Ecological impacts of arable intensification in Europe

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

Ecological impacts of arable intensificationin 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, LeicestershireLE7 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 effectsof 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 farmingcommunity, 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 faunaand 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 withincrops, with consequent disruption to food chains and declines in many farmland species. Abandonment of arablemanagement has also led to the replacement of such wildlife with more common and widespread species. Soils havedeteriorated as a result of erosion, compaction, loss of organic matter and contamination with pesticides, and in someareas, heavy metals. Impacts on water are closely related to those on soils as nutrient and pesticide pollution of waterresults from surface runoff and subsurface flow, often associated with soil particles, which themselves have economic andecological impacts. Nitrates and some pesticides also enter groundwater following leaching from arable land. Greatestimpacts are associated with simplified, high input arable systems. Intensification of arable farming has been associatedwith pollution of air by pesticides, NO2 and CO2, while the loss of soil organic matter has reduced the system’s capacity forcarbon sequestration. International trade contributes to global climate change through long distance transport of arableinputs and products. The EU Rural Development Regulation (1257/99) provides an opportunity to implement measures foralleviating ecological impacts of arable management through a combination of cross-compliance and agri-environmentschemes. To alleviate the problems described in this paper, such measures should take account of opportunities forpublic/private partnerships and should integrate social, cultural, economic and ecological objectives for multifunctionalland 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 originsin the eastern Mediterranean 10 000 years ago, to

1 Corresponding author: C. Stoate, The Allerton Research andEducational Trust, Loddington House, Loddington LE7 9XE,UK. Email: [email protected] Present address N.D.B.: CSL. Sand Hutton, York YO41 1LZ,UK.

one of the most widespread forms of landuse inEurope today. Table 1 shows the areas coveredby cereal, oilseed and protein (COP) crops in theEuropean Union. This development and spread hasbeen associated with the evolution of a uniquelyadapted and diverse fauna and flora (Potts, 1991,1997). However, the second half of the twenti-eth century has seen increasing concern over theimpacts of modern arable farming on agricultural

0301–4797/01/120337C29 $35.00/0 2001 Academic Press

338 C. Stoate et al.

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

1993/4 1994/5 1995/6 1996/7

B/Lux 521 522 514 524Denmark 1779 1720 1754 1819Germany 8696 8692 8640 8962Greece 1362 1346 1225 1302Spain 8830 8252 8191 8296France 12 108 11 873 11 921 12 506Ireland 296 278 281 294Italy 4746 4789 4998 5181Netherlands 425 436 434 434Portugal 855 828 750 764UK 3854 3813 3881 4020Austria 1032 1055Finland 1068 1143Sweden 1199 1275EU-12 43 472 42 548 42 589 44 102EU-15 45 888 47 575

ecosystems, and on the sustainability of arable sys-tems themselves. Environmental problems arisingfrom modern arable management are now asso-ciated with changes to landscapes and plant andanimal communities, and a deterioration in soil,water and air quality.

Few of these consequences are confined to thefarm on which they arise, the majority being ‘exter-nalised’ to become a cost to society as a whole.On-farm impacts of changes in arable manage-ment include a local simplification of landscapeand biodiversity, and a deterioration in soil char-acteristics, but these, together with many otherconsequences are more widely apparent as off-farmimpacts on biodiversity, landscape, water and air.These are often felt outside the country in whichthey originated. Such externalities are increasinglyregarded as unacceptable on both economic andethical grounds (Colman, 1994). For the UnitedKingdom (UK) alone, these costs to the commu-nity have been estimated at £2343 million per year(£208 ha�1 yr�1; Pretty et al., 2000).

Such arable intensification is exemplified byincreased adoption of external inputs such asfertilisers and pesticides, increasing scale of opera-tion, and simplification of farming systems (Meeus,1993). Although agricultural intensification hasbeen a feature of arable systems throughout west-ern Europe, it has predominated in the north,with small-scale, extensive mixed farming systemssurviving in many parts of the south. Extensivesystems also survive in many parts of easternEurope, but rapid moves towards both intensi-fication and abandonment have taken place inthe 1990s. Proposed integration of eastern Euro-pean countries into the European Union (EU)

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

Environmental measures under the CommonAgricultural Policy (CAP) were originally sup-ported under Regulation 797/85, article 19. Inthe 1992 reform of the CAP, member stateswere required under the accompanying measures(Regulation 2078/92) to develop agri-environmentschemes, with 50% of funding provided by theEuropean Community (75% in Objective 1 areas).Under the recent ‘Agenda 2000’ reform, agri-environment schemes are now supported under theRural Development Regulation (1257/99). Thesemeasures have experienced varying degrees of suc-cess in terms of their adoption by farmers, withinadequate funding, resistance to long-term obliga-tions, and reluctance to abandon traditional prac-tices being given as reasons for low adoption (Fay,1998). While some success has been reported (Bor-ralho et al., 1999; Peach et al., 2001), in other cases,conflicting objectives have led to some detrimentalconsequences being observed (Wakeham-Dawsonet al., 1998).

This paper provides an overview of the impactsof recent (post 1950) arable agriculture in Europe.We illustrate the major issues using one countryfrom northern Europe (the UK) and one fromthe south (Portugal), making additional referencesto other European countries where appropriate.Environmental impacts of arable management areoften considered in isolation, but in fact are highlyintegrated. We have treated each issue separatelyas far as possible in the sections below, describingthe environmental impacts of key managementpractices, but the aim of this paper is to provide abroad overview of the impacts of arable farming onterrestrial and aquatic ecosystems.

Throughout much of Europe, arable systems areoften highly integrated with livestock and forestry,and where this is the case we include these inour discussion of the arable system. There is atendency for such multiple land use to be moresustainable, and to be associated with higherbiodiversity and landscape value, than purelyarable systems. The combination of price supportand capital grants under the CAP has encouragedabandonment of such integrated systems, as well asencouraging environmentally damaging practicesthrough agricultural intensification.

We start this review with a description of thearable landscape and the impacts on it of changesin arable management. We then describe the influ-ence of landscape change and other consequences

Impacts of arable intensification 339

of intensification, on terrestrial wildlife. We thendevelop the review to include impacts on theecology of soils and aquatic systems, and on cli-mate through gaseous emissions. We end the paperwith a brief account of possible measures for thealleviation of these ecological impacts of arableintensification.

Cultural landscapes

Because it occupies such a large proportion ofthe European countryside, arable land makesa major aesthetic contribution to cultural land-scapes. Landscape diversity has declined in Europeduring the period of agricultural intensification(Meeus, 1993), with a tendency for the most pro-gressive farmers to create the simplest landscapes(Nassauer and Westmacott, 1987). In Britain andFrance for example, capital grants available underthe CAP have resulted in the loss of non-cropfeatures such as hedges and ditches, and thereplacement of traditional buildings with modernstructures (Burel and Baudry, 1995).

The rate of British hedge removal was greaterin the 1970s and 1980s than subsequently (Barret al., 1994; Westmacott and Worthington, 1997).More recently, changes have been more evident inthe structure of hedges, resulting from abandon-ment, than in their removal. Barr et al. (1994)reported annual rates of 5Ð2% of hedges aban-doned, compared with 1Ð3% restored, and perennialfield boundary flora, perceived by many to bean attractive landscape feature, have been lostfrom most arable farms (Boatman, 1989). How-ever, Westmacott and Worthington (1997) foundan ‘improvement’ in hedge ‘quality’ in some areas.As with many landscape assessments, perceptionsof ‘quality’ are likely to influence reported changes.Haines-Young et al. (2000) reported that by 1998,the declines in length of hedges and walls hadgenerally halted. Rates of hedge planting weresimilar to the 1980s, but rates of removal hadfallen considerably. Restoration management hadalso increased. However, there was evidence thatspecies richness of vegetation in linear habitatshad declined.

Throughout much of western Europe, mixed live-stock and arable farms have been greatly reducedin number, creating a less diverse landscape.With farm amalgamation, farm size has increased(Table 2) and farms have become more specialisedin their cropping systems, adopting simpler rota-tions than in the past. Economies of scale have ledto the creation of larger fields and ‘block cropping’,

Table 2. Average cereal area/holding (ha) (Eurostat, 1998)

1993/4 1995/6 1996/7

Belgium 8Ð7 8Ð7 10Ð0Denmark 21Ð7 23Ð2 25Ð6Germany 15Ð1 17Ð2 21Ð9Greece 3Ð3 3Ð3 4Ð1Spain 13Ð5 13Ð5 17Ð7France 17Ð8 19Ð4 24Ð3Ireland 13Ð8 13Ð8 15Ð4Italy 4Ð4 4Ð4 5Ð9Luxembourg 13Ð1 13Ð8 16Ð6Netherlands 9Ð7 9Ð7 8Ð0Portugal 2Ð5 2Ð7 4Ð3UK 40Ð8 42Ð6 52Ð0Austria 6Ð7 6Ð7 7Ð8Finland 12Ð3 12Ð3 13Ð1Sweden 19Ð1 19Ð1 20Ð3EU-12 10Ð1 10Ð8 14Ð4EU-15 10Ð4 10Ð8 14Ð2

resulting in a uniform, featureless landscape in themost intensively farmed arable areas such as partsof East Anglia in England and the Beauce region inFrance. The occurrence of colourful arable weedssuch as poppies has been reduced and some areclose to extinction (see next section).

Entec (1995) compared lowland arable landscapefeatures on British organic farms with those onconventionally managed farms. Crop diversity waslower on organic farms and fields were smaller.Hedges were taller and less intensively managedthan on conventional farms but of the farmersremoving hedges, 43% were organic. However,of farmers actively ‘improving’ hedges (or per-ceived by themselves or others to be doing so)only 11% were in conventional arable systems.Such differences relate more to the attitudes ofthe farmers themselves than to the type of farm-ing system (Entec, 1995). Chamberlain and Wilson(2000) found similar differences in hedge manage-ment between conventional and organic farms, andhigher crop diversity on organic farms.

In Portugal intensification of arable farming,often associated with the introduction of irriga-tion has led to the loss of fallows from arablesteppes, creating a more uniform landscape lack-ing the flowering plants which are often abundantin fallows and low-input arable crops (Pineda andMontalvo, 1995). Such intensification is also asso-ciated with the presence of irrigation pivots andelectricity pylons, both of which become prominentfeatures in an otherwise open landscape.

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


where they remain widespread (Harrison, 1996). InPortugal the montado area has remained relativelystable at the national scale (Direccao-Geral das Flo-restas, 1998) but arable intensification, especiallyif accompanied by irrigation, can result in the localloss of traditionally managed montado landscapes(Yellachich, 1993). Such an impact is currentlyoccurring as a result of irrigated agriculture inthe area of the Alqueva Dam, a major developmentproject funded by EU Cohesion Funds (Eden, 1996).

