situation of pioneer farms in the project or on an area-wide scale in different countries (Thailand, China, Vietnam, Myanmar, Mexico, E1 Salvador) and to furnish input data for GIS

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procedures to produce maps of the nutrientbalance situation. In these applications, the awareness-raising effect was the most i m p o r t a n t outcome. N u t r i e n t - b a l a n c e calculations have without doubt c o n t r i b u t e d greatly to the impressive increase in awareness of the potential environmental risks of intensive livestock production, an increase that was observed in the course of the project activities in the different countries, both at the policy-maker level and at the farmer and agricultural extension service level. This corresponds with the observation of Scoones & Toulmin (1998) that n u t r i e n t budgets and balances can serve as useful devices to encourage d e b a t e about soil-fertility issues among policy-makers, scientists and farmers. With this tool, in collaboration with specialists, farmers will be able to estimate how much animal waste they could recycle to their crops in a sustainable way. Monitoring the collaboration of livestock and crop farms for manure recycling using the model will be complicated, because of the large n u m b e r of small crop farms needed to recycle the manure of a large pig farm having, e.g. over 10 000 pigs. Practical use of the model to estimate the amounts and composition of m a n u r e s would only be possible after a thorough validation and improvement of the corresponding calculation modules. A n important application of the model could be the assessment of the livestock carryingcapacity of an area when granting permits for new livestock operations. In the different countries involved in the project, it is well recognized that in the future new livestock operations should be guided away from current areas of high animal densities. New farms should be created in areas which still have a low livestock density and a large nutrient demand from crop production.

Conclusions Nutrient balances can be a valuable tool to assess and visualize w h e t h e r or not farming systems are in equilibrium with respect to nutrient inputs and outputs. They can increase stakeholders' awareness with respect to potential nutrient deficits or surpluses situations both in developed and in developing countries. By supporting n u t r i e n t recycling, n u t r i e n t balances also contribute to a more efficient use of non-renewable resources in agriculture, such as fossil-fuel and phosphorus resources. Nevertheless, to be used for such purposes, the general approach must be relatively simple, so that it is understood by farmers or policy-makers.

Nutrient balances can help to identify hot-spot areas which merit special interventions by policymakers and research. Furthermore, they can directly support policy formulation (e.g. spatial planning - zoning and land/livestock balances) and policy implementation (results monitoring). In view of t h e s e p o t e n t i a l uses, n u t r i e n t balances can be considered as a promising tool to guide agricultural production, and especially livestock production, into a sustainable direction, with regard to soil fertility, water and air pollution control and generally the efficient use of resources.

References BINDRABAN, RS., STOORVOGEL, J.J., JANSEN, D.M., VLAMING, J. • GROOT, J.J.R. 2000. Land quality indicators for sustainable land management: proposed method for yield gap and soil nutrient balance. Agriculture, Ecosystems and Environment, 81, 103-112. BROUWER, EM., GODESCHALK, EL., HELLEGERS, EJ.G.J. & KELHOLT,H.J. 1995. Mineral Balances at Farm Level in the European Union. Agricultural Economics Research Institute (LEI-DLO), The Hague, The Netherlands. DAN, Y.T., HOA, T.A. ET AL. 2005. Animal waste management in Vietnam - problems and solutions. In: BERNAL, M.P., MORAL, R., CLEMENTE, R. & PAREDES,C. (eds) Proceedings, llth Conference of the 'Recycling Agricultural, Municipal and Industrial Residues in Agriculture Network' (RAMIRAN), Murcia, Spain, 6-9 October 2004, Vol. 2, 237-340. DE HAAN, C., STEINFELD,H. & BLACKBURN,H. 1998. Livestock and the Environment, Finding a Balance. European Commission Directorate-General for Development, Food and Agricultural Organization of the United Nations, Washington DC. DELGADO, C., ROSEGRANT,M., STE1NFELD,n., EHUI, S. & COURBOIS,C. 1999. Livestock to 2020. The Next Food Revolution. International Food Policy Research Institute, Food and Agriculture Organisation of the United Nations, International Livestock Research Institute, Washington DC. FAL & RAC (Eidg. Forschungsanstalt ftir Agrar/3kologie und Landbau, Zurich and Eidg. Forschungsanstalt for Pflanzenbau, Nyon) 2001. Grundlagen fiir die Duengung im Acker- und Futterbau. Agrarforschung, 8 (6, special issue), 80 pp. FANG, Y., YANG,J.S., KJAER, S.S., GERBER, P., KE, B.S. & MENZI, H. 2000. Area-wide integration of specialised livestock and crop production in Jiangsu Province, China. Proceedings, Conference of the F A O / E S C O R E N A Network on Recycling Agricultural, Municipal and Industrial Residues in Agriculture (RAMIRAN), Gargnano, Italy, 6-9 September 2000, 287-293. FAO 2001. Global Forest Resources Assessment 2000. FAO, Rome.

NUTRIENT BALANCES TO IMPROVE EFFICIENCY FROSSARD, E., JULIEN, P., NEYROUD, J.-A. & SINAJ, S. 2004. Phosphor in BOden: Standortbestimmung Schweiz. Schriftenreihe Umwelt Nr., 368, Bundesamt for Umwelt, Wald und Landschaft (BUWAL), Berne. GERBER, P., CILONDA, P., FRANCESCHINI,G. & MENZI, H. 2005. Geographical determinants and environmental implications of livestock production intensification in Asia. Bioresource Technology, 96, 263-276. GOODLASS, G., HALBERG, N. & VERSCHUUR, G. 2003. Input output accounting systems in the European community - an appraisal of their usefulness in raising awareness of environmental problems. European Journal of Agronomy, 20, 17-24. HERRING, J.R. & FANTEL, R.J. 1993. Phosphate rock demand into the next century: impact on world food supply. Nonrenewable Resources, 2 (3), 226-246. JARVIS, S.C. & MENZI, H. 2004. Optimising best practice for N management in livestock systems: meeting production and environmental targets. Grassland Science in Europe, 9, 361-372. KELLER, A., ROSSIER, N. & DESAULES,A. 2005. Heavy Metal Balances of Agricultural Soil Monitoring Sites NABO - Swiss Soil Monitoring Network (English summary; text in German). Schriftenreihe der FAL, Nr. 54. Agroscope FAL Reckenholz, Eidg. Forschungsanstalt far Agrar6kologie und Landbau, Zurich. MENZI, H. 2005. Need and implications of good manure and nutrient management in extensive livestock production in developing countries. In: Livestock Waste Management in East Asia Project Preparation Report, FAG, Rome, Italy. MENZI, H. & BESSON, J.-M. 1995. Bases des nouvelles valeurs indicatives sur la production et la composition des engrais de ferme. Revue Suisse d'Agriculture, 27, 57-62. MENZI, H., ROTTIMANN, L. & GERBER, P. 2002. NuFlux-AWI: a calculation model to quantify nutrient fluxes and balances of intensive livestock production in developing countries. Proceedings, lOth Conference of the FAO/ESCORENA Network on Recycling Agricultural Municipal and Industrial Residues in Agriculture (RAMIRAN), Strbske Pleso, Slovak Republic, 14-18 May, 137-142. OECD 2001. OECD National Soil Surface Nitrogen Balances, OECD. OENEMA, O., KROS, H. & DE VRIES, W. 2003.

