The relationship between local extinctions of grassland butterflies and increased soil nitrogen levels

Transcription

1 BIOLOGICAL CONSERVATION 128 (2006) available at journal homepage: The relationship between local extinctions of grassland butterflies and increased soil nitrogen levels Erik Öckinger a, *, Olle Hammarstedt b, Sven G. Nilsson a, Henrik G. Smith a a Department of Ecology, Animal Ecology, Lund University, SE Lund, Sweden b Svärmarevägen 28, SE Södra Sandby, Sweden ARTICLE INFO ABSTRACT Article history: Received 25 April 2005 Received in revised form 6 October 2005 Accepted 14 October 2005 Available online 1 December 2005 Keywords: Eutrophication Lepidoptera Turnover Vegetation succesion Extinction Colonization In contrast to many other western European countries, the knowledge about trends in the Swedish butterfly fauna is poor. We studied the effects of habitat and species characteristics on species turnover in 13 grassland sites in southern Sweden by comparing species richness and compositions at two separate occasions with an interval of 21 years. The mean number of species per site decreased from 30 to 24, with a large variation between sites. The number of extinctions was highest in sites where the proportion of trees and shrubs had increased most, but there was no detectable effect of area or of the composition of landscapes surrounding the sites. Areas protected as nature reserves had lost as many species as unprotected areas had, indicating both the importance of proper management of nature reserves and that nature reserves alone may not be enough to inhibit regional extinction of butterfly species. Species dependent on nutrient-poor conditions tended to decrease while species dependent on nutrient-rich conditions tended to increase, indicating a negative effect of increased soil nitrogen levels resulting from active fertilizing of pastures and/or atmospheric nitrogen deposition. A regular monitoring program could show whether our results are representative for Sweden or Northern Europe. Ó 2005 Elsevier Ltd. All rights reserved. 1. Introduction Butterflies have been shown to respond rapidly to changes in their environment, compared to other taxa, such as birds or vascular plants (Erhardt and Thomas, 1991; Thomas et al., 2004), and are thus suitable indicator organisms for monitoring. Since they can easily be monitored (Pollard and Yates, 1993), they are suitable indicator organisms for the state of the environment. The presence of certain butterfly species or a high butterfly species richness may also act as an indicator of a high species richness of other groups (New, 1997; Kerr et al., 2000). In several European countries (e.g., Britain: Pollard and Yates, 1993; The Netherlands: van Swaay et al., 1997; Finland: Saarinen et al., 2003), there exist regular monitoring programs for butterflies. Such programs provide valuable information on population trends (Pollard et al., 1995; van Swaay et al., 1997; Saarinen et al., 2003) and population fluctuations (e.g., Sutcliffe et al., 1996). In addition, systematic mapping of species distributions on national grid systems can be used to monitor large-scale population and distributional trends (Stoltze, 1996; Huldén, 2000; Asher et al., 2001). Although most such monitoring and mapping programs are young, they have revealed significant losses of butterfly species from several countries in western Europe, including Britain (Asher et al., 2001), Belgium (Maes and van Dyck, 2001) and The Netherlands (van Swaay, 1990; van Swaay et al., 1997) during the last decades, while a few species have increased in abundance or expanded their ranges (van Swaay, 1990; Asher et al., 2001; Saarinen et al., 2003). In Sweden, however, no such regular monitoring scheme or mapping exists, and the knowledge * Corresponding author: Tel.: address: erik.ockinger@zooekol.lu.se (E. Öckinger) /$ - see front matter Ó 2005 Elsevier Ltd. All rights reserved. doi: /j.biocon

2 BIOLOGICAL CONSERVATION 128 (2006) about quantitative changes in the butterfly fauna is therefore relatively poor. Environmental changes on different spatial scales may cause changes in species composition, local extinctions and range expansions of species. At a local scale, factors such as changed grazing regimes and increased soil nutrients as a result of fertilizing are likely to change the plant community (Smith et al., 2000), with cascading effects on higher trophic levels (e.g., Pywell et al., 2004). Mono- or oligophagous herbivores, such as the larvae of most butterfly species, are strongly dependent on the local abundance of host plants (Warren, 1992). Nectar-feeding insects, such as the adults of many butterflies, usually occur at higher densities where there is a high availability of nectar-rich flowers (Warren, 1992). Changes in vegetation height and the cover of trees and shrubs may alter