Because montados are a diverse and integratedsystem involving arable farming, sylviculture andlivestock, restriction of current support measuresto one component can result in the breakdownof the montado system with consequent damage tolandscape features. For example, livestock headagepayments contribute to abandonment of arable cul-tivation and to overgrazing of swards and the lossof regenerating trees in montados, resulting in anageing oak tree population structure and suscepti-bility to disease (Eden, 1996). Another example ofloss of habitat diversity through economic supportcomes from Spain where subsidies for almond cul-tivation are conditional on no more than 10% of thearea being used for another crop (Pretty, 1998).

Total abandonment of agriculture results inrapid development of shrub cover with the loss ofthe diverse flora associated with traditional arablerotations. While greater vegetative cover protectssoil from erosion, the risk of fire is considerablyincreased, once again exposing the soil to erosion(Andreu et al., 1995). Such abandonment alsoresults in the emigration of rural people, theloss of traditional skills and the deterioration ordestruction of traditional farm buildings whichthemselves form landscape features. Land thatceases to be used for arable cultivation is frequentlyplanted with Eucalyptus or Pinus species for pulpproduction, creating a uniform landscape whichis poor in terms of landscape features, and forbiodiversity.

Biodiversity

Simplification of cropping systems results inreduced crop diversity and loss of non-crop habitatssuch as grassland, field boundaries, water-coursesand trees, all of which can form an integral com-ponent of arable ecosystems. These, and the lossof livestock from arable systems, have contributedto a decline in biodiversity. Within the croppingsystem, increased application of fertilisers andpesticides, often accompanying drainage and irri-gation, has caused substantial damage to arable

ecosystems, with consequent implications for bio-diversity.

Birds provide good indicators of environmen-tal change as they are easily monitored, wellresearched, long-lived and high in the food chain(Furness and Greenwood, 1993). Many bird specieswhich live in arable landscapes have declined. Forexample, percentage declines in the 1970–1998period in UK populations given by Gregory et al.(2000) are grey partridge (Perdix perdix) (82%), lap-wing (Vanellus vanellus) (52%), turtle dove (Strep-topelia turtur) (77%), skylark (Alauda arvensis)(52%), song thrush (Turdus philomelos) (55%), treesparrow (Passer montanus) (87%), reed bunting(Emberiza schoeniclus) (52%) and corn bunting(Miliaria calandra) (85%). Similar declines in farm-land species have been experienced across Europe,with 42% of declining species being affected by agri-cultural intensification (Tucker and Heath, 1994).Such severe declines are not occurring for speciesassociated with other habitats (Gibbons et al., 1993;Crick et al., 1997; Gregory et al., 2000).

Declines have also been documented in arableinvertebrates and plants (Eggers, 1984; Potts,1991; Andreasen et al., 1996; Donald, 1998). Agood example of the interrelationship betweenthese groups is provided by Game ConservancyTrust work in Sussex, England, where impactsof agricultural change have been demonstratedthroughout the food chain (Potts, 1986; Ewald andAebischer, 1999).

Agri-environmental measures introduced underRegulation 2078/92, and later under the RuralDevelopment Regulation 1257/99, have attemptedto alleviate some of these adverse consequences ofintensification within arable systems. For example,in England a pilot Arable Stewardship Scheme hasprovided opportunities to restore field boundaryhabitats, reduce pesticide inputs, plant crop mix-tures specifically designed for wildlife and main-tain cereal stubbles through the winter, therebyimproving breeding habitats, foraging habitats andwinter food supplies for birds. Such incentives areintended to replace previously widespread arablehabitats that have been lost through changes incultivation, cropping patterns and arable inputs.Set-aside has been associated with some conserva-tion benefits for birds (Wilson et al., 1995) but suchbenefits are well below those possible under moreappropriate management, as agricultural priori-ties have determined the management of set-aside(Winter and Gaskell, 1998).

Organic farming has also been shown to benefitsome species. Recent studies in England suggestthat organic systems support more broad-leaved


plants than conventional systems which are asso-ciated with more grass weeds and cleavers (Galiumaparine) (Kay and Gregory, 1999; Hald, 1999;Brooks et al., 1995). Kay and Gregory (1999) foundthat, out of 23 rare or declining arable plant species,18 were more abundant on organic farms, with 13of them being absent on conventional farms. How-ever, increasing efficacy of mechanical weed controltechnology could reduce these differences in plantabundance and species richness between the twosystems (van Elsen, 2000).

Some studies have found that aphid abundanceis higher in conventional systems, with someother invertebrate groups being more abundantin organic systems (Redderson, 1997; Morebyet al., 1994). Brooks et al. (1995) found higherearthworm numbers in organic fields, and Feberet al. (1997) reported higher numbers of non pestbutterflies in organic systems (pest species wereequally abundant in both systems). There is someevidence for higher numbers of birds in organicsystems, especially skylarks (Wilson et al., 1997).However, ground-nesting species such as this couldbe adversely affected by the frequent mechanicalcontrol of weeds in organic crops when the birdsare nesting (Hansen et al., 2001). Abundance ofother species is influenced mainly by field boundarycharacteristics (Chamberlain and Wilson, 2000)which are influenced by farmers conservationattitudes (Entec, 1995).

Cultivation and crop rotation

Cropping systems have been simplified and becomegeographically polarised in northern Europe (deBoer and Reyrink, 1989), with major consequencesfor biodiversity within arable ecosystems. Highcrop diversity is necessary to the ecological require-ments of many species. For example, brown hares(Lepus capensis) graze different crops at differ-ent times of year (Tapper and Barnes, 1986),while skylarks move breeding territories from onecrop to another through the breeding season (Wil-son et al., 1997), and yellowhammers (Emberizacitrinella) switch from one crop to another as for-aging habitats during the breeding season (Stoateet al., 1998). Lapwings require cereals in which tonest and adjacent pasture on which to feed newlyhatched young (Tucker et al., 1994) while littlebustard (Tetrax tetrax) males and females havedifferent habitat requirements during the breed-ing season (Salamolard and Moreau, 1999). Aswell as this, inter-species differences in habitatrequirements result in higher numbers of species

in landscapes with low intensity farming and highcrop and structural diversity, as demonstrated forthe Alentejo region of Portugal by Stoate et al. (inpress). Monitoring of a Zonal Programme intro-duced to this region under 2078/92 suggests thatthis measure is maintaining bird species diversity(Borralho et al., 1999).

In Britain, geographical polarisation of arableand livestock farming has reduced the number offarms with high crop diversity. Pastures grazedby livestock are associated with large numbersof invertebrates which provide food for birds andother animals, and livestock feed sites in winterprovide a source of food for seed-eating birds. Theloss of livestock from farms in eastern Britainhas removed these components from the arablelandscape.

Intensification of grassland management andconversion of grass to arable cultivation resulted ina 92% decline in the area of ‘unimproved’ grasslandin the UK between the 1930s and 1980s (Fuller,1987), with a consequent decline in biodiversity,especially that of plants and invertebrates associ-ated with semi-natural grasslands (Green, 1990).In the Netherlands, plants and invertebrates (espe-cially dragonflies (Odonata) and butterflies (Lepi-doptera)) associated with semi-natural grasslandand forest edges are also declining (IKC-NBLF,1994) and desiccation of grasslands has resultedin declines of wading birds (Klumpers and Haart-sen, 1998; Reyrink, 1989). Newbold (1989) reportedlosses or serious damage to 80% of British calcare-ous grassland between 1949 and 1984. Loss ofdiverse grassland ecosystems to arable crops inthe UK has continued since then. Flax crops qual-ify for subsidies even when grown on land that isnot eligible for arable area payments, and thesesubsidies exceed the value of payments for protec-tion of grassland habitats. Previously uncultivatedgrassland is also currently being lost to potato pro-duction, as disease-free and pesticide-free land isoften required for this crop (CPRE, 1999). Produc-tion of potatoes is discouraged on arable land asthis crop is not eligible for area payments.

In Portugal and other Mediterranean countries,livestock plays an important ecological role inarable systems grazing fallows and influencingtheir botanical and invertebrate composition. Inparticular, fallow periods of several years providerelatively stable habitats which contribute to themaintenance of plants, invertebrates and birdscharacteristic of steppic arable systems in Portugal(Beaufoy et al., 1994; Moreira and Leitao, 1996;Stoate et al., in press). The recently disturbedsoil associated with first year fallow represents an


important foraging habitat for many seed-eatingbirds in winter (Diaz and Telleria, 1994). Fallowperiods vary with soil type, and this disparityhas increased since the 1970s. Intensification ofarable cropping on the best soils has led tothe complete loss of fallows from some systems,while cultivation has been abandoned in the leastfertile areas. In the latter, the loss of crops andencroachment of scrub have resulted in the lossof fauna associated with the extensive arablelandscapes, and in frequent cases land use convertsto forestry with financial support through EUregulation 2080/92. The area of Portuguese arableland converted to forestry under this regulationbetween 1994 and 1998 amounted to 138 715 ha(Portuguese Ministry of Agriculture, unpublisheddata). Such abandonment and changes in land useare currently a greater threat to biodiversity ofarable systems in southern Europe than the risk ofintensification.

Arable steppe landscapes support many of themost threatened species, and are therefore criticalfor maintaining species diversity at the nationaland European level. Such species include the glob-ally threatened bird species, great bustard (Otistarda) and lesser kestrel (Falco naumanii) aswell as many others of current conservation con-cern in Europe (Tucker and Heath, 1994). Theloss of fallows and other consequences of arableintensification are a direct threat to these species(Peris et al., 1992). However, at the local levelbird species diversity is higher in lightly woodedMontado (Portugal) and Dehesa (Spain) in whichholm oak (Quercus ilex) and cork oak (Q. suber)have traditionally been managed as part of thearable system. This habitat has been threatened bywheat-growing campaigns (especially in Portugal)since the 1930s, increased mechanisation, reducedregeneration of trees due to increased livestockdensities, and reduced demand for tree products(Yellachich, 1993; Diaz et al., 1997). Increasing useof plastic ‘corks’ in wine bottles could potentiallydamage the market for natural cork from Monta-dos. This could lead to the loss or deteriorationof Montados which support high levels of wildlifediversity and abundance (Araujo et al., 1996).

Livestock densities have also increased in arablesteppe landscapes, resulting in destruction ofthe vegetation used by invertebrates and birdsassociated with fallows. Such stocking densitieshave been encouraged by headage payments onsheep and cattle, and by increased labour costsleading to replacement of traditional herders bywire fences. In some parts of Spain the loss of thepredator control role of herders has resulted in

unsustainable levels of nest predation for scarcebreeding lark species (Suarez et al., 1993). Higherstocking densities on arable fallows are likelyto reduce the orthopteran food of bustards andother already threatened species (van Wingerdenet al., 1997). Such livestock densities could alsoincrease the area of cereals grown for fodder thanencouraging earlier harvesting.

