Insects, plants and succession: advantages of long-term set-aside

_AwricuUture r~cosystems eg Environment ELSEVI ER

Agriculture,Ecosystentgand Environment53 ( 1995) 201-217

Insects, plants and succession: advantages of long-term set-aside Sarah A. C o r b e t I Zoology Department, Downing Street, Cambridge CB2 3EJ, UK

Accepted 7 December 1994

Abstract Countryside quality over much of Britain depends on agricultural policies, important among which is the set-aside programme in which area based paynvmts are made to encourage farmers to take arable land out of food production. Decisions between long-term and short-:erm set-aside will influence the future age profile of communities on cnculiivated land, affecting their conservation value and their role as sources of pests, natural enemies and pollinators for nearby crops. Because long-tern1 studies of set-aside in Britain are not yet available, other evidence is u~d to review the successional changes to be expected in a 10 year period of natural regeneration after ploughing. Succession is considered in terms of ecological attributes of plant and insect communities, rather than lists of individual species, so that the concepts are more generally applicaEe. In a landscape increasingly affected by disturbance, in which some groups of plant and insect species of long-established vegetation are already decreasing, Jong-term unploughed set-aside offers an opportunity to establish and protect patches of undisturbed perennial herbaceous vegetation and their associated fauna, helpit;g common species to remain common and ecologically important species to remain functional. Keyword.¢:Conservation;Ploughing;Perennials;Pollinators:Naturalenem,e:.

1. Introduction To reduce surplus food production, agricultural payments have been structured to encourage farmers to set aside a proportion of their arable land and reduce the area under food crops. Set-aside is a recent innovation in Britain (Firbank et al., 1993). Studies in USA, where set-aside has a longer history (Ervin, 1992), give useful clues about general patterns, but aspects of succession that depend on the characteristics of individual species are better revealed by research in Europe, where the flora shares more species with Britain and with the list of British plant species whose ecological attributes are recorded in the data base of the Natural EnvironTel. 0 2 2 3 336680; F a x . 0 2 2 3 336676; e-mail [email protected]. 0167-8809/95/$09.50 © 1995El~vier Science B.V. All fights re~rved SSDI 0167-8809( 94 )00581-8

ment Research Council Unit of Comparative Plant Ecology at the University of Sheffield (Grime et al., 1988). The British arable landscape comprises a mosaic of annual crops in a matrix of uncultivated areas. This uncultivated land harbours communities of animals and plants that contribute to the quality of the landscape and affect the performance of adjacent crops. The nature of the community on uncultivated land, and the dynamics of the transfer of organisms between that community and adjacent crops, depend in part on the time that has elapsed since disturbance. Many insect species benefit crops by acting as natural enemies or pollinators. For these, short-lived vegetation such as an annual crop or I-year set-aside may act as a sink rather than a source if the numbers within the crop are main-

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tained net by in situ breeding but by immigration from uncultivated areas that support population growth (Fry, 1994) Some first-year set-.aside communities may be of little interest to natural historians and may harbour organisms regarded as weeds and pests, bringing many of the problems associated with annual arable cropping. If regulations permit, natural regeneration on long-term set-aside can produce perennial vegetation which may be of considerable botanical and entomological interest and r~lay benefit nearby crops by acting as a source of both pollinators and the natural enemies of pests. Although set-aside is designed primarily to reduce food production, it can be managed to meet other objectives, some of which are considered by Firbank et al. (1993). Management options depend on the period of undisturbed vegetation development after ploughing. This period is up to I or 2 years in annual rotational setaside, and also in certain non-rotational set-aside options for which annual tillage is recommended (Firbank et al., 1993). It is substantially longer in c~ther types of non-rotational set-aside. One-year set-aside, like the annually tilled conservation headlands recommended by the Game Conservancy (Sotherton, 1991 ), has a special value for rare arable weeds (Firbank et ai., 1993) and, as a complement to adiacent areas of longer-established vegetation, for gamebird chicks (Firbank et al., 1993; Aebischer et al., 1994). With these well-publicised exceptions, the other wildlife objectives listed by Firbank et al. (1993) are best met by allowing vegetation to develop over a longer period. The options considered by Firbank et al. (1993) ate mostly designed to favour particular rare s~cies or species groups, or to extend existing vegetation type,.;, and are appropriate only in localities where the releva~Lt species or vegetation types occur or have recently occurred. In contrast, the long-term option of natural regeneration, sometimes supplemented with sown pereanials, is expected to be applicable in many situations, including those where there is no case for more specific alternatives, because it is not limited to particular sites or to annually tilled soil by the requirements of particular species. This paper reviews the potential advantages of longer-term management options. Firbank et al. (1993) and Smith et al. (1993) describe the management required to establish perennial herbaceous vegetation. The expected benefits have been considered in relation to particular taxa (e.g. Feber et al., 1994.',, but if an informed choice is to be

made between this long-term optioh and shorter-term altei'natives, wildlife benefits should be documented in terms of whole communities rather than individual taxa. Such documentation is handicapped by a shortage of information iratwo major ways. First, existing studies necessarily focus on a limited range of taxa, and because of site-specific factors no clear overall successional picture emerges in terms of the species present. Secondly, existing empirical studies of natural regeneration on set-aside in Britain are short term (Clarke, 1992). All set-aside taken up in Britain in 1992 was annual rotational. The advent in 1993 of non-rotational setaside lasting at least 5 years and in 1994 of a Habitat Scheme lasting at least 20 years (not currently counting as set-aside) requires more information on the longerterm development of vegetation and its associated fauna. Because long-term options fit less readily into the annual farm management programme, and require more long-term financial commitment from the government, they are unlikely to be widely used unless their advantages are clearly seen. Decisions about setaside and comparable schemes cannot wait for the results of general, long-term studies. Here an attempt is made to explore the expected ecological impact of perennial herbaceous vegetation and its associated insect fauna in terms of both conservation value and interaction with adjacent or succeeding crops, paying more attenuon to the broadleaved herbs on which bees forage than to grasses. The problem of the absence of long-term set-aside studies in Britain is met by basing infere,nces on existing studies of comparable successions. The problem of site-to-site species differences is tackled by considering succession in terms of general ecological attributes, as given for some British plant species by Grime et al. ( 1988 ) rather than of individual species (see also Hills et al., 1994). Although a valuable perspective comes from the United States, where old-field succession has received much attention (e.g. Mellinger and McNaughton, 1975), studies involving European species are emphasised here. The discussion is illustrated by an analysis, in terms of plant species attributes, of a list of species in a IO-year succession on heat-sterilised soil in Germany (Fig. 1; Schmidt, 1976, listed in Ellenberg, 1988). The effects of setaside policy on the distribution and age profile of uncultivated vegetation in the landscape may not be straightforward, but some predictions can be made

