Ecological impacts of arable intensification

Journal of Environmental Management (2001) 63, 337–365 doi:10.1006/jema.2001.0473, available online at http://www.idealibrary.com on

Ecological impacts of arable intensification in Europe C. Stoate†1 , N. D. Boatman†2 , R. J. Borralho‡ , C. Rio Carvalho‡ , G. R. de Snoo§ and P. Eden¶ † The

Allerton Research and Educational Trust, Loddington House, Loddington, Leicestershire LE7 9XE, UK ‡ ERENA, Av. Visconde Valmor, 11-3, Lisboa, Portugal § Centrum voor Milieukunde Leiden (Centre of Environmental Science), Leiden University, P.O. Box 9518, 2300 RA Leiden, The Netherlands ¶ Box no. 237, Quinta dos Penedos, 7350 Elvas, Portugal Received 17 July 2000; accepted 12 May 2001

Although arable landscapes have a long history, environmental problems have accelerated in recent decades. The effects of these changes are usually externalised, being greater for society as a whole than for the farms on which they operate, and incentives to correct them are therefore largely lacking. Arable landscapes are valued by society beyond the farming community, but increased mechanisation and farm size, simplification of crop rotations, and loss of non-crop features, have led to a reduction in landscape diversity. Low intensity arable systems have evolved a characteristic and diverse fauna and flora, but development of high input, simplified arable systems has been associated with a decline in biodiversity.

Arable intensification has resulted in loss of non-crop habitats and simplification of plant and animal communities within crops, with consequent disruption to food chains and declines in many farmland species. Abandonment of arable management has also led to the replacement of such wildlife with more common and widespread species. Soils have deteriorated as a result of erosion, compaction, loss of organic matter and contamination with pesticides, and in some areas, heavy metals. Impacts on water are closely related to those on soils as nutrient and pesticide pollution of water results from surface runoff and subsurface flow, often associated with soil particles, which themselves have economic and ecological impacts. Nitrates and some pesticides also enter groundwater following leaching from arable land. Greatest impacts are associated with simplified, high input arable systems. Intensification of arable farming has been associated with pollution of air by pesticides, NO2 and CO2 , while the loss of soil organic matter has reduced the system’s capacity for carbon sequestration. International trade contributes to global climate change through long distance transport of arable inputs and products. The EU Rural Development Regulation (1257/99) provides an opportunity to implement measures for alleviating ecological impacts of arable management through a combination of cross-compliance and agri-environment schemes. To alleviate the problems described in this paper, such measures should take account of opportunities for public/private partnerships and should integrate social, cultural, economic and ecological objectives for multifunctional land use.  2001 Academic Press

Keywords: Common Agricultural Policy, arable ecosystems, soil erosion, pollution, pesticides, fertiliser, biodiversity, landscape, climate change, tillage, agri-environment, rural development.

Introduction Arable farming has a long history, from its origins in the eastern Mediterranean 10 000 years ago, to 1

Corresponding author: C. Stoate, The Allerton Research and Educational Trust, Loddington House, Loddington LE7 9XE, UK. Email: chris.stoate@ukonline.co.uk 2 Present address N.D.B.: CSL. Sand Hutton, York YO41 1LZ, UK. 0301–4797/01/120337C29 $35.00/0

one of the most widespread forms of landuse in Europe today. Table 1 shows the areas covered by cereal, oilseed and protein (COP) crops in the European Union. This development and spread has been associated with the evolution of a uniquely adapted and diverse fauna and flora (Potts, 1991, 1997). However, the second half of the twentieth century has seen increasing concern over the impacts of modern arable farming on agricultural  2001 Academic Press

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Table 1. Total arable areas (cereal, oilseed and protein crops) by country (000 ha) (Eurostat, 1998)

B/Lux Denmark Germany Greece Spain France Ireland Italy Netherlands Portugal UK Austria Finland Sweden EU-12 EU-15

1993/4

1994/5

1995/6

1996/7

521 1779 8696 1362 8830 12 108 296 4746 425 855 3854

522 1720 8692 1346 8252 11 873 278 4789 436 828 3813

43 472

42 548

514 1754 8640 1225 8191 11 921 281 4998 434 750 3881 1032 1068 1199 42 589 45 888

524 1819 8962 1302 8296 12 506 294 5181 434 764 4020 1055 1143 1275 44 102 47 575

ecosystems, and on the sustainability of arable systems themselves. Environmental problems arising from modern arable management are now associated with changes to landscapes and plant and animal communities, and a deterioration in soil, water and air quality.

Few of these consequences are confined to the farm on which they arise, the majority being ‘externalised’ to become a cost to society as a whole.

On-farm impacts of changes in arable management include a local simplification of landscape and biodiversity, and a deterioration in soil characteristics, but these, together with many other consequences are more widely apparent as off-farm impacts on biodiversity, landscape, water and air.

These are often felt outside the country in which they originated. Such externalities are increasingly regarded as unacceptable on both economic and ethical grounds (Colman, 1994). For the United Kingdom (UK) alone, these costs to the community have been estimated at £2343 million per year (£208 ha 1 yr 1 ; Pretty et al., 2000).

Such arable intensification is exemplified by increased adoption of external inputs such as fertilisers and pesticides, increasing scale of operation, and simplification of farming systems (Meeus, 1993). Although agricultural intensification has been a feature of arable systems throughout western Europe, it has predominated in the north, with small-scale, extensive mixed farming systems surviving in many parts of the south. Extensive systems also survive in many parts of eastern Europe, but rapid moves towards both intensification and abandonment have taken place in the 1990s. Proposed integration of eastern European countries into the European Union (EU)

is currently influencing the rate of this change, and environmental considerations are incorporated into agricultural policy in most eastern countries, as well as within the existing EU member states.

Environmental measures under the Common Agricultural Policy (CAP) were originally supported under Regulation 797/85, article 19. In the 1992 reform of the CAP, member states were required under the accompanying measures (Regulation 2078/92) to develop agri-environment schemes, with 50% of funding provided by the European Community (75% in Objective 1 areas).

Under the recent ‘Agenda 2000’ reform, agrienvironment schemes are now supported under the Rural Development Regulation (1257/99). These measures have experienced varying degrees of success in terms of their adoption by farmers, with inadequate funding, resistance to long-term obligations, and reluctance to abandon traditional practices being given as reasons for low adoption (Fay, 1998). While some success has been reported (Borralho et al., 1999; Peach et al., 2001), in other cases, conflicting objectives have led to some detrimental consequences being observed (Wakeham-Dawson et al., 1998).

This paper provides an overview of the impacts of recent (post 1950) arable agriculture in Europe.

We illustrate the major issues using one country from northern Europe (the UK) and one from the south (Portugal), making additional references to other European countries where appropriate.

Environmental impacts of arable management are often considered in isolation, but in fact are highly integrated. We have treated each issue separately as far as possible in the sections below, describing the environmental impacts of key management practices, but the aim of this paper is to provide a broad overview of the impacts of arable farming on terrestrial and aquatic ecosystems.

Throughout much of Europe, arable systems are often highly integrated with livestock and forestry, and where this is the case we include these in our discussion of the arable system. There is a tendency for such multiple land use to be more sustainable, and to be associated with higher biodiversity and landscape value, than purely arable systems. The combination of price support and capital grants under the CAP has encouraged abandonment of such integrated systems, as well as encouraging environmentally damaging practices through agricultural intensification.

We start this review with a description of the arable landscape and the impacts on it of changes in arable management. We then describe the influence of landscape change and other consequences

Impacts of arable intensification

of intensification, on terrestrial wildlife. We then develop the review to include impacts on the ecology of soils and aquatic systems, and on climate through gaseous emissions. We end the paper with a brief account of possible measures for the alleviation of these ecological impacts of arable intensification.

Cultural landscapes Because it occupies such a large proportion of the European countryside, arable land makes a major aesthetic contribution to cultural landscapes. Landscape diversity has declined in Europe during the period of agricultural intensification (Meeus, 1993), with a tendency for the most progressive farmers to create the simplest landscapes (Nassauer and Westmacott, 1987). In Britain and France for example, capital grants available under the CAP have resulted in the loss of non-crop features such as hedges and ditches, and the replacement of traditional buildings with modern structures (Burel and Baudry, 1995).

The rate of British hedge removal was greater in the 1970s and 1980s than subsequently (Barr et al., 1994; Westmacott and Worthington, 1997).

More recently, changes have been more evident in the structure of hedges, resulting from abandonment, than in their removal. Barr et al. (1994) reported annual rates of 5Ð2% of hedges abandoned, compared with 1Ð3% restored, and perennial field boundary flora, perceived by many to be an attractive landscape feature, have been lost from most arable farms (Boatman, 1989). However, Westmacott and Worthington (1997) found an ‘improvement’ in hedge ‘quality’ in some areas.

As with many landscape assessments, perceptions of ‘quality’ are likely to influence reported changes.

Haines-Young et al. (2000) reported that by 1998, the declines in length of hedges and walls had generally halted. Rates of hedge planting were similar to the 1980s, but rates of removal had fallen considerably. Restoration management had also increased. However, there was evidence that species richness of vegetation in linear habitats had declined.

Throughout much of western Europe, mixed livestock and arable farms have been greatly reduced in number, creating a less diverse landscape.

With farm amalgamation, farm size has increased (Table 2) and farms have become more specialised in their cropping systems, adopting simpler rotations than in the past. Economies of scale have led to the creation of larger fields and ‘block cropping’,

Table 2.

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Average cereal area/holding (ha) (Eurostat, 1998)

Belgium Denmark Germany Greece Spain France Ireland Italy Luxembourg Netherlands Portugal UK Austria Finland Sweden EU-12 EU-15

1993/4

1995/6

1996/7

8Ð7 21Ð7 15Ð1 3Ð3 13Ð5 17Ð8 13Ð8 4Ð4 13Ð1 9Ð7 2Ð5 40Ð8 6Ð7 12Ð3 19Ð1 10Ð1 10Ð4

8Ð7 23Ð2 17Ð2 3Ð3 13Ð5 19Ð4 13Ð8 4Ð4 13Ð8 9Ð7 2Ð7 42Ð6 6Ð7 12Ð3 19Ð1 10Ð8 10Ð8

10Ð0 25Ð6 21Ð9 4Ð1 17Ð7 24Ð3 15Ð4 5Ð9 16Ð6 8Ð0 4Ð3 52Ð0 7Ð8 13Ð1 20Ð3 14Ð4 14Ð2

resulting in a uniform, featureless landscape in the most intensively farmed arable areas such as parts of East Anglia in England and the Beauce region in France. The occurrence of colourful arable weeds such as poppies has been reduced and some are close to extinction (see next section).