Timing of the management of crops can alsoinfluence their suitability for invertebrate andbird species. Lapwings and skylarks breeding innorthern Europe favour spring-sown cereals, inpart because of their structure and less inten-sive management and these species have thereforebeen affected by the reduced area of these crops(Tucker et al., 1994; Odderskær et al., 1997). Wheregrass for grazing, hay or silage is a componentof the arable rotation, spring-sown cereals havehistorically been associated with undersowing ofgrass leys. This management practice encouragessawflies (Hymenoptera: Symphyta) which over-winter as pupae in the undisturbed soil and whoselarvae provide an important food sources for breed-ing birds such as grey partridge, skylark and cornbunting (Ewald and Aebischer, 1999). In Sussex(England) the distribution of breeding partridgesand corn buntings is closely related to that ofundersown arable leys (Potts, 1997; Aebischer andWard, 1997) and availability of invertebrates suchas sawflies is related to productivity of these species(Potts,1997; Brickle et al., 2000).

Spring sowing on light soils in Britain is tra-ditionally associated with the survival of cerealstubbles into the winter, thereby providing a foodsource for seed-eating birds. Modern machinerypermits rapid harvesting and ploughing of landfor subsequent crops, with the result that stub-bles now remain available to wildlife for a veryshort period. This loss of stubbles has been associ-ated with declines in numbers of many finches andbuntings (Chamberlain et al., 2000). In the Nether-lands the near extinction of hamsters (Crisetuscrisetus) is thought to be due to the short periodin which this species may gather grain beforehibernation (van Oorschot and van Mansvelt,1998). Changes from spring- to autumn-sowing arealso thought to have contributed to declines inmany formerly common spring-germinating arableplants such as corn marigold (Chrysanthemumsegetum) and night-flowering catchfly (Silene noc-tiflora) (Wilson, 1994a). Botanical composition ofarable crops is also influenced by cultivation meth-ods. Recent economic pressures have encouragedincreased adoption of minimum tillage and directdrilling in preference to conventional ploughing


(Jordan et al., 2000a). While ploughing encour-ages broad-leaved weed species that are importantinvertebrate and bird food, direct drilling encour-ages grass weeds and the pernicious broad-leavedweed, cleavers (Galium aparine). However, directdrilling leaves weed seeds near the soil surface, pro-viding more food for birds in winter (Higginbothomet al., 2000).

Cultivation methods can also affect the com-position of invertebrate communities in arableecosystems. Ploughing is the most destructive,affecting invertebrate populations through phys-ical destruction, desiccation, depletion of food andincreased exposure to predators. Earthworms aremost abundant in arable systems with minimalsoil disturbance such as ‘integrated crop manage-ment’ systems and organic systems that do notinclude potatoes in the rotation (Higginbothamet al., 2000). Shallow-burrowing species however,are more abundant in integrated crop manage-ment than either conventional or organic systems.As well as being an important component of thesoil fauna, earthworms are a major component ofthe diet of some birds and mammals. Large cara-bid beetles are often more abundant in ploughedfields than where minimal cultivation is used, withsmaller 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 methodson arthropods are currently the subject of researchin Britain. Little is also known about the effectsof minimal tillage on vertebrates, although Bel-monte (1993) suggests that such methods might bebeneficial to some birds in Spain.

Fertilisers

Use of fertilisers on arable land increased sub-stantially during the second half of the twentiethcentury, but there has been a decline in use morerecently. In France the use of nitrogen fell by 10%between 1986 and 1994, phosphates by 20% andpotassium by 13%. Total fertiliser use has alsodecreased in the Netherlands and the UK althoughin Portugal the trends are less obvious (Table 3).As market prices of arable crops have fallen,economic pressure for more accurate linking of fer-tiliser application rates to crop needs has increased.However, as modern arable crop varieties requirerelatively high rates of fertiliser application, fer-tiliser use remains high. ‘Precision farming’, i.e.varying application spatially in response to spatialvariations in soil fertility or other parameters, may

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

1980 1985 1990 1994

United Kingdom 2054 2524 2413 2219Netherlands 679 701 558 534Portugal 259 241 278 248

allow more accurate targeting in future, but stillrequires further development (Lu et al., 1997; God-win, 2000).

Modern crop varieties grow vigorously underhigh rates of fertiliser application, out-competingother arable plants, and increases in the useof fertilisers have contributed to a change inthe arable flora (Wilson, 1994a). The dense cropstructure associated with high levels of fertiliserapplication is also unsuitable for some birds as ahabitat for nesting and foraging (Wilson et al.,1997). The increased use of mineral fertilisersduring the second half of the twentieth centuryhas also influenced non-crop habitats associatedwith the arable system. Deposition of fertiliserin perennial vegetation at the field edge hascontributed to a change in botanical compositiontowards annual weeds such as cleavers and barrenbrome (Bromus sterilis) (Boatman et al., 1994).Such changes encourage the perception amongfarmers of field boundaries as a source of weeds,leading to the further destruction of this habitatthrough deliberate or accidental use of herbicides,ploughing into the field edge, and in many casescomplete removal (Boatman, 1989). Loss of thisperennial herbaceous field boundary habitat hassevere implications for wildlife associated with it(e.g. whitethroat (Sylvia communis) Stoate andSzczur, 2001).

Pesticides

Although pesticide use per hectare is lower onarable land than many other crops (e.g. mush-rooms, flowers, bulbs and fruit), the large areacovered by arable crops results in this land-usebeing the largest user overall (European Commis-sion, 1999). Pesticide use increased substantiallyduring the second half of the twentieth century, buthas declined slightly in recent years. Between 1991and 1996 the largest decreases in pesticide saleswere seen in those countries which have specificpolicies on pesticide reduction: Finland (�46%),the Netherlands (�43%), Denmark (�21%) andSweden (�17%) (Eurostat, 1998).


However, as methods of monitoring pesticideuse differ between countries, and as pesticideuse can fluctuate from year to year in responseto climatic and other factors, short-term assess-ments and comparisons are difficult. In 1996 therewas an increase in sales volumes, particularly inSpain (C19%), France (C11%) and the UK (C6%)because of seasonal conditions such as weather andpest pressure. Overall, sales were divided betweenfungicides (41% weight of active ingredient), her-bicides (39%), insecticides (12%) and others (8%)(European Commission, 1999). However, there issome variation between regions, with for examplefungicides forming a larger proportion of pesticidesused in southern Europe and herbicides predom-inating in the north. France is the largest EUmarket 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 havebeen greatly reduced since the phasing out oforganochlorines, although rodenticides continue tobe a cause of secondary poisoning of barn owls (Tytoalba) in areas of warfarin resistance (Shawyer,1987). The molluscicide methiocarb has also beenshown to cause mortality of wood mice (Apodemussylvaticus) under field conditions (Johnson et al.,1991). De Snoo et al. (1999) report few poisoningincidents resulting from arable use of pesticides inEurope but suggest that the efficacy of monitoringis uncertain and variable between countries.

The arable flora has changed in response toarable intensification. In Britain Wilson (1994a),using 10 km squares as sampling units, reportedthat the arable flora included 25 species that wererecorded from fewer than 100 squares, at least26 that were recorded from fewer than 15, anda further seven that had recently become extinct.Several of the rarest species no longer occur inarable habitats, but for others, the most suitableareas remain the calcareous and sandy soils ofsouth and east England (Wilson, 1994b). Similardeclines in the arable flora have been reportedfrom Germany (Eggers, 1984; Agra Europe, 1991),The Netherlands (Joenje and Kleijn, 1994) andDenmark (Andreasen et al., 1996). Arable floraare highly concentrated in field margins (Wilsonand Aebischer, 1995; Joenje and Kleijn, 1994; DeSnoo, 1997) where biomass, density and speciesdiversity are reduced by herbicide use (Chivertonand Sotherton, 1991). Leaving the outer few metresof a crop unsprayed with herbicide can have apositive effect on the presence and abundance ofplant species (Schumacher, 1984; Chiverton andSotherton, 1991; Hald et al., 1994; De Snoo, 1997).

Set-aside contributes to a reduction in over-all pesticide use on arable land, but an increasein the use of non-selective herbicides (especiallyglyphosate), which are known to affect field bound-ary vegetation structure and botanical composition,and the composition of associated invertebrate com-munities (Haughton et al., 1999).

Pesticide use in southern Europe is well belowthat in the north, but declines in some arable plantshave been reported following herbicide use, coupledwith increased fertiliser use and abandonment offallows (e.g. the Alentejo, Portugal [Moreira et al.,1996a]). Species threatened by such intensificationinclude Linaria ricardoi and Euphorbia transta-gana, both of which are included on the Directive92/43/EEC (Conservation of natural habitats andwild fauna and flora).

Herbicide use in arable crops is known to havea negative impact on invertebrate abundance andspecies diversity (Chiverton and Sotherton, 1991;Moreby et al., 1994; Moreby, 1997). Direct effects ofinsecticides are a major influence on invertebratecommunities (e.g. Moreby et al., 1994), althoughthe effects differ between species, depending in parton their ecology (Grieg-Smith et al., 1995). Somefungicides have also been implicated in influencinginvertebrate abundance (Sotherton et al., 1987;Reddersen et al., 1998).

In Sussex (England) there was a negative rela-tionship between broad-leaved weed abundanceand the use of dicotyledon-specific herbicides, andbetween grass weeds and broad-spectrum herbi-cides (Ewald and Aebischer, 1999). Spring andsummer use of herbicides was particularly effec-tive at reducing broad-leaved weed abundance. Ofthe five invertebrate groups studied, all showeda negative relationship between abundance andthe use of insecticides, and declines of four ofthem were associated with fungicide use. Partic-ularly strong effects were noted for the pyrethroidand organophosphate insecticides, but none of thegroups showed a negative relationship with useof the more selective insecticide, pirimicarb, sug-gesting that broad-spectrum insecticides are mostdamaging to cereal ecosystems. Broad-spectruminsecticides such as dimethoate continue to be themost widely used (Potts, 1997) and can cause sub-stantial damage to populations of beneficial arableinvertebrates and honeybees (Greig-Smith et al.,1995). Declines in British bumblebees have alsobeen linked to use of pesticides in arable crops(Williams, 1982). The molluscicide methiocarb istoxic to a wide range of fauna, and has been shownto affect populations of carabid beetles in the field(Purvis and Bannon, 1992).


