Agriculture, Ecosystems and Environment 102 (2004) 263–277
Developing grassland conservation headlands: response of carabid assemblage to different cutting regimes in a silage field edge K.A. Haysom a,∗ , D.I. McCracken a , G.N. Foster a , N.W. Sotherton b b
a SAC Auchincruive, Ayr KA6 5HW, UK The Game Conservancy Trust, Fordingbridge, Hampshire SP6 1EF, UK
Received 16 July 2002; received in revised form 12 September 2003; accepted 16 September 2003
Abstract Prescriptions for the management of arable field edges are an established component of many agri-environment schemes that aim to enhance the wildlife value of farmed land, yet there are few analogous recommendations for the margins of intensively managed grassland. An experiment examined the response of a carabid beetle assemblage to three cutting regimes (uncut, one cut, three cuts per annum) applied in the headland area of a silage field. In the third year of management, diversity in the uncut plots was significantly greater than in the one and three cuts plots. Ordinations showed distance from field edge and cutting regime effects in species composition. Relaxation of cutting management increased the effective area of the inter-crop habitat and species that responded positively were those that preferred shaded or moist conditions. Results are discussed in the context of individual species requirements and in relation to management options for agricultural grassland. © 2003 Elsevier B.V. All rights reserved. Keywords: Carabid; Conservation headland; Cutting; Diversity; Field margin; Grassland management
1. Introduction When properly managed, agricultural land can be a diverse habitat for wildlife, and the substantial area that this land-use represents within Europe has immense potential to create areas rich in farmland species (Sotherton, 1998). However since World War II, many regions have undergone agricultural intensification including increased use of nitrogenous fertilisers, wide-scale application of pesticides, removal of boundary features, such as hedges, resulting ∗ Corresponding author. Present address: CABI Bioscience UK Centre (Egham), Bakeham Lane, Egham, Surrey TW20 9TY, UK. Tel.: +44-1491-829045; fax: +44-1491-829100. E-mail address: k.haysom@cabi.org (K.A. Haysom).
in coarse-grained landscapes, and the exchange of small-scale mixed-farming systems for larger, specialist farms resulting in regional biases to arable (e.g. eastern Britain) or livestock production (e.g. western Britain) (Stoate, 1996; Vickery et al., 2001). Some of these changes occurred as responses to European policy demands for increased food production and have been linked to the prolonged decline of many species of wildflowers, invertebrates, birds and mammals associated with farmland (Wilson, 1994; Aebischer, 1990, 1991; Chamberlain et al., 2000; Tew et al., 1994). The UK has a suite of agri-environment schemes targeted nationally (e.g. Countryside Stewardship Scheme (UK), Rural Stewardship Scheme (Scotland), Tir Gofal (Wales)) and regionally (e.g. the various
0167-8809/$ – see front matter © 2003 Elsevier B.V. All rights reserved. doi:10.1016/j.agee.2003.09.014
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Environmentally Sensitive Area Schemes). Such schemes include wide-ranging management prescriptions for field edge habitats aimed at restoring landscape features (e.g. dry stone walls, hedgerows), protecting water courses and hedgerows from agrochemicals, preventing bankside erosion, creating overwintering sites for invertebrate predators of pests, and “wildlife corridors” to link features of conservation interest in the farmed landscape, promoting biodiversity and providing chick-food insects for ground-feeding birds (SEERAD, 2000; MAFF, 1999a; Countryside Council for Wales, 1999). These prescriptions have proved popular both with farmers and wildlife advisors. With the exception of water margins, however, the majority of these boundary prescriptions have focused on arable land. Less than 50% of UK agricultural land is arable and in regions of southern Scotland more than 80% of farmed land is under grass (Hopkins and Hopkins, 1994). Although problems associated with arable intensification have received more attention, intensive processes such as drainage, reseeding, application of fertilisers, earlier and more frequent cutting have reduced the diversity of plants and invertebrates on agricultural grassland (Fuller, 1987; Hopkins and Hopkins, 1994; Di Giullio et al., 2001). The total area occupied by grasslands has declined, through conversion to arable and other uses, and only 4% of conservation-value grasslands remain (Fuller, 1987). Agri-environment schemes that make provisions for grassland usually aim to protect existing rare semi-natural habitats, by encouraging traditional management practices or to restore species-rich grassland on a whole field-scale. While important, these approaches do nothing to increase the biodiversity associated with the majority of UK agricultural grasslands that are subject to more intensive management, where farmers are only likely to consider accomodating some intermediate form of conservation management. Field edge prescriptions in agri-environment schemes for arable land fall into two broad types: the “conservation headland/unsprayed edge” and “grass margin/wildlife strip” approaches. In conservation headlands the edge under agreement remains part of the crop but is subject to a modified management regime, i.e. no insecticide during summer and limited inputs of selective herbicides and fungicides. Conservation headlands were developed in the UK
to increase the supply of food-insects for gamebird chicks (Rands, 1985; Sotherton, 1990; Chiverton and Sotherton, 1991), and in Germany for the conservation of rare arable weed flora (e.g. Schumacher, 1987). Ground beetles, butterflies and small mammals have also shown positive responses (Hawthorne et al., 1998; Dover, 1997; Tew et al., 1994). In grass margin and wildlife strip approaches the edge under agreement is established by sowing a grass/grass-herb mix or by allowing natural regeneration from the seedbank. Such margins are distinct from the cropped area since they have different vegetation composition, vegetation structure and are relatively free of disturbance as they tend to remain in place over a period of 5–10 years. The use of grass margins by a range of invertebrates (particularly those which require cover for overwintering), songbirds and small mammals, and the impact of margin management such as cutting upon different species is well documented (e.g. Stoate and Szczur, 1994; Bence et al., 1999; Feber et al., 1999; Thomas and Marshall, 1999; Meek et al., 2002). Within agri-environment schemes such as Countryside Stewardship (MAFF, 1999b), land managers may elect to operate these approaches singly or in combination along crop edges. This paper concerns the conversion of such approaches to grassland situations. The rationale was to modify the management of the edge of the grass crop (i.e. conservation headland approach), rather than to replace the crop with a non-cropped area (i.e. grass margin approach), in order to enhance biodiversity whilst retaining the agricultural function for part of the year. Despite the importance of grassland for livestock production in the UK, no technique analogous to those prescribed for arable land has been described for improved grassland. The work described here is part of a larger experiment on the responses of plants and invertebrates to headland management in silage and pasture fields. The cutting or grazing regime was manipulated in a manner that, if effective, could be implemented easily by farmers. This paper examines the response of a ground beetle assemblage (Coleoptera: Carabidae) to three cutting regimes at the edge of a silage field, and variation in the assemblage with distance from the field boundary. The carabid fauna of farmed land is well known (Thiele, 1977) and some species are sensitive indicators of environmental quality (Luff, 1996). Sev-
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eral species prey on agricultural pests (Speight and Lawton, 1976; Luff, 1987; Hance, 1990; Winder, 1990), while some of the smaller ones are included in the diet of various birds (Sotherton and Moreby, 1992; Blake and Foster, 1998; McCracken and Bignal, 1998). Most investigations of agricultural grassland carabids have not included experimental manipulations of management intensity. The response of invertebrates to different cutting regimes has been studied mainly on semi-natural herb-rich grassland (e.g. Morris, 1979, 2000; Morris and Rispin, 1988). The feasibility of introducing new edge management techniques on agricultural grassland to enhance biodiversity is discussed with reference to an alternative approach already available in agri-environment schemes, the restoration of species-rich grassland. Since the restoration of plant diversity, and the maintenance of particular plant assemblages are known to be affected by soil nutrient conditions (Berendse et al., 1992; Mountford et al., 1996; Critchley et al., 2002), pertinent information on the different plots and their changes through the experiment are provided.