Entec (1995) compared lowland arable landscape features on British organic farms with those on conventionally managed farms. Crop diversity was lower on organic farms and fields were smaller.

Hedges were taller and less intensively managed than on conventional farms but of the farmers removing hedges, 43% were organic. However, of farmers actively ‘improving’ hedges (or perceived by themselves or others to be doing so) only 11% were in conventional arable systems.

Such differences relate more to the attitudes of the farmers themselves than to the type of farming system (Entec, 1995). Chamberlain and Wilson (2000) found similar differences in hedge management between conventional and organic farms, and higher crop diversity on organic farms.

In Portugal intensification of arable farming, often associated with the introduction of irrigation has led to the loss of fallows from arable steppes, creating a more uniform landscape lacking the flowering plants which are often abundant in fallows and low-input arable crops (Pineda and Montalvo, 1995). Such intensification is also associated with the presence of irrigation pivots and electricity pylons, both of which become prominent features in an otherwise open landscape.

Although once present throughout much of southern Europe, the open oak woodland/pasture/ cropping systems known as montado and dehesa are now largely confined to Portugal and Spain

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C. Stoate et al.

where they remain widespread (Harrison, 1996). In Portugal the montado area has remained relatively ˜ stable at the national scale (Direc¸cao-Geral das Florestas, 1998) but arable intensification, especially if accompanied by irrigation, can result in the local loss of traditionally managed montado landscapes (Yellachich, 1993). Such an impact is currently occurring as a result of irrigated agriculture in the area of the Alqueva Dam, a major development project funded by EU Cohesion Funds (Eden, 1996).

Because montados are a diverse and integrated system involving arable farming, sylviculture and livestock, restriction of current support measures to one component can result in the breakdown of the montado system with consequent damage to landscape features. For example, livestock headage payments contribute to abandonment of arable cultivation and to overgrazing of swards and the loss of regenerating trees in montados, resulting in an ageing oak tree population structure and susceptibility to disease (Eden, 1996). Another example of loss of habitat diversity through economic support comes from Spain where subsidies for almond cultivation are conditional on no more than 10% of the area being used for another crop (Pretty, 1998).

Total abandonment of agriculture results in rapid development of shrub cover with the loss of the diverse flora associated with traditional arable rotations. While greater vegetative cover protects soil from erosion, the risk of fire is considerably increased, once again exposing the soil to erosion (Andreu et al., 1995). Such abandonment also results in the emigration of rural people, the loss of traditional skills and the deterioration or destruction of traditional farm buildings which themselves form landscape features. Land that ceases to be used for arable cultivation is frequently planted with Eucalyptus or Pinus species for pulp production, creating a uniform landscape which is poor in terms of landscape features, and for biodiversity.

Biodiversity Simplification of cropping systems results in reduced crop diversity and loss of non-crop habitats such as grassland, field boundaries, water-courses and trees, all of which can form an integral component of arable ecosystems. These, and the loss of livestock from arable systems, have contributed to a decline in biodiversity. Within the cropping system, increased application of fertilisers and pesticides, often accompanying drainage and irrigation, has caused substantial damage to arable

ecosystems, with consequent implications for biodiversity.

Birds provide good indicators of environmental change as they are easily monitored, well researched, long-lived and high in the food chain (Furness and Greenwood, 1993). Many bird species which live in arable landscapes have declined. For example, percentage declines in the 1970–1998 period in UK populations given by Gregory et al.

(2000) are grey partridge (Perdix perdix) (82%), lapwing (Vanellus vanellus) (52%), turtle dove (Streptopelia turtur) (77%), skylark (Alauda arvensis) (52%), song thrush (Turdus philomelos) (55%), tree sparrow (Passer montanus) (87%), reed bunting (Emberiza schoeniclus) (52%) and corn bunting (Miliaria calandra) (85%). Similar declines in farmland species have been experienced across Europe, with 42% of declining species being affected by agricultural intensification (Tucker and Heath, 1994).

Such severe declines are not occurring for species associated with other habitats (Gibbons et al., 1993; Crick et al., 1997; Gregory et al., 2000).

Declines have also been documented in arable invertebrates and plants (Eggers, 1984; Potts, 1991; Andreasen et al., 1996; Donald, 1998). A good example of the interrelationship between these groups is provided by Game Conservancy Trust work in Sussex, England, where impacts of agricultural change have been demonstrated throughout the food chain (Potts, 1986; Ewald and Aebischer, 1999).

Agri-environmental measures introduced under Regulation 2078/92, and later under the Rural Development Regulation 1257/99, have attempted to alleviate some of these adverse consequences of intensification within arable systems. For example, in England a pilot Arable Stewardship Scheme has provided opportunities to restore field boundary habitats, reduce pesticide inputs, plant crop mixtures specifically designed for wildlife and maintain cereal stubbles through the winter, thereby improving breeding habitats, foraging habitats and winter food supplies for birds. Such incentives are intended to replace previously widespread arable habitats that have been lost through changes in cultivation, cropping patterns and arable inputs.

Set-aside has been associated with some conservation benefits for birds (Wilson et al., 1995) but such benefits are well below those possible under more appropriate management, as agricultural priorities have determined the management of set-aside (Winter and Gaskell, 1998).

Organic farming has also been shown to benefit some species. Recent studies in England suggest that organic systems support more broad-leaved

Impacts of arable intensification

plants than conventional systems which are associated with more grass weeds and cleavers (Galium aparine) (Kay and Gregory, 1999; Hald, 1999; Brooks et al., 1995). Kay and Gregory (1999) found that, out of 23 rare or declining arable plant species, 18 were more abundant on organic farms, with 13 of them being absent on conventional farms. However, increasing efficacy of mechanical weed control technology could reduce these differences in plant abundance and species richness between the two systems (van Elsen, 2000).

Some studies have found that aphid abundance is higher in conventional systems, with some other invertebrate groups being more abundant in organic systems (Redderson, 1997; Moreby et al., 1994). Brooks et al. (1995) found higher earthworm numbers in organic fields, and Feber et al. (1997) reported higher numbers of non pest butterflies in organic systems (pest species were equally abundant in both systems). There is some evidence for higher numbers of birds in organic systems, especially skylarks (Wilson et al., 1997).

However, ground-nesting species such as this could be adversely affected by the frequent mechanical control of weeds in organic crops when the birds are nesting (Hansen et al., 2001). Abundance of other species is influenced mainly by field boundary characteristics (Chamberlain and Wilson, 2000) which are influenced by farmers conservation attitudes (Entec, 1995).

Cultivation and crop rotation Cropping systems have been simplified and become geographically polarised in northern Europe (de Boer and Reyrink, 1989), with major consequences for biodiversity within arable ecosystems. High crop diversity is necessary to the ecological requirements of many species. For example, brown hares (Lepus capensis) graze different crops at different times of year (Tapper and Barnes, 1986), while skylarks move breeding territories from one crop to another through the breeding season (Wilson et al., 1997), and yellowhammers (Emberiza citrinella) switch from one crop to another as foraging habitats during the breeding season (Stoate et al., 1998). Lapwings require cereals in which to nest and adjacent pasture on which to feed newly hatched young (Tucker et al., 1994) while little bustard (Tetrax tetrax) males and females have different habitat requirements during the breeding season (Salamolard and Moreau, 1999). As well as this, inter-species differences in habitat requirements result in higher numbers of species

341

in landscapes with low intensity farming and high crop and structural diversity, as demonstrated for the Alentejo region of Portugal by Stoate et al. (in press). Monitoring of a Zonal Programme introduced to this region under 2078/92 suggests that this measure is maintaining bird species diversity (Borralho et al., 1999).

In Britain, geographical polarisation of arable and livestock farming has reduced the number of farms with high crop diversity. Pastures grazed by livestock are associated with large numbers of invertebrates which provide food for birds and other animals, and livestock feed sites in winter provide a source of food for seed-eating birds. The loss of livestock from farms in eastern Britain has removed these components from the arable landscape.

Intensification of grassland management and conversion of grass to arable cultivation resulted in a 92% decline in the area of ‘unimproved’ grassland in the UK between the 1930s and 1980s (Fuller, 1987), with a consequent decline in biodiversity, especially that of plants and invertebrates associated with semi-natural grasslands (Green, 1990).

In the Netherlands, plants and invertebrates (especially dragonflies (Odonata) and butterflies (Lepidoptera)) associated with semi-natural grassland and forest edges are also declining (IKC-NBLF, 1994) and desiccation of grasslands has resulted in declines of wading birds (Klumpers and Haartsen, 1998; Reyrink, 1989). Newbold (1989) reported losses or serious damage to 80% of British calcareous grassland between 1949 and 1984. Loss of diverse grassland ecosystems to arable crops in the UK has continued since then. Flax crops qualify for subsidies even when grown on land that is not eligible for arable area payments, and these subsidies exceed the value of payments for protection of grassland habitats. Previously uncultivated grassland is also currently being lost to potato production, as disease-free and pesticide-free land is often required for this crop (CPRE, 1999). Production of potatoes is discouraged on arable land as this crop is not eligible for area payments.

In Portugal and other Mediterranean countries, livestock plays an important ecological role in arable systems grazing fallows and influencing their botanical and invertebrate composition. In particular, fallow periods of several years provide relatively stable habitats which contribute to the maintenance of plants, invertebrates and birds characteristic of steppic arable systems in Portugal ˜ 1996; (Beaufoy et al., 1994; Moreira and Leitao, Stoate et al., in press). The recently disturbed soil associated with first year fallow represents an

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important foraging habitat for many seed-eating birds in winter (Diaz and Telleria, 1994). Fallow periods vary with soil type, and this disparity has increased since the 1970s. Intensification of arable cropping on the best soils has led to the complete loss of fallows from some systems, while cultivation has been abandoned in the least fertile areas. In the latter, the loss of crops and encroachment of scrub have resulted in the loss of fauna associated with the extensive arable landscapes, and in frequent cases land use converts to forestry with financial support through EU regulation 2080/92. The area of Portuguese arable land converted to forestry under this regulation between 1994 and 1998 amounted to 138 715 ha (Portuguese Ministry of Agriculture, unpublished data). Such abandonment and changes in land use are currently a greater threat to biodiversity of arable systems in southern Europe than the risk of intensification.