Arable invertebrates are an essential compo-nent of the diet of many farmland birds duringthe breeding season and their decline has beenlinked most convincingly to the substantial declinein grey partridge numbers in Britain (Potts, 1986;Potts and Aebischer, 1991). In Sussex, partridgedensity was inversely related to the number ofherbicide applications, and positively related tothe number of weed taxa (Ewald and Aebischer,1999). The abundance of corn buntings and sky-larks was also inversely related to herbicide use,and that of corn buntings with fungicide use. Allthree species have declined substantially through-out their range in northern Europe (Tucker andHeath, 1994; Flade and Steiof, 1990; Fuller et al.,1995) and the Sussex results provide further evi-dence for the impact of pesticides on the abundanceof farmland birds as well as invertebrates. Nestlingsurvival of corn buntings is inversely related toinvertebrate abundance, especially Symphyta andLepidoptera (Brickle et al., 2000), which occur athighest densities in relatively low-input arable sys-tems incorporating undersown leys (Aebischer andWard, 1997). Corn buntings are also strongly asso-ciated with low input arable systems in Portugal(Stoate et al., 2000).

Current use of herbicides and cropping prac-tices combine to reduce production of weed seedson arable land (Jones et al., 1997), and for someseed-eating birds a reduction in the area of weedywinter stubbles is thought to have contributed toincreased winter mortality and subsequent pop-ulation declines (Campbell et al., 1997). Wherethey still occur, weedy winter stubbles are stronglyfavoured as a foraging habitat by finches andbuntings such as cirl buntings (Emberiza cirlus)(Evans and Smith, 1994) and corn buntings (Don-ald and Evans, 1994).

Herbicide tolerance in genetically modified (GM)arable crops could, in the near future, lead toincreased use of very-broad spectrum herbicides,and more complete removal of arable plants and theother wildlife dependent on them. The use of broad-spectrum herbicides on such genetically modifiedcrops could also result in even greater damageto adjacent habitats than is currently the case.There is also the danger that arable weeds couldthemselves acquire herbicide tolerance throughcross-pollination with GM crops. Insect resistancein GM crops could reduce the use of insecticideson arable land, with consequent benefits for otherwildlife, but the impact of such GM crops on thenatural predators of crop pests is not well under-stood. There is currently inadequate information

available on the environmental impacts of GMcrops.

Drainage and irrigation

In northern Europe, large areas of grasslandhave been drained for conversion to arable cropproduction since the 1940s, but remaining wetgrassland habitats have also been severely affectedby drainage of adjacent arable land (Mountford andSheail, 1984; Williams and Bowers, 1987). In theNetherlands about 60% of the lowering of watertables is caused by draining of adjacent arablefields (RIVM, 1998). As a result there have beensubstantial declines in abundance and diversityof birds, plants and invertebrates associated withwet grassland habitats in northern Europe (deBoer and Reyrink, 1989). Baldock (1990) highlightsrapid changes to wet grassland habitats in Franceresulting from agricultural intensification.

In southern Europe the area of irrigated cropshas increased considerably since the 1960s, tak-ing the form of pivot irrigation in formerly dryareas, and flooded rice in low-lying areas (Suarezet al., 1997; Fasola and Ruız, 1997). Pivot irri-gation of crops such as maize is associated withincreased fertiliser and pesticide applications andthe environmental impacts of irrigation are there-fore largely those of these inputs, including the lossof fallows in crop rotations.

Irrigation can result in the loss of wetland habi-tats. For example, in Las Tablas de Daimiel ofSpain subsidised expansion of the irrigated areahas resulted in a 60% reduction in the size ofthis designated Ramsar wetland of internationalimportance, Special Protection Area and Natura2000 site (WWF, 2000). Drainage for irrigationhas resulted directly in the local extinction ofarable plants such as Armeria arcuata (Moreiraet al., 1996b) and subsequent use of herbicidesand fertilisers have a wider impact on the arableflora (Moreira et al., 1996a). Intensification asso-ciated with irrigation can also reduce invertebrateabundance. Although Barranco and Pascual (1992)reported more Orthopteran species in irrigatedthan dryland cereals in Almeria, Stoate et al.(2000) found higher numbers of Orthoptera inPortuguese extensively managed cereals than inintensive systems.

In contrast, irrigated rice can increase the localdiversity of birds and the aquatic invertebrateson which they feed (Fasola and Ruız, 1997)if pesticide use is not high. It can serve aparticularly valuable role in the conservation of


wetland wildlife, including breeding, winteringand migratory birds, where rice is grown closeto estuary habitats, as is the case in the Ebrodelta (Spain) (Pain, 1994). Breeding abundanceof six species of heron, as well as white stork, isrelated to area of flooded rice close to their breedingsites, and rice fields and the irrigation channelsassociated with them are favoured foraging areas inPortugal and elsewhere in southern Europe (Pain,1994; Coelho, 1998). When flooded in the winter,to encourage ducks for hunting, they provideimportant feeding areas for wading birds (Pain,1994; Fasola and Ruız, 1997). Substantial declinesin the rice crop area, and a recent trend towards drycultivation of rice in southern Europe could reducethe current ecological role of this crop (Fasola andRuız, 1997).

In Italy, the EU member state with the largestrice-growing area, the increasing scale of rice-growing operations has had substantial impactson rice fields as a habitat for wildlife (Pain, 1994).Fields have been enlarged, reducing the habitatdiversity and area of uncropped habitat, and theuse of lasers enables farmers to produce levelground, lacking wet patches in which aquaticanimals can survive when fields are drained.Such precision equipment also enables rice to begrown in shallower water which reaches highertemperatures and is less suitable for aquaticanimals (Pain, 1994).

Soil

Soils are affected by the method, timing andfrequency of cultivation, and by soil type andtopography. For example, compaction resultingfrom the use of heavy machinery and decliningorganic matter resulting from frequent cultivationsaffect soil structure and composition. Inputs inthe form of pesticides and organic and inorganicfertilisers also influence soil structure directly, andthrough their impact on the soil fauna.

Simplification of cropping systems, increases infield size and increased use of heavy machineryand pesticides have all contributed to higher levelsof soil erosion (Evans, 1996), as described below.

Evans (1996) reports mean annual rates of soilloss from arable land of 3Ð6 t ha�1 in Belgium and6Ð1 t ha�1 and 5Ð1 t ha�1 in the English counties ofSomerset and Hampshire respectively. EuropeanUnion Environment Directorate (DGXI) estimatesof 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 asingle storm (De la Rosa et al., 2000).

Soil structure

Sandy and peaty soils are particularly susceptibleto erosion by wind and the peat area of the EastAnglia (UK) arable area has declined considerablyas a result of erosion, with the peat depthin remaining areas being substantially depleted.Erosion of cultivated soils resulting from the actionof wind and water leads to loss of nutrients and croprooting depth as well as to pollution, eutrophicationand sedimentation of aquatic habitats (see below).Off-farm costs of erosion resulting from damageto property, roads and communications, pollutionof waterways and drinking water, sedimentation ofreservoirs, and damage to fisheries are externalisedto society and are considerably greater than on-farm costs incurred by farmers.

Loveland et al. (2000) recorded a decreasein mean soil organic carbon from arable leysites of 0Ð49% over a 15 year period (organiccarbon 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, amountingto 1Ð42 t ha�1 yr�1 (expressed as carbon). Walling(1990) suggests that organic matter is eroded fromarable land disproportionately to its availability inthe soil. Such loss of organic matter has severeimplications for the soil. It can lead to reducedwater retention within the soil and to consequentdrought in dry regions, and to reduced drainagein wet ones (Benckiser, 1997). Organic matterwithin the soil also serves an important functionin reducing leaching of pesticides to water throughadsorption and higher microbial activity.

Organic matter levels are higher in arablesystems incorporating livestock or legume-basedleys than purely arable systems (Drinkwater et al.,1998). However, Fox (2000) has demonstrated thatnon-inversion tillage within purely arable systemscan increase soil organic carbon and microbialbiomass. Although recent adoption of this reducedcultivation has been driven primarily by economicconsiderations (Jordan et al., 2000a), it may alsohave environmental benefits for the ecology ofboth soil and aquatic ecosystems. Jordan et al.(2000b) indicate that sediment loss from arableland can be reduced by 68% and phosphate lossby 81% under minimum tillage, compared withconventional ploughing.


Owing to drought susceptibility, shallow soils areassociated with lower cereal yields (Loveland et al.,2000). De la Rosa et al. (2000) have estimated thatover the next 100 years the loss in crop yield innorthern Europe would be around 0–5%, with mostsites being at the lower end of this range. Directeffects on yields in the shorter term are not feltunless soil organic carbon levels fall below 1%, acondition currently associated with, for example,only about 5% of UK arable land (Loveland et al.,2000). Changes in soil are generally slight duringthe period of a farmer’s life and environmentalproblems associated with erosion are externalised.The incentive for preventative action on the part ofthe farmer is therefore low.

An increased area of autumn cultivation,increases in field size and associated loss ofhedges, and continuous arable cropping all increasethe exposure of soil to wind and water (Evans,1996). Lack of crop cover and the presence oftramlines and wheelings also increase erosionrates on arable land (Chambers et al., 1992),and late harvested spring-sown crops such asmaize (increasingly planted as a silage crop), sugarbeet, potatoes and other vegetables are associatedwith relatively high levels of erosion becausesoils are exposed to rain during the autumnand winter (Evans, 1996). Recent incorporation ofoutdoor pigs into arable systems has also becomea major local source of soil sediment. Rainfall,slope and soil type can all be major influenceson erosion risk (MAFF, 1999a). Alstrom andBergman (1990), working in Sweden, emphasisedthe influence of slope, slope length and areawhile highest rates of soil erosion are generallyassociated with storm events (e.g. Boardman,1990).

In southern Europe, erosion rates are much morevariable, but generally higher than in the northand influences of slope, rainfall and crop typeare greater. De la Rosa et al. (2000), workingin Portugal, Spain, Italy and Greece, reportederosion rates of 2Ð8–150Ð0 t ha�1 yr�1, dependingon site and crop type (Table 4). These rates couldresult in a severe decline in crop yield, as wellas having a deleterious impact on terrestrial andaquatic ecosystems. In the most severe case (inthe Algarve, Portugal) De la Rosa et al. predicta 97% yield reduction within 100 years, althoughabandonment is likely considerably earlier thanthis. Much arable land is already abandoned inthis region through the interaction of shallow soils,declining rainfall and economic and structuralconstraints.