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Fig. I Changes over i0 years in plant species eolonising heat-sierilised soil (Schmidt, 1976, as listed in Ellenberg, 1988). In ( a ) - ( f ) , specie~ attnbntes ore taken from Grime et at. (1988) and the data base of the NBRC Unit of Comparative Plant Ecology, University of Sheffield (J.G. Hedg~on, personal communication, 1993 ). (a) Numbers of species of annuals (including summer and winter annuals and species that sometimes behave as annuals) and perennials (including monocorpic perennials). (b) Mean value for the spnc~es present in each year with respect to the three functional types of Grime ( 1974.1979). The extreme strategies ( pure competitors (C). stress-tolerators (S) and ruderals ( R )) are located at the comers of a triangular framework within which intermedia:e strategic types are represented by C-S-R coordinates. A radius value of 5 represents an extreme position most like the pure C. S or R strategy, and a value of I represents the opposite position, most unlike the pure strategy. (c) Percentages of species of annuals and perennials classified as decreasing in abundance. (d) Percentage of species associated with species-rich vegetation (over 22 species m-2). (e) Percentage of species associated with habitats classified as arable, spoil, or wasteland plus pasture. (f) Percentage of species in whi':h flowering begins in spring (M~trch and April) or summer ( May, June and July). (g) Pereentage of species with small seeds (less than 0.5 rag). and percentage with a long.term seed bank (over 5 years). (h) Number of species (all perennials) classified as 'bumblebee flowers' by virtue of their inclusion in a list of the top 20 species ranked by an index of group-specific selectivity for at least one colour group of bumblehees in Britain ( Fussell and Corbet. 1992a). Species of Crepis. Hierociumand Solidago are included because they qualify at the generic level.

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about the nature of the communities to be expected in vegetation of a given age. If iand has not been ploughed after harvesting an arable crop, a period of set-aside may begin in autumn or winter, not with bare soil, but with a limited cover of amble weeds and crop volunteers. This review focuses on succession from bare soil following ploughing or herbicide treatment after harvest. Disturbance, by the definition of Grime (1979), limits plant biomass by causing its partial or total destruction. The type of disturbance considered here is destruction of the vegetation by ploughing or largescale herbicide treatment. Unlike mowing or grazing, which remove above-ground biomass leaving the turf intact up to soil level, ploughing destroys soil-surface and underground structures as well, by mechanical damage or by altering the depth of burial, and also destroys the soil profile (Edwards, 1984; Davis al., 1992).

2. Vegetation In the first months after an area has been ploughed, its plants and insects are the survivors of cultivation. Plants survive in the soil as seeds in the seed bank or as vegetative fragments in the 'bud bank' (Harper, 1977). Insects survive as resistant stages such as pupae in the soil. Other individuals move in, largely in the seed rain (natural or man-made) or as airborne adult insects in the aerial plankton. These incomers may or may not leave progeny before dying or moving out. Thus cultivation filters the local species pool, destroying some of the species originally present and limiting the initial post.cultivation community to a subset of species that are resistant enough to survive ploughing or mobile enough tc invade soon afterwards. Over succeeding years there will be a gradual accumulation of further species that are less mobile and slower to mature and multiply (see below). The vegetation in year 1 is likely to comprise widespread, mobile species that tolerate a wide range of different soils and microclimates (Grubb, 1987); it is very different from the communities on the same site in later years (Brown and Southwood, 1987). Plants of arable habitats typically regenerate from a seed bank or from vegetative fragments (Hodgson and Grime, 1990), and are short-lived and monocarpic

(Grime, 1979). Important in the first year after cultivation are annual species (Fig. I (a); Hokkanen and Raatikainen, 1977a; Trrmiil~i, 1982; Praeh, 19~5; Gross, 1987; Brown, 1991 ), many of which are small, fast-growing plants with numerous small seeds capable of prolonged survival in a seed bank (Fig. l ( g ) ; Harper, 1977; Grime, 1979; Fenner, 1987; Trrmiil~i, 1982). Annuals with large~ ~eeds are expected to be less mobile ,~unless the seeds are dispersed by wind or by animals) but are more competitive during establishment. In the first year they may be represented by volunteer plants of the recent crop (Davies et al., 1992). Atthough there is a low total resource allocation to reproduction, associated with their small size and the short time available for accumulation of assimilate, ruderal plants may show a high proportional allocation to seeds (Grime, 1979; Brown and Southwood, 1987; but see Symonides, 1988), both between and within species (Stewart and Thompson, 1982; Falinska, 1991 ). They are often self-pollinated (Cruden, 1977; Grime, 1979; Brown and Burdon, 1987; Symonides, 1988) with a low pollen/ovule ratio (Cruden, 1977). The allocation to vegetative reproduction or under ground storage is usually small (Gross, 1987; Falinska, 1991). Also maturing early are clonal perennials from the bud bank, capable of lateral vegetative spread (Leakey, i981 ). Few polycarpic perennials attain reproductive maturity in their first year, and most species flowering in year ! will be annuals or bud bank perennials such as Cirsium arvense. Other perennials, typically originating from the seed rain and having small or winddispersed seeds or achenes, may be present in year 1 only as immature plants. From the second year onwards, as the area of bare ground decreases, annuals will become less important, with the maturation and flowering of herbaceous perennials (Fig. l(a); TOrm~l~i, 1982; Prach, 1985; Brown, 199 !; Davies et al., 1992), the representation of which depends on the proximity of seed sources in established vegetation (Smith and Macdonald, 1919; Wright and Bonser, 1992). Monocarpie perennials flower in the second and subsequent years, and thereafter soon give place to polycarpic perennials (Brown, 1991 ). From year 2 onwards, the vegetative spread of tall, fast-growing polycarpic perennials will create clonal units large enough to accommodate insects requiring a large ecological neighbourhood (Addicott