Arable steppe landscapes support many of the most threatened species, and are therefore critical for maintaining species diversity at the national and European level. Such species include the globally threatened bird species, great bustard (Otis tarda) and lesser kestrel (Falco naumanii) as well as many others of current conservation concern in Europe (Tucker and Heath, 1994). The loss of fallows and other consequences of arable intensification are a direct threat to these species (Peris et al., 1992). However, at the local level bird species diversity is higher in lightly wooded Montado (Portugal) and Dehesa (Spain) in which holm oak (Quercus ilex) and cork oak (Q. suber) have traditionally been managed as part of the arable system. This habitat has been threatened by wheat-growing campaigns (especially in Portugal) since the 1930s, increased mechanisation, reduced regeneration of trees due to increased livestock densities, and reduced demand for tree products ´ et al., 1997). Increasing use (Yellachich, 1993; Diaz of plastic ‘corks’ in wine bottles could potentially damage the market for natural cork from Montados. This could lead to the loss or deterioration of Montados which support high levels of wildlife ´ et al., 1996). diversity and abundance (Araujo Livestock densities have also increased in arable steppe landscapes, resulting in destruction of the vegetation used by invertebrates and birds associated with fallows. Such stocking densities have been encouraged by headage payments on sheep and cattle, and by increased labour costs leading to replacement of traditional herders by wire fences. In some parts of Spain the loss of the predator control role of herders has resulted in

unsustainable levels of nest predation for scarce ´ breeding lark species (Suarez et al., 1993). Higher stocking densities on arable fallows are likely to reduce the orthopteran food of bustards and other already threatened species (van Wingerden et al., 1997). Such livestock densities could also increase the area of cereals grown for fodder than encouraging earlier harvesting.

Timing of the management of crops can also influence their suitability for invertebrate and bird species. Lapwings and skylarks breeding in northern Europe favour spring-sown cereals, in part because of their structure and less intensive management and these species have therefore been affected by the reduced area of these crops (Tucker et al., 1994; Odderskær et al., 1997). Where grass for grazing, hay or silage is a component of the arable rotation, spring-sown cereals have historically been associated with undersowing of grass leys. This management practice encourages sawflies (Hymenoptera: Symphyta) which overwinter as pupae in the undisturbed soil and whose larvae provide an important food sources for breeding birds such as grey partridge, skylark and corn bunting (Ewald and Aebischer, 1999). In Sussex (England) the distribution of breeding partridges and corn buntings is closely related to that of undersown arable leys (Potts, 1997; Aebischer and Ward, 1997) and availability of invertebrates such as sawflies is related to productivity of these species (Potts,1997; Brickle et al., 2000).

Spring sowing on light soils in Britain is traditionally associated with the survival of cereal stubbles into the winter, thereby providing a food source for seed-eating birds. Modern machinery permits rapid harvesting and ploughing of land for subsequent crops, with the result that stubbles now remain available to wildlife for a very short period. This loss of stubbles has been associated with declines in numbers of many finches and buntings (Chamberlain et al., 2000). In the Netherlands the near extinction of hamsters (Crisetus crisetus) is thought to be due to the short period in which this species may gather grain before hibernation (van Oorschot and van Mansvelt, 1998). Changes from spring- to autumn-sowing are also thought to have contributed to declines in many formerly common spring-germinating arable plants such as corn marigold (Chrysanthemum segetum) and night-flowering catchfly (Silene noctiflora) (Wilson, 1994a). Botanical composition of arable crops is also influenced by cultivation methods. Recent economic pressures have encouraged increased adoption of minimum tillage and direct drilling in preference to conventional ploughing

Impacts of arable intensification

(Jordan et al., 2000a). While ploughing encourages broad-leaved weed species that are important invertebrate and bird food, direct drilling encourages grass weeds and the pernicious broad-leaved weed, cleavers (Galium aparine). However, direct drilling leaves weed seeds near the soil surface, providing more food for birds in winter (Higginbothom et al., 2000).

Cultivation methods can also affect the composition of invertebrate communities in arable ecosystems. Ploughing is the most destructive, affecting invertebrate populations through physical destruction, desiccation, depletion of food and increased exposure to predators. Earthworms are most abundant in arable systems with minimal soil disturbance such as ‘integrated crop management’ systems and organic systems that do not include potatoes in the rotation (Higginbotham et al., 2000). Shallow-burrowing species however, are more abundant in integrated crop management than either conventional or organic systems.

As well as being an important component of the soil fauna, earthworms are a major component of the diet of some birds and mammals. Large carabid beetles are often more abundant in ploughed fields than where minimal cultivation is used, with smaller species being more numerous in the lat´ ter (Baguette and Hance, 1997; Carcamo, 1995), but these findings are not consistent across studies (Kendall et al., 1995). Effects of tillage methods on arthropods are currently the subject of research in Britain. Little is also known about the effects of minimal tillage on vertebrates, although Belmonte (1993) suggests that such methods might be beneficial to some birds in Spain.

Fertilisers Use of fertilisers on arable land increased substantially during the second half of the twentieth century, but there has been a decline in use more recently. In France the use of nitrogen fell by 10% between 1986 and 1994, phosphates by 20% and potassium by 13%. Total fertiliser use has also decreased in the Netherlands and the UK although in Portugal the trends are less obvious (Table 3).

As market prices of arable crops have fallen, economic pressure for more accurate linking of fertiliser application rates to crop needs has increased.

However, as modern arable crop varieties require relatively high rates of fertiliser application, fertiliser use remains high. ‘Precision farming’, i.e. varying application spatially in response to spatial variations in soil fertility or other parameters, may

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Table 3. Consumption of all commercial fertilisers (‘000 t) (Eurostat, 1997)

United Kingdom Netherlands Portugal

1980

1985

1990

1994

2054 679 259

2524 701 241

2413 558 278

2219 534 248

allow more accurate targeting in future, but still requires further development (Lu et al., 1997; Godwin, 2000).

Modern crop varieties grow vigorously under high rates of fertiliser application, out-competing other arable plants, and increases in the use of fertilisers have contributed to a change in the arable flora (Wilson, 1994a). The dense crop structure associated with high levels of fertiliser application is also unsuitable for some birds as a habitat for nesting and foraging (Wilson et al., 1997). The increased use of mineral fertilisers during the second half of the twentieth century has also influenced non-crop habitats associated with the arable system. Deposition of fertiliser in perennial vegetation at the field edge has contributed to a change in botanical composition towards annual weeds such as cleavers and barren brome (Bromus sterilis) (Boatman et al., 1994).

Such changes encourage the perception among farmers of field boundaries as a source of weeds, leading to the further destruction of this habitat through deliberate or accidental use of herbicides, ploughing into the field edge, and in many cases complete removal (Boatman, 1989). Loss of this perennial herbaceous field boundary habitat has severe implications for wildlife associated with it (e.g. whitethroat (Sylvia communis) Stoate and Szczur, 2001).

Pesticides Although pesticide use per hectare is lower on arable land than many other crops (e.g. mushrooms, flowers, bulbs and fruit), the large area covered by arable crops results in this land-use being the largest user overall (European Commission, 1999). Pesticide use increased substantially during the second half of the twentieth century, but has declined slightly in recent years. Between 1991 and 1996 the largest decreases in pesticide sales were seen in those countries which have specific policies on pesticide reduction: Finland ( 46%), the Netherlands ( 43%), Denmark ( 21%) and Sweden ( 17%) (Eurostat, 1998).

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However, as methods of monitoring pesticide use differ between countries, and as pesticide use can fluctuate from year to year in response to climatic and other factors, short-term assessments and comparisons are difficult. In 1996 there was an increase in sales volumes, particularly in Spain (C19%), France (C11%) and the UK (C6%) because of seasonal conditions such as weather and pest pressure. Overall, sales were divided between fungicides (41% weight of active ingredient), herbicides (39%), insecticides (12%) and others (8%) (European Commission, 1999). However, there is some variation between regions, with for example fungicides forming a larger proportion of pesticides used in southern Europe and herbicides predominating in the north. France is the largest EU market for pesticides with 31%, followed by Italy (16%), the UK (12%), Germany (12%) and Spain (11%) (European Commission, 1999).

Direct effects of pesticides on vertebrates have been greatly reduced since the phasing out of organochlorines, although rodenticides continue to be a cause of secondary poisoning of barn owls (Tyto alba) in areas of warfarin resistance (Shawyer, 1987). The molluscicide methiocarb has also been shown to cause mortality of wood mice (Apodemus sylvaticus) under field conditions (Johnson et al., 1991). De Snoo et al. (1999) report few poisoning incidents resulting from arable use of pesticides in Europe but suggest that the efficacy of monitoring is uncertain and variable between countries.

The arable flora has changed in response to arable intensification. In Britain Wilson (1994a), using 10 km squares as sampling units, reported that the arable flora included 25 species that were recorded from fewer than 100 squares, at least 26 that were recorded from fewer than 15, and a further seven that had recently become extinct.

Several of the rarest species no longer occur in arable habitats, but for others, the most suitable areas remain the calcareous and sandy soils of south and east England (Wilson, 1994b). Similar declines in the arable flora have been reported from Germany (Eggers, 1984; Agra Europe, 1991), The Netherlands (Joenje and Kleijn, 1994) and Denmark (Andreasen et al., 1996). Arable flora are highly concentrated in field margins (Wilson and Aebischer, 1995; Joenje and Kleijn, 1994; De Snoo, 1997) where biomass, density and species diversity are reduced by herbicide use (Chiverton and Sotherton, 1991). Leaving the outer few metres of a crop unsprayed with herbicide can have a positive effect on the presence and abundance of plant species (Schumacher, 1984; Chiverton and Sotherton, 1991; Hald et al., 1994; De Snoo, 1997).

Set-aside contributes to a reduction in overall pesticide use on arable land, but an increase in the use of non-selective herbicides (especially glyphosate), which are known to affect field boundary vegetation structure and botanical composition, and the composition of associated invertebrate communities (Haughton et al., 1999).

Pesticide use in southern Europe is well below that in the north, but declines in some arable plants have been reported following herbicide use, coupled with increased fertiliser use and abandonment of fallows (e.g. the Alentejo, Portugal [Moreira et al., 1996a]). Species threatened by such intensification include Linaria ricardoi and Euphorbia transtagana, both of which are included on the Directive 92/43/EEC (Conservation of natural habitats and wild fauna and flora).

Herbicide use in arable crops is known to have a negative impact on invertebrate abundance and species diversity (Chiverton and Sotherton, 1991; Moreby et al., 1994; Moreby, 1997). Direct effects of insecticides are a major influence on invertebrate communities (e.g. Moreby et al., 1994), although the effects differ between species, depending in part on their ecology (Grieg-Smith et al., 1995). Some fungicides have also been implicated in influencing invertebrate abundance (Sotherton et al., 1987; Reddersen et al., 1998).