Table 4. Estimates of soil erosion rates (t ha�1 yr�1)Ł for20 European sites and three crop types, derived fromImpelERO soil erosion models (De la Rosa et al., 2000)

Wheat Sugar beet Sunflower

MediterraneanPortugal 67Ð0 87Ð3 99Ð9Portugal 147Ð9 150Ð0 150Ð0Portugal 3Ð9 4Ð9 6Ð0Spain 2Ð8 3Ð8 4Ð6Greece 3Ð9 4Ð9 6Ð0Italy 4Ð4 5Ð9 7Ð1Italy 78Ð0 100Ð1 115Ð7AtlanticIreland 9Ð8 18Ð4 29Ð4England 1Ð4 1Ð4 1Ð8Scotland 1Ð4 1Ð4 1Ð8Netherlands 1Ð4 1Ð6 2Ð0France 1Ð4 1Ð6 2Ð0ContinentalFrance 1Ð4 1Ð6 2Ð0France 1Ð4 1Ð4 1Ð8Germany 1Ð4 1Ð6 2Ð0Germany 4Ð0 4Ð6 5Ð7Denmark 3Ð8 4Ð0 4Ð9Denmark 1Ð4 1Ð4 1Ð8Luxembourg 1Ð4 1Ð6 2Ð0Luxembourg 1Ð4 1Ð6 2Ð0Ł tons per square mileD.x/1Ð016)/259

Autumn ploughing in the Alentejo (Portugal),before the main rainfall period, can account fora soil loss of 6 t ha�1 per tillage operation onslopes (Bergkamp et al., 1997). Ploughing of fal-lows continues through the spring in order tostimulate mineralisation of organic matter, priorto sowing the next crop. Barreiros et al. (1996a, b)demonstrated that no-tillage systems can decreaserunoff and increase soil bulk density, reducing ero-sion by 60% that of ploughed land. However, theapplicability of such management will vary withsoil type.

Kosmas et al. (1997), reporting on the MEDALUSstudy in Portugal, Spain, France, Italy andGreece, found that erosion rates were relativelylow in shrubland and other areas where peren-nial and herbaceous vegetation cover exceeded90%. However, erosion rates under Eucalyptusplantations (often advocated as an alternativeland use for degraded arable land) were slightlyhigher than on rainfed arable land. Roxo andCortesao Casimiro (1997) also found that ero-sion was lower under naturally regenerating shrubor herbaceous vegetation than arable land. How-ever, rainfall has declined since the 1930s andhas become increasingly erratic, both within yearsand months (Bergkamp et al., 1997), and this


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

Throughout Europe, the use of heavy machineryand frequent passes with cultivating equipmentcan cause soil compaction, increasing runoff at thesoil surface and creating a soil pan within thesoil. The latter inhibits drainage, causing water-logging of crop plants on some soils and creatinga physical barrier for their roots, making themmore susceptible to drought in dry conditions. Soilcompaction is a particular problem on soils withlow organic matter where earthworm abundanceand activity is low (Makeschin, 1997).

Earthworms help to maintain soil structurein soils that are not compacted and maintainhigh organic matter, improving aeration, croproot growth and drainage (Marinissen, 1992;Makeschin, 1997). Worms also distribute beneficialprotozoa and mycorrhizal fungi (Makeschin, 1997),and can reduce leaching of nitrogen by increasingnitrification of soluble nitrates (Elliott et al., 1990),but they increase N flux in the autumn (when cropuptake is low), as well as in the spring (Whalen andParmelee, 2000). Soil compaction reduces abun-dance of microfauna, especially in deep-tilled soils(Schrader and Lingnau, 1997) and causes anaero-bic conditions and changes in microbial communitystructure (Bamford, 1997).

Although direct drilling and minimum tillage areassociated with some crop pests, they also favourthe soil fauna responsible for breakdown andmineralisation of soil organic matter, especiallydeep burrowing worm species such as Lumbricusterrestris (Edwards, 1984; Jordan et al., 2000b).Shrader and Lingnau (1997) found higher earth-worm densities in integrated (reduced cultivation,pesticide and mineral fertiliser applications) thanin conventional arable systems. Conventional cul-tivation is especially damaging to soil fauna insemi-arid low organic matter soils (Bamford, 1997).

Nutrients

In southern Europe, soil nutrients have becomeseverely depleted since the beginning of the cen-tury following the adoption of cropping systemsusing little or no manure or fertiliser (Esselink andVangilis, 1994), contributing to wide scale aban-donment. In northern Europe nutrients are lostfrom soils which continue to receive high levelsof nutrient inputs. Erosion is the main cause ofphosphate loss from the soil. Leaching of nitrogencan result from applications of mineral fertiliser at

rates above those required by the crop at the timeof application. However, much of the nitrogen lostfrom soil is now known to be associated with min-eralisation of soil organic matter at a time whenthere is no crop cover to exploit the mineral nitro-gen made available, normally the period followingharvest (Bloem et al., 1994). Soil cultivation inwarm wet conditions after harvest maximises min-eralisation and loss of nitrate during the periodbefore the following crop becomes established.

Organic matter reduces soil erosion, improvessoil moisture retention and supports soil meso-fauna that maintain appropriate soil structurefor crop growth (Benckiser, 1997). Crop rotationswhich incorporate grass leys improve soil organicmatter and reduce loss of nitrogen through leach-ing, but can be associated with high levels ofleaching when they are ploughed (Young, 1986).The use of farmyard manure increases soil organicmatter and releases nitrogen more gradually thanan application of mineral fertiliser, but the min-eralisation and subsequent availability of nitrogendoes not necessarily match the requirements ofthe crop, with the result that leaching can occur(Laanbroek and Gerards, 1991). Manure from poul-try is especially associated with excessive ratesof mineralisation (Benckiser, 1997), while applica-tion of slurry from intensive livestock systems canbe toxic to some earthworm species (Makeschin,1997). However, nitrate leaching can be lower inarable systems incorporating livestock or legume-based leys than in conventional arable systems(Drinkwater et al., 1998).

In the Netherlands ‘integrated’ systems withreduced cultivation, nitrogen and pesticide appli-cation were found to maintain organic matter(whereas this declined in conventional systems)and have crop yields which were 90% of thoseobtained from conventional systems (Lebbink et al.,1994). Most groups of organisms also had higherbiomass in integrated than conventional systems(Zwart et al., 1994). Protozoa and nematodes weremore abundant, and mineralisation of nitrogenwas therefore higher in the integrated system, butexcessively high in both after harvest (Bloem et al.,1994). Didden et al. (1994) reported that faunalmineralisation was 49% and 87% of the total min-eralisation in conventional and integrated systemsrespectively.

Regional production of sewage and manure areincreased by 20% as a result of imported foodand fodder (Benckiser, 1997). In 1999 42% ofUK sewage production was applied to agricul-tural land, and the quantity used was predicted todouble by 2006, following the ban on sewage sludge


disposal at sea in 1998 (Chambers, 1998). Sludgecake applications supplying 250 kg ha�1 total N willtypically provide about 4 t ha�1 of organic matter,and liquid sludge about 3 t ha�1, to the soil, reduc-ing erosion risk and increasing moisture retention(Chambers, 1998). However, current usage repre-sents less than 1% of British agricultural land(Chambers, 1997). Application of sewage sludge isassociated with high soil phosphate accumulationsand with problems of heavy metal accumulationin the soil (Ribeiro and Serrao, 1996; Benckiser,1997). These heavy metals reduce and changethe composition of the microfauna, hampering soilmetabolism and reducing degradation of pesticides(Benckiser, 1997).

Pesticides and heavy metals

In the Netherlands 123 000 kg of pesticide (activeingredient) is known to reach the soil in arablefields, mainly via droplet drift (MJP-G, 1997). Pes-ticides can influence soil structure and nutrientstatus through their action on soil fauna and flora,while some pesticides are themselves degradedby soil fauna (AFRC, 1990), the rate at whichthis occurs varying between compounds (Table 5).Although most herbicides are not toxic to soilfauna (Bamford, 1997), those that are include thetriazines such as atrazine (Edwards, 1984) andbipyridyl herbicides such as paraquat. Althoughparaquat has been shown to have insecticidal prop-erties in the laboratory, impacts on invertebratepopulations in the field are probably low (Wrightet al., 1985). However, Edwards (1984) suggestedthat herbicides can indirectly reduce soil organicmatter and the organisms associated with it by

preventing the growth and eventual decay of weedswithin the crop.

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

Sewage sludge, and to a lesser extent, animalmanure, can contribute heavy metals such as zinc,copper and cadmium to the soil. In the Netherlands,where the use of these manures is high, heavymetal concentrations on many arable field soils arehigher than the target value (Dutch environmentalstandard) (CBS, 1997).

Irrigation can also influence soil chemistry.Intensive production of irrigated horticulturalcrops can result in serious and permanent soildegradation through increased salinity (EuropeanEnvironment Agency, 1998). Contamination ofsoils by pesticides, heavy metals and nutrientsin Spain varies with soil type, crop type and man-agement practices, and is highest under intensiveirrigated systems (de la Rosa and Crompvoets,1998). Copper-based fungicides have long beenused in vineyards and Dias and Soveral-Dias (1997)reported high levels of soil contamination by cop-per in ground previously occupied by vines andconverted to arable use, with highest levels in soilfrom older vineyards. Copper is little used in arablesystems, with the exception of organically grownpotatoes, where it’s use continues to be permit-ted but increasingly questioned as it is recognisedas a potential long-term environmental problem(Redman, 1992).

Table 5. Pesticide degradation rates in soil (data supplied by PesticideSafety Directorate, York, UK; F. Hutson, personal communication)

Active Situation Days Mean Sourceingredient (DT50) days

Atrazine (laboratory) 41–146 81 EU Review(field) 5–60 29 EU Review

Isoproturon (laboratory) 10–20 UK Review(field) 13–25 UK Review

Pirimicarb (laboratory) 10–263 UK Review(field) 15–21 UK Review

Dimethoate (laboratory) 4–16 UK ReviewMetaldehyde (laboratory) 1–71 UK Review

(laboratory) 67–1662 UK ReviewCarbendazim (laboratory) 20–365 EU Review

(field) 100–180 EU Review

DT50 - time for 50% loss; half-life.1German soils.2US soils.


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

Water

Although the proportion of total water useattributable to agriculture is relatively small, thefact that most of this water is used in the summermeans that ecological impact can be considerable.There is in any case increasing concern aboutthe sustainability of water use in England (Envi-ronment Agency, 1999a) because water abstrac-tion lowers groundwater, reduces wetland habitatareas, slows flow rates of rivers and exacerbatesproblems with pesticides and nutrients in water-courses (European Environment Agency, 1998).The problem is more acute in southern Europewhere over-exploitation of groundwater for agri-culture can lead to a decline in the quality andquantity 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 andsurface waters, seriously affecting the quality ofdrinking water, and the cost of its treatment. Theirpresence in surface water also can have seriousconsequences for aquatic life. Erosion of arablesoils results in sedimentation of watercourses anddeterioration in the quality of water and aquaticecosystems. Economic support under CAP regimesfor intensification of arable systems has increasedthe arable area, field sizes and fertiliser use,resulting in deterioration in the quality of aquaticenvironments.