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Experimental plots were established in the headland area of the northern edge of the field. The headland was backed by a post and wire fence adjacent to a grassy bank and a wet ditch. This headland was on a peaty loam soil, with organic matter some 30% and a pH between 5.5 and 5.7. Nine adjacent 10 m ×10 m plots were established in May 1996. Three cutting regimes were assigned, one per plot, in a randomised block design that comprised three blocks. Each cutting regime was replicated three times and appeared only once per block, i.e.: (1) sward left uncut; (2) sward cut once a year in August; and (3) sward cut three times a year in May, June and August. Each headland cutting was synchronised with the corresponding main field cut. The sward was cut to 5 cm above ground level using a reciprocating blade cutting machine (Agria) and clippings were removed the same day. No fertiliser, slurry, or pesticide applications occurred in the headland. The plots were fenced from April to October and sheep grazed intermittently from October to February.
2. Methods 2.1. Soil nutrient composition The study site was a 5.9 ha silage field at SAC Crichton Royal Farm (55◦ 3 N, 3◦ 35 W) near Dumfries in southwest Scotland. The field had a northeasterly slope and was bordered by rough pasture to the north, a young plantation to the east, and silage/grazing land to the south and west. The sward was a perennial ryegrass-clover ley, approximately 6 years old. Since its establishment the field received, on average, 300 kg N: 60 kg P: 60 kg K ha−1 per annum, in three applications. During the 3 years of this study, N application ranged between 300 and 310 kg ha−1 per annum. In the same period, 130–220 m3 slurry ha−1 per annum were spread in three to five applications of approximately 45 m3 ha−1 . No herbicide or insectide application occurred during the study. The number of silage cuts per annum varied from three cuts (mid-May, early July, mid-August) in 1996 to four in 1997 (mid-May, early July, mid-August, mid-September) and two in 1998 (mid-May and early July). The field was grazed by young stock after the final cut until October. Sheep grazed the field intermittently from October to February.
Soil cores were collected from the headland plots to monitor nutrient composition. Twenty 0.024 m3 cores were taken at random from each plot in late January/early February each year. Cores were mixed to produce one sample per plot. Extractable magnesium, phosphorus and potassium levels were assessed by extraction with modified Morgan’s solution followed by colorimetry (Smith, 1983). 2.2. Vegetation measurements The average biomass of vegetation in each plot was assessed each summer. When each of the three cuts was made, a 0.35 m2 strip of vegetation was cut in every plot with a sheen alpina (small reciprocating blade cutting machine). The wet mass of each sample was recorded and a sub-sample was dried to constant mass in an oven at 105 ◦ C. The percentage dry mass value was used to calculate the dry mass of the whole sample and the biomass per plot (kg m−2 ). The average value from the three cutting dates was
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used to compare vegetation biomass between cutting regimes. 2.3. Carabid sampling Carabid beetles were studied using pitfall traps containing some 2 cm monopropylene-glycol preservative (Luff, 1996). A row of three traps, 2 m apart, was set parallel to the field edge in the centre of every plot. Each outer trap in the row was 3 m from the edge of the adjacent plot. A second row of traps was set 1 m behind the headland fence within the field boundary. Three rows of three traps were also set in the main field area, parallel to the headland, at a distance of approximately 80 m from the field edge. Traps were set from late May to mid-July and from late August to early October, and emptied every 3 weeks. Carabid beetles were counted and identified to species in the laboratory using Lindroth (1974). Results from autumn 1996 and 1997 have already been analysed (Haysom et al., 1999) and this paper presents data collected in 1997 and 1998. 2.4. Data analyses Mean soil nutrient concentrations were compared in the different headland plots, between years and cutting × year interaction, by split-plot analysis of variance (ANOVA), where cutting regime and year were the main plot and subplot treatments, respectively (Genstat 5 Committee, 1993). For crop biomass, data for each year were analysed separately as a randomised block design, because the block × treatment error term varied significantly between years. Total carabids caught in the pitfall traps were compared between cutting regimes, and at two distances from the field edge (1 m into the boundary and 5 m into the headland plot), using split-plot ANOVA. The cutting regime and distance factors were used as main and sub-plot treatments, respectively. To satisfy the normality assumption of analysis of variance a ln(x + 1) transformation was applied to the count data. Data for 1997 and 1998 were analysed separately as a randomised block design because the error term varied significantly between years. Variation in the abundance-activity of individual species was examined following the same procedure, for all species
for which more than 40 individuals were recorded in any 1 year. The species composition of the carabid assemblage in different headland cutting regimes and at different distances from the field edge was analysed using detrended correspondence analysis (DCA) (ter Braak, 1987). Data for each row of three traps were pooled. For each species, actual counts were transformed to a percentage of the total carabid catch at each position. This provided an estimate of the relative abundance-activity of each species at each location and adjusted for large variations in the size of catches between rows. This allowed comparison of the relative importance of each species within the carabid assemblage, local distribution or ubiquity, and the dominance patterns of species between locations. Separate DCA ordinations were prepared using 1997 and 1998 data to examine how beetle assemblages varied with location and cutting regime in each year. A third DCA sample score ordination was prepared by amalgamating 1997 and 1998 prior to analysis. The preparation of a single ordination diagram based on data from both years, facilitated the identification of temporal trends in assemblage change. The counts of the different carabid species were used to calculate an index of effective diversity for each row location. The index used was the complement of Simpson’s index: 1 − D = 1 − Σ(pi )2 , and was chosen because it is relatively robust to changes in sample size (Simpson, 1949) and because its distribution pattern allowed comparison of diversity at the two distances from the field edge, and between cutting regimes, using split-plot ANOVA. The cutting regime and distance factors were used as main and sub-plot treatments respectively. Data for 1997 and 1998 were analysed separately as a randomised block design.
3. Results 3.1. Soil nutrients The level of extractable magnesium was generally high (range 290–376 mg l−1 ) with a significantly lower level in 1998 than in 1996 and 1997 (P < 0.001, Table 1a). There were no significant differences between cutting regimes. Extractable phosphorus levels increased progressively between 1996 and 1998,
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Table 1 Mean concentration of soil extractable magnesium, phosphorus and potassium in (a) three successive years after headland plot establishment and (b) under three different cutting treatments (no significant block effects and no significant treatment × year interaction) Nutrient (mg l−1 )
1996
1997
1998
S.E.D.