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Approaches and uncertainties in nutrient budgets: implications for nutrient management and environmental policies. European Journal of Agronomy, 20, 3-16. RATTANARAJCHARKUL,R., RUCHA, W., SOMMER, S. MENZI, H. 2000. Area-wide integration of specialised livestock and crop production in Eastern Thailand. Proceedings, Conference of the FAO/ESCORENA Network on Recycling Agricultural, Municipal and Industrial Residues in Agriculture (RAMIRAN), Gargnano, Italy, 6-9 September 2000, 295-300. SALEEM, M.A.M. 1998. Nutrient balance patterns in African livestock systems. Agriculture, Ecosystems and Environment, 71, 241-254. SCHUEPBACH,H. 2002. Calculating P balances at farm level - a comparison of methods. Proceedings of 'The Development of a Risk Assessment Methodology for Predicting Phosphorus Losses at the Field Scale'. Joint Meeting of the working groups 1 and 2 of the COST action 832 from 16-19 October, Swiss Federal Institute of Technology (ETH) Zurich, Zurich. SCOONES, I. & TOULMIN, C. 1998. Soil nutrient balances: what use for policy? Agriculture, Ecosystems and Environment, 71, 255-267. SMALING, E.M.A. & OENEMA, O. 1997. Estimating nutrient balances in agro-ecosystems at different spatial scales. In: LAL, R., BLUM,W.H., VALENTINE, C. & STEWART,B.A. (eds) Methods for Assessment of Soil Degradation. CRC, Boca Raton, FL, 229-252. STEWART, W.M., HAMMOND, L.L. & VAN KAUWENBERGH, S.J. 2005. Phosphorous as a natural resource. In: SIMS, J.T. & SHARPLEY, A.N. (eds) Phosphorus: Agriculture and the Environment. American Society of Agronomy, Monographs, 46. TILMAN,D., CASSMAN,K.G., MATSON,EA., NAYLOR,R. 8z POLASKY,S. 2002. Agricultural sustainability and intensive production practices. Nature, 418, 671-677. UEBERSAX, A. & SCHUEPBACH, H. 2004. Nutrient balancing at the farm level: the Swiss experience. Grassland Science in Europe, 9, 1193-1195. VON STEIGER, B., KELLER, A. c~z SCHULIN, R. 1998. Regional mass flux balancing for controlling gentle soil remediation operations. Nutrient Cycling in Agroecosystems, 50, 303-306. WALTHER, U., MENZi, H. ET At. 1994. Grundlagen fiir die Dgmgung im Acker- und Futterbau. Agrarforschung, 1/7, 1-40.

Perspectives on the relationship between soil science and geology E C. U G O L I N I 1 & B. P. W A R K E N T I N

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1University of Florence, Department of Soil Science and Plant Nutrition, Florence, Italy (e-mail." fugolini@unifi, iO 20regon State University, Department of Crop and Soil Science, Corvallis, Oregon, USA Abstract: Soil science came into its own only in the 20th century. Before this, the study of

soils was dominated by geologists, agronomists and chemists. It was Dokuchaev in 1886, who recognized soil as a physical entity with properties acquired from the impact of soil-forming factors, among which the geological substrate was only one. This vision resulted in the establishment of a new discipline, called pedology. With time, geologists began to appreciate soil in a pedological context. In fact, palaeosols in particular have been utilized to interpret the stratigraphy of metamorphic and sedimentary rocks and Quaternary deposits. Also, palaeosols have been used for correlating unconsolidated sediments, faults and neotectonics, or for the relative dating of deposits or surfaces. Weathering is a field where soil chemists have interacted with geochemists to evaluate chemical denudation and landscape evolution. Geological engineering in terms of water storage, pollutant transport, and critical load, in addition to location, design and construction of roads, is another area of interaction between soil researchers and geologists. The exploration of the planets of the solar system is a field which has assembled soil chemists and geochemists to collect, analyse and interpret data sent by space vehicles. Future interactions between geology and soil science will occur on issues such as: water in the vadose zone; risks due to Earth movements; and functions of soils in ecosystems. We predict and also welcome more communication between the two disciplines, as solutions to some of these problems are demanded by society.

Interaction between soil science and geology, as for other disciplines, has been influenced by the historical development of the sciences. During the formative stages of the natural disciplines, mostly in the e i g h t e e n t h and n i n e t e e n t h centuries, there has been a tendency for disciplines to evolve into separate entities. This trend, now called reductionism, has been very successful, and the individual disciplines: biology, physics, chemistry, geology and others were able to achieve r e m a r k a b l e advances. However, this progress was o b t a i n e d at the expense of an integration of the E a r t h sciences. Soil science is a relatively young discipline that came into its own only in the twentieth century. Until the end of the 1800s, the prevailing concepts of soil science were dominated by the geological, agronomical and chemical points of view (Joffe 1949). The geologists considered soils as an admixture of weathered rocks and organic debris; and the agronomists saw soil as a medium for plant growth, while the chemists saw it as a chemical laboratory where chemical decomposition and synthesis occurred. We had to wait until 1886 when Dokuchaev, a Russian mineralogist, recognized soil as a physical entity on the surface of the Earth, with properties acquired through the impact of the soil-forming

factors, among which the geological substrate was only one (Dokuchaev 1886). This vision was responsible for the establishment of a new discipline called pedology. Because of the language barrier, the ideas of D o k u c h a e v penetrated very slowly in the Western world. In fact, in Europe and North America, in the absence of a specifc school of pedology, the teaching of soil science was in the hands of geologists. The book written by Shaler (1891), a geologist, had a great impact in N o r t h America, while in Western E u r o p e the instruction of R a m a n n (1905) prevailed. In the U S A , eminent soil scientists, such as Hilgard (1893) (who, independently from Dokuchaev, recognized the role of climate in soil formation) and Marbut, Director of the US Soil Survey Division, were b o t h geologists. Initially, soil maps for use in agriculture were based on surface geology; in E u r o p e the relationship between soil and surficial deposits was stronger than in Russia and North America, where the climate was more important. The turning point in the establishment of soil science as an i n d e p e n d e n t discipline, came in 1927 in connection with the First International Soil Congress held in Washington, DC. O n this occasion, the R u s s i a n ideas, d o c u m e n t e d by p a p e r s written in E n g l i s h on the R u s s i a n

FROM:FROSSARD,E., BLUM,W. E. H. & WARKEYrIN,B. P. (eds) 2006. Functionof Soilsfor Human Societiesand the Environment. Geological Society, London, Special Publications, 266,183-190. 0305-8719/06/$15 9 The Geological Society of London 2006.

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contributions made to pedology, were distributed widely, and the pedological concept of soil became universally acknowledged (Academy of Sciences of the USSR 1927). The separation of the two disciplines of soil science and geology did not, however, represent a negative event for the interaction between these two fundamental fields of Earth science. As will be discussed later, the pedological concept of soil became a valuable tool for geologists to decipher time of exposure, palaeoclimate and vegetation assemblage in the Quaternary and older deposits. Also, soil became important for correlating unconsolidated sediments, faults and neotectonics. Soil chronosequences have benefited both pedologists and glacial geologists when assessing the rate of soil formation and correlating the relative age of glacial events. Geologists also have utilized palaeosols in metamorphic and sedimentary rocks for reconstructing the environment of the Earth since Archaean times (Retallack 1990, 2001). Interaction among soil specialists and geologists has also occurred in other areas, such as geochemistry, geological engineering and planetary sciences.