Cultivation

Soil erosion has increasing environmental conse-quences for aquatic habitats as well as for the soilitself, but in the case of aquatic habitats, theseconsequences are externalised from the farmingsystem. Late cultivations associated with prepar-ing autumn seedbeds and late harvesting of cropssuch as maize, potatoes and sugar beet contributeto high silt loads in streams and rivers because soilis exposed during periods of high rainfall. Earlyploughing of rape residues can also lead to highlevels of nitrogen leaching to water. Loss of soilto watercourses is greatest on sandy soils where

infiltration capacity has been reduced by surfacecapping, and on clay soils where surface runoff isalso high. Sub-soil fissures in clay soils during dryweather can lead to rapid movement of water andclay particles during subsequent rains. Subsoil-ing, mole ploughing and other drainage practicesoperate in a similar way, increasing movement ofsoil-derived sediment (and nutrients) into water-courses and bypassing the natural filtering effectof undrained soils. Continuous cultivation of arablecrops and removal of hedges are thought to havecontributed to increased silt loads in rivers (Skin-ner and Chambers, 1996; Van Oost et al., 2000). Theassociated economic costs could be considerable. InBritain, Evans (1996) estimated the annual costof removing soil-derived impurities from drinkingwater at £3Ð6–£30 million.

Ecological impacts of sedimentation inwatercourses are best documented for salmonids(Theurer et al., 1998) but also affect otheraquatic organisms and macrophytes. Salmonideggs require a period of 60–180 days in gravel‘redds’ on the river bed. Sedimentation into reddsduring this period rapidly reduces dissolved oxygenavailable in the water with the result thatdeveloping embryos are killed. This is thoughtto have been a major factor in the decline innumbers of economically important salmonids inBritish waters and to be closely linked with arable,rather than grassland systems. In the River Test(Hampshire, England), hatching rates of fry fromredds in a section of the river fed by grasslandstreams was substantially higher than that wherethe river was fed by streams flowing through arableland (Anon, 1999).

Cultivation of arable land in the Iberian Penin-sula coincides with the main autumn rainfallperiod. Rainfall is erratic and discharge of waterinto rivers, and erosion of soil, can be high duringrainfall events (Bergkamp et al., 1997). Bergkampet al., reported declining annual rainfall in theAlentejo (Portugal) since the 1930s, and there hasbeen a decline in spring rainfall over the pastdecade (Battencourt, 1999). Reduced flows couldmean that sedimentation in streams and riversis higher now than in the past. Sedimentationof reservoirs can be considerable, reducing theircapacity for water storage (Kosmas et al., 1997).

Nutrients

Nitrate and phosphate are the main nutrient pollu-tants of watercourses arising from arable farming.Nitrate levels in drinking water are regulated by


the EU Drinking Water Directive (80/778/EEC),which established a maximum admissible con-centration of 50 mg l�1, based on World HealthOrganisation guidelines. The EU Nitrates Direc-tive (91/676/EEC) requires members states toreduce nitrate pollution by introducing controls onagriculture in water catchments where the nitrateconcentration exceeds 50 mg l�1 or is in danger ofdoing so.

Nitrogen and phosphate reach watercoursesfrom the soil by leaching, surface run-off, sub-surface flow and soil erosion. Nitrate is solubleand enters water via leaching and run-off whilephosphate molecules bind to eroded soil particlesand enter watercourses as run-off. Most nitrateleaching occurs during the autumn when nitratepasses through the root zone faster than the cropis able to exploit it, and following ploughing ofgrassland, when organic nitrogen is mineralised(Young, 1986). Leaching is greater under cerealsthan under permanent grass (Croll and Hayes,1988), but can also be high under rotational set-aside (Meissner et al., 1998). Drainage of heavysoils increases the rate at which nitrates andphosphates enter surface water.

In a UK experiment, nitrate lost from directdrilled land was 24% lower than from ploughedland, but still higher than the EC DrinkingWater limit (Skinner et al., 1997). The relativecontribution to leached nitrate of nitrate frommineralisation of organic nitrogen and that fromapplied fertiliser is unclear, but the former isknown to account for a substantial proportion ofnitrate lost from cultivated and uncultivated soils(Addiscott, 1991; Sylvester-Bradley and Powlson,1993). Although short-term responses to changes infertiliser applications to cereal crops have also beenreported (NRA, 1992), Addiscott, (1991) reportedthat only 6% of labelled fertiliser nitrogen appliedto winter wheat was lost directly by leaching, andploughing of grassland in the 1940s and 1950sis thought to have contributed to current highlevels of nitrate in groundwater (Johnston, 1993).Nitrate levels remain highest in the south and eastof England where rainfall and therefore dilutionof nitrates, is lowest. The period between leachingand appearance in the saturated zone of the aquiferdepends on geology and can exceed 40 years onsandstone and chalk but is considerably less thanthis on more permeable rocks such as limestone.

Phosphates enter surface water following periodsof rain, when soil particles are eroded from exposedsoil, especially where fields on slopes are ploughed.Phosphorus may enter water by surface runoff orby sub-surface flow through soil cracks and drains,

and may be in the form of dissolved P or adsorbedto soil particles. Losses are greatest during stormevents when transport through the soil is toofast for P to become adsorbed to stable particleswithin the soil. For example, Heathwaite (1997)recorded that particulate P formed the bulk of totalP from sub-surface flow following a storm event.Readily drained sandy soils, and clay soils proneto cracking, especially those with field drains, arethe most prone to loss of P in both dissolved andparticulate form (Heathwaite, 1997). In Denmark,Kronvang (1990) found that 60% of annual P fluxesconsisted of particulate P, with 70–90% resultingfrom short-term storm events. Studies in the UKsuggest that farming is a major contributor ofphosphates to surface water and that outputs fromboth agricultural and sewage treatment sourcesmust be reduced if limiting levels of phosphatein water are not to be exceeded (Johnes andO’Sullivan, 1989).

Both nitrate and phosphate can cause severeeutrophication of water, nitrates affecting mainlycoastal waters such as the North Sea (Baldocket al., 1996) and Baltic (Saull, 1990) and phosphatesaffecting rivers and lakes, including those ofconservation importance (e.g. Slapton Ley, Devon,England) (Tytherleigh, 1997). Eutrophication inboth coastal and inland waters can result, throughexcessive growth of phytoplankton, in depletionof oxygen from water bodies, and subsequentdeath of aquatic invertebrates, fish and otheraquatic animals. Blue-green algae associated witheutrophication produce toxins to which fish andterrestrial animals are susceptible.

Eutrophication and changes in the fauna andflora of Loch Leven (Scotland) have been associ-ated with high levels of phosphate derived fromfarmland (Castle et al., 1999). For Loch Levenin 1992, summer algal blooms were estimated tohave cost the area up to £783 000 in lost busi-ness, and increased production costs to the down-stream industries by £160 000 (Castle et al., 1999).Eutrophication is an international problem. Forexample, discharge of phosphates from Spanisharable and industrial sources into the Guadianariver have resulted in phosphate concentrations ofup to 5Ð36 mg l�1 and excessive growth of the waterfern, Azolla in the Portuguese section of the river(Carrapi et al., 1996).

There have also been concerns about possiblerisks to human health arising from nitrates inwater, though the evidence for these risks hasrecently 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


the next 20 years (Skinner et al., 1997). Addi-tional costs of removing algal growths resultingfrom eutrophication could be considerable (Mag-arara and Kunikane, 1986). Pretty et al. (2000)put the cost of nitrate use to UK water consumersat £16Ð4 million per year, with additional costs of£52Ð3 million arising from soil erosion and phos-phate pollution. Kraemer and Kahlenborn (1998)claim that encouraging farmers in Munich to adoptorganic farming and other modified cropping sys-tems protects groundwater from nitrate leachingand (at DM 1 million) is cheaper than removingnitrates from drinking water.

The Dutch agricultural sector is responsible forapproximately 64% of the N-load in surface waters,and 38% of the P-load (NMP-3, 1998 ). Legislationto control application rates, methods and timing,aims to reduce these environmental impacts. Otherparts of Europe where similar problems occurinclude much of southern England, low-lying areasof Belgium and France, parts of Germany and thenorthern plains of Italy, but the greatest impact isoften derived from intensive livestock, rather thanarable systems (Gardner, 1996).

Rates of nitrogen and phosphate fertiliser appli-cation to crops in Portugal are considerably lowerthan those in northern Europe and nitrate lev-els in water are generally not a serious problem(de Sequeira, 1991; Ribeiro and Serrao, 1996).Eutrophication of inland waters is rare, butphosphorus accumulation can occur where sed-imentation results from erosion of arable land(Soveral-Dias and Sequeira, 1992). Where inten-sively managed crops such as maize are grown,e.g. in the Ribatejo region (Ribeiro and Serrao1996) applications of 300 kg N and 140 kg P2O5per hectare have been common (Soveral-Dias andSequeira, 1992). Here, levels of nitrate in ground-water have been found to exceed maximum accept-able concentrations defined by the European Com-munity, with highest concentrations reflecting tim-ing of crop irrigation (Cerejeira et al., 1995).

Pesticides

Pesticides enter surface water from point sourcecontamination following spillage incidents (e.g.Anon, 1999), and from diffuse sources followingtheir application to crops. They can be toxic toaquatic organisms and some are potentially car-cinogenic (Cartwright et al., 1991). While 70% ofthe EU’s drinking water is derived from groundwa-ter sources, maximum allowable concentrations forpesticides are well below drinking water standards

so as to reduce damage to aquatic invertebratecommunities (Cartwright et al., 1991). Examplesof countries in which such levels are exceeded aregiven in Table 6.

The presence of pesticides as pollutants of waterdepends on their mobility, solubility and rateof degradation. Highly persistent organochlorinepesticides are no longer used in arable systems,reducing the risks of pollution incidents resultingfrom arable operations. Many modern pesticidesbreak down quickly in soil or sunlight but aremore likely to persist if they reach subsoil orgroundwater because of reduced microbial activity,absence of light and lower temperatures (Environ-ment Agency, 1999b).

Many pesticides are supplied in high concen-trations of active ingredient and there is a highrisk of pollution incidents resulting from spillages,inappropriate disposal and washing of sprayers.Diffuse pollution of water by pesticides resultsmainly from surface run-off following spraying,rather than from pesticides entering aquifers. Fordryland crops drainage can increase movement ofpesticides from field to surface water, by-passingthe soil profile where such pesticides might other-wise be degraded (Cartwright et al., 1991).