F(2,12)
(a) Mg P K
351.7 6.33 136.0
353.1 7.27 106.0
311.4 7.58 136.4
7.72 0.252 11.96
18.8 13.2 4.27
Nutrient (mg l−1 )
Uncut
One cut per annum
Three cuts per annum
S.E.D.
F(2,4)
(b) Mg P K
346.0 7.54 159.5
354.7 6.86 120.6
315.6 6.78 98.4
36.92 0.476 14.56
0.62 1.57 8.96
P-value ∗∗∗ ∗∗∗ ∗
P-value NS NS ∗
NS, not significant. ∗ P < 0.05 (ANOVA). ∗∗∗ P < 0.001 (ANOVA).
with levels significantly lower in 1996 than in either of the following years (P < 0.001, Table 1a). There were no significant cutting regime effects. Phosphorus levels (range 5.77–7.83 mg l−1 ) were moderate, and within the range recommended for agricultural grassland (Dyson, 1992). Levels of extractable potassium ranged between 80 and 175 mg l−1 and there were significant differences between cutting regimes and between years (P < 0.05, Table 1a and b). The lowest level was recorded in 1997. Extractable potassium decreased as cutting intensity increased, with a significant difference between the uncut and three cuts per year plot soils. 3.2. Vegetation There was no significant difference between cutting regimes in vegetation biomass in 1996. In 1997 and 1998, the average biomass of uncut and one cut per year plots was similar, that of three cuts per year plots
was significantly lower than the others (P < 0.05, Table 2). 3.3. Carabid composition A total of 8811 carabid beetles were trapped (4862 in 1997, 3949 in 1998), comprising 46 species from 22 genera. The total abundance-activity, as recorded by counts of carabids in pitfall traps, was approximately two-fold higher 5 m from the field edge, than in the field boundary (P < 0.001, Table 3). The abundance-activity of some species varied significantly among cutting regimes, whereas that of the carabid assemblage as a whole did not differ significantly between treatments. Table 4 summarises data on species trapped in different headland treatments and the adjacent boundary, and Table 5 data on species trapped at different distances from the boundary abutting the three cuts per annum plots. Both Tables order species data according
Table 2 Mean standing crop biomass (kg m−2 ) during three successive summers in grass headland plots subjected to three different cutting regimes (no significant block effect) Year
Uncut
One cut per annum
Three cuts per annum
S.E.D.
F(2,4)
P-value
1996 1997 1998
0.575 0.552 0.480
0.537 0.521 0.576
0.428 0.344 0.283
0.071 0.041 0.063
2.31 15.4 11.1
NS
NS, not significant. ∗ P < 0.05 (ANOVA).
∗ ∗
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Table 3 Mean number of ground beetles collected per plot during trapping periods 1997 and 1998 in (a) grass headland plots subjected to three cutting regimes and (b) at two distances from the field edge (after ln(x + 1) transformation) Year
Uncut
One cut per annum
Three cuts per annum
S.E.D.
F(2,4)
(a) 1997 1998
4.20 3.85
4.21 3.71
4.44 4.15
0.154 0.137
1.53 5.24
Year
Boundary
5m
S.E.D.
F(1,6)
(b) 1997 1998
3.98 3.49
4.59 4.32
0.077 0.099
64.6 69.5
P-value NS NS P-value ∗∗∗ ∗∗∗
NS, not significant. ∗∗∗ P < 0.001 (ANOVA).
to eco-group following Cole et al. (2002), excepting that group e (small nocturnal predators) fused group 5 (small, nocturnal predators) and group 6 (ecologically similar with greater variation in size and diel activity). The assemblage was dominated by small and medium sized nocturnal predators and small diurnal Collembola feeders. In 1997, 44 species were recorded and 39 in 1998 but Pterostichus melanarius, Bembidion lampros and Nebria brevicollis dominated the assemblage in both years, making up 46–87% of carabid records at any location in either year. Most species (80%) were trapped in both years. The nine species that were trapped in a single year, were recorded as less than five specimens. In total, 41 species (89%) were recorded in the field boundary, 38 (83%) in the headland plots, and 23 (50%) in the field centre, Harpalus affinis being recorded exclusively in the mid-field. P. melanarius, B. lampros and N. brevicollis accounted for approximately 86% of the catch in the field centre, where 11 species were represented by more than five individuals per year; their contribution was reduced to 46% at the field edge because of a higher number of abundantly occurring species. Within the headland area, 21 species (55%) occurred in all three cutting regimes and 11 species (29%) in only one regime. None of the species confined to a single cutting regime represented more than 0.2% of a year’s carabid catch within that regime, and these rare individuals were distributed among the three different regimes. The first two axes of DCA ordinations of the carabid species data explained 42 and 44% of variation in 1997 and 1998, respectively. In both years, the strongest separation occurred along axis one (approx-
imately 32% of variation explained) and represented mainly the effect of distance from the field edge. In 1997, samples in the boundary and headland area separated completely along axis one. Both headland and boundary samples were distinct from field samples on axis two, and the only between-distance overlap that occurred along axis one was between the three cuts per year headland plots and the field (Fig. 1). Evidence of slight differences in beetle species composition between cutting regimes was apparent along axis one, with the uncut plots separating from the cut plots in the headland and the boundary. In 1998, the main distance-related species composition groupings were still apparent along the first axis, with field samples clearly separated from all other sites. Species composition changes related to cutting regime had become more distinct in the headland, with the three cuts per annum sample separated completely from the uncut and one cut plots along both axes (Fig. 2). The position of the three cuts per year plots indicated that species composition was more similar to the field samples. The uncut and one cut per annum plots overlapped along axis one, with the uncut plots lying closest to field boundary samples. Boundary samples showed wider variation in species composition than in 1997, yet like the headland samples, the greatest overlap occurred between the two less frequently cut plots. One boundary sample lying behind a three cuts per year plot was more like the uncut headland plots in species composition. In a DCA ordination combining data from both years, the first two axes explained 35% of variation, with distance and cutting regime effects occurring
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269
Table 4 Relative abundance-activity (percentage of total carabid catch at each location) of the carabid species recorded in headland plots under different cutting regimes in 1997 and 1998 Species
1997
1998
Uncut headland 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13. 14. 15. 16. 17. 18. 19. 20. 21. 22. 23. 24. 25. 26. 27. 28. 29. 30. 31. 32. 33. 34. 35. 36. 37. 38. 39. 40. 41. 42. 43. 45. 46.