Palaeosols in metamorphic and sedimentary rocks Investigation of soils or of palaeosols in metamorphic and sedimentary rocks has been comprehensively presented by Retallack in the two editions, 1990 and 2001 of his book Soils of the Past. In this work, pedology is supplemented by other disciplines such as palaeontology, palaeobotany, geochemistry and planetary sciences for reconstructing the geological history of the Earth's landscape from the Archaean to the present day. The study and interpretation of palaeosols in metamorphic and sedimentary rocks is not an easy task for either pedologists or geologists, because, upon burial, the original soils have been altered by compaction, cementation, metamorphism, groundwater, hydrothermal fluids, erosion and other processes or events. Upon burial, the material of these palaeosols has been exposed to regimes different from the near-surface conditions existing when the soil was originally forming. When interpreting lithified geological strata that were once a soil, it is crucial to recognize soil features. There are specific soil characteristics that betray soil-forming episodes. According to Retallack (1990), these are root traces, soil horizons and soil structure. Fossil roots can be used only for those geological periods where

plants existed; they cannot be found in palaeosols older than the Silurian. Soil horizons are valid tools for reconstructing palaeopedological processes and events. Field techniques used by pedologists to describe soil horizons and profiles of modern soils are also valid for palaeosols. The contact with other horizons is important; from this observation, normal soil development or truncation due to erosion can be deduced. Retallack (1990) has always found that these features are preserved even under moderate metamorphism. In palaeosols, generally the upper part of the soil profile is truncated by erosion, which creates particular problems when classifying these soils according to modern soil classifications, where characteristics of the epipedon are necessary to assign the soil to different taxa (Birkeland 1999). Other characteristics such as clay skins, concretions, nodules and their composition can help in the environmental reconstruction. What cannot be determined in the field must be obtained through laboratory analyses. On the other hand, the influx of supergenic or hypogenic fluids may vitiate the significance of the analytical data. Despite these limitations, well-identified palaeosols represent good stratigraphic markers for recognizing unconformities over large areas, as is the case for a Cretaceous palaeosol discussed by Sigleo & Reinhardt (1988). In Germany, the Wurzelboden, a palaeosol in the Ruhr region, represents a useful guide to the stratigraphy of Carboniferous rocks (Roeschmann 1971). The Violet horizon, a palaeosol in the Permian and Triassic formations, is an important marker all over southern Germany and the Vosges (Ortlam 1971). In NW Nigeria, a palaeosol consisting of indurated ptinthite is present in early Tertiary and Cretaceous sediments (Sombroek 1971).

Use of Soils in Quaternary stratigraphic studies Soil used in the subdivision of Quaternary sediments has proven to be of great help for distinguishing and correlating climatic events and associated environmental changes. Soils represent hiatusus in the deposition of sediments, while their profiles and properties recapitulate the conditions under which they were formed. Soil stratigraphy is defined by Birkeland (1999) as: 'the use of soils in defining a local stratigraphy succession, estimating the ages of units associated with the soils and suggesting short-or long-range correlation'. For the geomorphologist, deciphering the processes responsible for

INTERACTION IN SOIL SCIENCE AND GEOLOGY the formation of buried soils requires a deep knowledge of soils, in addition to the ability to recognize the nature of the sediments, their mode of transport and deposition, and the geomorphological setting where the soil was formed. Historically, the presence of weathering profiles was recorded by geologists in the early 1900s in the Pleistocene glacial tills, loess and other sediments of the Upper Mississippi Valley (Ruhe 1969). A systematic study of buried soils in the pedological context came only later. For these investigations, we owe much to a number of geomorphologists for having introduced soils for separating and correlating Quaternary deposits. The foremost names that come to mind for this research include: Frye and his associates at the Illinois Geological Survey; Ruhe in Iowa; Richmond in the Rocky Mountains; and Morrison in the western US; Richmond in the La Sal Mountains, Utah, used soils and their degree of development to delineate major and minor interglacial periods (Richmond 1962). In addition to the contribution to glacial history, the work of Richmond represents valuable research in pedology. Given that the deposits were arranged along an altitudinal transect, he was able to filter the effect of altitude from the effect of time. Morrison (1964,1967) subdivided Pleistocene deposits of the western US on the basis of soils. To distinguish the soil in the context of a stratigraphic sequence, Morrison (1967) introduced the term geosol. The geosols are buried soils that must occupy a given stratigraphic position and must be traceable laterally on the landscape. Another term, pedoderm, was also suggested - being equivalent to geosol, but having the advantage that it could be relict and exhumed (Birkeland 1999). The Illinois State Geological Survey pioneered the use of soils for interpreting the Quaternary stratigraphy of the State (William & Frye 1970). The field observations were amply supported by analytical data from the buried soils and sediments, such as particle size, carbonates and clay mineralogy, in order to establish the degree of soil development and the correlation among the deposits. The major soils intercalated in deposits appeared to be depleted in carbonates and enriched with expanding clay minerals. The work of Ruhe is summarized in his book: Quaternary Landscape in Iowa (Ruhe 1969). Ruhe was a Quaternary geologist with an excellent knowledge of soils, and was thoroughly familiar with the mid-continent glacial history. As a good Quaternary geologist, he reconstructed the environmental conditions of Iowa during the Pleistocene using not only

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soils but also palaeobotanical evidence within a framework of radiocarbon dates (Ruhe 1969). In time-controlled soil stratigraphy, pedologists can learn about the evolution of certain soil parameters with time and the possible effect of climatic changes. In turn, they can also help the geomorphologists to solve apparent discrepancies in the degree of soil development and time of exposure. A n example of this interplay between geomorphologist and pedologists is the interpretation of the Churchill geosol and a geosol in Lake Bonneville Basin (Morrison 1991; Scott et al. 1983; Chadwick & Davis 1990). A good combination of expertise in both pedology and Quaternary stratigraphy is needed for long-range correlation of geosols. Soils, as a pedologist knows, may show considerable variability within a landscape. Even in a relatively short distance, one can find welldrained, poorly drained, eroded and depositional soils. The same can be expected in buried soils belonging to the same stratigraphic unit, even if of about the same age or time of exposure to soil formation, but developed under different positions on the landscape. Faulting in Quaternary deposits attracted considerable attention regarding the suitability of sites for construction of nuclear plants, for proposed nuclear-waste repository locations, and for controlling the stability of terrain near dams (Douglas 1980; Birkeland 1999). In these situations, soils played an important role for identifying stratigraphic units; for relative dating of the deposits according to their degree of development; and for estimating the recurrence of episodes of faulting (Machette 1978, 1988; Birkeland 1999). The interpretation of Quaternary soils is also undermined by possible post-burial alteration, although the problem is less acute than for the pre-Quaternary soils. Still, percolating solutions can recharge the palaeohorizons with bases and carbonates, while erosion can truncate the profiles. Buried soils, even if well preserved, can only give the relative age of the deposits. Radiometric measurements, amino acids, cosmogenic isotopes, or other suitable methods can provide numerical data.