the micro-climate, which is also likely to be a limiting factor for some species (Thomas, 1993). At a regional scale, the extent and spatial configuration of suitable habitat affect the probability of population persistence for species dependent on metapopulation dynamics (Hanski, 1999), and changes in the quality of the matrix may affect source-sink dynamics (Dias, 1996) or create ecological traps (Battin, 2004) which may drive populations to extinction. Often, both local and regional factors and their interactions may be crucial for local population persistence (Thomas et al., 2001). Global climate change may cause expansions or contractions of species geographical ranges (Parmesan et al., 1999). Effects of climate change may also be detected at regional or local scales, for example as altitudinal shifts (Konvicka et al., 2003) or changes in survival rates due to changed micro-climate, which is an important aspect of habitat quality for many insect species, especially at the extremes of their range (Bourn and Thomas, 2002). Similarly, global or continent-wide processes of air pollution may affect local habitat quality for butterflies. Atmospheric deposition of N and S changes the nutrient status and the acidity of the soil, respectively, which in turn are likely to change the composition of the plant community (Burke et al., 1998). Species must be assumed to respond differently to environmental changes (e.g., Reynolds, 2003; Kotiaho et al., 2005). Habitat specialization (Krauss et al., 2003; Kotiaho et al., 2005), mobility (Thomas, 2000; Maes and van Dyck, 2001; Kotiaho et al., 2005), geographical distribution of larval host plants (Kotiaho et al., 2005), nutrient requirements of larval food plants (Oostermeijer and van Swaay, 1998) and nutrient status of the breeding habitat (Maes and van Dyck, 2001), flight period (Kotiaho et al., 2005) and general life history strategies (Hodgson, 1993; Dennis et al., 2004) have all been shown to influence the status and/or population trend of butterfly species. Moreover, populations within a species may be differently affected depending on the distance to geographical range limit, since populations near the range margin can be assumed to have narrower habitat requirements, especially in terms of climatic conditions (Thomas et al., 1999). The aim of this study was to investigate changes in the butterfly fauna of southern Sweden during the last 20 years by re-surveying 13 sites, and to relate the possible changes to habitat and species characteristics. We predicted the loss of species to be greater in small and isolated sites, and in sites where habitat quality has decreased, compared to large, well-connected sites where habitat quality has not changed. We also predicted rare, specialized and less mobile species and species dependent on nutrient poor conditions to have been more negatively affected than common, mobile generalist species dependent on nutrient-rich conditions. Finally, we predicted that species with a southern distribution, which meet their northern range margin in southern Scandinavia, have increased as a result of a warmer climate. 2. Methods 2.1. Study area We studied butterfly species richness in 13 sites in southern Sweden (Fig. 1). They vary in size from 9 to 49 ha (Table 1). All of them are dominated by semi-natural grassland, with varying covers of trees and shrubs. All sites were grazed during both survey periods, most of them by cattle with one exception; Tromtö was grazed by sheep in but by cattle in Butterfly survey Thirteen sites in southern Sweden were surveyed for butterflies (Rhopalocera) and burnet moths (Zygaenidae) in (Hammarstedt, 1996 and unpublished), and re-surveyed in Burnet moths are mainly diurnal and are similar to butterflies in most ecological aspects (Naumann et al., 1999). Unless otherwise stated, butterflies refer to both groups. At each site, butterflies were surveyed using a modified version of Pollard s standardized transect counts (Pollard and Yates, 1993). In the field, the observers counted all butterflies within 5 m ahead and at both sides along the transect. Field work was only carried out in good weather, i.e., at least 17 C in sunny conditions and 20 C in cloudy weather with no rain and no strong wind. At all sites, the same methods and the same transects were used in both periods. During the first period, most of the sampling was done in 1981, with some preliminary sampling in 1980 and some additional sampling in During the second period, most of the sampling was done in 2002, with some complementary visits in 2003 in order to cover the same phenological periods (mid May to the end of August) as during the first survey. Each site was visited between 6 and 13 times during the first period and between 7 and 10 times during the second period. For convenience, we will refer to the first period as 1981 and the second period as Since the two species Plebejus idas and P. argus were not always possible to separate in the field, they are grouped as one species in the analyses. The nomenclature follows Karsholt and Razowski (1996) Classification of butterflies Butterfly (and burnet) species were classified according to several ecological and life history traits; mobility, host plant