In Sussex (England) there was a negative relationship between broad-leaved weed abundance and the use of dicotyledon-specific herbicides, and between grass weeds and broad-spectrum herbicides (Ewald and Aebischer, 1999). Spring and summer use of herbicides was particularly effective at reducing broad-leaved weed abundance. Of the five invertebrate groups studied, all showed a negative relationship between abundance and the use of insecticides, and declines of four of them were associated with fungicide use. Particularly strong effects were noted for the pyrethroid and organophosphate insecticides, but none of the groups showed a negative relationship with use of the more selective insecticide, pirimicarb, suggesting that broad-spectrum insecticides are most damaging to cereal ecosystems. Broad-spectrum insecticides such as dimethoate continue to be the most widely used (Potts, 1997) and can cause substantial damage to populations of beneficial arable invertebrates and honeybees (Greig-Smith et al., 1995). Declines in British bumblebees have also been linked to use of pesticides in arable crops (Williams, 1982). The molluscicide methiocarb is toxic to a wide range of fauna, and has been shown to affect populations of carabid beetles in the field (Purvis and Bannon, 1992).

Impacts of arable intensification

Arable invertebrates are an essential component of the diet of many farmland birds during the breeding season and their decline has been linked most convincingly to the substantial decline in grey partridge numbers in Britain (Potts, 1986; Potts and Aebischer, 1991). In Sussex, partridge density was inversely related to the number of herbicide applications, and positively related to the number of weed taxa (Ewald and Aebischer, 1999). The abundance of corn buntings and skylarks was also inversely related to herbicide use, and that of corn buntings with fungicide use. All three species have declined substantially throughout their range in northern Europe (Tucker and Heath, 1994; Flade and Steiof, 1990; Fuller et al., 1995) and the Sussex results provide further evidence for the impact of pesticides on the abundance of farmland birds as well as invertebrates. Nestling survival of corn buntings is inversely related to invertebrate abundance, especially Symphyta and Lepidoptera (Brickle et al., 2000), which occur at highest densities in relatively low-input arable systems incorporating undersown leys (Aebischer and Ward, 1997). Corn buntings are also strongly associated with low input arable systems in Portugal (Stoate et al., 2000).

Current use of herbicides and cropping practices combine to reduce production of weed seeds on arable land (Jones et al., 1997), and for some seed-eating birds a reduction in the area of weedy winter stubbles is thought to have contributed to increased winter mortality and subsequent population declines (Campbell et al., 1997). Where they still occur, weedy winter stubbles are strongly favoured as a foraging habitat by finches and buntings such as cirl buntings (Emberiza cirlus) (Evans and Smith, 1994) and corn buntings (Donald and Evans, 1994).

Herbicide tolerance in genetically modified (GM) arable crops could, in the near future, lead to increased use of very-broad spectrum herbicides, and more complete removal of arable plants and the other wildlife dependent on them. The use of broadspectrum herbicides on such genetically modified crops could also result in even greater damage to adjacent habitats than is currently the case.

There is also the danger that arable weeds could themselves acquire herbicide tolerance through cross-pollination with GM crops. Insect resistance in GM crops could reduce the use of insecticides on arable land, with consequent benefits for other wildlife, but the impact of such GM crops on the natural predators of crop pests is not well understood. There is currently inadequate information

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available on the environmental impacts of GM crops.

Drainage and irrigation In northern Europe, large areas of grassland have been drained for conversion to arable crop production since the 1940s, but remaining wet grassland habitats have also been severely affected by drainage of adjacent arable land (Mountford and Sheail, 1984; Williams and Bowers, 1987). In the Netherlands about 60% of the lowering of water tables is caused by draining of adjacent arable fields (RIVM, 1998). As a result there have been substantial declines in abundance and diversity of birds, plants and invertebrates associated with wet grassland habitats in northern Europe (de Boer and Reyrink, 1989). Baldock (1990) highlights rapid changes to wet grassland habitats in France resulting from agricultural intensification.

In southern Europe the area of irrigated crops has increased considerably since the 1960s, taking the form of pivot irrigation in formerly dry ´ areas, and flooded rice in low-lying areas (Suarez et al., 1997; Fasola and Ru´ız, 1997). Pivot irrigation of crops such as maize is associated with increased fertiliser and pesticide applications and the environmental impacts of irrigation are therefore largely those of these inputs, including the loss of fallows in crop rotations.

Irrigation can result in the loss of wetland habitats. For example, in Las Tablas de Daimiel of Spain subsidised expansion of the irrigated area has resulted in a 60% reduction in the size of this designated Ramsar wetland of international importance, Special Protection Area and Natura 2000 site (WWF, 2000). Drainage for irrigation has resulted directly in the local extinction of arable plants such as Armeria arcuata (Moreira et al., 1996b) and subsequent use of herbicides and fertilisers have a wider impact on the arable flora (Moreira et al., 1996a). Intensification associated with irrigation can also reduce invertebrate abundance. Although Barranco and Pascual (1992) reported more Orthopteran species in irrigated than dryland cereals in Almeria, Stoate et al.

(2000) found higher numbers of Orthoptera in Portuguese extensively managed cereals than in intensive systems.

In contrast, irrigated rice can increase the local diversity of birds and the aquatic invertebrates on which they feed (Fasola and Ru´ız, 1997) if pesticide use is not high. It can serve a particularly valuable role in the conservation of

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wetland wildlife, including breeding, wintering and migratory birds, where rice is grown close to estuary habitats, as is the case in the Ebro delta (Spain) (Pain, 1994). Breeding abundance of six species of heron, as well as white stork, is related to area of flooded rice close to their breeding sites, and rice fields and the irrigation channels associated with them are favoured foraging areas in Portugal and elsewhere in southern Europe (Pain, 1994; Coelho, 1998). When flooded in the winter, to encourage ducks for hunting, they provide important feeding areas for wading birds (Pain, 1994; Fasola and Ru´ız, 1997). Substantial declines in the rice crop area, and a recent trend towards dry cultivation of rice in southern Europe could reduce the current ecological role of this crop (Fasola and Ru´ız, 1997).

In Italy, the EU member state with the largest rice-growing area, the increasing scale of ricegrowing operations has had substantial impacts on rice fields as a habitat for wildlife (Pain, 1994).

Fields have been enlarged, reducing the habitat diversity and area of uncropped habitat, and the use of lasers enables farmers to produce level ground, lacking wet patches in which aquatic animals can survive when fields are drained.

Such precision equipment also enables rice to be grown in shallower water which reaches higher temperatures and is less suitable for aquatic animals (Pain, 1994).

Soil Soils are affected by the method, timing and frequency of cultivation, and by soil type and topography. For example, compaction resulting from the use of heavy machinery and declining organic matter resulting from frequent cultivations affect soil structure and composition. Inputs in the form of pesticides and organic and inorganic fertilisers also influence soil structure directly, and through their impact on the soil fauna.

Simplification of cropping systems, increases in field size and increased use of heavy machinery and pesticides have all contributed to higher levels of soil erosion (Evans, 1996), as described below.

Evans (1996) reports mean annual rates of soil loss from arable land of 3Ð6 t ha 1 in Belgium and 6Ð1 t ha 1 and 5Ð1 t ha 1 in the English counties of Somerset and Hampshire respectively. European Union Environment Directorate (DGXI) estimates of mean annual soil loss across northern Europe (cited by Gardner, 1996) are higher, at 8 t ha 1 . In

southern Europe 30 to 40 t ha 1 can be lost in a single storm (De la Rosa et al., 2000).

Soil structure Sandy and peaty soils are particularly susceptible to erosion by wind and the peat area of the East Anglia (UK) arable area has declined considerably as a result of erosion, with the peat depth in remaining areas being substantially depleted.

Erosion of cultivated soils resulting from the action of wind and water leads to loss of nutrients and crop rooting depth as well as to pollution, eutrophication and sedimentation of aquatic habitats (see below).

Off-farm costs of erosion resulting from damage to property, roads and communications, pollution of waterways and drinking water, sedimentation of reservoirs, and damage to fisheries are externalised to society and are considerably greater than onfarm costs incurred by farmers.

Loveland et al. (2000) recorded a decrease in mean soil organic carbon from arable ley sites of 0Ð49% over a 15 year period (organic carbon comprises 55% of total soil organic matter; Persson and Kirchmann, 1994). Pretty et al. (2000) estimate that 20% of UK arable soils lost 1Ð7% or more of organic matter since 1980, amounting to 1Ð42 t ha 1 yr 1 (expressed as carbon). Walling (1990) suggests that organic matter is eroded from arable land disproportionately to its availability in the soil. Such loss of organic matter has severe implications for the soil. It can lead to reduced water retention within the soil and to consequent drought in dry regions, and to reduced drainage in wet ones (Benckiser, 1997). Organic matter within the soil also serves an important function in reducing leaching of pesticides to water through adsorption and higher microbial activity.

Organic matter levels are higher in arable systems incorporating livestock or legume-based leys than purely arable systems (Drinkwater et al., 1998). However, Fox (2000) has demonstrated that non-inversion tillage within purely arable systems can increase soil organic carbon and microbial biomass. Although recent adoption of this reduced cultivation has been driven primarily by economic considerations (Jordan et al., 2000a), it may also have environmental benefits for the ecology of both soil and aquatic ecosystems. Jordan et al.

(2000b) indicate that sediment loss from arable land can be reduced by 68% and phosphate loss by 81% under minimum tillage, compared with conventional ploughing.

Impacts of arable intensification

Owing to drought susceptibility, shallow soils are associated with lower cereal yields (Loveland et al., 2000). De la Rosa et al. (2000) have estimated that over the next 100 years the loss in crop yield in northern Europe would be around 0–5%, with most sites being at the lower end of this range. Direct effects on yields in the shorter term are not felt unless soil organic carbon levels fall below 1%, a condition currently associated with, for example, only about 5% of UK arable land (Loveland et al., 2000). Changes in soil are generally slight during the period of a farmer’s life and environmental problems associated with erosion are externalised.

The incentive for preventative action on the part of the farmer is therefore low.

An increased area of autumn cultivation, increases in field size and associated loss of hedges, and continuous arable cropping all increase the exposure of soil to wind and water (Evans, 1996). Lack of crop cover and the presence of tramlines and wheelings also increase erosion rates on arable land (Chambers et al., 1992), and late harvested spring-sown crops such as maize (increasingly planted as a silage crop), sugar beet, potatoes and other vegetables are associated with relatively high levels of erosion because soils are exposed to rain during the autumn and winter (Evans, 1996). Recent incorporation of outdoor pigs into arable systems has also become a major local source of soil sediment. Rainfall, slope and soil type can all be major influences on erosion risk (MAFF, 1999a). Alstr¨om and Bergman (1990), working in Sweden, emphasised the influence of slope, slope length and area while highest rates of soil erosion are generally associated with storm events (e.g. Boardman, 1990).