Geochemistry and soil chemistry An area where soil science and geology have interacted and mutually benefited is weathering - specifically chemical weathering. The empirical approach to chemical weathering proposed by Goldich in 1938 (Goldich 1938) was followed by other geologists: e.g. Pettijohn (1957), Loughnan (1969) and Carroll (1970). Later, the

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Goldich concept was used by Jackson, a wellknown soil chemist, to arrange clay and claysized minerals in soils in a weathering sequence where the end-members represented the leastand the most-resistant minerals (Jackson et al. 1948). This approach, in time, appeared inadequate, above all because weathering rates and leaching rates were considered to be equivalent. Garrels & Christ (1965) placed the weathering reactions among the different minerals on a thermodynamic basis and inaugurated an entirely new field. Knowing the free energy of the reactions, it was possible to build lines or fields of stability of minerals in equilibrium with natural solutions of known composition. Varying the composition of the solution, it was possible to establish the thermodynamic stability or instability of the minerals in a given weathering environment. Kittrick, a soil chemist, spent time at Harvard with Garrels to explore the application of this method to the soil system. Later, Kittrick (1969) applied the new ideas to the stability of clay minerals in relation to soil-solution composition. Lindsay (1979) also worked on chemical equilibria in soils and on stability diagrams that depicted the theoretical area of stability of minerals. Other geochemists interested in soil, e.g. Tardy & Garrels (1974), used chemical reactions and a thermodynamic approach to study weathering. The limitations to this approach were, however, recognized by both geochemists and soil chemists. Metastable minerals, even if not thermodynamically stable, could persist in the soil and in the weathering environment because they were kinetically favoured (Stumm 1992). The mineral equilibrium approach, however, had a large impact in soil science and geology; it stimulated the determination and refinement of free-energy values for soil minerals and gave impetus to the collection and analyses of soilsolution-, river- and stream-water. The chemical composition of surficial waters allowed geochemists to calculate the present-day rate of weathering for watersheds, e.g. Clayton 1986, and to evaluate human disturbances (Likens et al. 1977). Lysimeters emplaced in the soil to collect soil solutions made it possible to establish, using the thermodynamic approach, the mineral stability of newly formed minerals and the rate of weathering of fresh Mount St Helen's 1980 volcanic ash (Dahlgren & Ugolini 1989; Zabowski & Ugolini 1992; Dahlgren et al. 1999). Denudation of land, both physical and chemical depends, other things being equal, on the rate of weathering and soil formation. The stream solid load is supplied by the eroded soils, while the chemical denudation is the result of soil-forming processes and weathering of the parent

material. Geochemists, soil chemists and hydrogeologists are all involved in these studies. In the early 1980s, Ronald Amundson at Berkeley, University of California, initiated a series of studies using stable C-isotopes to address soil problems that the conventional methods could not solve. In pursuing this work, he successfully collaborated with geochemists, geologists and chemists. He made important contributions to the transport of 14CO2 in soil and how the radioactive signal is transferred to pedogenic carbonates. Also, he investigated the isotopic composition of soil CO2 and provided a model showing that the ~513C of pedogenic carbonates reflects the 513C of the biomass (Amundson 2004).

Geological engineering The use of soil for its engineering qualities is perhaps as old as its use for farming. Early humans built mud houses and made ceramic wares using soil material. In the context of geological engineering, the soil is seen as an unconsolidated earthy material that can be moved by machinery. This concept is not too different from the term 'Regolith', a mantle of unconsolidated rock material, whether residual or transported, including soil that overlies solid bedrock (Bates & Jackson 1984). The Soil Science Society of America, Special Publication 34 provides an extensive discussion on the term regolith (Cremeens et al. 1994). Although, in the US, common interests were shared between soil specialists and geological engineers, their interaction came only 25 years after the first soil-survey report was published in 1899 (Allemeier 1973). At that time, the mapping of soils fulfilled the purpose of showing on maps the type of soils that differed in crop response (Simonson 1989). Departure from this mission occurred in 1925, in Michigan, where soil information was also used in the design and construction of roads (Allemeier 1973). The use of soil-survey data by State Highways Departments became widespread, and facilitated a cooperative programme between the Bureau of Roads and the USDA (United States Department of Agriculture; Stokstad 1958). As a result of this cooperation, engineering-test data for the mapped soils were provided and included in the Soil Survey Reports (Simonson 1989). At present, the mapped units in the Soil Survey Reports include suitability and limitations ratings for a number of non-agricultural uses of soils. These limitations are for road locations, low building construction, recreation, and trafficability off-road and also for the use of soil for fills and embankments. In the United States,

INTERACTION IN SOIL SCIENCE AND GEOLOGY particularly, because of the prevalence of detached homes in rural and residential areas, not as yet linked to a sewer system, soil-survey information is very useful for assessing the suitability of soils for accommodating septic tanks (Simonson 1989). The suitability is based on the saturated hydraulic conductivity values. In addition, the County Soil Survey reports are provided with analytical data that are directly applicable to geo-engineering problems. Modern soil surveys are often provided with important parameters such as liquid limit and plasticity index; if not provided, both can be obtained by knowing the type of clay present in the soil (Ugolini & Wolf 2006). Particle-size analyses are also presented in the reports, including the percentages of several sand fractions, silt and clay. Although the U S D A separates differ from the other systems, a table comparing the USDA, the International, the Unified, the A A S H T O (American Association of State Highway and Transportation Office) and the Modified Wentworth Scale is available in the Soil Survey M a n u a l (Soil Survey Division Staff 1993) and in the field book for describing and sampling soils (Schoeneberger et al. 2002). Other observations of the soil surveyor, although not numerically expressed, can alert future users to a problem. This is the case for the presence of cracks, especially the reversible trans-horizon cracks common in Vertisols; they can indicate some relative magnitude of the coefficient of linear extensibility (COLE). Of considerable relevance to the geo-engineer is the presence of horizons or layers that can show different degrees of cementation or consolidation. These horizons or layers may display the hardness of a rock; they have a thickness of several centimetres and are distributed over large areas. Examples are petrocalcic, petrogypsic, duripan and fragipan horizons. As part of classifying the soils either in the US Soil Taxonomy or in the World Reference Base for Soil Resources, the moisture and the temperature regime are identified, giving a general environmental setting to every classified soil. In conclusion, it appears that Soil Survey Reports have a wealth of information either promptly usable or of indirect application. Soil maps, although not made to provide geotechnical data, can assist in the identification of surficial deposits in areas where geological maps are not available.

Planetary geochemistry and soil chemistry With the beginning of unmanned and manned exploration of the terrestrial-type plants of the solar system, a debate has arisen on the use of

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the term 'soil' for the apparently lifeless substrates of the Moon, Mars and Venus. Ecologists asserted that the term soil should be used exclusively for substrates that support life. Geologists or planetary geologists, however, referred to lunar and Martian soils. Soil specialists have also agreed on the term soil, because life is being found on this planet in many harsh environments, such as the Antarctic soils (Ugolini 1970; Cameron et al. 1971). At the beginning of the N A S A missions on the planets, the formation of scientific teams brought together specialists in geology, geophysics and soil physical-chemistry, in addition to biologists. Anderson, a soil physicalchemist, was a member of the team working on NASA's Viking Mission to Mars, and was responsible for the construction of a soil-water detector for finding water at the landing sites (Anderson 1982). Other soil scientists have also interacted with geologists to interpret the data from past and present planetary missions, and have provided hypotheses for life on Mars (Banin 1986; Banin et al. 1997).