3 566 BIOLOGICAL CONSERVATION 128 (2006) Fig. 1 Position of the studied sites in southern Sweden. The position of the study region is shown on the small overview map in the top right corner. Table 1 Site characteristics, species richness and diversity in the two periods Site Reserve Area Landsc 1 Landsc 3 Grazing Open N 1981 N D D 2002 Extinctions Colonizations Bellinga Gummarp Humlarödshus Kronovall Oxhagen Skoghuset Skoghejdan Stenshuvud Svabesholm Taxås Tromtö Vambåsa Önneslöv Reserve refers to whether a site is protected as a nature reserve or not. Surveyed areas of semi-natural grasslands are given in hectares. Landsc 1 and Landsc 3 are the proportions of semi-natural grasslands within 1 and 3 km, respectively. Grazing is the change in rank order of grazing intensity, where high (positive) values indicate increased grazing intensity and low (negative) numbers indicate decreased grazing intensity. Open is the percentage change in open area (not covered by trees or shrubs). N 1981 and N 2002 are the numbers of species in the respective period and 1-D 1981 and 1-D 2002 are 1-Simpson s diversity index for the respective period. We present 1-D because the value of D is higher when the diversity is low, and lower when the diversity is high. specialization, northern and southern range limits, voltinism, nutrient associations of larval host plants, threat status and family. Mobility. We applied the classification of species as sedentary, intermediate or mobile according to Pollard and Yates (1993). This classification has been used successfully in previous studies (e.g., Thomas, 2000). Species occurring in our study area but not classified by Pollard and Yates (1993) were fit into the three categories according to data from Naumann et al. (1999) for Zygaenidae and Henriksen and Kreutzer (1982) and Stoltze (1996) for butterflies (Table 2). Host plant specialization. Based on data in Stoltze (1996) for butterflies and Naumann et al. (1999) for burnets we classified the host plant specificity of each species as 1 = dependent on one or a few host plant species, of which none is common; 2 = may use several species within one genus, where at least one species is widespread; 3 = may use host plant species from several genera.

4 BIOLOGICAL CONSERVATION 128 (2006) Table 2 Mobility classification of all recorded species, the number of sites (maximum 13) in which each species was found in the two periods, and the number of extinction and colonization events for each species Species Mobility Number of sites Extinctions Colonizations Adscita statices Sedentary Zygeana viciae Sedentary Zygeana filipendulae Sedentary Zygeana lonicerae Sedentary Erynnis tages Sedentary Pyrgus malvae Sedentary Thymelicus lineola Sedentary Hesperia comma Intermediate Ochlodes venata Sedentary Papilio machaon Intermediate Leptidea sinapis Intermediate Aporia crataegi Mobile Pieris brassicae Mobile Pieris rapae Mobile Pieris napi Intermediate Anthocaris cardamines Intermediate Colias hyale Mobile Gonopteryx rhamni Mobile Nymphalis antiopa Mobile Inachis io Mobile Vanessa atalanta Mobile Vanessa cardui Mobile Aglais urticae Mobile Polygonia c-album Intermediate Araschnia levana Mobile Argynnis paphia Intermediate Argynnis aglaja Intermediate Argynnis adippe Intermediate Argynnis niobe Intermediate Issoria lathonia Mobile Brenthis ino Sedentary Clossiana selene Sedentary Clossiana euphrosyne Sedentary Mellitaea cinxia Intermediate Mellitaea diamina Sedentary Mellitaea athalia Sedentary Hipparchia semele Sedentary Maniola jurtina Sedentary Aphantopus hyperantus Sedentary Coenonympha pamphilus Sedentary Coenonympha arcania Sedentary Pararge aegeria Sedentary Lasiommata megera Mobile Lasiommata maera Mobile Thecla betulae Intermediate Quercusia quercus Sedentary Satyrium w-album Sedentary Satyrium pruni Sedentary Satyrium ilicis Intermediate Callophrys rubi Sedentary Lycaena phlaeas Intermediate Lyceana virgaureae Intermediate Lycaena hippothoe Intermediate Celastrina agriolus Mobile Maculinea arion Sedentary Cyaniris semiargus Intermediate Polyommatus icarus Intermediate Polyommatus amanda Intermediate Aricia agestis Intermediate Vacciniina optilete Sedentary Plebejus idas Sedentary