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et al., 1987); such competitive perennials may support a rich insect fauna (Redfera, 1983; Davis, 1991). In succeeding years the number of perennial plant species may increase (Fig. 1(a)) (Brown, 1991), including larger-seeded forms that are slower to arrive but better able to colonise already-vegetated sites (Leishman and Westoby, 1994). In so far as large seeds are associated with large, nectar-rich flowers (see below), these perenniai~ can supply enough floral reward to repay large insects with high foraging costs associated with endothermy, a capacity for metabolic warming that enables them to forage in cool weather. Plant species of longundisturbed habitats typically lack obvious adaptations for seed dispersal in space or time, but are suspected of forming a bank of persistent juveniles (Hodgson and Grime, 1990). The characteristics that enable plants to colonise naturally regenerating first-year set-aside parallel those of successful weed: in annual arable crops (Baker, 1974), and first-year set-aside is expected to harbour weeds well adapted to transfer into nearby or following-crops via the seed bank, the bud bank, or the seed rain. So long as the sward is not disturbed by repeated ploughing or large-scale herbicide treatment, the second and subsequent years will see the annual weed problem diminish as mobile annual ruderal species are progressively displaced by competitive perennials (Hokkanen and Raatikainen, 1977a; T0rmiilii, 1982 Prach, 1985; Roebuck, 1987; Firbank et al., 1993; Smith et ai., 1993, 1994). Measures taken against annual weeds in year 1 include mowing to prevent seeding, suppression by sowing seed mixes or cover crops (typically largeseeded, competitive annual plants), or destruction by cultivation or herbicides. These destructive treatments, such as ploughing or any but the most sparing, localised and selective herbicide application, may promote resurgence of the annual weed problem in subsequent years, by preventing the development of the perennial sward that would otherwise eventually suppress annual weeds. The high nutrient status of recently cropped land will influence the pattern of regeneration. Species conspicuously increasingor decreasing in abundance in Britain are listed by Grime et at. (1988). In general, the species of frequently disturbed and productive habitats are common and increasing, whereas the rarer and decreasing species are not expected to establish until the fertility has diminished (Fig. l(c)) (Hodgson, 1986a;

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see below). To achieve a floristically diverse sward rich in plant species of~ o~aservationinterest, it is therefore desirable to accelerate nutrientremoval by grazing or by cutting and removing vegetation, if set-aside regulations permit, and to avoid increasing fertility by spray drift. Establishment of this diverse community will take several years at best, and may be impossible without intervention in places where nearby source communities are lacking, or where high fertility favours rank competitive species regarded as weeds (Smith et at., 1993, 1994). Here, long-term natural regeneration would produce a relatively small number of widespread, competitive perennial species. Although these may support a richer community of insects and other wildlife than the annuals of first-year set-aside (see below; Redfern, 1983; Davis, 1991), they are likely to be of little botanical interest. The succession illustrated in Fig. 1 differs from setaside in that preliminary heat sterilisation of the soil probably affected both the initial seed and bud bank and the nutrient status; and in the absence of mowing, shrubs shaded the herbaceous vegetation by year 10. In conventional periodically mown set-aside, seed-bank annuals and bud-bank perennials might be better represented in year 1, and the increase in floristic diversity, including perennial bumblebee flowers (see belong), might be sustained for longer. With these exceptions, Schmidt's findings ~Fig. 1) illustrate features expected in long-term set-aside under natural regeneration: initial dominance of annuals, many of them small-seeded, spring-flowering ruderal forms typical of productive arable habitats, givingplace tt, a communitydominated by perennials, with a progressive increase in floristic diversity and in the proportion of species flowering in summer and visited by 'oul~ble bees.

3. Microclimate

In year 1 of natural legeneration after cultivation, plant cover may be sparse, leaving areas of bare ground subject to extremes of temperature and relative humidity (Stuutjesdijk and Barkmar~, 1992). Many soil-surface insects, notably some predatory carabid beeries, avoid sunlit open ground and would be active only at night unless sheltered by vegetation (Speight and Lawton, 1976). In year I the high temperature of sunlit bare

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soil may make the soil surface niche unavailable to humidity-loving insects, and immobile insects exposed there may be killed by microclimatic extremes. Over the early years the vegetation hlcreases in height, biomass and architectural complexity (Hokkanen and Raatikainen, 1977a; Brown and Southwood, 1987; Brown, 1991 ),carrying the active surface, where exchange of radiation and water vapour leads to microclimatic extremes, upward from the soil surface. Below the canopy is a humid, shaded soil surface habitat that was patchy or absent in year !. Canopy closure can be effected in year ! by sowing a cover crop, which may create an open subcanopy habitat where gamebird chicks can feed without getting wet. Annual cover crops provide a short cut to a habitat that is microclimatically more favourable to insects than bare soil and Jess subject to pesticides than an arable crop, but much less complex than established perennial vegetation. Because annual cover crops involve annually repeated disturbance, they are incompatible with the development of perennial vegetation. A 'conservation headland" is subjected to annual tillage and other crop management operations but reduced pesticide application (Sotherton, 1991 ). Conservation headlands are recommended by the UK Game Conservancy to benefit gamebird chicks and the insects on which these feed. Because conservation headlands receive reduced pesticide input they can help to protect adjacent perennial vegetation from spray drift, but because they involve annual cultivation they cannot replace the perennial communities of Iong-te~'m setaside.