Future interactions The nature of the Earth-science issues now facing society indicates that geology and soil science will share even more cooperative work in the future. Four issues are briefly discussed here as examples: the processes occurring as water moves through the vadose zone; the risks to human safety from earth movements; the role of soils and subsoil layers in the functions of ecosystems; and the need to communicate Earth-science information to a wide audience beyond the two disciplines. The vadose zone, reaching from the soil surface down to the water table, is where both water and air occupy the void volume, the zone of partial water saturation. Below is the zone of water saturation. The properties of the vadose zone - the area where the disciplines of soil science, geology and geotechnical engineering meet, have become important in controlling groundwater quality (Bergstr6m & Djodjic (2006). Water and chemicals from the surface pass through this zone of partial saturation; how they are modified, and the rate of movement, are critical to measurements and prediction in pollution control. Contaminants at the surface pose a threat to groundwater - the nature and magnitude of the threat can be predicted from the soil and rock layers in the vadose zone. The distribution of size and continuity of voids determines the flow path and velocity of water and contaminants. The nature of the minerals in the vadose zone determines sorption and

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exchange reactions of specific c o n t a m i n a n t molecules as downward flow proceeds. This may be rapid (centimetres per hour), or slow (millimetres per 100 years). The need to find long-term storage places for spent nuclear fuel has l e n g t h e n e d the time that must be considered. A new area of subsurface science, with contributions from the several disciplines, is developing. New journals, such as the Vadose Zone Journal, report studies from the different disciplines involved in subsurface science. U n d e r s t a n d i n g and predicting risks to the safety of humans and structures from E a r t h movements requires knowledge of both geology and soil science in geological engineering. As the structures that we build become more complex and expensive, their protection from E a r t h movements becomes a more important issue in both the technical and cultural roles of Earth scientists. Unstable layers near the Earth's surface can lead to landslides. The options that we have for coping with these events are determined by the properties of the Earth's mantle, and the knowledge of soil and geology. Soils are characterized by a large diversity in their properties over short distances. They cannot be modelled as uniform Earth material for stability predictions. Understanding and prediction of landslides requires understanding how the strength of soil layers responds to changing water content. The effect of positive or neutral water-pressures on soil strength is adequately described using the concept of effective stress from soil mechanics. Negative soil-water pressures are more difficult to measure and to use in soil-strength predictions. The concepts are regularly used in soil science (e.g. Iwata et al. 1994), and the mechanics of partly saturated soils have been worked out. In many landslide situations, the soil-water content changes quickly and frequently in the range of unsaturation. This often causes instability. More generally, society's increasing concern about e n v i r o n m e n t a l p r o t e c t i o n brings soil science and geology together in meeting ecological concerns. Both disciplines had begun responding to these concerns in the last third of the twentieth century. The ecological functions (Blum 1990; Blum et al., Paper 1 of this volume) at the Earth's surface include distribution of incoming water. Water that infiltrates the soil plays a very different role in the hydrological cycle to the water that moves across the surface. This is d e t e r m i n e d by properties studied by geologists and soil scientists. The health of ecosystems depends upon how these processes work; this information is available in the two disciplines.

The increasing involvement of geology and soil science with issues i m p o r t a n t to large segments of society, requires that people understand what these disciplines can tell us about the E a r t h ' s surface. The disciplines have the responsibility to provide that understanding, and to provide education for all levels, from schools to adult education. This is best provided as Earth-science education, rather than from the separate disciplines. Recently, Earth-science curricula have been prepared for different levels in schools and different audiences. Geology has t a k e n the lead in developing Earth-science materials suitable for different levels. This will need to become a high priority, so that the knowledge available in the Earth sciences can be used for h u m a n benefit. The first author, E C. U. is grateful to S. Pelacani for computer assistance.

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MACHETTE, M.N. 1978. Dating Quaternary faults in the southwestern United States by using buried calcic paleosols. US Geological Survey Journal Research, 6, 369-381. MACHETrE, M.N. 1988. Quaternary movement along the La Jencia fault, central New Mexico. US Geological Survey Professional Paper, 1440, 82 pp. MORRISON, R.B. 1964. Lake Lahontan: geology of the southern Carson Desert, Nevada. US Geological Survey, Professional Paper, 401. MORRISON, R.B. 1967. Principles of Quaternary stratigraphy. In: MORRISON, R.B. & WRIGHT, H.E, JR (eds) Quaternary Soils. International Association of Quaternary Research, VII Congress Proceedings, Vol. 9. MORRISON, R.B. 1991. Quaternary stratigraphic, hydrologic, and climatic history of the Great Basin, with emphasis on Lake Lahontan, Boneville and Tokopa. In: MORRISON, R.B. (ed.) Quaternary Nonglacial Geology, Conterminious U.S., The Geology of North America, Vol. K-2, Geological Society of America, Boulder, Colorado, 283-320. ORTLAM,D. 1971. Paleosols and their significance in stratigraphy and applied geology in the Permian and Triassic of southern Germany. In: YAALON, D.H. (ed.) Paleopedology, International Society of Soil Science and Israel University Press, Jerusalem, 321-327. PETTIJOHN, EJ. 1957. Sedimentary Rocks, 2nd edn, Harper & Row, New York. RAMANN, E. 1905. Bodenkunde, 2nd edn, Julius Springer, Berlin. RETALLACK, G.J. 1990. Soils of the Past. Unwin Hyman, Winchester, MA. RETALLACK, G.J. 2001. Soils of the Past. 2nd edn, Blackwell Science, Oxford. RICHMOND,G.M. 1962. Quaternary stratigraphy of the La Sal Mountains, Utah. US Geological Survey, Professional Paper, 324, 135 pp. ROESCHMANN, G. 1971. Problems concerning investigations of paleosols in older sedimentary rocks, demonstrated by the example of Wurzelboden of Carboniferous system. In: YAALON, D.H. (ed.) Paleopedology, International Society of Soil Science and Israel University Press, Jerusalem, 311-320. RUHE, R.V. 1969. Quaternary Landscape in Iowa. Iowa State University Press, Ames, Iowa. SCHOENEBERGER, P.J., WYSOCKI, D.A., BENHAM, E.C. & BRODERSON, W.D. (eds) 2002. Field Book for Describing and Sampling Soils, Version 2.0, Natural Resources Conservation Service, National Soil Survey Center, Lincoln, NE. SCOTT,W.E., McCoY, W.D., SHROBA,R.R. & RUBIN,M. 1983. Reinterpretation of the exposed record of the last two cycles of Lake Boneville, western United States. Quaternary Research, 20, 261-285. SHALER, N.S. 1891. The origin and nature of soils. US Geological Survey, I2th Annual Report 1890-91 pt 1,213-245. SIGLEO,W. & REn~rHARDT,J. 1988. Paleosols from some Cretaceous environments in the southeastern United States. Geological Society of America, Special Paper, 216, 123-142.

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Index Page numbers in italic denote figures. Page numbers in bold denote tables.