5 568 BIOLOGICAL CONSERVATION 128 (2006) Geographical range. The position (in degrees) of the northern and southern range limits in western Europe, respectively, for each species were obtained from Kudrna (2002). Voltinism. The number of generations per year were taken from Stoltze (1996) for butterflies and Naumann et al. (1999) for burnets. Nutrient optimum of larval host plants. In order to detect if butterfly species associated with different soil nutrient conditions we applied the Ellenberg N indicator value (Ellenberg et al., 1991) of each butterfly species host plants. Based on the N indicator values, we classified the nutrient associations for each butterfly species as high (N-value >5) or low (N-value 65). For species with more than one possible host plant species, the mean of the respective N-values were used. Data on host plant use were recoded from Stoltze (1996) for butterflies and Naumann et al. (1999) for burnets. Family. In order to detect possible differences between taxonomic groups we included family (Hesperidae, Papilionidae, Pieridae, Lycaenidae, and the subfamilies Nymphalinae and Satyrinae) as a fixed factor. Satyrinae was separated from the rest of the Nymphalidae since it is a well-defined taxon, and ecologically different from the rest of the family. Threat status. To detect whether threatened species showed different trends from more common species, we included a binary variable indicating whether or not a species was included in the Swedish Red List (Gärdenfors, 2000) Classification of sites For each site, we calculated the area of semi-natural grassland and the proportion of semi-natural grasslands within 1 and 3 km radius, respectively. In order to detect changes in habitat quality we calculated a measure of change in the cover of woody plants and in grazing intensity. We also indicated whether a site was under legal protection or not. Site area. We calculated the censured area of semi-natural grassland of each site using GIS (ArcView 3.2), based on aerial photos. Landscape composition. As a measure of landscape composition we calculated the area of semi-natural grasslands in the surroundings of each study site. We used geographical data from a nation-wide survey of valuable grasslands (from a botanical point of view), conducted (Anonymus, 1997), and calculated the area of such grasslands within 1 and 3 km radius of the study sites, using GIS (ArcView 3.2). Changes in cover of trees and shrubs. In order to estimate the change in cover of the tree- and shrub layers, we compared aerial photographs (orthophotos) in scale 1:2000 of the sites between 1975 and , representing the two periods, respectively. We applied a m (5 5 mm) grid to the photos and counted the number of grid cells where the (pooled) cover of trees and shrubs was >50% of the grid cell area. The area of open land was calculated as the total area minus the area of the squares with >50% cover for each site and period. We then calculated the relative (%) change in open area; (A 2 A 1 )/A 1, where A 1 is the open area in 1975 and A 2 is the open area in , as a measure of overgrowth by trees and shrubs. Changes in grazing intensity. Grazing intensity was estimated at each visit and for each transect segment in a 6- graded scale were 0 = no grazing and 5 = intense grazing. However, since we used a subjective estimation of grazing intensity, we suspected that it would not be appropriate to compare the measures of grazing intensity directly. Instead, as a more robust measure, we ranked the sites based on the mean values of grazing intensity and used the change in rank order as a rough estimate of change in grazing intensity. Protected areas. To detect if the probability of population persistence generally was higher in nature reserves than in unprotected areas we applied a binary variable indicating whether a site was under legal protection (n = 8) or not (n = 5) Data analysis Data on species richness do not describe the dominance and relative abundance of individual species. In addition to species richness, we therefore calculated Simpson s diversity index (D) for each site and period D ¼ X ðn i ðn i 1Þ=ðNðN 1ÞÞÞ; where n i is the number of individuals of species i and N is the total number of observed individuals of all species. Simpson s diversity index was chosen because in contrast to many other diversity indices it is not sensitive to differences in sample size (Magurran, 2004). The differences in species richness and diversity between the two periods were analyzed using paired-samples t-test. We will refer to situations were a species was present in 1981 but absent in 2002 as extinctions and when a species was absent in 1982 but present in 2002 as colonizations. Extinctions and colonizations will be collectively referred to as turnover. The third possible outcome, no change, is the result either of presence by a certain species in both years or of absence in both years. The effects of site characteristics and species characteristics on the turnover were analyzed using Generalized Linear Mixed Models (SAS macro Glimmix) (Littell et al., 1996). To be able to handle colonizations and extinctions in