4. Herbivorous and predatory insects There are parallels between insects ana.. plants with respect to colonisation after disturbance. Analogous to small, many-seeded annual plants are r-selected herbivorous insects with high fecundity and a short generation time. Analogous to plants that persist in the seed bank or as buried fragments are insect species that survive cultivation in the soil. Again, similar to plant species with small or plumed seeds in the seed rain are insect species with small, weak-flying windborne adults in the aerial plankton such as aphids and thrips, which have been shown to settle in a crop in a pattern related to the wind shelter of hedges or trees (Lewis,

1970). A category with no obvious plant analc,gy (except perhaps p|ants with animal-dispersed seeds) comprises large-bodied, strong-flying insects with a large foraging range; in both trap-nested bees (Gathmann et al., 1994) and butterflies (Hodgson, 1993), larger species are ibund in Ibe earliest years after disturbance, and smaller species arrive later. Thus the size-frequency distribution of insects in a newly disturbed habitat may be expected to be bimodai, including both very small windborne insects and very large strongly flying species. In the years after disturbance "-,heresident insect community has been shown to increase in density and species richness (Hokkanen and Raatikainen, 1977b; Brown ~nd Southwood, 1987; Brown, 1990). As the period available for colonisation extends, perennial plant species richness and architectural and microclimatic diversity increase, offering a greater range of niches (Brown and Southwood, 1987 ) including those associated with stored reserves in roots and stems, and the large flowers and seeds of perennials (Brown, 1991). The proportional representation of different trophie groups of insects has been shown to change through succession (Brown and Southwood, ! 987 ). In the first year after cultivation the insect community was species-poor, and was dominated by sap-sucking and other phytophages. Many were generalist feeders, mad many were muitivoltine forms with a high rate of increase and high local population density (Brown, 1990,1991; Edwards-.iones and Brown, 1993). In subsequent years the abundance of most pest species decreased (Hokkanen and Raatikainen, 1977b), more speci~ists appeared among the phytophages, and this gaild became relatively less important overall, as parasitoids and predators formed an increasing proportion of the community (Hokkanen and Raatikainen, 1977b; Brown and Southwood, 1987). High parasitoid diversity is associated with floristically diverse vegetation, which may therefore act as a source of natural enemies for pests in adjacent crops (Altieri et al., 1993). Predators such a.s ladybirds and hoverflies, which have very mobile adults and larvae that feed on a concentrated resource, for example colonies of aphids, may build up numbers early. Among hoverfly species with predatory larvae the most abundant are those with a broad diet (Owen and Gilbert, 1989). Many predators are expected to colonise more slowly than these, both

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because of the need for prey populations and perhaps because predators are often larger than their prey, and may therefore have a longer life cycle and slower initial population growth (Sabelis, 1992). Important among polyphagous predators are carabid beetles. A~though some species overwintcr in arable fields, many overwinter in nearby uneuhivatexI areas (Sotberton, 1984) and invade arable fields progressively through the following season (Wratten a~d Thomas, 1990). Some species of carabids rarc!y or never fly. Larger species can run faster than small species (Evans, 1990), and might therefore be expected to invade sooner after cultivation, but most of the species that are good dispersers and characteristic of productive amble habitats are capable of flight, and relatively small (Turin and den Boer, 1988; Den Boer, 1990). Among the carabids of arable fields, small species tend to be dark or metallic, diurnal, with activity peaking in late June and July, and they typically overwinter as adults; whereas the adults of larger species tend to be paler, and to shelter by day and be active at night, peaking in August: these forms typically overwinter as larvae (Kegei, 1990). In sites remote from established vegetation, small diurnal species are expected to arrive first whereas larger, nocturnal species may eolonise more slowly over the years. ~n the first year of natural regeneration after ploughing, the insect community will resemble that of an arable crop in some respects, particularly if volunteer plants are present from a previous crop. It will include mobile, pulyphagous r-selected herbivorous insects some of which may be pests capable of moving into adjacent crops (Hokkanen and Raatikainen, 1977b). Its restricted complement of natural enemies will include some parasitoids and muitivoltine aphid predators, the adults of which may seek nectar from perennials in established vegetation (Van Emden, 1962), but there will be few effective crop pollinators or large, long-cycle polyphagous predators.

5. Pollination and flower-visiting insects Successional changes in plant species composition are accompanied by changes in the predominant pollination syndrome and in the assemblage of flower-visiting insects, and these changes affect the role of

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uncropped habitats as a source of pollinators for adjacent crops. In an array of plant communities in the United States, Ostler and Harper (1978) shelved that nigh floristic diversity correlated with a high percentage of animalpollinated flowers, the diversity of which was correlated with the relative abund~ce of plant species with attributes associated with pollination by long-tongued bees: blue colour, zygomorphic form and restriction of access to the nectary, usualli,: by a tube or spur (Faegri and Van der Pijl, 1979; M~w~z~land Shmida, 1993). It was the communities with. high floristic diversity that harboured the greatest numLers of bee-pollinated flowers. A similar pattern is revealed in Schmidt's succession, which also shows a r~rogressive increase in the proportion of plant species Lhatbegin flowering in summer, when worker bumbl~ bees arc numerous (Fig. l(f)). Parallel successional changes in the assemblage of flower-visiting insects have been shown, In an American did-field succession, Parrish and Bazzaz (1979) found that the flowers of ruderal annuals in the first year were visited by short-tongued generalist insects such as flies and small bees. Some such insects may be crop pests, parasitoids or predators, but they are not large enough ~r mobile enough to be useful pollinators of many entomophilous crops. The perennials that appeared later in succession were visited by larger bees with longer tongues (Parfish and Bazzaz, 1979). Corresponding observations have been made in Britain. In a national survey of the plants visited by bumble bees, these large bees were found to forage preferentially on perennials (especially monocarpic perennials) rather than on annuals (Fussell and Corbet, 1991, 1992a; Saville, 1993). Honey bees (Saville, 1993) and butterflies (Feber, 1993; Feber et al., 1994; Smith et al., 1994) also visited perennials preferentially over annuals. With respect to their value to bees, the critical difference between ruderal annuals and perennials is probably related to the size of individual flowers and the amount of nectar they secrete. The amount of nectar secreted will depend on the total amount of assimilate available in the plant at the time of flowering; the proportion of that resource that is allocated to sexual reproduction; and the number of individual flowers among which that allocation to sexual reproduction is shared.