A-horizon, geochemistry, Portugal 161-169 Acaulospora laevis 95, 98 Acaulospora mellea 95 Acaulospora paulinae 91 Acaulospora scrobiculata 99 Acaulospora spp. 93, 95, 98,100, 101 Acaulospora trappei 98 acid sulphate soil 144-145 mosquitoes 145 acidity 144-145 aggregate, slaking 143 agriculture and civilization 9,12 conservation 31 ecological 13-14 historic, cultural soilscapes 127-129 historical perspective 10-11 intensive, N20 emissions 32-34 land-use change 26-27 sustainable 11-13 water reuse, San Joachin Valley, California 80-87 agroforestry 27, 29 see also Integrated on-Farm Drainage Management (IFDM) system agrogeology 133 alfalfa 81, 84 Alpine Convention 153-154 aluminium, Portuguese soil 162,163 ammonium 34 Anthrosol 126-127 arbuscular mycorrhizal fungi 89-105 and agricultural practices crop rotation 92-94,105 fertilizers 94-96,105 grazing and burning 99-100,105 irrigation 97-99,105 pesticides 96-97,105 soil sterilization 96-97,105 soil tillage 91,105 'cheaters' 93 distribution 90 functions 90 and nutrient uptake 90, 92, 97 pollution and heavy metals 100-102 reversal of human impact 103-104 soil compaction 102-103 and soil redox potential 98 soil stabilization 90 topsoil movement 103 archaeology and cultural soilscapes 126-127 landscape research 125 Arenosol, Portugal 160,161 Aristotle (384-322BC) Four Elements Theory 10

Atriplex lentiformis L. 81, 84 see also saltbush Australia, soil properties and environment 142-146

B-horizon 3 bacteria 6 methanogenic 25 Balfour, E.B. 13 The Living Soil 12 Below Ground Biodiversity Project 156 benomyl fungicide 96, 97 benzimidazole fungicide 96 biofue185, 86 biogas 47 biomass 4 biota, soil 1, 2-3, 4 Black, Joseph (1728-1799), discovery of CO2 15 bogs, CH 4 flux 35 boron in irrigated soil, San Joachin Valley 79-87 phytotoxicity 86 Brassica napus 81, 84, 85, 86 see also canola brown earths see Cambisol buffering 5 soil organic matter 13 burning see fire cadmium effect on arbuscular mycorrhizal fungi 101 toxicity, Belgium 145 caesium, effect on arbuscular mycorrhizal fungi 102 Calcic sol, Portugal 160,161 California, agricultural irrigation 79-87 Cambisol Portugal 160,161 Switzerland 63, 64, 67 Gleyic 64, 67 heavy metal EDPs 68-74 canola 81, 82, 84, 85, 86 captan fungicide 92, 96 carbendazim 96, 97 carbon black 28, 25 soil dynamics 16-17 soil organic matter 24-25 carbon balance 24-25 carbon isotope ratio, prehistoric agricultural studies 128 carbon nutrition, historical perspectives 10-11 carbon saturation 24 carbon sequestration 14-15, 30-31, 39 economic value 17-18 Carthamus tinctorius 81, 84

FROM: FROSSARD,E., BLUM,W. E. H. & WARKENTIN,B. P. (eds) 2006. Function of Soils for Human Societies and the Environment. Geological Society, London, Special Publications, 266,191-196. 0305-8719/06/$15 9 The Geological Society of London 2006.

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INDEX

see also safflower catch crops, N leaching control 50 CH 4 14-15, 23, 25-26 global flux 34-36, 37-39 measurement 24 transport 26 charcoal 25 'cheaters', mycorrhiza193 chloropicrin fumigation 96 chlorothaloni197 chromium effect on arbuscular mycorrhizal fungi 101 Swiss forest soil 63, 64, 65, 67-73, 76 enrichment depletion patterns 68

restoration 30 and SOC storage 29-30 denitrification 25 desertification 29 Dialeimmaso1121 dinitrogen monoxide 34 dispersion 143 Distichlis spicata L. 81, 84, 85 see also saltgrass drainage and N20 emmission 33 San Joaquin Valley 80, 81 see also wetlands dust, global 29

civilization, and agricultural failure 9,12 clay dispersion 143 in pore systems 2, 3 shrink-swell 143-144 climate change 7, 39 CO2 efflux 23 flux 15-16, 23 and land-use change 27-29 free air experiment (FACE) 16 global emission 37-39 greenhouse gas 14-15, 23 measurement 24 historical 15-16 Coast Ranges, California 79 Coccidioides 134 compost 12 conservation, soil 6-7, 30-31 construction industry effect on soil 141-142 SOC 6-7, 30 shrink-swell problems 144 soil sealing 117-119 see also engineering, geological contamination 135 and human health 145 copper effect on arbuscular myeorrhizal fungi 101 Swiss forest soil 63, 64, 65, 70, 73-74, 76 vine mortality 58-59 cord grass 81, 84, 85 cotton 81, 84, 85 Critical Zone studies 134-135 crop rotation effect on arbuscular mycorrhizal fungi 92-94 see also rotation, agricultural crusting 117, 142 see also soil sealing cultural heritage 6 cultural soilscape 6, 125-129 archaeology 126-127 soil chemistry 127-129

earth sciences, education college and university 136-139 school 135-136 ecohydrology 139 education, earth sciences college and university 136-138 school 135-136 Ekrano-pseudogley 121 Ekrano-reductosol 121 Ekranolith 121 Ekranoso1121,122,123 element balance 172-174 eluviation 58 engineering, geological 186-187 see also construction industry England, First Soil Action Plan, D E F R A 2004 152 enrichment, trace element 66 enrichment depletion patterns, Swiss forest soils 67-74 E n t r o p h o s p o r a infrequens 91 erosion sodic soil 142 topsoil 7, 29, 30, 31 ethnopedology 139 eucalyptus 81,84 European Union, Thematic Strategy for Soil Protection 2002 154-156 evaporation ponds, San Joaquin Valley 80 evaporator, solar 80, 81, 85

decomposition 4 anaerobic 25 deforestation 26-27, 28 D E F R A , First Soil Action Plan for England 2004 152 degradation 149 soil 6, 7

FACE (free air CO2 experiment) 16 FAO, Global Forest Resource Assessment 2000 26, 27 farm-gate balance 172 farming intensive 11-12 organic 12,13, 31, 46 and soil conservation 31 fenamiphos 97 fens, CH 4 flux 35 Ferralsol 2 fertility decline 12 ecosystem 10-19 fertilizer chemical 11-12 nitrogen 32-34 controlled-release 50 effect on arbuscular mycorrhizal fungi 94-96 nutrient balance 171-180