In southern Europe, erosion rates are much more variable, but generally higher than in the north and influences of slope, rainfall and crop type are greater. De la Rosa et al. (2000), working in Portugal, Spain, Italy and Greece, reported erosion rates of 2Ð8–150Ð0 t ha 1 yr 1 , depending on site and crop type (Table 4). These rates could result in a severe decline in crop yield, as well as having a deleterious impact on terrestrial and aquatic ecosystems. In the most severe case (in the Algarve, Portugal) De la Rosa et al. predict a 97% yield reduction within 100 years, although abandonment is likely considerably earlier than this. Much arable land is already abandoned in this region through the interaction of shallow soils, declining rainfall and economic and structural constraints.

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Table 4. Estimates of soil erosion rates (t ha 1 yr 1 )Ł for 20 European sites and three crop types, derived from ImpelERO soil erosion models (De la Rosa et al., 2000) Wheat

Sugar beet

Sunflower

67Ð0 147Ð9 3Ð9 2Ð8 3Ð9 4Ð4 78Ð0

87Ð3 150Ð0 4Ð9 3Ð8 4Ð9 5Ð9 100Ð1

99Ð9 150Ð0 6Ð0 4Ð6 6Ð0 7Ð1 115Ð7

Atlantic Ireland England Scotland Netherlands France

9Ð8 1Ð4 1Ð4 1Ð4 1Ð4

18Ð4 1Ð4 1Ð4 1Ð6 1Ð6

29Ð4 1Ð8 1Ð8 2Ð0 2Ð0

Continental France France Germany Germany Denmark Denmark Luxembourg Luxembourg

1Ð4 1Ð4 1Ð4 4Ð0 3Ð8 1Ð4 1Ð4 1Ð4

1Ð6 1Ð4 1Ð6 4Ð6 4Ð0 1Ð4 1Ð6 1Ð6

2Ð0 1Ð8 2Ð0 5Ð7 4Ð9 1Ð8 2Ð0 2Ð0

Mediterranean Portugal Portugal Portugal Spain Greece Italy Italy

Ł

tons per square mileD.x/1Ð016)/259

Autumn ploughing in the Alentejo (Portugal), before the main rainfall period, can account for a soil loss of 6 t ha 1 per tillage operation on slopes (Bergkamp et al., 1997). Ploughing of fallows continues through the spring in order to stimulate mineralisation of organic matter, prior to sowing the next crop. Barreiros et al. (1996a, b) demonstrated that no-tillage systems can decrease runoff and increase soil bulk density, reducing erosion by 60% that of ploughed land. However, the applicability of such management will vary with soil type.

Kosmas et al. (1997), reporting on the MEDALUS study in Portugal, Spain, France, Italy and Greece, found that erosion rates were relatively low in shrubland and other areas where perennial and herbaceous vegetation cover exceeded 90%. However, erosion rates under Eucalyptus plantations (often advocated as an alternative land use for degraded arable land) were slightly higher than on rainfed arable land. Roxo and ˜ Casimiro (1997) also found that eroCortesao sion was lower under naturally regenerating shrub or herbaceous vegetation than arable land. However, rainfall has declined since the 1930s and has become increasingly erratic, both within years and months (Bergkamp et al., 1997), and this

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can limit growth of natural vegetation and its capacity to minimise erosion (Kosmas et al., 1997).

Throughout Europe, the use of heavy machinery and frequent passes with cultivating equipment can cause soil compaction, increasing runoff at the soil surface and creating a soil pan within the soil. The latter inhibits drainage, causing waterlogging of crop plants on some soils and creating a physical barrier for their roots, making them more susceptible to drought in dry conditions. Soil compaction is a particular problem on soils with low organic matter where earthworm abundance and activity is low (Makeschin, 1997).

Earthworms help to maintain soil structure in soils that are not compacted and maintain high organic matter, improving aeration, crop root growth and drainage (Marinissen, 1992; Makeschin, 1997). Worms also distribute beneficial protozoa and mycorrhizal fungi (Makeschin, 1997), and can reduce leaching of nitrogen by increasing nitrification of soluble nitrates (Elliott et al., 1990), but they increase N flux in the autumn (when crop uptake is low), as well as in the spring (Whalen and Parmelee, 2000). Soil compaction reduces abundance of microfauna, especially in deep-tilled soils (Schrader and Lingnau, 1997) and causes anaerobic conditions and changes in microbial community structure (Bamford, 1997).

Although direct drilling and minimum tillage are associated with some crop pests, they also favour the soil fauna responsible for breakdown and mineralisation of soil organic matter, especially deep burrowing worm species such as Lumbricus terrestris (Edwards, 1984; Jordan et al., 2000b).

Shrader and Lingnau (1997) found higher earthworm densities in integrated (reduced cultivation, pesticide and mineral fertiliser applications) than in conventional arable systems. Conventional cultivation is especially damaging to soil fauna in semi-arid low organic matter soils (Bamford, 1997).

Nutrients In southern Europe, soil nutrients have become severely depleted since the beginning of the century following the adoption of cropping systems using little or no manure or fertiliser (Esselink and Vangilis, 1994), contributing to wide scale abandonment. In northern Europe nutrients are lost from soils which continue to receive high levels of nutrient inputs. Erosion is the main cause of phosphate loss from the soil. Leaching of nitrogen can result from applications of mineral fertiliser at

rates above those required by the crop at the time of application. However, much of the nitrogen lost from soil is now known to be associated with mineralisation of soil organic matter at a time when there is no crop cover to exploit the mineral nitrogen made available, normally the period following harvest (Bloem et al., 1994). Soil cultivation in warm wet conditions after harvest maximises mineralisation and loss of nitrate during the period before the following crop becomes established.

Organic matter reduces soil erosion, improves soil moisture retention and supports soil mesofauna that maintain appropriate soil structure for crop growth (Benckiser, 1997). Crop rotations which incorporate grass leys improve soil organic matter and reduce loss of nitrogen through leaching, but can be associated with high levels of leaching when they are ploughed (Young, 1986).

The use of farmyard manure increases soil organic matter and releases nitrogen more gradually than an application of mineral fertiliser, but the mineralisation and subsequent availability of nitrogen does not necessarily match the requirements of the crop, with the result that leaching can occur (Laanbroek and Gerards, 1991). Manure from poultry is especially associated with excessive rates of mineralisation (Benckiser, 1997), while application of slurry from intensive livestock systems can be toxic to some earthworm species (Makeschin, 1997). However, nitrate leaching can be lower in arable systems incorporating livestock or legumebased leys than in conventional arable systems (Drinkwater et al., 1998).

In the Netherlands ‘integrated’ systems with reduced cultivation, nitrogen and pesticide application were found to maintain organic matter (whereas this declined in conventional systems) and have crop yields which were 90% of those obtained from conventional systems (Lebbink et al., 1994). Most groups of organisms also had higher biomass in integrated than conventional systems (Zwart et al., 1994). Protozoa and nematodes were more abundant, and mineralisation of nitrogen was therefore higher in the integrated system, but excessively high in both after harvest (Bloem et al., 1994). Didden et al. (1994) reported that faunal mineralisation was 49% and 87% of the total mineralisation in conventional and integrated systems respectively.

Regional production of sewage and manure are increased by 20% as a result of imported food and fodder (Benckiser, 1997). In 1999 42% of UK sewage production was applied to agricultural land, and the quantity used was predicted to double by 2006, following the ban on sewage sludge

Impacts of arable intensification

disposal at sea in 1998 (Chambers, 1998). Sludge cake applications supplying 250 kg ha 1 total N will typically provide about 4 t ha 1 of organic matter, and liquid sludge about 3 t ha 1 , to the soil, reducing erosion risk and increasing moisture retention (Chambers, 1998). However, current usage represents less than 1% of British agricultural land (Chambers, 1997). Application of sewage sludge is associated with high soil phosphate accumulations and with problems of heavy metal accumulation ˜ 1996; Benckiser, in the soil (Ribeiro and Serrao, 1997). These heavy metals reduce and change the composition of the microfauna, hampering soil metabolism and reducing degradation of pesticides (Benckiser, 1997).

Pesticides and heavy metals In the Netherlands 123 000 kg of pesticide (active ingredient) is known to reach the soil in arable fields, mainly via droplet drift (MJP-G, 1997). Pesticides can influence soil structure and nutrient status through their action on soil fauna and flora, while some pesticides are themselves degraded by soil fauna (AFRC, 1990), the rate at which this occurs varying between compounds (Table 5).

Although most herbicides are not toxic to soil fauna (Bamford, 1997), those that are include the triazines such as atrazine (Edwards, 1984) and bipyridyl herbicides such as paraquat. Although paraquat has been shown to have insecticidal properties in the laboratory, impacts on invertebrate populations in the field are probably low (Wright et al., 1985). However, Edwards (1984) suggested that herbicides can indirectly reduce soil organic matter and the organisms associated with it by

preventing the growth and eventual decay of weeds within the crop.

Insecticides have a more direct impact on soil fauna. Organophosphate insecticides have been shown to change the ratio of predatory mites (Acari) to springtails (Collembola), while carbamates are more persistent and have more broad-spectrum toxic and sublethal effects on soil organisms, including earthworms (Edwards, 1984; Makeschin, 1997).

Sewage sludge, and to a lesser extent, animal manure, can contribute heavy metals such as zinc, copper and cadmium to the soil. In the Netherlands, where the use of these manures is high, heavy metal concentrations on many arable field soils are higher than the target value (Dutch environmental standard) (CBS, 1997).

Irrigation can also influence soil chemistry.

Intensive production of irrigated horticultural crops can result in serious and permanent soil degradation through increased salinity (European Environment Agency, 1998). Contamination of soils by pesticides, heavy metals and nutrients in Spain varies with soil type, crop type and management practices, and is highest under intensive irrigated systems (de la Rosa and Crompvoets, 1998). Copper-based fungicides have long been used in vineyards and Dias and Soveral-Dias (1997) reported high levels of soil contamination by copper in ground previously occupied by vines and converted to arable use, with highest levels in soil from older vineyards. Copper is little used in arable systems, with the exception of organically grown potatoes, where it’s use continues to be permitted but increasingly questioned as it is recognised as a potential long-term environmental problem (Redman, 1992).

Table 5. Pesticide degradation rates in soil (data supplied by Pesticide Safety Directorate, York, UK; F. Hutson, personal communication) Active ingredient Atrazine Isoproturon Pirimicarb Dimethoate Metaldehyde Carbendazim

Situation

Days (DT50 )

Mean days

Source

(laboratory) (field) (laboratory) (field) (laboratory) (field) (laboratory) (laboratory) (laboratory) (laboratory) (field)

41–146 5–60 10–20 13–25 10–263 15–21 4–16 1–71 67–1662 20–365 100–180

81 29

EU Review EU Review UK Review UK Review UK Review UK Review UK Review UK Review UK Review EU Review EU Review

DT50 - time for 50% loss; half-life.