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The total amount of assimilate allocated to reproduction may be greater in perennials than in ruderal annuals because of the longer period that has been available for the accumulation of assimilate and their generally larger sL'e (Brown, 1991), and perhaps because of their abiiity to store and reallocate assimilate, associated with the capacity for vegetative spread and storage overwinter (Fussell and C:Jrbet, 1992a). Even if polycarpic perennials allocate a smaller proportion of this total assimilate pool to sexual reproduction as opposed to vegetative growth than ruderal , annuals do ( see above), the allocation to sexual repro~ duction per unit area of ground may be greater in perennials, because plant size and photosynthetic area are so much greater. The proportional a~l~ation is expected to be particularly high in monocarpic peren::ials, because these commit all available assimilate to a single episode of flowering, whereas polycarpic per,mnials share it between flowering, vegetative reproduction and reserves for overwintering. Evolutionary trade-offs between size and number have been considered in relation to both flowers and seeds (Primack, 1987; Lloyd, 1989). An individual flower is likely to receive a larger share of the assimilate available for reproduction in a species with few large flowers than in a species with many small ones (Harder and Cruzan, 1990). Competitive perennials generally have larger seeds than ruderal annuals (Harper, 1977). and they probably have larger flowers too. When analysed by phylogenetic regression, seed (or achene) size correlated positively with an index of corolla depth (style length measured to the internal base of the flower), in interspecifie comparisons both within the Compositae (Kirk, 1993) and in a larger set of 208 native species from the Warwickshire flora for which seed weight categories are given by Grime et al. (1988) (W.D.J. Kirk. personal communication, 1993). Larger flowers have been found to secrete more nectar than small flowers, in interspecific comparisons within the flora of Israel (Menzel and Shmida, 1993) and among Ericaeeae (Harder and Cruzan, 1990). Although most of the d e e r nectar-rich flowers visited by large bees are perennials, there is a small but interesting category of special annuals that sometimes receive visits from large bees. It includes a group of hemiparasites (species of Rhinanthus, Melampyrum (Kwak, 1979) and Odontites (Saville, 1993)), in which assimilate drawn from the host may increase the

total resource available for allocation to reproduction (Press et al., 1991); a very few ruderals with smai! seeds ( Lamium purpureum, Viola arvensis and Kickxia spuria (Saville, 1993) ); and a ,.-roup of large-seeded forms, some with spurs (lmpat~'.~r,s glandulifera, Consolida sp.) and some with nectar concealed in a deep tube or pocket (Phacelia tanacetifolia, Borago officinalis, Viciafaba (Fussell and Corbet, 1992a), and Galeopsis tetrahit (Saville, 1993)) in which large seeds are associated with large flowers (Primack, i 987 ), giving the potential for a large per-flower nectar resource allocation (Harder and Cruzan, 1990). Many annual entomophiious crops are of this type. With these few exceptions, annuals do not help to sustain populations of the larger pollinators. Mature vegetation with a high diversity of broad-leaved perennial species is important to provide the seasonal succession of forage on which bumblebee colony development depends. Fig. l ( h ) illustrates the incidence in Schmidt's (1976) succession of flowers much visited by bumble bees in a national survey (Fussell and Corbet, 1992a). All were perennials. The soil treatment that destroyed the seed and bud banks may have restricted bumblebee flowers to those available in the seed rain. All but one of these bumblebee flower species were composites and willowherbs with plumed seeds. Cereals are wind pollinated, and for so long as cereal crops have dominated arable farming in Britain insect pollinators have received little sustenance from flowering crops and little attention from farmers or policymakers. There is no mention of pollinators in the Explanatory Guide for Arable Area Payments (Ministry of Agriculture, Fisheries and Food, 1993) or in the Guide to Alternative Combinable Crops (Royal Agricultural Society of England/Agricultural Development and Advisory Service, 1985). For many non-cereal crops, bees are valuable pollinators because each worker forages for the colony as well as for herself, and flower visits are numerous and rapid. Among bees, social species have an advantage over solitary species because of their broad diet and their long foraging season (Corbet et al., 1991); large bees have an advantage over small bees because their greater thermoregulatory ability (Stone and Willmer, 1989) enables them to forage at lower ambient temperatures. Among social bees in Britain, all of which are large, bumble bees are often more effective than honey bees because they forage faster (Free, 1993)

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and at lower temperatures (Corbet et al., 1993), and because behavioural and morphological features enable them to forage profitably on some flowers unprofitable for honey bees. Long-tongued bumblebee species such ~ts Bombus hortorum and Bombus pascuorum are particularly valuable because they can take nectar (legally, from the front of the flower, allowing pollen transfer) from deep-flowered crops like field bean and red clover. Bumble bees have sometimes been shown to transfer more pollen per visit than honey bees, whose contribution to pollen transfer is sometimes small (Westerkamp, 1991; Wilson and Thomson, 199 ! ). Although bumble bees are known to play an important role in the pollination of some entomophilous crops (such as red clover and field bean; Corbet et al., 1991) and wild flowers (Kwak et al., 1991 ), their role remains unquantiffed for most crop and wild flower species. A reduction in the numbers of these pollinators is expected ~o reduce seed set in some self-incompatible crops and wild flowers. Some leguminous and oilseed crops have large seeds and large, nectar-rich, deep flowers, in contrast to ruderal annual weeds which typically have many small seeds and small, nectar-poor, open flowers visited by a variety of small, short-tongued flies, beetles and hymenopterans (the 'insect rift-raft' of Allen ( 1891 ) ). The insects that visit these entomophiious crops are typically large endothermic species with ,l large energy requirement and a long tongue, and as complementary forage sources when the crop is not flowering they need a seasonal succession of deep, nectar-rich flowers comparable in the amount and accessibility of the reward with those of the crops for which pollinators are needed. This can be provided by species-rich perennial herbaceous vegetation. The wild pollinator community is potentially important to agriculture and conservation, and its erosion may bring problems that cannot be solved by bringing in honeybee colonies. Several species of bumble bee have already been lost from arable regions of Britain; the species lost are those associated elsewhere with the unploughed vegetation of old meadows, saltmarshes, sand dunes and shingle (Williams, 1986). In east central England further losses, particularly of the few remaining long-tongued species, could reduce floristie diversity and restrict future agricultural options by depriving certain wild flowers and potential crops of