the same analysis (since they are not independent), we analyzed presence and absence of species in a repeated measures analysis. The dependent variable was presence (1 or 0) in each period. Hence, each combination of species, site and period constituted one data point. Site was included as a random factor and period ( and , respectively) as a fixed factor. Effects of site characteristics and species characteristics on the turnover were analyzed in two separate models. Variables to be included in the final model were selected by stepwise forward selection. Fixed factors and interactions between fixed factors and year were included one by one in order of significance. Interactions were only included in the model if the main factors were already included and significant. We were mainly interested in factors that influence the changes (colonizations, extinctions or no change), which are given by the interactions of each factor with the year-factor. Hence, we only report these interactions. Since the dataset was unbalanced, we used the Satterthwaite method to estimate the denominator degrees of freedom, which may

6 BIOLOGICAL CONSERVATION 128 (2006) not necessarily produce whole numbers (Littell et al., 1996). All statistical analyses were performed in SAS 8.2 for Windows. 3. Results In total 59 species were found during and 53 species during The changes in species richness showed a large variation between sites (Table 1) The mean number of species per site decreased from 29.8 to 24.2, which is a statistically significant difference (T-test: t 12 = 3.01, p = 0.011) but there was no difference in Simpson s diversity between the two periods (T-test: t 12 = 0.72, p = 0.49). The number of species decreased at seven sites, increased at four sites and did not change at two sites. However, even at those sites where the number of species did not change species were lost, but the same numbers of species had also colonized the sites. In total, there were 136 local extinctions and 68 colonization events. Eight species found in at least one site in were not observed at all in Mellitaea cinxia and Erynnis tages were both found in four sites; Leptidea sinapis and Maculinea arion were found in three sites and Aporia crataegi was found in two sites in , but were not observed in any site in (Table 2). The species that were lost from the largest numbers of sites were Pyrgus malvae, Argynnis adippe and Lycaena virgaureae which went extinct from seven sites each and Hesperia comma and Satyrium pruni which were lost from six sites each (Table 2). The two species that decreased most in total abundance were Lycaena virgaureae which decreased from 2079 to 37 observed individuals and Brenthis ino which decreased from 804 to 35 observed individuals across all sites. In contrast, Araschina levana was not found in any of the sites in , but in six sites in (Table 2) Site characteristics Turnover was significantly affected by the changes in treeand shrub cover (F 1,1504 = 10.4, p = 0.001, Table 3), with an increasing loss of species with increasing cover of trees and shrubs (Fig. 2). The turnover rate was not affected by differences in relative grazing pressure, the proportion of seminatural grasslands within 1 km or within 3 km, distance to nearest neighbor or if the site was a reserve or not (all p > 0.1). There was a positive relation between area and presence by a species in each period, but not between area and turnover (Table 3) Species characteristics Only one of the species characteristics, larval host plant nutrient optimum (F 2,1509 = 21.2, p < 0.001) affected turnover (Table 3). The effect of nutrient optimum of host plants was the result of both a higher number of extinctions of species with food plants with low nutrient values and a higher number of colonizations of species with high values (Fig. 3). None of the other species characteristics (threat status, family, host plant specificity, mobility, generations per year or position of northern or southern range limit) had any significant effects on species turnover (all p > 0.05, Table 3). 4. Discussion Decreasing species richness of butterflies has been reported from several countries in Western Europe (van Swaay and Table 3 Outcome of the Glmm models Variable F df p Site variables Area Overgrowth Grazing change Landscape 3 km Landscape 1 km Reserve Period overgrowth Period area Species variables Nutrient optimum N limit Generations Threat status Host specialization Family S limit Mobility Period nutrient optimum <0.001 Period N limit df denotes the denominator degrees of freedom; variables included in the final models are indicated with bold text. The interaction terms between period and each variable indicate whether there was any relation between turnover and the respective variable. Interaction terms were only included for significant main terms. See text for details.