Isoproturon (IPU) is the most widely usedherbicide in the UK and known to be susceptible toentering surface waters via runoff and movementthrough soil cracks (White et al., 1997). In aCambridgeshire (UK) study, levels of mecoprop andother herbicides were highest at times of high riverflow (Hennings et al., 1990; Clark et al., 1991).

Evans (1996) cites one study in which pesticidelevels were up to 680 times higher during floodthan under normal flow conditions. Croll (1988)found mecoprop in 35% of surface water samplesfrom the English Anglian Region, with lindaneand dimethoate (insecticides toxic to aquatic inver-tebrates) present in 16% and 14% respectively.Atrazine and simazine were the most frequentlyoccurring pesticides (58% and 42% respectively)but these herbicides are extensively used in Britainby users other than arable farmers and the rate atwhich they enter leached or surface water varieswith soil type, climate and cultivation methods(Hall et al., 1991).

In a wider survey of 3500 sites in Englandand Wales, 100 of the 120 pesticides targetedwere detected and five herbicides (atrazine, diuron,bentazone, isoproturon and mecoprop) regularlyexceeded EC Drinking Water Directive limits(NRA, 1995). Pesticides differ considerably in theirtoxicity to aquatic life, and the UK EnvironmentAgency sets Maximum Allowable Concentrations

Imp

actso

farab

leintensifi

cation

353

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

Austria Denmark France Germany Spain Luxembourg Norway UK Czecho- Slovakia Sloveniaslovakia

Atrazine 16Ð3 0Ð9 8Ð2 4Ð3 0 13 32Ð1(1660) (1006) (85) (12 000) (28) (355) (84)

Simazine 0Ð2 0Ð5 0 0Ð9 0 4Ð8(1248) (1006) (81) (11 437) (28) (84)

Lindane 0 0Ð2 0 0 25(72) (994) (116) (215) (8)

Desethyl-atrazine 24Ð5 1Ð4 7Ð5 47Ð6(1666) (292) (10 972) (84)

Heptachlor 0 0 0(72) (4) (12)

Metolachlor 1Ð1 0 4Ð8(1248) (28) (84)

Bentazone 0 80(28) (5)

DDT 0 0(215) (12)

Dichlorprop 1Ð4 83Ð3(1006) (6)

Methoxy-chlorMCPA 0Ð2 100

(1006) (2)Desisopropylatrazine 1Ð3 1Ð4

(1666) (292)Hexazinon 0Ð4 2Ð6

(277) (2234)


Table 7. Pesticide environmental quality standards in UKsurface water (Environment Agency, 1999)

Annual average Maximum allowable(µg l�1) concentration (µg l�1)

HerbicidesAtrazine 0Ð01Isoproturon 2 20MCPA 2 20Mecaprop 20 200Simazine 2 10Triallate 0Ð25 5

InsecticidesChlorfenvinphos 0Ð01 0Ð1Cypermethrin 0Ð0001 0Ð001Dimethoate 1Malathion 0Ð01 0Ð5Methiocarb 0Ð01 0Ð16Pirimicarb 1 5

that tend to be higher for herbicides than insecti-cides. These concentrations (and the annual aver-age concentrations recorded) are given in Table 7for a selection of commonly used arable pesticides(Environment Agency, 1999b). Pesticide applica-tion is associated with considerable costs to societyin terms of water treatment. Pretty et al. (2000)put the cost to UK water consumers of pesticideapplication at £120 million.

In the Netherlands large quantities of pesticidereach ground and surface water as a result of drift,evaporation and runoff. In a survey of large surfacewaters in the Netherlands, all contained pesticides.For approximately 25% of the substances, theconcentrations exceeded the maximum allowableconcentrations (RIZA, 1996). Some pesticides arefound in groundwater in concentrations above EUstandards for drinking water (CBS, 1997). Heavymetals have also been reported from groundwaterbelow arable fields (CBS, 1997).

Rates of pesticide application to cereal crops inPortugal and other southern European countriesare generally considerably lower than those innorthern Europe, and pesticide levels in water areprobably not a serious problem in most catchments.However, where intensively managed crops such asmaize are grown, levels of atrazine in groundwaterhave been found to exceed maximum allowable con-centrations, with highest concentrations occurringfollowing crop irrigation (Cerejeira et al., 1995). Inthe Po Valley of Northern Italy use of the herbicidesatrazine and molinate on irrigated rice and maizeover permeable gravel aquifers has resulted in thepresence of these pesticides in the groundwater(Cartwright et al., 1991).

In parts of north-east Europe levels of aluminiumand heavy metal ions in soils are high, and acid rainresulting from industry and transport pollutionto the air can cause increasing solubility andmobilisation of these metals, and their subsequentoccurrence in groundwater (Bouma et al., 1998;Kraemer and Kahlenborn, 1998).

Air

Gaseous emissions from arable farming includethe greenhouse gasses N2O and CO2 and aerosolsNOx and NH3. In addition, use of fossil fuelsfor the manufacture of arable inputs for cropproduction, and for transport of arable inputsand products, is a major contributor of CO2 andSO2 to the atmosphere (European EnvironmentAgency (EEA), 1998). Nitrogen and sulphur areassociated with acidification of terrestrial andaquatic ecosystems following the transport anddeposition of these pollutants.

CO2 is the main greenhouse gas, representing65% of the anthropogenic contribution to globalwarming, while N2O contributes an additional5% (EEA, 1998). CH4 contributes 20%, but isassociated with livestock systems, rather thanarable land. SO2, NOx and NH3 in the atmo-sphere offset global warming to some extent buttheir effect is short-lived and localised, and emis-sions of these gases are declining (EEA, 1998).European annual mean air temperatures haverisen by 0Ð3°C to 0Ð6 °C since 1900 (EEA, 1998),and climate change resulting from anthropogenicgaseous emissions is increasingly thought to havea substantial impact on terrestrial and aquaticecosystems. These include the destruction of coralreefs, changes in the abundance, distribution andphenology of many plant and animal species, andlocal extinction of Arctic and alpine plants (Hughes,2000).

Production of the gasses NO and NO2 canresult from denitrification within arable systems(Benckiser, 1997) and 70–80% of emitted nitro-gen is deposited back onto the land resultingin eutrophication and acidification of some envi-ronments (Goulding et al., 1998). Deposition ofatmospheric nitrogen can lead to eutrophicationand acidification of semi-natural environments,resulting in reduction in botanical species diver-sity and changes to soil processes, as demonstratedat Rothamsted (UK) by Goulding et al. (1998).Similar changes in soil processes and botanicalcomposition resulting from atmospheric nitrogendeposition have been reported in British upland


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

Crop type, soil moisture and nitrate levels allinfluence rates of denitrification (Svensson, 1991a,b; Addiscott and Powlsen, 1992) and irrigated cropsare known to increase NO2 emissions (Armstrong-Brown et al., 1995). An estimated 3Ð.2% of appliedN is thought to be lost to the atmosphere as NO2(Armstrong-Brown et al., 1995). In addition, over90% of emissions from estuaries and rivers of thegreenhouse gas, N2O are in the northern hemi-sphere where they are thought to be derived fromanthropogenic sources such as nitrogen fertilisers(Seitzinger et al., 2000).

The agriculture sector world-wide accounts forabout 5% of anthropogenic CO2 emissions (ECAF,undated), while cultivation of arable soils is esti-mated to release 30 Mt yr�1 carbon to the atmo-sphere globally, through oxidation of carbon alone(Armstrong-Brown et al., 1995). However, soilorganic matter and bioenergy crops have consider-able potential as a carbon sink for CO2 emitted byarable operations (Armstrong-Brown et al., 1995;Smith et al., 2000).

The process of manufacturing nitrogen fertiliserscontributes substantial emissions of NH3 to the air.Within the UK, this process accounts for 37% of allindustrial emissions of ammonia (Sutton et al.,2000). The process is also associated with highenergy consumption, with the result that emissionsof CO2 are also high. Cormack and Metcalfe (2000)estimate that 50% of total energy input into cerealcrops is accounted for by the manufacture andtransport of fertiliser and pesticides.

Long distance transport, especially that associ-ated with international trade of arable inputs andproducts, contributes to emission of CO2 and SO2 tothe atmosphere. Transporting crops from southernEurope to the north increases the energy costs by352% over those associated with local delivery (Cor-mack and Metcalfe, 2000). As well as CO2 and SO2emissions, shipping results in substantial marinepollution (Collins, 1994), while NOx emissions fromaircraft have been found to be responsible for upto 10% of upper tropospheric O3 (Stevenson et al.,1997). Pretty et al. (2000) estimate that air pol-lution and greenhouse gas emissions arising fromagriculture account for 48% of the industry’s exter-nalised costs, amounting to £738 million for N2O,and £47 million for CO2, even without includingpollution associated with international trade.

Emissions of pesticides to the air are consid-erable where they have been documented in theNetherlands. Approximately 3Ð1 million kg (active

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

Reducing ecological impacts

Agricultural intensification is a feature of arablesystems in both northern and southern Europe,while in the south, abandonment of arable andlivestock systems is an additional threat to arableecosystems. Wider environmental costs of inten-sification include soil degradation, siltation andpollution of watercourses, reduced biodiversity,contributions to global climate change and sim-plification of cultural landscapes. Few of theseact directly on farm businesses in the mainarable areas, and economic incentives are currentlymainly for their continuation through production-led support. However, rural depopulation is anincreasing problem in many formerly agriculturalareas and a combination of social, ecological andother environmental objectives increasingly formsthe focus for agricultural policy.

As indicated in the previous sections, a numberof changes to arable systems have been investi-gated in order to reduce environmental impacts.Organic farming differs from conventional inten-sive systems in that external inputs are reducedand a more integrated approach is taken to cropproduction as a system. ‘Integrated Crop Manage-ment’ reduces external inputs, but to a lesser extentthan organic farming, and seeks to exploit nat-urally occurring predators of crop pests, and toadopt minimum tillage. Both organic farming andintegrated crop management have potential bene-fits over unmodified conventional arable productionin terms of landscape structure, biodiversity, soil,water and air quality. Table 8 summarises theprevious sections in relation to conventional ‘inten-sive’, ‘organic’ and ‘integrated crop management’,as well as the ‘extensive’ systems surviving insouthern Europe.

A number of European Commission Directivesconcerned with improving or protecting the envi-ronment have implications for arable agriculture.In addition to the Drinking Water Directive andthe Nitrates Directive already referred to, a WaterFramework Directive, (2000/60/EC) has recentlybeen adopted which will extend the existing provi-sions on water quality and promote pricing as anincentive for the sustainable use of water resources.

356C

.Sto

ateet

al.