1 German soils.

2 US soils.

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While heavy metals do not therefore generally have a widespread impact on arable soil ecology, the impact of pesticides is much more widespread and related to levels of soil organic matter through the activity of soil fauna.

Water Although the proportion of total water use attributable to agriculture is relatively small, the fact that most of this water is used in the summer means that ecological impact can be considerable.

There is in any case increasing concern about the sustainability of water use in England (Environment Agency, 1999a) because water abstraction lowers groundwater, reduces wetland habitat areas, slows flow rates of rivers and exacerbates problems with pesticides and nutrients in watercourses (European Environment Agency, 1998).

The problem is more acute in southern Europe where over-exploitation of groundwater for agriculture can lead to a decline in the quality and quantity of water, an increase in the economic cost, and an increase in water table depth (Mairota et al., 1997).

Pesticides and nutrients can enter ground and surface waters, seriously affecting the quality of drinking water, and the cost of its treatment. Their presence in surface water also can have serious consequences for aquatic life. Erosion of arable soils results in sedimentation of watercourses and deterioration in the quality of water and aquatic ecosystems. Economic support under CAP regimes for intensification of arable systems has increased the arable area, field sizes and fertiliser use, resulting in deterioration in the quality of aquatic environments.

Cultivation Soil erosion has increasing environmental consequences for aquatic habitats as well as for the soil itself, but in the case of aquatic habitats, these consequences are externalised from the farming system. Late cultivations associated with preparing autumn seedbeds and late harvesting of crops such as maize, potatoes and sugar beet contribute to high silt loads in streams and rivers because soil is exposed during periods of high rainfall. Early ploughing of rape residues can also lead to high levels of nitrogen leaching to water. Loss of soil to watercourses is greatest on sandy soils where

infiltration capacity has been reduced by surface capping, and on clay soils where surface runoff is also high. Sub-soil fissures in clay soils during dry weather can lead to rapid movement of water and clay particles during subsequent rains. Subsoiling, mole ploughing and other drainage practices operate in a similar way, increasing movement of soil-derived sediment (and nutrients) into watercourses and bypassing the natural filtering effect of undrained soils. Continuous cultivation of arable crops and removal of hedges are thought to have contributed to increased silt loads in rivers (Skinner and Chambers, 1996; Van Oost et al., 2000). The associated economic costs could be considerable. In Britain, Evans (1996) estimated the annual cost of removing soil-derived impurities from drinking water at £3Ð6–£30 million.

Ecological impacts of sedimentation in watercourses are best documented for salmonids (Theurer et al., 1998) but also affect other aquatic organisms and macrophytes. Salmonid eggs require a period of 60–180 days in gravel ‘redds’ on the river bed. Sedimentation into redds during this period rapidly reduces dissolved oxygen available in the water with the result that developing embryos are killed. This is thought to have been a major factor in the decline in numbers of economically important salmonids in British waters and to be closely linked with arable, rather than grassland systems. In the River Test (Hampshire, England), hatching rates of fry from redds in a section of the river fed by grassland streams was substantially higher than that where the river was fed by streams flowing through arable land (Anon, 1999).

Cultivation of arable land in the Iberian Peninsula coincides with the main autumn rainfall period. Rainfall is erratic and discharge of water into rivers, and erosion of soil, can be high during rainfall events (Bergkamp et al., 1997). Bergkamp et al., reported declining annual rainfall in the Alentejo (Portugal) since the 1930s, and there has been a decline in spring rainfall over the past decade (Battencourt, 1999). Reduced flows could mean that sedimentation in streams and rivers is higher now than in the past. Sedimentation of reservoirs can be considerable, reducing their capacity for water storage (Kosmas et al., 1997).

Nutrients Nitrate and phosphate are the main nutrient pollutants of watercourses arising from arable farming.

Nitrate levels in drinking water are regulated by

Impacts of arable intensification

the EU Drinking Water Directive (80/778/EEC), which established a maximum admissible concentration of 50 mg l 1 , based on World Health Organisation guidelines. The EU Nitrates Directive (91/676/EEC) requires members states to reduce nitrate pollution by introducing controls on agriculture in water catchments where the nitrate concentration exceeds 50 mg l 1 or is in danger of doing so.

Nitrogen and phosphate reach watercourses from the soil by leaching, surface run-off, subsurface flow and soil erosion. Nitrate is soluble and enters water via leaching and run-off while phosphate molecules bind to eroded soil particles and enter watercourses as run-off. Most nitrate leaching occurs during the autumn when nitrate passes through the root zone faster than the crop is able to exploit it, and following ploughing of grassland, when organic nitrogen is mineralised (Young, 1986). Leaching is greater under cereals than under permanent grass (Croll and Hayes, 1988), but can also be high under rotational setaside (Meissner et al., 1998). Drainage of heavy soils increases the rate at which nitrates and phosphates enter surface water.

In a UK experiment, nitrate lost from direct drilled land was 24% lower than from ploughed land, but still higher than the EC Drinking Water limit (Skinner et al., 1997). The relative contribution to leached nitrate of nitrate from mineralisation of organic nitrogen and that from applied fertiliser is unclear, but the former is known to account for a substantial proportion of nitrate lost from cultivated and uncultivated soils (Addiscott, 1991; Sylvester-Bradley and Powlson, 1993). Although short-term responses to changes in fertiliser applications to cereal crops have also been reported (NRA, 1992), Addiscott, (1991) reported that only 6% of labelled fertiliser nitrogen applied to winter wheat was lost directly by leaching, and ploughing of grassland in the 1940s and 1950s is thought to have contributed to current high levels of nitrate in groundwater (Johnston, 1993).

Nitrate levels remain highest in the south and east of England where rainfall and therefore dilution of nitrates, is lowest. The period between leaching and appearance in the saturated zone of the aquifer depends on geology and can exceed 40 years on sandstone and chalk but is considerably less than this on more permeable rocks such as limestone.

Phosphates enter surface water following periods of rain, when soil particles are eroded from exposed soil, especially where fields on slopes are ploughed.

Phosphorus may enter water by surface runoff or by sub-surface flow through soil cracks and drains,

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and may be in the form of dissolved P or adsorbed to soil particles. Losses are greatest during storm events when transport through the soil is too fast for P to become adsorbed to stable particles within the soil. For example, Heathwaite (1997) recorded that particulate P formed the bulk of total P from sub-surface flow following a storm event.

Readily drained sandy soils, and clay soils prone to cracking, especially those with field drains, are the most prone to loss of P in both dissolved and particulate form (Heathwaite, 1997). In Denmark, Kronvang (1990) found that 60% of annual P fluxes consisted of particulate P, with 70–90% resulting from short-term storm events. Studies in the UK suggest that farming is a major contributor of phosphates to surface water and that outputs from both agricultural and sewage treatment sources must be reduced if limiting levels of phosphate in water are not to be exceeded (Johnes and O’Sullivan, 1989).

Both nitrate and phosphate can cause severe eutrophication of water, nitrates affecting mainly coastal waters such as the North Sea (Baldock et al., 1996) and Baltic (Saull, 1990) and phosphates affecting rivers and lakes, including those of conservation importance (e.g. Slapton Ley, Devon, England) (Tytherleigh, 1997). Eutrophication in both coastal and inland waters can result, through excessive growth of phytoplankton, in depletion of oxygen from water bodies, and subsequent death of aquatic invertebrates, fish and other aquatic animals. Blue-green algae associated with eutrophication produce toxins to which fish and terrestrial animals are susceptible.

Eutrophication and changes in the fauna and flora of Loch Leven (Scotland) have been associated with high levels of phosphate derived from farmland (Castle et al., 1999). For Loch Leven in 1992, summer algal blooms were estimated to have cost the area up to £783 000 in lost business, and increased production costs to the downstream industries by £160 000 (Castle et al., 1999).

Eutrophication is an international problem. For example, discharge of phosphates from Spanish arable and industrial sources into the Guadiana river have resulted in phosphate concentrations of up to 5Ð36 mg l 1 and excessive growth of the water fern, Azolla in the Portuguese section of the river (Carrapi et al., 1996).

There have also been concerns about possible risks to human health arising from nitrates in water, though the evidence for these risks has recently been called into question (Addiscott, 2000).

The total UK costs of achieving the 50 mg NO3 l 1 standard have been estimated at £199 million over

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the next 20 years (Skinner et al., 1997). Additional costs of removing algal growths resulting from eutrophication could be considerable (Magarara and Kunikane, 1986). Pretty et al. (2000) put the cost of nitrate use to UK water consumers at £16Ð4 million per year, with additional costs of £52Ð3 million arising from soil erosion and phosphate pollution. Kraemer and Kahlenborn (1998) claim that encouraging farmers in Munich to adopt organic farming and other modified cropping systems protects groundwater from nitrate leaching and (at DM 1 million) is cheaper than removing nitrates from drinking water.

The Dutch agricultural sector is responsible for approximately 64% of the N-load in surface waters, and 38% of the P-load (NMP-3, 1998 ). Legislation to control application rates, methods and timing, aims to reduce these environmental impacts. Other parts of Europe where similar problems occur include much of southern England, low-lying areas of Belgium and France, parts of Germany and the northern plains of Italy, but the greatest impact is often derived from intensive livestock, rather than arable systems (Gardner, 1996).

Rates of nitrogen and phosphate fertiliser application to crops in Portugal are considerably lower than those in northern Europe and nitrate levels in water are generally not a serious problem ˜ (de Sequeira, 1991; Ribeiro and Serrao, 1996).

Eutrophication of inland waters is rare, but phosphorus accumulation can occur where sedimentation results from erosion of arable land (Soveral-Dias and Sequeira, 1992). Where intensively managed crops such as maize are grown, ˜ e.g. in the Ribatejo region (Ribeiro and Serrao 1996) applications of 300 kg N and 140 kg P2 O5 per hectare have been common (Soveral-Dias and Sequeira, 1992). Here, levels of nitrate in groundwater have been found to exceed maximum acceptable concentrations defined by the European Community, with highest concentrations reflecting timing of crop irrigation (Cerejeira et al., 1995).

Pesticides Pesticides enter surface water from point source contamination following spillage incidents (e.g.

Anon, 1999), and from diffuse sources following their application to crops. They can be toxic to aquatic organisms and some are potentially carcinogenic (Cartwright et al., 1991). While 70% of the EU’s drinking water is derived from groundwater sources, maximum allowable concentrations for pesticides are well below drinking water standards

so as to reduce damage to aquatic invertebrate communities (Cartwright et al., 1991). Examples of countries in which such levels are exceeded are given in Table 6.

The presence of pesticides as pollutants of water depends on their mobility, solubility and rate of degradation. Highly persistent organochlorine pesticides are no longer used in arable systems, reducing the risks of pollution incidents resulting from arable operations. Many modern pesticides break down quickly in soil or sunlight but are more likely to persist if they reach subsoil or groundwater because of reduced microbial activity, absence of light and lower temperatures (Environment Agency, 1999b).