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effective pollinators. Such pollinators have been described as keystone species (Kevan, 1991) in tha~. their presence is crucial in maintaining the integrity of communities. Keystone species are considered further below. The repeated cultivation requiJed by an annual crop prevents the development of a diverse perennial sward, but if annual cover crops are to be sown to suppress weeds they might be selected to contribute to bee forage (Corbet et at., 1994). Large-seeded annuals with !arge, deep, nectar-rich flowers, such as Phacelia tanacetifolia and borage, iJorago officinalis (Williams and Christian, 1991; Fussell and Corbet, 1992a; Saville, 1993; J.L. Osborne, personal communication, 1994) provide good bee forage. The 'Tiibinger Mischung' is a seed mixture designed to provide forage for t datively short-tongued wild bees (Engels et al., 1994). Some perennial legumes, such as red clover (Trifolium pratense ) , white clover ( Trifolium repens ~, sainfoin (Onobrychis viciifolia) and birdsfoot trefoil (Lotus cornfculatus), are valuable for bumble bees (Fussell and Corbet, 19~.2a; Saville, 1993). Set-_aside regulations may limit the sowing of legumes, alone or in seed mixtures, but the importance of legumes for wild bees should not be disregarded. Rasmont (1988) attributed losses of several long-tongued bumblebee species from areas of Belgium and France to a decrease in the area of leguminoes crops, partly attributable to the shift from horses to tractors. It may be appropriate to by-pass the slow colonisation phase and accelerate the development of a diverse perennial sward by sowing a wildflower seed mixture, if there is no suitable semi-natural vegetation nearby to act as a source (Smith and Macdonald, 1989; Smith et al., 1993, 1994). If it is to improve forage for pollinating bumble bees, the seed mix should not be too heavily dominated by grasses and should include a seasonal spread of large-flowered broad-leaved species flowering ttom early spring to late summer. Seed threshed from a meadow in late summer may lack the earlyflowering species so important for bumblebee colony establishment. To avoid the risk of contaminating seminatural vegetation with alien genotypes, Akeroyd (1994) has emphasised the need to use authentic native seed. The larger bees may be unable to forage ia first-year set-aside areas because of the absence of reward-rich flowers, but their foraging r~,~ge is large in relation to

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the scale of set-aside, and they can fly into these areas from adjacent habitats. Gathmann et at. (1994) found that it was the larger species of solitary bees, with their greater migratory ability, that were the first to colonise wooden trap nests in newly ploughed habitats. However, in the absence of trap nests, wild bees are unlikely to establish nests in first-year set-aside, where soil is likely to be the only nesting place available. Because of fidelity to established nesting places, most soil-nesting solitary bees are expected to nest in habitats free from large-scale disturbance (although small-scale rabbit-dil~ging~ provide the sunlit bare ~oil that some species need (Falk, 1991; Wesserling and Tscharntke, 1994)). Bumble bees, too, rarely nest in newly disturbed land. Instead, they often occupy the disused nests of small mammals, or nest among tussocky grass in established vegetation where moss is available (Fussell and Corbet, 1992b). Without special management. first-year set-aside is unlikely to provide either nesting places or forage for bees. Banaszak (1992) suggested that uncultivated refuges should occupy at least a quarter of the agricultural landscape if they are to support enough bees to perform an adequate pollination service. Long-term unploughed set-aside could contribute to these refuges during the successional time-window between replacement of small-flowered ruderal annuals by nectar-rich perennials and shading by shrubs and trees. Occasional mowing or grazing of naturally regenerating set-aside to prevent scrub invasion benefits bumble bees, because they usually forage in open sunlit habitats; they may nest under the forest canopy, but they do not forage there (Saviile, 1993). Because seed production in many crops and perennial wild flawer~ 1. . . ~. ,. ... . . . f::,. .. .. .a. . .h,,mhle . . . . .do,,oadc r . . . . . .on . . . .... bees, the distribution of which is already declining, management to favour these insects is expected to bring ecological benefits (or to avert losses) that go beyond the conservation of the species for their own sake.

6. Commonness, rarity and diversity With increasing rates of soil disturbance in Britain, productive and disturbed habitats have spread rapidly (Ratclfffe, 1984), and habitats free from disturbance or eutrophication have become rarer and are decreasing. The flora has become less diverse and is now

largely dominated by rodemls of productive, newly disturbed habitats (Hodgson, 1986b). In general, except for a high proportion of rare annuals in arable habitats, rare and decreasing plant species are associated with established vegetation in unproductive habitats, whereas common and increasing species are associated with productive, disturbed habitats (Hodgson, 1986a). Analysis ofSchmidt's (1976) data (Figs. 1(a) and 1(c)) showed that among perennials both species number and the proportion of decreasing species increased year by year. Interestingly, annuals showed at, inverse pattern; species number and the proportion of decreasing species fell year by year. Over ~he first 10 years after disturbance, then, the diversity and botanical interest of a mown sward is expecte~ to increase with resl~ct to perennials, but decrease with respect to annuals. The overall pattern of change may depend on the balance between annuals and perennials; species richness may increase steadily with time (Hokkanen and Raatikainen, 1977a; Brown, 1991 ), or show a bimodal pattern with an early peak when the ruderals overlap with the early perennial colonists (Bazzaz, 1975; Prach, 1985; Osbornova et al., 1990; T. Tscharntke, personal communication, 1993). Among butterflies, the commoner species use larval food plant specie5 ofmderal communities; rarer species are associated with food plants of Iong-unploughed vegetation in habitats of low fertility (Hodgson, 1993). The ruderal annual plants of first-year set-aside do not support the rarer butterfly species, whose larval food plants grow in Iong-unploughed habitats. The only butterfly species likely to benefit from first-yeur set-aside are already very common, and some are pests. A microclimatic requirement may restrict some rarer butterfly species to close-grazed, sun-warmed turf (Thomas, 1991 ), but this may be centuries old; the rudeml annuals that follow cultivation of the soil are not the larval food plants of rare butterflies. The plant species that adult butterflies visit for nectar are not necessarily those used as larval food plants, but, as in the case of larval food plants, many are characteristic of established vegetation. Like bumble bees, butterflies make more flower visits to perennials than to annuals in relation to the relative abundance of annuals and perennials in the local flora (Feber, 1993; Feber et at., 1994). The positive effect on butterfly numbers claimed for conservation headlands probably results largely from protection of perennials in the adjacent