Calathus fuscipesa Harpalus rufipesa Laemostenus terricolaa Nebria brevicollisa Patrobus atrorufusa Pterostichus cupreusa Pterostichus madidusa Pterostichus melanariusa Pterostichus nigera Carabus granulatusb Carabus nemoralisb Amara communisc Amara plebejac Trechus microsc Trechus quadristriatusc Asaphidion flavipesd Bembidion lamprosd Dyschirius globosusd Loricera pilicornisd Notiophilus biguttatusd Agonum dorsalee Agonum fuliginosume Agonum gracilee Agonum muellerie Badister bipustulatuse Bembidion aeneume Bembidion guttulae Bembidion mannerheimie Bembidion tetracolume Calathus melanocephaluse Calathus piceuse Clivina fossore Elaphrus cupreuse Leistus fulvibarbise Leistus rufescense Pterostichus diligense Pterostichus minore Pterostichus nigritae Pterostichus rhaeticuse Pterostichus strenuuse Pterostichus vernalise Stomis pumicatuse Trechus obtususe Synuchus nivalisf Trichocellus placidusf
Total no. of species Total no. of individuals
One cut per annum headland
Three cuts per annum headland
4.9
(4.2)
9.8
(4.1)
0.1 2.3
11.8
(9.9) (0.8)
10.4
(8.6)
18.0
19.6 0.9
(0.4) (12.9) (1.3)
0.1 33.6 2.1 0.1
(14.5) (0.2)
0.1 30.0 1.7
6.2 1.0 0.3 6.2
(1.1) (3.0) (0.4) (0.2) (4.2)
35.9 0.9 0.4 0.4
0.2 3.1 0.3 3.3
(0.2) (2.7)
(16.3) (0.2) (1.5)
26.9 0.3
(3.6) (0.2) (24.9) (0.2) (1.6)
(0.4) (2.1) (0.4)
0.2 0.2
(1.8) (0.7)
(3.6) (0.2) (13.1) (0.3)
Uncut headland 0.2 4.1
(0.8) (8.2)
0.6 1.6
0.5
(2.5) (0.4) (3.6) (0.2)
0.8
2.0
1.8 0.1
(1.4) (4.8) (0.5) (4.1) (0.5) (3.4)
(0.3) (2.5)
1.5
(2.7)
1.1
14.0 0.2 0.2
(20.2) (3.4)
22.7
(25.2) (1.8)
12.9
(13.1) (3.6)
(8.4) (2.6)
25.4 1.3
(12.6) (1.4)
(10.6)
32.7 1.3
(21.2) (3.6)
0.3 0.1 1.5 0.2 4.9
(0.9) (1.0) (0.2) (0.2) (2.2)
0.7 1.6 0.2 0.2 6.6
(4.5) (1.4)
(2.2) (0.4)
0.1 0.5 1.0 0.4 2.3
(1.7)
31.5
(30.4)
23.4
(10.6)
0.2
(1.5)
0.9 0.2
(0.6)
0.5 0.3 0.5 0.2 2.8 0.2 27.8 0.2 0.5
0.2
(0.7) (2.1)
(0.3)
0.2 0.5
(0.4) (1.8)
0.1
0.5
(0.8) (1.4)
(3.6)
0.2 26.5 1.1
(2.2) (8.8)
38.7 0.1 0.4
(0.3) (2.5) (0.3) (13.9) (0.3)
(0.3) (0.3)
(1.1)
0.6 2.0
Three cuts per annum headland
(2.8)
(0.2) (2.5)
One cut per annum headland
0.1 1.9 0.9 0.1 1.2
(1.4) (7.6) (0.7) (2.1)
0.4 5.5 0.5 0.4
(1.1) (10.6) (1.4) (1.1)
0.2 4.4 0.7 1.1
(0.4) (15.9) (3.1) (1.3)
0.6 2.3 0.9 1.0
(0.3) (16.2) (1.9) (1.1)
2.4
(0.7)
2.9
(3.6)
2.9
(4.9)
1.6
(2.8)
0.1 0.6
(0.3) (0.7) (1.2) (0.9)
(0.4) (0.4) 0.1 0.4 3.3
1.3 1.5 20 791
(3.6) (1.3) (0.2) (2.7) (0.4) (2.3) (8.0) (2.3) (0.8) (34) (473)
0.8 1.6 0.1 1.0 0.7 0.1 26 890
(0.5) (0.2) (2.9) (2.3) (0.5) (1.8) (1.8) (8.4) (2.7) (29) (442)
0.8 0.2 0.1 0.1 0.4 27 1053
(3.3) (0.3) (1.2) (9.3) (2.7) (0.2) (31) (582)
0.4 0.2 5.9 0.2 1.6 2.5 27 559
(0.8) (0.6) (1.1) (1.7) (0.3) (2.2) (0.6) (2.0) (5.0) (5.3) (0.6) (29) (357)
(0.3)
0.7 2.9 0.2 1.0 0.3 26 611
(0.9) (3.1) (1.8) (0.4) (2.7) (0.4) (1.8) (3.5) (3.1) (25) (226)
0.1 0.3
(2.8) (0.3)
0.8
(0.8) (0.3) (1.7) (4.7) (0.8)
0.4 0.3 23 927
(28) (359)
Figures in parentheses are relative abundance-activity of carabid species in the boundary adjacent to the headland. Numbers in superscript following species name are eco-groups following Cole et al. (2002) (Harpalus affinis was not recorded in the headland plots and has been omitted). a Medium sized nocturnal predators. b Large wingless Carabus spp. c Small mainly diurnal plant feeders. d Small diurnal Collembola feeders. e Small nocturnal predators. f Very small to medium sized nocturnal plant feeders.
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Table 5 Relative abundance-activity (percentage of total carabid catch at each location) of the carabid species recorded at different distances from the field boundary adjacent to the three cuts per annum plots in 1997 and 1998 Species
1997 Boundary
1. 2. 3. 4. 5. 7. 8. 9. 11. 12. 13. 14. 15. 16. 17. 18. 19. 20. 21. 22. 24. 26. 27. 28. 29. 30. 32. 35. 36. 37. 38. 40. 41. 42. 43. 44. 45. 46.