7 570 BIOLOGICAL CONSERVATION 128 (2006) Fig. 2 The number of extinctions and colonizations in relation to the relative change in open area. The turnover was significantly related to the change in open area (F 1,1504 = 10.4, p = 0.001, see text for details). Fig. 3 Number of extinctions and colonizations of species whose larval host plants are dependent on low and high nutrient conditions. The turnover was significantly related to the nutrient optimum of host plants (F 2,1509 = 21.2, p < 0.001). Mean values ± SE. Warren, 1999; Asher et al., 2001; Maes and van Dyck, 2001). Habitat loss through exploitation and agricultural intensification is assumed to be the main cause to these extinctions (Maes and van Dyck, 2001). In this study, we have shown that significant negative changes in butterfly species richness have also occurred in Sweden, even though all sites did not show a net decrease in species numbers. Although some sites had a large number of extinctions, the average rate of species loss observed in this study was relatively moderate compared to those reported elsewhere (e.g., Maes and van Dyck, 2001; Thomas et al., 2004). However, most of our sites were relatively large and well-connected, and were continuously managed between the two study periods. The mean number of species per site during the second study period was significantly higher (T-test: t 36 < 0.001) than the mean number of species in 24 smaller (6 10 ha) semi-natural grassland sites (E. Öckinger and H.G. Smith, unpublished data) in the same region. Moreover, eight of the 13 sites were legally protected as nature reserves or similar. Taking this into account, the observed declines must be considered to be rather large and nay be a conservative estimate of species decline across Sweden. We found no support for several of our predictions regarding which species would have declined and from which sites they would have been lost, perhaps due to the small number of sites. However, we did detect negative effects of changed habitat quality, in the form of increased cover of trees and shrubs, and strong negative effects on species dependent on nutrient-poor conditions, compared to species dependent on nutrient-rich conditions. The increasing cover of trees and shrubs may have affected some species directly, especially those dependent on open conditions and a warm micro-climate, as an increase in tree and shrub cover is likely to create a colder microclimate, and a decrease in the area of suitable habitat for such species (Thomas, 1993; Greatorex-Davies et al., 1993). Most species, however, may have been more indirectly affected by an altered composition and structure of the field layer plant community (Erhardt, 1985; Balmer and Erhardt, 2000), and the increasing tree and shrub cover could rather be seen as an indicator of significant changes in habitat quality more generally. Typically, butterfly abundance and species richness are highest at extensively grazed pastures or abandoned sites at early stages of succession (Erhardt, 1985; Balmer and Erhardt, 2000; Franzén and Ranius, 2004), but as the succession proceeds and the cover of trees and shrubs increases, the number of butterfly species falls again (Erhardt, 1985; Balmer and Erhardt, 2000). Surprisingly, we detected no effects of the changes in grazing intensity. However, even with constant grazing intensity, increased soil nitrogen levels may have contributed to an increasing tree

8 BIOLOGICAL CONSERVATION 128 (2006) and shrub cover in some of the sites, by increasing the rate of vegetation succession. Also, we may have failed to detect changes in grazing intensity, especially if there where general trends in grazing intensity across most of the sites, since our measure of changes in grazing intensity was rather coarse and we only used the change in rank order between the sites. Temporarily increased, decreased or ceased grazing during one or a few years may also have caused extinctions of butterfly populations (cf. Thomas et al., 1986) but we are not able to detect such effects due to the long interval between the surveys. During the 20 years covered by our study, species whose host plants are connected with nutrient-poor conditions tended to decrease and those whose host plants are related to nutrient-rich conditions tended to increase. It was also obvious that many of the species that had the highest numbers of extinctions are species associated with dry, nutrientpoor grasslands, e.g., Pyrgus malvae and Hesperia comma. This is not very surprising, since Tyler and Olsson (1997) found similar results for vascular plants in southern Sweden. Our results also agree well with those of Oostermeijer and van Swaay (1998), who found significant correlations between occurrence patterns and soil properties such as nutrient conditions, acidity and moisture for several butterfly species, and with Pollard et al. (1998) who reported increasing abundance of species feeding of coarse-leaved grasses, probably as a result of increased nutrient