INDEX filtration 5 fingering 47 fire 25 effect on arbuscular mycorrhizal fungi 99-100 forest 28 First Soil Action Plan for England, DEFRA 2004 152 flow finger 47 pathways 47-49 preferential 47-49, 50 Fluvisol Portugal 160,161 Switzerland 64 forests conversion to pasture 28-29 fire 28 see also deforestation formaldehyde fumigation 96 fumigation 96 fungi 6 arbuscular mycorrhiza189-105 fungicide 96, 97 Gasparin, Adrien de (1783-1862), Cours d'Agriculture 11 geo-archaeology 126 geo-engineering see engineering, geological geochemistry 185-186 Portugal survey 159-169 geology agricultural 133 and soil science 183-188 Geosol 185 Germany Federal Soil Protection Act 1998 151-152 soil sealing 118,119 Gigaspora gigantea 93, 95 Gigaspora margarita 90, 95, 97 Gigaspora rosea 92, 97 Gigaspora spp. 93, 95, 96,101,103 gilgai soil 144 Gleyso128 Switzerland 63, 64, 67 heavy metal EDPs 68-74 global warming 14-15, 23 global warming potential 37 glomalin 90, 91, 92, 95 Glomeromycota 89 Glomus aggregatum 95 Glomus albidum 93 Glomus caledonium 91, 93 Glomus claroideum 92 Glornus clarum 90, 98 Glomus coronatum 97 Glomus deserticola 98 Glomus etunicatum 91, 93, 95, 97, 98,102 Glomus fasciculatum 95, 98 Glomus geosporum 97 Glomus gerdemannii 101 Glomus intraradices 92, 95, 98,101,102,103 Glomus leptotichum 95, 98 Glomus macrocarpum 93 Glomus manihotis 92 Glomus microcarpum 93 Glomus mosseae 93, 95, 97, 98, 99,101,102

193

Glomus occultum 91, 93, 95, 98 Glomus spp. 92, 98,100,101 Gossypium hirsutum 81, 84, 85 see also cotton

grass, salt-tolerant 81 grassland, land-use change 27, 29 grazing, effect on arbuscular mycorrhizal fungi 99-100 greenhouse gas 14-15 global warming potential 15 reduction strategies 39-40 and soils 23-41 integrated assessment 36-37 groundwater heavy metal contamination 63 nutrient contamination 46 San Joaquin Valley, California 79 halophytes 80, 81 Hassenfratz, Jean Henri (1755-1827) 10 health, human effect of contamination 145 and soil fertility 12 heavy metals 5, 7,145 effect on arbuscular mycorrhizal fungi 100-102 Swiss forest soil 63-77 effect on micro-organisms 75, 76 effect on soil fertility 75, 76, 77 effect on water quality 74-75 Helianthus anuus L. 81, 84, 85 see also sunflower Helmont, Johannes Baptista van (1577-1644) 10 herbicide 96, 97 Histoso128 holism 12-13 Howard, A., The Soil and Health 12 humus 4 historical perspectives 10-11, 16 hydrocarbon, polyaromatic 100,101 hydrogeophysics 139 hydropedology 139 illuviation 58 insecticide 96-97 Integrated on-Farm Drainage Management system 80 Intergovernmental Panel on Climate Change 23, 34 Guidelines for National GHG inventories 1996 32, 33 Intergovernmental Panel on Land and Soils 157 irrigation effect on arbuscular mycorrhizal fungi 97-99 San Joaquin Valley, water reuse 79--87 Kesterton Reservoir, Se toxicity 80 Kyoto Protocol 1997 15,156 land-use, effect on soil 141-142 land-use change 26-27, 141-142,145 CO2 flux 27-29 land-use, land-use change and forestry see LULUCF landscape research 125-126 Latuca sativa 81, 84, 85 see also lettuce Law of Return 12

194

INDEX

leaching 6 Massif de la Clape vineyards 58 nutrient 46-51 soil salts, San Joaquin Valley 79-80 lead effect on arbuscular mycorrhizal fungi 101 Swiss forest soil 63, 64, 65, 74, 76 enrichment depletion patterns 72 legumes, nitrogen fixing 32 Leptosol Portugal 160,161 Switzerland 63, 64, 67 heavy metal EDPs 68-74 lettuce 81, 84, 85 Liebig, Justus von (1803-1873) 10,11 Die organische Chemie in ihrer Anwendung auf Agrikultur und Physiologic 11

lightning, nitrogen fixing 32 livestock production, Southeast Asia 176-180 LULUCF (land-use, land-use change and forestry) 16 Lundegfirdh, Henrik (1888-1969) 15-16 Luvisol Portugal 160,161 Switzerland 63, 64, 67 heavy metal EDPs 68-74 Lycopersicon esculentum 81,84 see also tomatoes macropores 2, 3, 4 flow 48, 50 magnesium, effect on arbuscular mycorrhizal fungi 94 management, soil 7 manganese, effect on arbuscular mycorrhizal fungi 90, 102 manure 31, 33 anaerobic digestion 46--47 effect on arbuscular mycorrhizal fungi 94, 95 farmyard, effect on CO2 concentration 15 green 47 effect on arbuscular mycorrhizal fungi 94 N leaching 46-51 poultry 46 Southeast Asia 176-180 Suisse Bilanz nutrient balance 174-175 ultramicrofiltration 47 Massif de la Clape 54 copper concentration 58-59 soil system 56-59, 57 rehabilitation 59-60 vine mortality 53-60, 55, 60 Medicago sativa L. 81, 84 see also alfalfa metalaxyl fungicide 97 methamsodium fumigation 96, 97 methane see CH4 methanogenesis 25 methylbromide fumigation 96, 97 micro-organisms 6 mosquitoes, acid sulphate soil 145 N20 14-15, 23, 25 global emission 32-33, 37-39 intensive agriculture 32-34 measurement 24

reduction 34 rice agriculture 36 nanogeoscience 139 nematicide 96-97 neogeomorphology 139 nickel Portuguese soil 162-163,164,169 Swiss forest soil 63, 64, 65, 73, 76 enrichment depletion patterns 69 nitrification 25 inhibitors 34 nitrogen element balance 172,173,174-175 fertilizer 32-34, 46,171 controlled-release 50 effect on arbuscular rnycorrhizal fungi 94-96, 97 leaching 46 fixation 32 Suisse Bilanz nutrient balance 174-176 in terrestrial ecosystems 32 nitrous oxide see N20 nutrient balance 7,171-180 Southeast Asia 176-180 Suisse Bilanz 174-176 nutrient uptake, mycorrhizal symbiosis 90, 92, 97 nutrient-flux model, Southeast Asia 178-180 O-horizon, geochemistry, Portugal 161-169 ocean, terrestrial organic matter 39 organic matter 1,4 soil C stock 24-25 economic value 17-18 in ecosystem 9-19 functions 9,17 Oxisol 2, 3 oxygen, effect on arbuscular mycorrhizal fungi 98-99 oxytetracycline 92 Palaeosols 184 Palissy, Bernard (1510-1589) 10 Pararendzina see Regosol peatlands 27 accumulation 25, 27 burning 28 C stocks 28 CH 4 flux 35, 37-39, 40 CO2 flux 35, 37, 40 pedology 183-185 pesticides effect on arbuscular mycorrhizal fungi 96-97 solute transport 47, 48 phosphate 7 nutrient balance, Southeast Asia 177-178,177 Suisse Bilanz nutrient balance 174-176 phosphorus in Anthrosol 126-128 effect on arbuscular mycorrhizal fungi 94-99 element balance 172,173 leaching 4749, 48 limitation 33