Calathus fuscipesa Harpalus rufipesa Laemostenus terricolaa Nebria brevicollisa Patrobus atrorufusa Pterostichus madidusa Pterostichus melanariusa Pterostichus nigera Carabus nemoralisb Amara communisc Amara plebejac Trechus microsc Trechus quadristriatusc Asaphidion flavipesd Bembidion lamprosd Dyschirius globosusd Loricera pilicornisd Notiophilus biguttatusd Agonum dorsalee Agonum fuliginosume Agonum muellerie Bembidion aeneume Bembidion guttulae Bembidion mannerheimie Bembidion tetracolume Calathus melanocephaluse Clivina fossore Leistus rufescense Pterostichus diligense Pterostichus minore Pterostichus nigritae Pterostichus strenuuse Pterostichus vernalise Stomis pumicatuse Trechus obtususe Harpalus affinisf Synuchus nivalisf Trichocellus placidusf
Total no. of species Total no. of individuals
3.6 0.2 13.1 0.3
1998 Headland
Field
0.1 2.3
1.4
18.0
25.8
16.0
8.4 2.6 0.9 1.0 0.2 0.2 2.2
0.1 30.0 1.7 0.3 0.1 1.5 0.2 4.9
30.4
31.5
43.3
1.5
0.2
0.7 2.1 0.2
0.2
2.7 0.3 0.2 0.2 0.6
1.4 7.6 0.7 2.1 0.7 0.3 0.7 1.2 0.9 3.3 0.3 1.2 9.3
0.1 1.9 0.9 0.1 1.2 2.4
0.1 0.6 0.8 0.2 0.1 0.1
Boundary 0.3 2.5
0.2
1.3
0.3
Field
1.1
0.1 0.5
13.1 3.6
12.9
21.8
21.2 3.6 1.7
32.7 1.3 0.1 0.5 1.0 0.4 2.3
3.6 0.5 3.3
Headland
0.3 2.5 0.3 13.9 0.6 0.3 0.8 1.4
38.7 0.1 0.4 0.1
1.5 44.8
2.4 2.1 0.5 19.9 2.6 0.3 0.3 0.2
0.3 16.2 1.9 1.1 2.8 0.3
0.6 2.3 0.9 1.0 1.6
2.8 0.3 0.8 0.3 1.7 4.7
0.1 0.3 0.8
0.8
0.3
0.1 0.9 0.5 0.1
0.9 0.1
0.4
0.3 2.7 0.2 31 582
0.4 27 1053
16 631
28 359
23 927
0.1 20 910
Letters in superscript following species name are eco-groups following Cole et al. (2002) (Pterostichus cupreus, Carabus granulatus, Agonum gracile, Badister bipustulatus, Calathus piceus, Elaphrus cupreus, Leistus fulvibarbis, Pterostichus rhaeticus were not recorded at these positions and have been omitted). a Medium sized nocturnal predators. b Large wingless Carabus spp. c Small mainly diurnal plant feeders. d Small diurnal Collembola feeders. e Small nocturnal predators. f Very small to medium sized nocturnal plant feeders.
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271
160 1997 Field
140
Uncut One cut / year Three cuts / year
120
Axis 2
100
80
Headland
60
40
Boundary
20 Field 0 0
20
40
60
80
100
120
140
160
180
Axis 1
Fig. 1. Diagram of the first two axes of a DCA ordination of carabid assemblage data at 21 sampling locations in 1997 (eigenvalue axis 1 = 0.2; axis 2 = 0.059).
mainly along the first axis (27% of variation explained). In 11 of the 12 plots that were either uncut, cut once per annum, or located in the boundary directly adjacent to these treatments, the location of plots on the DCA in 1998 fell to the left of the 1997 positions (Fig. 3), i.e. the carabid species composition became more “boundary” like. The species composition in four of the six three cuts per annum and corresponding adjacent boundary plots became more like that of the field. This effect was strongest in the headland, where all three plots shifted towards a field species composition, and was more ambiguous in the boundary area that lay behind the cut plots. These patterns reflect changes in species diversity, in overall number of species and in sample heterogeneity and the tendency of certain species to become relatively more abundant-active in particular treatments. In both years, diversity was significantly higher in the field boundary than in the headland (1997 F1,6 = 77.4, P < 0.001; 1998 F1,6 = 57.2, P < 0.001, Fig. 4). Amara communis, B. lampros, Clivina fos-
sor, Harpalus rufipes, N. brevicollis, P. melanarius, P. strenuus and Trechus quadristriatus were significantly more abundant-active in the headland than in the field boundary in at least one of the study years (P < 0.05– < 0.001). In contrast, Bembidion mannerheimi, Synuchus nivalis and Trechus obtusus were significantly more abundant-active in the boundary than in the headland (P < 0.05– < 0.001). Among the most similar treatments (field, three cuts per annum headland and adjacent boundary) there was a highly significant variation in diversity with distance from the field edge (F2,12 = 33.2, P < 0.001). An ANOVA with treatment and year as factors and an interaction term revealed that Simpson’s diversity index was significantly higher in the boundary than in the headland in both 1997 and 1998 (mean Simpson’s 1−D = 0.85 and 0.73, respectively). There was a similar overall effect between the headland and the field, examination indicating a significant difference in 1997 but not in 1998, year and interaction terms being not significant.
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K.A. Haysom et al. / Agriculture, Ecosystems and Environment 102 (2004) 263â&#x20AC;&#x201C;277 160 1998 Field Uncut One cut / year Three cuts / year
140
120
Axis 2
100 Headland
Field 80
60 Boundary 40
20
0 0
20
40
60
80
100
120
140
160
180
Axis 1
Fig. 2. Diagram of the first two axes of a DCA ordination of carabid assemblage data at 21 sampling locations in 1998 (eigenvalue axis 1 = 0.257; axis 2 = 0.094).
Carabid diversity did not differ between cutting regimes in 1997, and was significantly greater in 1998 in the uncut plots than in the one cut and three cuts plots (F2,4 = 9.43, P < 0.05, Fig. 4). Some species were trapped significantly more frequently in, or adjacent to, particular cutting regimes. In 1997, headland catches of P. strenuus and A. communis were greatest in the uncut plots. Catches of A. communis were significantly larger in the uncut than in either the one cut or three cuts regimes (F2,4 = 20.5, P < 0.01). Catches of P. strenuus decreased as cutting intensity increased, with a significant difference between the three cut and uncut plots (F2,4 = 11.1, P < 0.05). In 1998, the catch of T. quadristriatus was largest in the uncut headland plots and in the boundary adjacent to the uncut plots. There were significant differences between the uncut and both cut headland regimes, and between the traps adjacent to uncut and adjacent to one cut in the boundary (F2,4 = 9.06, P < 0.05). Catches of N. brevicollis, P. melanarius and P. niger were greatest in the most
intense cutting regime in at least 1 year of the study. In 1998, the relative contribution of P. melanarius to the assemblage (Tables 4 and 5) increased both with the distance from the field edge, and with increasing frequency of cutting. Catches of N. brevicollis (1997 F2,4 = 13.6, P < 0.05) and P. melanarius (1998 F2,4 = 10.4, P < 0.05) were significantly higher in the three cuts per annum than in either the one cut or uncut treatments. For P. niger (1998 F2,4 = 8.5, P < 0.05) there was a significant difference between the three cut and the one cut treatments.
4. Discussion The carabid distribution changed in relation to cutting intensity. Grass cutting is a catastrophic process, whereby structural changes to the vegetation occur rapidly, often over large areas (Morris, 2000). Modern silage machinery poses a particular hazard in this respect, due to its extremely efficient suction method.
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273
Fig. 3. Diagram of the first two axes of a DCA of carabid assemblage data at 21 sampling locations (data of 1997 and 1998 combined). Arrows link data collected in 2 years and point towards 1998 assemblage (eigenvalue axis 1 = 0.219; axis 2 = 0.065).