Table 8. Summary of effects of four European arable systems on soil, water, biodiversity, landscape and air

Soil Water Air Biodiversity Landscape

Erosion Organic Structure Soil Nutrient Pesticide Sediment Terrestrial Aquaticmatter fauna pollution pollution

Intensive Ploughingaggravateserosion

OM levels lowand oftendeclining

Compactioncommon.Plough pans

Fewer worms.Impover-ished microfauna andflora

Pollution fromfertilis-ers–leaching,drains andrunoff

Pollution fromspray drift,runoff andleaching

Surfaceerosion anddrainagelead tosedimentpollution

Pesticide pollution.Greater energyuse fromagrochemicalmanufacture andapplication

Simplified cropsystems,high fertiliserandpesticideuse reducehabitatdiversity andfood supply

Eutrophication,pesticidepollution andsedimen-tation

Larger fields,blockcropping,lessnon-croppedland lead tohomoge-neouslandscape

Organic Ploughinglessfrequentunlessstockless

Higher OM dueto use of leysand organicmanures

Leys andmanureimprovestructure

Rotations andmanuresencouragemicrobes.Nopesticides.

Ploughing ofgrassreleasesnitrates. Maybeaggravatedwithlegumes

Few pesticidesused

Surfaceerosion anddrainagelead tosedimentpollution

Low energy use.Possibly greaternitratemineralisation

Greater cropandnon-crophabitatdiversitythanintensive.Fewpesticides.Mechanicalweed controlcan bedetrimental.

Little or nopesticidepollution.Mixedfarmingsystem mayreducesedimentpollution.Nutrientpollution stilloccurs.

Mixed farmingproducesmore diverselandscape.Higherproportion ofnon-cropfeatures.

Integrated Reduced cul-tivationsreduceerosion

More OM thanintensive

Reducedcultivationimprovessoil structure

Reducedcultivationincreaseswormnumbers.Reducedpesticideimpact.

Lower, moretargeted useof fertiliserreducespollution

Reducedpesticideuse

Reducedcultivationsminimiseerosion

Reducedcultivation andagrochemicaluse may reducepollution?

Higherdensities ofsome inver-tebrates.Limitedevidence forother taxa

Pesticide,nutrient andsedimentpollutionlikely to bereduced

Crop rotationmayincreaselandscapediversity.May includegrassmargins etc.

Extensive(mainlysouthernEurope)

Erosioncommondue toploughing

OM levels lowand oftendeclining

Probably poor Reducedpesticideimpact

Low rates offertiliser use

Low pesticideimpact

Frequentcultivationsencourageerosion

Low energy use.Possibly greaternitratemineralisation

Use of fallowsand lowerfertiliser andpesticideuse lead tohigherbiodiversityand charac-teristicspecies notfound inothersystems

Limitedevidence

Floweringplants incrops andfallows.Livestockgrazefallows.Oaksactivelymaintainedin someareas


The EC Habitats Directive (92/43/EEC) requireseach member state to compile a list of areas con-taining the habitat types and species listed in theDirective. These areas will become ‘Special Areasof Conservation’ (SACs), complementing the ‘Spe-cial Protection Areas’ (SPAs) established under theEC Directive on the Conservation of Wild Birds(79/409/EEC), which together form a network ofsites collectively known as ‘Natura 2000’. Direc-tive 85/337/EEC requires environmental impactassessment to be made for projects likely to have asignificant effect on the environment, which couldin some cases include agricultural projects whichentail major changes in land use. There is alsoEU legislation concerning plant protection prod-ucts and organic farming.

Further attempts to reduce specific environmen-tal impacts of arable farming come in the form ofnational Codes of Good Agricultural Practice andstatutory arrangements, such as the UK’s LocalEnvironmental Risk Assessment for Pesticideswhich places restrictions on the use of pesticidesat field margins on arable land (LERAP) (MAFF,1999b). In some European countries attempts havebeen made to reduce pesticide use by governmentmandate, including the withdrawal of approval forproducts considered to be high risk and encour-agement of farmers to use lower doses by researchand advice (Bellinder et al., 1994; Thonke, 1991).More recently, some countries have introduced pes-ticide taxes (Daugbjerg, 1998) although there isa danger that such taxes encourage the use ofcheaper broad-spectrum, and therefore more dam-aging, compounds, and there are also concernsabout effects on farm competitiveness and prof-itablility (Falconer and Hodge, 2001).

In some EU countries, arable area payments areconditional on compliance with certain environ-mental restrictions on arable management, a sys-tem known as cross-compliance. Cross-compliancewas introduced under the Common Rules Regula-tion (1259/99) of the ‘Agenda 2000’ reform of theCAP. However, similar conditions were applied insome countries prior to Agenda 2000, for examplethe Republic of Ireland has used cross-complianceto reduce overgrazing since 1998. The UK has alsoapplied conditions to livestock headage paymentsto prevent overgrazing, and has made receipt ofarable area payments conditional upon observingmanagement conditions for set-aside land to pro-tect species and habitats. Other countries haveintroduced cross-compliance in the arable sector totackle pollution problems; for example, Denmarkis using cross-compliance to enforce national reg-ulations including compulsory fertiliser plans, the

maintenance of green cover over winter, and aban on cultivating closer than 2 m from a water-course. Finland has imposed new environmentalconditions including 60 cm wide riparian bufferzones, green cover, and regulations on fertiliserand organic waste use. France is using cross-compliance to enforce water abstraction legislation,whilst The Netherlands has introduced conditionsfor two intensive crops: maize and potatoes, toincrease the use of mechanical weed control as analternative to herbicides. These are examples ofthe varied approaches taken by different memberstates of the EU; several countries have plans tofurther extend such conditionality in the future.For cross-compliance to be effective, measuresshould avoid penalising farmers already adopt-ing good environmental practices (e.g. UK farmswith high hedge density and small field size) andrewarding those who do not. They should also avoidperverse effects such as piping watercourses toavoid adoption of buffer zones adjacent to them.

The Rural Development Regulation (1257/99)provides voluntary opportunities for the adoption offurther ecologically, economically and socially sus-tainable management practices and systems withinEuropean arable systems, extending the provisionsof Regulation 2078/92 under the 1992 reform ofthe CAP. Agri-environment options are designedto have positive impacts to counter specific long-term environmental problems. They aim to addressproblems that farmers can not realistically beexpected to approach under cross-compliance andhave an additional cost to the farmer who iscompensated accordingly. As for cross-compliance,different countries have used agri-environmentschemes in contrasting ways, some emphasisingreductions in, or controls on resource use whilstothers have encouraged the conservation of habi-tats and features. Examples are: the reduction inuse of nitrogen fertilisers (Denmark) and protec-tion of flower species accompanying arable land(Nordrhein-Westfalen, Germany). The UK hasrecently completed a three year pilot for an ArableStewardship Scheme specifically designed to con-serve wildlife and habitats on arable land.

Where possible, management practices shouldaddress more than one environmental problem.For example, agri-environment options intended torestore soil organic matter could reduce erosionrates by increasing infiltration rates, increasebreakdown of pesticides by increasing microbialactivity and act as a sink for atmospheric carbon.Riparian buffer zones have been shown to reduceloss of nutrients from arable land to watercourses(Castle et al., 1999), but can also be managed as


wetland habitats, increasing landscape diversityand biodiversity.

Greatest environmental benefits are likely toresult from the integration of a number of mea-sures, according to local conditions. Inclusion of anagri-environment option for the provision of adviceand training would enable farmers to integratefully the relevant measures, in order to max-imise the environmental benefits and to accommo-date compatible marketing of agricultural produce.Agri-environment schemes can in themselves per-form an important function as a training exercisefor environmentally positive management (Morrisand Potter, 1995). Environmental benefits of main-taining or restoring the integration of livestock orsylviculture into arable systems can be as great as,or greater than, modification of practices withineither livestock, forestry or arable systems, andenvironmental measures must be sufficiently flex-ible to accommodate this approach. Whole farmplans are useful in this regard. A local scaleapproach promotes the integration of measures inorder to achieve environmental, social and natureconservation objectives.

Such incentives are likely to be most effective assupport for diversification of sustainable resourcemanagement on farmland, including tourism andproduction of regional and other niche products, asis currently the case in the Alentejo of Portugal.Here, regional sheep cheeses and pork products,cork and charcoal production are integrated intoarable systems.

Cultural interests of farmers, including land-scape design and wildlife management for hunting,fishing and other recreation also have an importantpart to play in the development of multifunc-tional management of natural resources (Stoate,in press). Potential for influencing managementpractices through the market are also being devel-oped in the form of eco-labelling of food producedunder specified ecologically sensitive guidelines,and environmental certification of companies alongthe production chain (e.g. Udo De Haes and DeSnoo, 1997), as well as an established market fororganic food. Such a combination of public and pri-vate support for sustainable management of arablelandscapes will avoid both the negative impactsof abandonment of arable systems (especially insouthern Europe), and of continued intensification.

Demand for organic produce has increased sub-stantially, most recently in the UK where thenumber of registered organic farmers increasedfrom 828 in 1997 to 1568 in 1999 (Soil Associa-tion, 1999). However, 70% of organic cereals, and82% of fruit and vegetables consumed in the UK

is currently imported (Soil Association, 1999) andenvironmental costs associated with transport areexternalised. For example organic wheat is regu-larly imported from as far away as Australia, andthe long distance transport involved contributes togaseous emissions that may damage the environ-ment on a large (and unnecessary) scale.

Consumption of locally produced food has envi-ronmental benefits in terms of reducing air pol-lution associated with transport, and increasinghabitat diversity, as well as restoring social inte-gration of rural communities (Pretty, 1998). Suchlocal marketing is occasionally applied to arableproducts (e.g. Wookey, 1987), but is more gen-erally associated with horticultural and livestockproducts, either within, or independently of, arablesystems. It is exemplified by box schemes and farm-ers’ markets (Pretty, 1998). Multifunctional use offarmland, without production support, is alreadyadopted in Switzerland (not an EU member), fol-lowing a referendum on the issue in 1996 (SAEFL,2000). Specific practical solutions to environmentalproblems will have greatest and most efficient envi-ronmental benefits if adopted within the context ofsuch wider policies.

Acknowledgements

This review is adapted from part of a commissionedreport to the European Union Directorate General, Envi-ronment, Nuclear Safety and Civil Protection (DGXI)(Contract no. B4-3040/98/000703/MAR/D1)) (Boatmanet al., 1999). We are grateful to Michael Hamell, PeterBilling and Luis Carazo Jimenez (DGXI) for their contri-butions to the manuscript. However, the views expressedare those of the authors and do not necessarily representthose of the Commission or the Environment Directorate.Edward Urbansky and two anonymous reviewers helpedto improve the manuscript.

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Ecological impacts of arable intensification in Europe

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