Many pesticides are supplied in high concentrations of active ingredient and there is a high risk of pollution incidents resulting from spillages, inappropriate disposal and washing of sprayers.

Diffuse pollution of water by pesticides results mainly from surface run-off following spraying, rather than from pesticides entering aquifers. For dryland crops drainage can increase movement of pesticides from field to surface water, by-passing the soil profile where such pesticides might otherwise be degraded (Cartwright et al., 1991).

Isoproturon (IPU) is the most widely used herbicide in the UK and known to be susceptible to entering surface waters via runoff and movement through soil cracks (White et al., 1997). In a Cambridgeshire (UK) study, levels of mecoprop and other herbicides were highest at times of high river flow (Hennings et al., 1990; Clark et al., 1991).

Evans (1996) cites one study in which pesticide levels were up to 680 times higher during flood than under normal flow conditions. Croll (1988) found mecoprop in 35% of surface water samples from the English Anglian Region, with lindane and dimethoate (insecticides toxic to aquatic invertebrates) present in 16% and 14% respectively.

Atrazine and simazine were the most frequently occurring pesticides (58% and 42% respectively) but these herbicides are extensively used in Britain by users other than arable farmers and the rate at which they enter leached or surface water varies with soil type, climate and cultivation methods (Hall et al., 1991).

In a wider survey of 3500 sites in England and Wales, 100 of the 120 pesticides targeted were detected and five herbicides (atrazine, diuron, bentazone, isoproturon and mecoprop) regularly exceeded EC Drinking Water Directive limits (NRA, 1995). Pesticides differ considerably in their toxicity to aquatic life, and the UK Environment Agency sets Maximum Allowable Concentrations

Table 6. Results of pesticide sampling of groundwater in European countries: Percentage of sampling sites with pesticide concentrations >0Ð1 µg l 1 . Number of sampling sites in brackets. Source: EEA (1998)

Atrazine Simazine

Austria

Denmark

France

Germany

16Ð3 (1660) 0Ð2 (1248)

0Ð9 (1006) 0Ð5 (1006)

8Ð2 (85) 0 (81) 0 (72)

4Ð3 (12 000) 0Ð9 (11 437) 0Ð2 (994) 7Ð5 (10 972)

Lindane Desethyl-atrazine

24Ð5 (1666)

1Ð4 (292)

Heptachlor Metolachlor

0 (72)

Spain

Luxembourg

Norway

0 (28) 0 (28)

1Ð1 (1248)

0 (116)

0 (215)

47Ð6

4Ð8 (84) 80 (5)

1Ð3 (1666)

0Ð2 (1006) 1Ð4 (292) 0Ð4 (277)

83Ð3 (6) 100 (2) 2Ð6 (2234)

0 (12) Impacts of arable intensification

Methoxy-chlor MCPA

25 (8) (84) 0 (12)

0 (28) 0 (28) 1Ð4 (1006)

Slovenia 32Ð1 (84) 4Ð8 (84)

0 (215)

Dichlorprop

Slovakia

13 (355)

DDT

Hexazinon

Czechoslovakia

0 (4)

Bentazone

Desisopropylatrazine

UK

353

354

C. Stoate et al.

Table 7. Pesticide environmental quality standards in UK surface water (Environment Agency, 1999) Annual average (µ g l 1 ) Herbicides Atrazine Isoproturon MCPA Mecaprop Simazine Triallate Insecticides Chlorfenvinphos Cypermethrin Dimethoate Malathion Methiocarb Pirimicarb

0Ð01 2 2 20 2 0Ð25 0Ð01 0Ð0001 1 0Ð01 0Ð01 1

Maximum allowable concentration (µg l 1 )

20 20 200 10 5 0Ð1 0Ð001 0Ð5 0Ð16 5

that tend to be higher for herbicides than insecticides. These concentrations (and the annual average concentrations recorded) are given in Table 7 for a selection of commonly used arable pesticides (Environment Agency, 1999b). Pesticide application is associated with considerable costs to society in terms of water treatment. Pretty et al. (2000) put the cost to UK water consumers of pesticide application at £120 million.

In the Netherlands large quantities of pesticide reach ground and surface water as a result of drift, evaporation and runoff. In a survey of large surface waters in the Netherlands, all contained pesticides.

For approximately 25% of the substances, the concentrations exceeded the maximum allowable concentrations (RIZA, 1996). Some pesticides are found in groundwater in concentrations above EU standards for drinking water (CBS, 1997). Heavy metals have also been reported from groundwater below arable fields (CBS, 1997).

Rates of pesticide application to cereal crops in Portugal and other southern European countries are generally considerably lower than those in northern Europe, and pesticide levels in water are probably not a serious problem in most catchments.

However, where intensively managed crops such as maize are grown, levels of atrazine in groundwater have been found to exceed maximum allowable concentrations, with highest concentrations occurring following crop irrigation (Cerejeira et al., 1995). In the Po Valley of Northern Italy use of the herbicides atrazine and molinate on irrigated rice and maize over permeable gravel aquifers has resulted in the presence of these pesticides in the groundwater (Cartwright et al., 1991).

In parts of north-east Europe levels of aluminium and heavy metal ions in soils are high, and acid rain resulting from industry and transport pollution to the air can cause increasing solubility and mobilisation of these metals, and their subsequent occurrence in groundwater (Bouma et al., 1998; Kraemer and Kahlenborn, 1998).

Air Gaseous emissions from arable farming include the greenhouse gasses N2 O and CO2 and aerosols NOx and NH3 . In addition, use of fossil fuels for the manufacture of arable inputs for crop production, and for transport of arable inputs and products, is a major contributor of CO2 and SO2 to the atmosphere (European Environment Agency (EEA), 1998). Nitrogen and sulphur are associated with acidification of terrestrial and aquatic ecosystems following the transport and deposition of these pollutants.

CO2 is the main greenhouse gas, representing 65% of the anthropogenic contribution to global warming, while N2 O contributes an additional 5% (EEA, 1998). CH4 contributes 20%, but is associated with livestock systems, rather than arable land. SO2 , NOx and NH3 in the atmosphere offset global warming to some extent but their effect is short-lived and localised, and emissions of these gases are declining (EEA, 1998).

European annual mean air temperatures have risen by 0Ð3° C to 0Ð6 ° C since 1900 (EEA, 1998), and climate change resulting from anthropogenic gaseous emissions is increasingly thought to have a substantial impact on terrestrial and aquatic ecosystems. These include the destruction of coral reefs, changes in the abundance, distribution and phenology of many plant and animal species, and local extinction of Arctic and alpine plants (Hughes, 2000).

Production of the gasses NO and NO2 can result from denitrification within arable systems (Benckiser, 1997) and 70–80% of emitted nitrogen is deposited back onto the land resulting in eutrophication and acidification of some environments (Goulding et al., 1998). Deposition of atmospheric nitrogen can lead to eutrophication and acidification of semi-natural environments, resulting in reduction in botanical species diversity and changes to soil processes, as demonstrated at Rothamsted (UK) by Goulding et al. (1998).

Similar changes in soil processes and botanical composition resulting from atmospheric nitrogen deposition have been reported in British upland

Impacts of arable intensification

heath and calcareous grassland (Lee and Capron, 1998), and in heathland in The Netherlands (Prins et al., 1991).

Crop type, soil moisture and nitrate levels all influence rates of denitrification (Svensson, 1991a, b; Addiscott and Powlsen, 1992) and irrigated crops are known to increase NO2 emissions (ArmstrongBrown et al., 1995). An estimated 3Ð.2% of applied N is thought to be lost to the atmosphere as NO2 (Armstrong-Brown et al., 1995). In addition, over 90% of emissions from estuaries and rivers of the greenhouse gas, N2 O are in the northern hemisphere where they are thought to be derived from anthropogenic sources such as nitrogen fertilisers (Seitzinger et al., 2000).

The agriculture sector world-wide accounts for about 5% of anthropogenic CO2 emissions (ECAF, undated), while cultivation of arable soils is estimated to release 30 Mt yr 1 carbon to the atmosphere globally, through oxidation of carbon alone (Armstrong-Brown et al., 1995). However, soil organic matter and bioenergy crops have considerable potential as a carbon sink for CO2 emitted by arable operations (Armstrong-Brown et al., 1995; Smith et al., 2000).

The process of manufacturing nitrogen fertilisers contributes substantial emissions of NH3 to the air.

Within the UK, this process accounts for 37% of all industrial emissions of ammonia (Sutton et al., 2000). The process is also associated with high energy consumption, with the result that emissions of CO2 are also high. Cormack and Metcalfe (2000) estimate that 50% of total energy input into cereal crops is accounted for by the manufacture and transport of fertiliser and pesticides.

Long distance transport, especially that associated with international trade of arable inputs and products, contributes to emission of CO2 and SO2 to the atmosphere. Transporting crops from southern Europe to the north increases the energy costs by 352% over those associated with local delivery (Cormack and Metcalfe, 2000). As well as CO2 and SO2 emissions, shipping results in substantial marine pollution (Collins, 1994), while NOx emissions from aircraft have been found to be responsible for up to 10% of upper tropospheric O3 (Stevenson et al., 1997). Pretty et al. (2000) estimate that air pollution and greenhouse gas emissions arising from agriculture account for 48% of the industry’s externalised costs, amounting to £738 million for N2 O, and £47 million for CO2 , even without including pollution associated with international trade.

Emissions of pesticides to the air are considerable where they have been documented in the Netherlands. Approximately 3Ð1 million kg (active

355

ingredient) of pesticides were emitted to the air in 1995 (MJP-G, 1995). This is 24% of the total amount of pesticide used, and more than 90% of the total emissions to the environment. Pearce and Mackenzie (1999) report increasing concentrations of some pesticides in rainwater in Europe.

Reducing ecological impacts Agricultural intensification is a feature of arable systems in both northern and southern Europe, while in the south, abandonment of arable and livestock systems is an additional threat to arable ecosystems. Wider environmental costs of intensification include soil degradation, siltation and pollution of watercourses, reduced biodiversity, contributions to global climate change and simplification of cultural landscapes. Few of these act directly on farm businesses in the main arable areas, and economic incentives are currently mainly for their continuation through productionled support. However, rural depopulation is an increasing problem in many formerly agricultural areas and a combination of social, ecological and other environmental objectives increasingly forms the focus for agricultural policy.

As indicated in the previous sections, a number of changes to arable systems have been investigated in order to reduce environmental impacts.