INDEX pickleweed 81, 84, 85, 86 Planosol, Portugal 160,161 plant nutrition, historical perspective 10-11 plant-soil system 10, 11 Podsol Portugal 160,161 Switzerland 63, 64, 67 heavy metal EDPs 68-74 pollutants, effect on arbuscular mycorrhizal fungi 99-100 pollution agricultural 6, 7 reduction 49-50 polyaromatic hydrocarbons, effect on arbuscular mycorrhizal fungi 100, 101 pore systems 2-4 inner soil surface 3 pore size 3 soil processes 2-3 Portugal geochemical mapping 159-169 geochemical survey 159-169 soil types 160,161 potassium, effect on arbuscular mycorrhizal fungi 94 precipitation, and SOC 29 propiconazole 97 Quaternary stratigraphy 184-185 redox potential, effect on arbuscular mycorrhizal fungi 98 redox reactions 25 Regosol Switzerland 63, 64, 67 heavy metal EDPs 68-74 Rendzina see Leptosol rice agriculture, CH 4 flux 35-36, 37-39, 39 roots, and arbuscular mycorrhizal fungi 89 rotation crop 29, 30, 31 effect on arbuscular mycorrhizal fungi 92-94 safflower 81, 84 Salicornia bigelovii Torr. 81, 84, 85, 86 see also pickleweed salinity 143,144 see also salts salt-sensitive crops 80, 81, 82, 85 salt-tolerant crops 80, 81, 82, 85 saltbush 81, 84 saltgrass 81, 84, 85 salts, in irrigated soil, San Joaquin Valley 79-80 San Joachin Valley, California, toxic trace elements 79-87 Saussure, Horace B6n6dict de (1740-1799), Voyages clans les Alpes 16,16-17 Saussure, Nicolas Th6odore de (1767-1845), Recherches Chimiques sur la V~gdmtion 16 Sclerocystis rubiformis 100 Scutellospora calospora 95, 96 Scutellospora heterogama 93, 95, 98, 100 Scutellospora pellucida 91,103 Scutellospora spp. 92, 95,101,103 sealing see soil sealing

195

seeding, direct 31 selenium deficiency 86 in irrigated soil, San Joachin Valley 79-87 toxicity 86 sequestration, carbon 14-15, 30-31, 39 sewage, effect on arbuscular mycorrhizal fungi 101 shrink-swell 143-144 slaking 143 SOC see soil organic carbon sodicity 142-143,144 sodium, exchangeable cations 142 soil chemistry 185-186 in archaeological research 127,127-129 compaction, effect on arbuscular mycorrhizal fungi 102-103 conservation 6-7, 30-31 cultural functions 6 degradation 6, 7, 30-31 ecological function 4-6 as environmental interface 141 extra-terrestrial 187 formation 1 horizons 1, 2 geochemistry, Portugal 161-169 human impact 45,141-142 infiltration rate 85-86 inner surface 3 memory 126 properties 142 protection 149-157 Alpine Convention 153-154 England 152 European Union 154-156 Germany 151-152 international 152-157 USA 149-151 Quaternary stratigraphy 184-185 stabilization, arbuscular mycorrhizal fungi 90 sterilization effect on arbuscular mycorrhizal fungi 96-97 strength 142-143 surface balance 172 soil organic carbon effect of degradation 29-30 global stock 28 measurement 23 stock 16-17,17 storage 15 soil organic matter 4 agronomic function 10-13 buffering 13 C stock 15, 24-25 measurement 24 saturation 24 economic value 17-18 in ecosystem 9-19 environmental function 14-17 soil science 133-134, 183 education college and university 136-139 school 135-136 extra-terrestrial 187

196 and geology 183-188 professional societies 138 recruitment 138 soil sealing 30,117-123 effects 119 mitigation 120 sealed soil properties 120-123 soilscape, cultural 125-129 solarization 96 solutes, transport 47-49 SOM see soil organic matter Southeast Asia nutrient balance 176-180 nutrient-flux model 178-180 Spartina gracilis Trin. 81, 84, 85 see also cord grass sporulation 96 steaming, soil sterilization 96 stratigraphy, Quaternary 184-185 Streptomyces griseus 6 streptomycin 6 Striga hermonthica 90 Suisse Bilanz, nutrient balance 174-175 sulphate, effect on arbuscular mycorrhizal fungi 94 sunflower 81, 84, 85 surface area, inner 3 sustainability, agricultural 11-12 Switzerland forest soil 64 heavy metals 63-77 effect on micro-organisms 75, 76 enrichment depletion patterns 67-74 soil fertility 75, 76, 77 water quality 74-75 Suisse Bilanz nutrient balance 174-176 symbiosis, mycorrhizal 89, 95, 96, 98, 99, 100,101 Tha~r, Albrecht Daniel (1752-1828), humus theory 16,18 Principles of Rational Agriculture 10-11 Thematic Strategy for Soil Protection, European Union 2002 154-156 tillage effect on arbuscular mycorrhizal fungi 91-92 and greenhouse gas emission 15, 37 and soil conservation 30-31,34, 50 tomatoes 81, 84 topsoil erosion 7, 29 movement effect on arbuscular mycorrhizal fungi 103 in residential development 142 transformation, soil 5 Tutzing Initiative 157

INDEX United Nations Convention on Biological Diversity 1992 156 Convention to Combat Desertification 1994 156 Framework Convention on Climate Change 1992 15,156 uranium, effect on arbuscular mycorrhizal fungi 101 urbanization acid sulphate soil 145 effect on soil 141-142 SOC 30 soil degradation 149 soil sealing 117-119 US Soil Conservation Act 1935 149-151 vadose zone 4,187 flow pathways 47-49 solute transport 46-47 Vertic sol, Portugal 160,161 vine mortality Massif de la Clape 53-60 copper concentration 58-59 soil system 56-59, 57, 60 rehabilitation 59-60 wastewater disposal 135 water effect on arbuscular mycorrhizal fungi 97 erosion 29 retention 2 reuse San Joachin Valley, California 80-87 B and Se accumulation 82, 83, 84 plant species 80-81, 81 plant tolerance 82 saturation, and soil organic matter 25 weathering 1,2,135 biological 1 chemical 1,185-186 physical 1 wetlands 27, 28 CH 4 flux 34-36 cultivation 27 wind erosion 29 World Commission on Environment and Development 13 zinc effect on arbuscular mycorrhizal fungi 97,101,102 Swiss forest soil 63, 64, 65, 74, 76 enrichment depletion patterns 71 zirconium, enrichment comparison 66

Function of Soils for Human Societies and the Environment Edited by E. Frossard, W. E. H. Blum and B. R Warkentin

Earth sciences are becoming ever more concerned with how their disciplines, their research and teaching, need to become directly related to environmental and social concerns. The biology of the surface layers and at depth is increasingly important in the geosciences. A knowledge of biological and physical-chemical ~.~,,~~::?~ '~WZI~ functions in terrestrial ecosystems (such as biomass production, [~,..,..~,~.~. ';~, ....... ,~:....": ~ filtering, buffering and transformation, water routing, and maintenance of biodiversity) that are studied in soil science provides a background for Earth sciences.

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The papers in this volume address issues of soil formation, soil management, soil protection and the role of biodiversity that must be considered for a sustainable soil use. The papers are aimed at geoscientists in the broadest sense, and others concerned with soil use who will also find chapters relevant to their interests. Soils knowledge used within other Earth sciences is essential for maintaining healthy ecosystems, for the solutions of problems in environmental quality and for sustainable use of soils by humans.

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Cover illustration: ISBN 1-86239-207-2

'Preferential flow pattern in a sandy soil', painted by Gerd Wesso]ek in 2004. Technique:sandy soil material and oil co]our on canvas.

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