The degree to which different taxonomic groups and individual species are affected by mechanical cutting is influenced by their stratification within the vegetation, the way they use different grassland resources, their mobility, the scale of the cutting operation, life-history stage and the timing of cutting relative to time of breeding. Individual species respond differently to cutting (Morris, 1979, 2000; Di Giullio et al., 2001), but Morris and Rispin (1988) predicted that large-scale cutting operations repeated at the same time every year would select assemblages comprising species with appropriate breeding strategies, or change the balance of phytophagous, saprophagous, fungivorous and detritivorous strategists. Carabid responses to the three headland cutting regimes took the form of changes in the relative abundance-activity of individual species rather than presence or absence from particular regimes. The small size of the plots (10 m Ă&#x2014; 10 m) has to be considered, A. dorsale, P. cupreus and P. melanarius being capable of travelling 100â&#x20AC;&#x201C;300 m distances during a
2-week period (Welling, 1990). Indeed Thomas et al. (1998) found that P. melanarius dispersed further than 100 m in 48 h. Even the smaller species recorded in this study would have been capable of crawling across several adjacent plots during the trapping periods. The fact that multivariate techniques separated assemblages trapped in different management regimes or field locations despite the small plot dimensions, implies that the trends identified were genuine. Luff (1990) also commented that differences between treatments were more likely to reflect differential activity or habitat preferences, rather than population differences caused by reproduction in each habitat. Species with abundance-activity significantly greater in the uncut plots (P. strenuus, T. quadristriatus and A. communis), all had a life history requiring tall, less disturbed vegetation. P. strenuus is associated with shady places and damp deciduous soils (Lindroth, 1974; Luff, 1998), and A. communis with open habitats such as grasslands, cultivated gardens and open woodlands (Luff, 1998).
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Fig. 4. Mean species diversity (Simpson’s 1 − D) of the carabid assemblage recorded in three cutting regimes, at two distances from the field edge in 1997 and 1998. Diversity was significantly higher in the boundary than in the headland in both years (P < 0.001), significantly different between treatments in 1998, and significantly higher in uncut plots than in one cut and three cuts (P < 0.05) (no significant block, or cutting treatment × distance effects in either year).
Three species increased significantly in abundanceactivity with increasing cutting frequency: P. melanarius, P. niger and N. brevicollis were all rapidly moving predators. Active hunting may be facilitated by open habitats but the greater trap catches of these species in the three cuts per year plots may be due to individuals being forced to make more foraging movements in response to lower densities of prey (Chiverton, 1984). Lower densities of potential prey groups such as slugs, caterpillars, and sawfly larvae were recorded in the most frequently cut plots in pitfall and sweepnet surveys in 1998 (Haysom, unpublished data). In addition to direct changes in vegetation structure, management can modify substrate characteristics, and indirectly affect invertebrates that are sensitive to soil conditions (e.g. organic matter, moisture gradients). Some changes in soil nutrient concentrations occurred during this experiment. Year to year variation in the level of soil nutrients is influenced greatly by weather
conditions, soil texture and topography, which together mediate microbial activity and the process of mineralisation. Potassium was the only nutrient to show a trend related to cutting regime, its concentration decreasing at higher cutting intensity. Repeated cutting and removal of grass in the frequently cut plots was sufficient to deplete the soil concentration over the course of each season. The main influence on carabid species composition in the plots appeared to be in terms of vegetation biomass, and may have acted on the carabids by altering local temperature and moisture conditions. There was a strong distance-from-boundary effect on Leistus fulvibarbis, L. rufescens, Calathus piceus, Laemostenus terricola, Agonum gracile and Badister bipustulatus, six species that were recorded only in the boundary. Leistus spp., C. piceus and A. gracile all prefer shaded, moist environments (Lindroth, 1974; Luff, 1998),whereas P. minor is linked to damp habitats and water margins (Lindroth, 1974; Luff, 1998). Carabus
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nemoralis, a large and slow moving species, was never recorded more than 5 m from the field boundary. The reduction in the complexity of the beetle assemblage with increased distance from the field edge suggests that conservation management techniques on agricultural grasslands are likely to be of greatest overall benefit to carabids when applied at the margin. Bell et al. (1996) trapped the greatest number of carabid species on an agricultural grassland within 1 m of a hedge boundary, and decreases in carabid species richness with increased distance from a field boundary have also been recorded in several arable studies (Fournier et al., 1998). Few studies have used manipulative experiments to ascertain the response of carabids to variations of particular treatments on improved grassland (Rushton et al., 1990). Instead there has been a tendency to rank the management intensity of locations in multi-site studies (e.g. Eyre et al., 1989; Rushton et al., 1990; Luff et al., 1992). Eyre et al. (1989) found that management was the most important factor influencing carabid distribution in lowland northern England where all sites had similar substrates. Young leys managed intensively for silage, hay or grazing were characterised by an abundance of very active species (Loricera pilicornis, Bembidion aeneum and P. melanarius) together with colonisers of recently disturbed land (Notiophilus biguttatus). Permanent pastures trapped more Calathus melanocephalus and Pterostichus strenuus but Carabus violaceus, a large species, was an indicator of unmanaged grasslands. Blake et al. (1994) found that such variations in species composition could affect carabid biomass characteristics at a site, with average body size decreasing as the intensity of management increased. The restoration of species-rich grassland is an alternative approach to increasing the biodiversity of agricultural land, including improved grass, that is already prescribed in agri-environment schemes at field-scale (MAFF, 1999b). However, high soil fertility, especially high P levels, is a frequent obstacle. Fertiliser applications as low as 25 kg N ha−1 can reduce plant diversity on peat soils, mainly due to the associated P (Tallowin, 1996). In the headland of our study field, which did not receive fertiliser, extractable P levels were still moderate by agricultural standards (Dyson, 1992), and due to progressive P release, soil levels became significantly higher over 3 years. Sowing wild-
275
flowers to create species-rich grassland would not have been appropriate at the headland-scale, since groundwater transfer of nutrients applied in the field would have restricted the degree of P reduction possible in the margin, constraining plant diversity. In this experiment the headland after 3 years was still essentially an improved agricultural grassland yet the diversity of the carabid assemblage increased in plots left uncut or cut only once each summer. The net result of low intensity management was an increase in the effective area of the inter-crop habitat, i.e. beetle species that normally utilised the boundary also used the less frequently cut plots and the value of inter-crop habitats as wildlife corridors, refuges, and overwintering sites is well documented (e.g. Thomas et al., 1992; Dennis et al., 1994; Lee et al., 2001). Many of the species studied are common on farmland and P. melanarius, H. rufipes, B. lampros, P. cupreus, A. dorsale, A. muelleri, H. affinis and Trechus quadristriatus are known to occur on over 67% of agricultural areas studied throughout northern Europe and Russia (Thiele, 1977). The present results supply information for potential future applications. In addition to the various “edge” management prescriptions for arable land the development of prescriptions for the headlands of intensively managed grassland could prove a useful addition to the suite of management tools currently available to farmers.