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uncultivated field margin from spray damage. The annuals and bud-bank perennials that flower on annually cultivated conservation headlands include few species that are butterfly nectaring flowers, and few species of butterfly visit them. Of the flower visits reported from conservation headlands, over 95% of those made by three common pierid species were to charlock (Sinapis arvensis, an annual) and creeping thistle (Cirsium arvense, a bud-bank perennial), and over 90% of visits by two common satyrids were to creeping thistle and scentless mayweed (Tripleurospermum inodorum, an annual) (Dover, 1989). Other insects are less well documented at the species level, but the species richiiess of several othe_r insect groups has been found to increase after the first year of succession (Sou'hwood et al., 1979; T6rm~il~i, 1982). Established vegetation is therefore of greater conservation interest than newly disturbed land with respect to butterflies, bumble bees and other insects as well as perennial plants, being more likely to su~port both breeding populations and foraging individuals of the rarer species. Colonising plant species tolerate a wide range of conditions with respect to nutrient and water status, whereas species that arrive later have more specific requirements (Bazzaz, 1987; Grubb, 1987). There is evidence for progressive narrowing of uiches through succession for the host ranges of sap-feeding insects (Brown and Southwood, 1987; Edwards-Jones and Brown, 1993), and for pollination; both the flower pollination syndrome and the insect visitors became more specialised (Parrish and Bazzaz, 1979). For this reason, mature communities should be more characteristic of a region or soil type than young communities. The first species to establish in an uncontested site will be wide-ranging, tolerant species sharing only their ability to colonise effectively through space or time. At this stage site-to-site differences depend largely on site history (Hokkanen and Raatikainen, 1977a). As the numbers of individuals and species increase, dispersibility will become less important, competition will favour species well suited to local conditions, and substrate differences will be increasingly reflected in the species composition of plant (and therefore insect) communities. Site-to-site differences now depend largely on habitat type (Hokkenen and Raatikalnen, 1977a). Diversity will increase at the community level (Bakker, 1989). lnouye et al. (1987) showed that different sites

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of the same successional age were floristically similar just after disturbance but diverged in the following years. Thus increased successional age brings an increase in community diversity between sites as well as an increase in species diversity within sites.

7~ Evaluation One measure of the ecological impo.,'tanee of a species in a community is the extent to which other species depend on it. Although the term 'keystone species' lacks a clear operational definition, and it is hard to demonstrate the crucial role of individual wide-ranging species experimentally without eliminating them over large areas (Mills et al., 1993), policy makers needle distinguish between pivotal species of this kind, whose loss has far-reaching ecological repercussions, and rare species with weaker ecological interactions. The ecologic-al importance of a species depends on the frequency, as well as the nature, of its interspecific interactions. Common species generally participate in inL:erspecific interactions that are numerous (because tht~ species are abundant) and involve many species (because abundant host plants generally support more insect herbivore species than rare ones (see Strong et al., 1984) ). As a species becomes rarer its ecological importance declines, but the interest it attracts from conservationists increases, often peaking when the species approaches the verge of extinction. Rare species sometimes attract more funding than decreasing common species of much greater ecological value. Although many ecologists appreciate the extent to which the continued functioning of semi-natural ecosystems depel'ids on keeping common species common, public perception is based on different, simpler criteria. Diversity and rarity value are not always the best criteria for habitat quality. Certain communities meriting conservation are characteristic and irreplaceable but not species rich, and would be degraded by Iocalised disturbance even though this would increase their diversity. Equally, management designed to postpone the extinction of rare species is sometimes less important and often less cost-effective than management to ensure that common species remain common and keystone species remain functional. Ambivalent attitudes towards introduction of rare species show that the case

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for preserving rarities goes beyond the practical effort to maintain the interactions on which overall ecosystem dynamics depends. An analogy might be made with coins. If the minting of fresh coins were to cease, and casual losses eroded stocks of current and obsolete coins Mike, some people would focus on protecting the coins still in current use. Others would elect to protect the few remaining obsolete coins because of their antiquarian value. The prospect of forging new ones would bring them no more consolation than gardening with locally extinc~ plant species brings to botanical purists. Evaluation of communities is often based on plants, because they are easier to identify than insects. Decisions about the future fate of long-term set-aside are iikely to be based on evaluation after the first 4 or 5 years. At this stage perennials are likely to dominate, but few rare and decreasing plant species will have established (Figs. l(a) and I ( c ) ) . The decision whether to allow succession to continue should depend on the present and future ecological value of the community, not only on the rarity value of the individual species. There is a temptation to disregard ecological value because it is more difficult to quantify than rarity value. In another context, a self-monitoring procedure developed for ditch-bank plants by Kruk et al. (1994) was based on indicator species selected not because they are rare, but because they are conspicuous, easily recognised and associated with species-rich vegetation. Plants were the basis of this monitoring procedure, which served the dual functions of stimulating the farmers' interest and limiting the requirement for site visits to confirmation of farmers" positive records. Evaluation ,.~r~the basis of overall insect diversity requires entomological expertise (Disney, 1986 ). A ~elf-moni~.oring procedure similar to Kruk's might be designed for insects, using indicator species selected because they are conspicuous, easily recognised and associated with species-rich faunas and belonging to relevant groups such as bumble bees, butterflies and carabid beetles, for example (Forsythe, 1989; PrOs-Jones and Corbet, 1991; Thomas and Lewington, 1991 ).