Organic farming differs from conventional intensive systems in that external inputs are reduced and a more integrated approach is taken to crop production as a system. ‘Integrated Crop Management’ reduces external inputs, but to a lesser extent than organic farming, and seeks to exploit naturally occurring predators of crop pests, and to adopt minimum tillage. Both organic farming and integrated crop management have potential benefits over unmodified conventional arable production in terms of landscape structure, biodiversity, soil, water and air quality. Table 8 summarises the previous sections in relation to conventional ‘intensive’, ‘organic’ and ‘integrated crop management’, as well as the ‘extensive’ systems surviving in southern Europe.

A number of European Commission Directives concerned with improving or protecting the environment have implications for arable agriculture.

In addition to the Drinking Water Directive and the Nitrates Directive already referred to, a Water Framework Directive, (2000/60/EC) has recently been adopted which will extend the existing provisions on water quality and promote pricing as an incentive for the sustainable use of water resources.

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Table 8.

Summary of effects of four European arable systems on soil, water, biodiversity, landscape and air Soil Organic matter

Structure

Soil fauna

Nutrient pollution

Pesticide pollution

Air Sediment

Biodiversity Terrestrial

Aquatic

Landscape

Intensive

Ploughing aggravates erosion

OM levels low and often declining

Compaction common.

Plough pans

Fewer worms.

Impoverished micro fauna and flora

Pollution from fertilisers–leaching, drains and runoff

Pollution from spray drift, runoff and leaching

Surface erosion and drainage lead to sediment pollution

Pesticide pollution.

Greater energy use from agrochemical manufacture and application

Simplified crop systems, high fertiliser and pesticide use reduce habitat diversity and food supply

Eutrophication, pesticide pollution and sedimentation

Larger fields, block cropping, less non-cropped land lead to homogeneous landscape

Organic

Ploughing less frequent unless stockless

Higher OM due to use of leys and organic manures

Leys and manure improve structure

Rotations and manures encourage microbes.

No pesticides.

Ploughing of grass releases nitrates. May be aggravated with legumes

Few pesticides used

Surface erosion and drainage lead to sediment pollution

Low energy use.

Possibly greater nitrate mineralisation

Greater crop and non-crop habitat diversity than intensive.

Few pesticides.

Mechanical weed control can be detrimental.

Little or no pesticide pollution.

Mixed farming system may reduce sediment pollution.

Nutrient pollution still occurs.

Mixed farming produces more diverse landscape.

Higher proportion of non-crop features.

Integrated

Reduced cultivations reduce erosion

More OM than intensive

Reduced cultivation improves soil structure

Reduced cultivation increases worm numbers.

Reduced pesticide impact.

Lower, more targeted use of fertiliser reduces pollution

Reduced pesticide use

Reduced cultivations minimise erosion

Reduced cultivation and agrochemical use may reduce pollution?

Higher densities of some invertebrates.

Limited evidence for other taxa

Pesticide, nutrient and sediment pollution likely to be reduced

Crop rotation may increase landscape diversity.

May include grass margins etc.

Extensive (mainly southern Europe)

Erosion common due to ploughing

OM levels low and often declining

Probably poor

Reduced pesticide impact

Low rates of fertiliser use

Low pesticide impact

Frequent cultivations encourage erosion

Low energy use.

Possibly greater nitrate mineralisation

Use of fallows and lower fertiliser and pesticide use lead to higher biodiversity and characteristic species not found in other systems

Limited evidence

Flowering plants in crops and fallows.

Livestock graze fallows.

Oaks actively maintained in some areas

C. Stoate et al.

Erosion

Water

Impacts of arable intensification

The EC Habitats Directive (92/43/EEC) requires each member state to compile a list of areas containing the habitat types and species listed in the Directive. These areas will become ‘Special Areas of Conservation’ (SACs), complementing the ‘Special Protection Areas’ (SPAs) established under the EC Directive on the Conservation of Wild Birds (79/409/EEC), which together form a network of sites collectively known as ‘Natura 2000’. Directive 85/337/EEC requires environmental impact assessment to be made for projects likely to have a significant effect on the environment, which could in some cases include agricultural projects which entail major changes in land use. There is also EU legislation concerning plant protection products and organic farming.

Further attempts to reduce specific environmental impacts of arable farming come in the form of national Codes of Good Agricultural Practice and statutory arrangements, such as the UK’s Local Environmental Risk Assessment for Pesticides which places restrictions on the use of pesticides at field margins on arable land (LERAP) (MAFF, 1999b). In some European countries attempts have been made to reduce pesticide use by government mandate, including the withdrawal of approval for products considered to be high risk and encouragement of farmers to use lower doses by research and advice (Bellinder et al., 1994; Thonke, 1991).

More recently, some countries have introduced pesticide taxes (Daugbjerg, 1998) although there is a danger that such taxes encourage the use of cheaper broad-spectrum, and therefore more damaging, compounds, and there are also concerns about effects on farm competitiveness and profitablility (Falconer and Hodge, 2001).

In some EU countries, arable area payments are conditional on compliance with certain environmental restrictions on arable management, a system known as cross-compliance. Cross-compliance was introduced under the Common Rules Regulation (1259/99) of the ‘Agenda 2000’ reform of the CAP. However, similar conditions were applied in some countries prior to Agenda 2000, for example the Republic of Ireland has used cross-compliance to reduce overgrazing since 1998. The UK has also applied conditions to livestock headage payments to prevent overgrazing, and has made receipt of arable area payments conditional upon observing management conditions for set-aside land to protect species and habitats. Other countries have introduced cross-compliance in the arable sector to tackle pollution problems; for example, Denmark is using cross-compliance to enforce national regulations including compulsory fertiliser plans, the

357

maintenance of green cover over winter, and a ban on cultivating closer than 2 m from a watercourse. Finland has imposed new environmental conditions including 60 cm wide riparian buffer zones, green cover, and regulations on fertiliser and organic waste use. France is using crosscompliance to enforce water abstraction legislation, whilst The Netherlands has introduced conditions for two intensive crops: maize and potatoes, to increase the use of mechanical weed control as an alternative to herbicides. These are examples of the varied approaches taken by different member states of the EU; several countries have plans to further extend such conditionality in the future.

For cross-compliance to be effective, measures should avoid penalising farmers already adopting good environmental practices (e.g. UK farms with high hedge density and small field size) and rewarding those who do not. They should also avoid perverse effects such as piping watercourses to avoid adoption of buffer zones adjacent to them.

The Rural Development Regulation (1257/99) provides voluntary opportunities for the adoption of further ecologically, economically and socially sustainable management practices and systems within European arable systems, extending the provisions of Regulation 2078/92 under the 1992 reform of the CAP. Agri-environment options are designed to have positive impacts to counter specific longterm environmental problems. They aim to address problems that farmers can not realistically be expected to approach under cross-compliance and have an additional cost to the farmer who is compensated accordingly. As for cross-compliance, different countries have used agri-environment schemes in contrasting ways, some emphasising reductions in, or controls on resource use whilst others have encouraged the conservation of habitats and features. Examples are: the reduction in use of nitrogen fertilisers (Denmark) and protection of flower species accompanying arable land (Nordrhein-Westfalen, Germany). The UK has recently completed a three year pilot for an Arable Stewardship Scheme specifically designed to conserve wildlife and habitats on arable land.

Where possible, management practices should address more than one environmental problem.

For example, agri-environment options intended to restore soil organic matter could reduce erosion rates by increasing infiltration rates, increase breakdown of pesticides by increasing microbial activity and act as a sink for atmospheric carbon.

Riparian buffer zones have been shown to reduce loss of nutrients from arable land to watercourses (Castle et al., 1999), but can also be managed as

358

C. Stoate et al.

wetland habitats, increasing landscape diversity and biodiversity.

Greatest environmental benefits are likely to result from the integration of a number of measures, according to local conditions. Inclusion of an agri-environment option for the provision of advice and training would enable farmers to integrate fully the relevant measures, in order to maximise the environmental benefits and to accommodate compatible marketing of agricultural produce.

Agri-environment schemes can in themselves perform an important function as a training exercise for environmentally positive management (Morris and Potter, 1995). Environmental benefits of maintaining or restoring the integration of livestock or sylviculture into arable systems can be as great as, or greater than, modification of practices within either livestock, forestry or arable systems, and environmental measures must be sufficiently flexible to accommodate this approach. Whole farm plans are useful in this regard. A local scale approach promotes the integration of measures in order to achieve environmental, social and nature conservation objectives.

Such incentives are likely to be most effective as support for diversification of sustainable resource management on farmland, including tourism and production of regional and other niche products, as is currently the case in the Alentejo of Portugal.

Here, regional sheep cheeses and pork products, cork and charcoal production are integrated into arable systems.

Cultural interests of farmers, including landscape design and wildlife management for hunting, fishing and other recreation also have an important part to play in the development of multifunctional management of natural resources (Stoate, in press). Potential for influencing management practices through the market are also being developed in the form of eco-labelling of food produced under specified ecologically sensitive guidelines, and environmental certification of companies along the production chain (e.g. Udo De Haes and De Snoo, 1997), as well as an established market for organic food. Such a combination of public and private support for sustainable management of arable landscapes will avoid both the negative impacts of abandonment of arable systems (especially in southern Europe), and of continued intensification.

Demand for organic produce has increased substantially, most recently in the UK where the number of registered organic farmers increased from 828 in 1997 to 1568 in 1999 (Soil Association, 1999). However, 70% of organic cereals, and 82% of fruit and vegetables consumed in the UK

is currently imported (Soil Association, 1999) and environmental costs associated with transport are externalised. For example organic wheat is regularly imported from as far away as Australia, and the long distance transport involved contributes to gaseous emissions that may damage the environment on a large (and unnecessary) scale.

Consumption of locally produced food has environmental benefits in terms of reducing air pollution associated with transport, and increasing habitat diversity, as well as restoring social integration of rural communities (Pretty, 1998). Such local marketing is occasionally applied to arable products (e.g. Wookey, 1987), but is more generally associated with horticultural and livestock products, either within, or independently of, arable systems. It is exemplified by box schemes and farmers’ markets (Pretty, 1998). Multifunctional use of farmland, without production support, is already adopted in Switzerland (not an EU member), following a referendum on the issue in 1996 (SAEFL, 2000). Specific practical solutions to environmental problems will have greatest and most efficient environmental benefits if adopted within the context of such wider policies.

Acknowledgements This review is adapted from part of a commissioned report to the European Union Directorate General, Environment, Nuclear Safety and Civil Protection (DGXI) (Contract no. B4-3040/98/000703/MAR/D1)) (Boatman et al., 1999). We are grateful to Michael Hamell, Peter Billing and Luis Carazo Jimenez (DGXI) for their contributions to the manuscript. However, the views expressed are those of the authors and do not necessarily represent those of the Commission or the Environment Directorate.

Edward Urbansky and two anonymous reviewers helped to improve the manuscript.

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