Acknowledgements This project was funded by Scottish Executive Environment and Rural Affairs Department (formerly Scottish Office Agriculture, Environment and Fisheries Department), Dumfries and Galloway European Partnership and The Game Conservancy Trust. We thank Shona Blake, Béatrice Rouget, Severino Opio Joseph, Jim McCleary, Kate Wenham, Heidi Cunningham, Alex Brook and Anne Dowdeswell for assistance with fieldwork and invertebrate sorting. Thanks are also due to Tim Bromilow for information on soils, to George Fisher for useful discussions and for project development, and to other SAC staff for maintaining the headland plots and general advice. Ian Nevison from BIOSS provided guidance on statistical analyses.
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References Aebischer, N.J., 1990. Assessing pesticide effects on non-target invertebrates using long-term monitoring and time-series modelling. Funct. Ecol. 4, 369–373. Aebischer, N.J., 1991. Twenty years of monitoring invertebrates and weeds in cereal fields in Sussex. In: Firbank, L.G., Carter, N., Darbyshire, J.F., Potts, G.R. (Eds.), The Ecology of Temperate Cereal Fields. 32nd Symposium of the British Ecological Society. Blackwell Scientific, Oxford, pp. 305–331. Bell, A.C., Henry, T., McAdam, J.H., 1996. Managing grass fields for wildlife value. In: Vegetation Management in Forestry, Amenity and Conservation Areas: Managing for Multiple Objectives. Aspects of Applied Biology 44, 251–256. Bence, S., Stander, K., Griffiths, M., 1999. Nest site selection by the harvest mouse (Micromys minutus) on arable farmland. In: Boatman, N.D., Davies, D.H.K., Chaney, K., Feber, R., deSnoo, G.R., Sparks, T.H. (Eds.), Field Margins and Buffer Zones: Ecology, Management and Policy. Aspects of Applied Biology 54, 197–202. Berendse, F., Oomes, M.J.M., Altena, H.J., Elberse, W.T., 1992. Experiments on the restoration of species-rich meadows in The Netherlands. Biol. Conserv. 62, 59–65. Blake, S., Foster, G.N., 1998. The influence of grassland management on body size in ground beetles and its bearing on the conservation of wading birds. In: Joyce, C.B., Wade, P.M. (Eds.), European Wet Grasslands: Biodiversity, Management and Restoration. Wiley, Chichester, pp. 163–169. Blake, S., Foster, G.N., Eyre, M.D., Luff, M.L., 1994. Effects of habitat type and grassland management practices on the body size distribution of carabid beetles. Pedobiologia 38, 502–512. Chamberlain, D.E., Fuller, R.J., Bunce, R.G.H., Duckworth, J.C., Shrubb, M., 2000. Changes in the abundance of farmland birds in relation to the timing of agricultural intensification in England and Wales. J. Appl. Ecol. 37, 771–788. Chiverton, P.A., 1984. Pitfall-trap catches of the carabid beetle Pterostichus melanarius in relation to gut contents and prey densities, in insecticide treated and untreated spring barley. Entomol. Exp. Appl. 36, 23–30. Chiverton, P.A., Sotherton, N.W., 1991. The effects on beneficial arthropods of the exclusion of herbicides from cereal crop edges. J. Appl. Ecol. 28, 1027–1039. Cole, L.J., McCracken, D.I., Dennis, P., Downie, I.S., Griffin, A.L., Foster, G.N., Murphy, K.J., Waterhouse, A., 2002. Relationships between agricultural management and ecological groups of ground beetles (Coleoptera: Carabidae) on Scottish farmland. Agric. Ecosyst. Environ. 93, 323–336. Countryside Council for Wales, 1999. Tir Gofal. An AgriEnvironment Scheme for Wales. Farmers Handbook. CCC 165. Critchley, C.N.R., Chambers, B.J., Fowbert, J.A., Sanderson, R.A., Bhogal, A., Rose, S.C., 2002. Association between lowland grassland communities and soil properties. Biol. Conserv. 105, 199–215. Dennis, P., Thomas, M.B., Sotherton, N.W., 1994. Structural features of field boundaries which influence the overwintering densities of beneficial arthropod predators. J. Appl. Ecol. 31, 361–370.
Di Giullio, M., Edwards, P.J., Meister, E., 2001. Enhancing insect diversity in agricultural grasslands: the roles of management and landscape structure. J. Appl. Ecol. 38, 310–319. Dover, J.W., 1997. Conservation headlands: effects on butterfly distribution and behaviour. Agric. Ecosyst. Environ. 63, 31–49. Dyson, P., 1992. Removal by crops and PK balance sheets. SAC TN T308 Fertiliser Series No. 13. Eyre, M.D., Luff, M.L., Rushton, S.P., Topping, C.J., 1989. Ground beetles and weevils (Carabidae and Curculinoidea) as indicators of grassland management practices. J. Appl. Entomol. 107, 508–517. Feber, R.E., Smith, H., Macdonald, D.W., 1999. The importance of spatially variable field margin management for two butterfly species. In: Boatman, N.D., Davies, D.H.K., Chaney, K., Feber, R., deSnoo, G.R., Sparks, T.H. (Eds.), Field Margins and Buffer Zones: Ecology, Management and Policy. Aspects of Applied Biology 54, 155–162. Fournier, E., Loreau, M., Havet, P., 1998. Effects of new agricultural management practices on the structure and diversity of ground beetle communities (Coleoptera, Carabidae). Gibier Faune Sauvage 15, 43–53. Fuller, R.J., 1987. The changing extent and conservation interest of lowland grasslands in England and Wales: a review of grassland surveys 1930–1984. Biol. Conserv. 40, 281–300. Genstat 5 Committee, 1993. Genstat 5 Release 3 Reference Manual. Clarendon Press, Oxford. Hance, T., 1990. Relationships between crop types, carabid phenology and aphid predation in agroecosystems. In: Stork, N.E. (Ed.), The Role of Ground Beetles in Ecological and Environmental Studies. Intercept, Andover, pp. 55–64. Hawthorne, A.J., Hassall, M., Sotherton, N.W., 1998. Effects of cereal headland treatments on the abundance and movements of three species of carabid beetles. Appl. Soil Ecol. 9, 417–422. Haysom, K.A., McCracken, D.I., Foster, G.N., Sotherton, N.W., 1999. Grass conservation headlands—adapting an arable technique for the grassland farmer. In: Boatman, N.D., Davies, D.H.K., Chaney, K., Feber, R., deSnoo, G.R., Sparks, T.H. (Eds.), Field Margins and Buffer Zones: Ecology, Management and Policy. Aspects of Applied Biology 54, 171–178. Hopkins, A., Hopkins, J.J., 1994. UK Grasslands now: agricultural production and nature conservation. In: Haggar, R.J., Peel, S. (Eds.), Grassland Management and Nature Conservation. BGS Occasional Symposium No. 28. Cambrian Printers, Aberystwyth, pp. 10–19. Lee, J.C., Menalled, F.D., Landis, D.A., 2001. Refuge habitats modify impact of insecticide disturbance on carabid communities. J. Appl. Ecol. 38, 472–483. Lindroth, C.H., 1974. Coleoptera: Carabidae. Handbooks for the Identification of British Insects, vol. 4. Royal Entomological Society, London, pp. 1–148. Luff, M.L., 1987. Biology of polyphagous ground beetles in agriculture. Agric. Zool. Rev. 2, 237–277. Luff, M.L., 1990. Spatial and temporal stability of carabid communities in a grass/arable mosaic. In: Stork, N.E. (Ed.), The Role of Ground Beetles in Ecological and Environmental Studies. Intercept, Andover, pp. 191–200. Luff, M.L., 1996. Environmental assessments using ground beetles (Carabidae) and pitfall traps. In: Eyre, M.D. (Ed.),