8. Management: pl,~ughing and policy The conserv~tion potential of set-aside depends on the frequency o|" disturbance at a field scale and at a

landscape scale. In an individual field, several years of natural regeneration without ploughing will bring an increase in the floristic diversity of perennials, and often also in overall plant and insect species richness (Fig. I (a)), in the botanical and entomological interest of the community, and in the availability of nesting places and forage for social bees and nectaring flowers for butterflies. Established vegetation may also support a diverse flora and fauna including some species that are rare and decreasing in Britain. Although the community may take decades to reach its full value, longterm set-aside can go some way to meet the conservation objective of arresting or reversing the decline in decreasing plant and insect species, whether ot not these have already achieved the status of rarity, and it can give a substantial start for a longer-term programme of restoration of semi-natural communities (Firbank et al., 1993). At the landscape scale, habitat diversity will increase with time since disturbance in set-aside, not only because of increased representation of habitat-specific species reflecting the characteristics of individual sites, but also because in many areas newly disturbed habitats are well represented elsewhere in the landscape, and are increasing at the expense of established vegetation ( Barr et al., 1993 ). Established vegetation will also act as a source of propagules, helping maintain the laudscape's capacity to recover from disturbance by accelerating succession in newly disturbed patches nearby, and alsc as a crop support system, supplying adjacent crops with pollinators and natural enemies of pests. In so far as management costs diminish and benefits to consep/ation and agriculture increase progressively over the years, short-term studies may give a misleadingly pessimistic impression of the cost-effectiveness of long-term natural regeneration. A study lasting only 3 years would reveal the early weed management problems but not the long-term benefits, and might not provide strong enough evidence to persuade farmers or their advisers of the value of maintaining natural regeneration through and beyond the first three relatively unrewarding years. That value is emphasised by Smith et al. (1993). Ancient permanent grassland in Britain was once highly valued for grazing and hay-making, and covenants forbidding the ploughing of ancient grassland were common. When artificial fertilisers and improvements in land drainage and seed mixtures enabled high

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productivities to be achieved on short-term leys, ancient grassland lost its agricultural primacy (Duffey et al., 1974). When the Second World War necessitated increased agricultural production, covenants protecting ancient grassland were annalled, and efforts were made to bring more land under the plough. Sir George Stapledon did much to dispel the farmers' 'reverential esteem' for 'sacrosanct pastures', encouraging them to plough established grassland (Stapledon, 1935, Waller, 1962). Even if this was merely an acceleration of a process already in train (Chapman and Sbeaii, 1994), its effects on the agricultural landscape were profound, leaving a permanent mark on the vegetational history of Britain. More recently, pressure to increase productivity has been replaced by pressure to decrease it, and today's aim is often to ntaximise net profit rather than production. Technological improvements have removed the need to keep land under the plough in case a rapid increase in food production is needed. Several wild pollinators have become extinct locally (Williams, 1986). Long-established semi-natural vegetation has become rarer, and harder to replace; sources for recolonisation are so sparse that when ancient vegetation has been ploughed, even without further disturbance it may take centuries, or be impossible, to restore the soil profile and the complement of uncommon plants, traditional nesting places for bees, and the uncharted diversity of insect herbivores, pollinators and predators. Most of the remaining refuges in arable regions are small, unmapped fragments of roadside, field edge or hedgerow, perhaps only 5 or 50 years old, easily destroyed by a bulldozer or a plough. Any policy that results, explicitly or accidentally, in destruction of long-unploughed vegetation will weaken the already tenuous network of uncultivated land that harhours the relics of Britai:~'s native flora and fauna and supplies crops with natural enemies and pollinators. Both agriculture and conservation will be impoverished if these communities are lost. Set-aside policy has been responsible for ~uch destruction in the past (Welch, 1990).

9. Conclusion

This review shows that if plato and qnimal commun;.ties are allowed to recover for several years after soil

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disturbance, their structure generally changes progressively in ways that improve their impact on nearby crops and semi-natural communities. In the years after cessation of ploughing, rnderal annual plants are replaced successively by monocarpic and then polycarpic perennials. There are increases in plant stature, microclimatie and architectural diversity, and in the representation of species with large, nectar-rich flowers and large seeds, and of rare and decreasing species of perennials. The plant community becomes more characteristic of the local site, contributing to diversity of the landscape. An insect community of small, mobile, polyphagous herbivores, including some pests, will be replaced by a richer and more diverse assemblage with more specialised herbivores, more predators, and larger, more specialised pollinators. First-year set-aside may harbour weeds and pests, but over the years natural regeneration, perhaps supplemented by introduction of selected wild plant species, may produce a species-rich community of botanical and entomological interest, supplying natural enemies and pollinators to adjacent crops and uncultivated land. A desirable objective for the future would be to preserve and extend established perennial vegetation, and its associated animal communities. A case has been made for ?,hort-term set-aside as a management tool encouraging particular species in particular circumstances (Firbank et al., 1993). Long-term set-aside should be seen not simply as ~ default option, but as an option that can produce communities of more general ecological value, as well as rarity value, and can help to sustain ecologically important species that are not only of conservation interest, but are essential for the effective functioning of agricultural ecosystems.

Acknowledgements For comments and suggestions that have greatly improved drafts of this paper the author thanks Eric Allen, Philip Corbet, Gary Fry, John Hodgson, William Kirk, Juliet Osborne, Derek Ratcliffe, Teja Tscharntke, Catherine Williams and an anonymous referee. The author is very grateful to John Hodgson and the NERC Unit of Comparative Ecology, University of Sheffield, for the analyses shown in Fig~. 1 ( a ) - I (g), and to William Kirk for examining the correlation between style length and seed weight.

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