K.A. Haysom et al. / Agriculture, Ecosystems and Environment 102 (2004) 263–277 Environmental Monitoring, Surveillance and Conservation Using Invertebrates. EMS, Newcastle Upon Tyne, pp. 42–47. Luff, M.L., 1998. Provisional Atlas of the Ground Beetles (Coloptera, Carabidae) of Britain. Biological Records Centre, Huntingdon. Luff, M.L., Eyre, M.D., Rushton, S.P., 1992. Classification and prediction of grassland habitats using ground beetles (Coleoptera, Carabidae). J. Environ. Manage. 35, 301–315. MAFF, 1999a. Arable Stewardship: Information and How to Apply. PB 3957A. MAFF, 1999b. The Countryside Stewardship Scheme: Information and How to Apply. PB 3950A. McCracken, D.I., Bignal, E.M., 1998. Applying the results of ecological studies to land-use policies and practices. J. Appl. Ecol. 35, 961–967. Meek, W., Loxton, R., Sparks, T., Pywell, R., Pickett, H., Nowakowski, M., 2002. The effect of arable field margin composition on invertebrate biodiversity. Biol. Conserv. 106, 259–271. Morris, M.G., 1979. Responses of grassland invertebrates to management by cutting. II. Heteroptera. J. Appl. Ecol. 16, 417– 432. Morris, M.G., 2000. The effects of structure and its dynamics on the ecology and conservation of arthropods in British grasslands. Biol. Conserv. 95, 129–142. Morris, M.G., Rispin, W.E., 1988. A beetle fauna of oolithic limestone grassland, and the responses of species to conservation management by different cutting régimes. Biol. Conserv. 43, 87–105. Mountford, J.O., Lakhani, K.H., Holland, R.J., 1996. Reversion of grassland vegetation following the cessation of fertiliser application. J. Veg. Sci. 7, 219–228. Rands, M.R.W., 1985. Pesticide use on cereals and the survival of grey partridge chicks: a field experiment. J. Appl. Ecol. 22, 49–54. Rushton, S.P., Eyre, M.D., Luff, M.L., 1990. The effects of management on the occurrence of some carabid species in grassland. In: Stork, N.E. (Ed.), The Role of Ground Beetles in Ecological and Environmental Studies. Intercept, Andover, pp. 209–216. Schumacher, W., 1987. Measures taken to preserve arable weeds and their associated communities. In: Way, J.M., Greig-Smith, P.W. (Eds.), Field Margins. BCPC Monograph No. 35. British Crop Protection Council, Thornton Heath, UK, pp. 109–112. SEERAD, 2000. The Rural Stewardship Scheme. Scottish Executive Rural Affairs Department, Edinburgh. Simpson, E.H., 1949. Measurement of diversity. Nature 163, 688. Smith, K.A. (Ed.), 1983. Soil Analysis: Instrumental Techniques and Related Procedures. Marcel Dekker, New York. Sotherton, N.W., 1990. Conservation headlands: a practical combination of intensive cereal farming and conservation. In: Firbank, L.G., Carter, N., Darbyshire, J.F., Potts, G.R. (Eds.), The Ecology of Temperate Cereal Fields. 32nd Symposium of the British Ecological Society. Blackwell, Oxford, pp. 373–397.
277
Sotherton, N.W., 1998. Land use changes and the decline of farmland wildlife: an appraisal of the set-aside approach. Biol. Conserv. 83, 259–268. Sotherton, N.W., Moreby, S.J., 1992. The importance of beneficial arthropods other than natural enemies in cereal fields. Asp. Appl. Biol. 31, 11–18. Speight, M.R., Lawton, J.H., 1976. The influence of weed-cover on the mortality imposed on artificial prey by predatory ground beetles in cereal fields. Oecologia 23, 211–223. Stoate, C., 1996. The changing face of lowland farming and wildlife. Part 2. 1945–1995. Br. Wildlife 7, 162–172. Stoate, C.S., Szczur, J., 1994. Nest site selection and territory distribution of yellowhammer (Emberiza citrinella) and whitethroat (Sylvia communis) in field margins. In: Boatman, N.D. (Ed.), Field Margins: Integrating Agriculture and Conservation. BCPC, Farnham, pp. 129–132. Tallowin, J.R.B., 1996. Effects of inorganic fertilizers on flower-rich hay meadows: a review using a case study on the Somerset Levels, UK. Grasslands Forage Abst. 66, 147–152. ter Braak, C.J.F., 1987. CANOCO—A FORTRAN program for canonical community ordination by (partial) (detrended) (canonical) correspondence analysis, principal components analysis and redundancy analysis. ITI-TNO, Wageningen. Tew, T.E., Todd, I.A., Macdonald, D.W., 1994. Field margins and small mammals. In: Boatman, N. (Ed.), Field Margins: Integrating Agriculture and Conservation. BCPC Monograph No. 58. British Crop Protection Council, Farnham, pp. 85–94. Thiele, H.U., 1977. Carabid Beetles in Their Environment. Springer-Verlag, Berlin. Thomas, C.F.G., Marshall, E.J.P., 1999. Arthropod abundance and diversity in differently vegetated margins of arable fields. Agric. Ecosyst. Environ. 72, 131–144. Thomas, M.B., Wratten, S.D., Sotherton, N.W., 1992. Creation of “island” habitats in farmland to manipulate populations of beneficial arthropods: predator densities and species composition. J. Appl. Ecol. 29, 524–531. Thomas, C.F.G., Parkinson, L., Marshall, E.J.P., 1998. Isolating the components of activity–density for the carabid beetle Pterostichus melanarius in farmland. Oecologia 116, 103–112. Vickery, J.A., Tallowin, J.R., Feber, R.E., Asteraki, E.J., Atkinson, P.W., Fuller, R.J., Brown, V.K., 2001. The management of lowland neutral grasslands in Britain: effects of agricultural practices on birds and their resources. J. Appl. Ecol. 38, 647– 664. Welling, M., 1990. Dispersal of ground beetles (Col. Carabidae) in arable land. Med. Fac. Landbouww. Rijksuniv. Gent. 55, 483–491. Wilson, P.J., 1994. Botanical diversity in arable field margins. In: Boatman, N. (Ed.), Field margins: Integrating Agriculture and Conservation. BCPC Monograph No. 58. British Crop Protection Council, Farnham, pp. 53–58. Winder, L., 1990. Predation of the cereal aphid Sitobion avenae by polyphagous predators on the ground. Ecol. Entomol. 15, 105–110.