levels. The rapid increase of A. levana, first observed in Sweden in 1982 (Palmqvist, 1983) and occurring in six of our 13 sites in , might also partly be explained by an increase in nutrients and a subsequent increase in the abundance of its nitrogen-favored host plant Urtica dioica. Since A. levana is expanding its range both at its northern and southern margins (Parmesan et al., 1999), an increased host plant abundance is a more likely explanation than climate change. The relation between population changes and the larval host plant nutrient optimum may be a result of active fertilizing in order to improve pasture productivity, atmospheric nitrogen deposition, or a combination of these. At some of the sites, active fertilizing occurred during the 1970s and 1980s (O.H., S.G.N.; personal observations). If active fertilizing was the main cause behind this effect, this could explain the large variation in change of species richness between the sites. However, at some sites with large decreases in species richness, there are no signs of recent active fertilization, indicating that atmospheric nitrogen deposition might be more important. The levels of atmospheric nitrogen deposition are high in the study region (wet deposition: kg ha 1 year 1 ) and are increasing (Bernes and Grundsten, 1992), so it can be assumed that the soil nutrient content has generally increased between the two study periods. However, since the levels of atmospheric deposition are likely to be similar across our sites, it seems unlikely that this alone could have caused the great differences between the sites, even though different soil types and different plant communities may differ in their sensitivity to increased nitrogen levels. Based on island biogeography (MacArthur and Wilson, 1967) and metapopulation theory (Hanski, 1999), we expected to find effects of site area and isolation on turnover rates. Studying significantly smaller grassland fragments than we did, Krauss et al. (2003) found significant effects of site area but not of isolation on butterfly species turnover, but other studies have demonstrated the importance of connectivity for butterfly species richness (Bergman et al., 2004; E. Öckinger and H.G. Smith, unpublished data). The influence of the surrounding landscape on population dynamics can be assumed to be more important in small sites than in larger ones. Since many of our sites are large, landscape-scale processes may be of rather limited importance, and this may be one reason for not detecting any effects of landscape composition. Moreover, the differences in both area and connectivity between sites may have been too small to detect such effects. 5. Conclusions and implications As in many other European countries (van Swaay and Warren, 1999; Asher et al., 2001; Maes and van Dyck, 2001; Saarinen et al., 2003), the species richness of butterflies in Sweden appears to be in decline, but perhaps at a slower rate than in some of the more densely populated countries of Western Europe. In the sites studied here, the local extinction rates were related to increased soil nitrogen levels and increased cover of trees and shrubs. The fact that many species were also lost from sites that are protected as nature reserves indicates that the management of these reserves is sub-optimal for butterflies. Comparable observations have been made in Britain (Warren, 1993), and there is a great risk that the situations for other less conspicuous and less studied groups of organisms are similar, or even worse. The negative effect of increasing tree and shrub cover should probably be seen mainly as an indication of general habitat quality change and perhaps as a result of increased soil nutrient status. Nevertheless, reserve management should ensure that trees and shrubs do not cover a too large proportion of the grassland area and take measures to control vegetation succession. The extinction rates were as high in large as in small sites. This and the fact that there were a substantial number of local extinctions even in sites where no net decreases in species richness were observed (Table 1) indicates that the establishment of nature reserves is often not enough to save a species from local extinction. Instead, a landscape approach may be more fruitful. Increased connectivity between local populations and a larger number of suitable habitat patches increases the probability of long term population persistence (Hanski, 1999; Cabeza et al., 2004). However, such an approach is likely to be insufficient to counteract the detrimental effects of large-scale phenomena, such as atmospheric nitrogen deposition. Since no regular butterfly monitoring program exists in Sweden, we do not know to what extent the observed declines in species richness are representative for the region, for Sweden or for Scandinavia. Most of the sites are rather large and of relatively high quality compared to other grasslands in the region, and at least initially most of them had relatively rich butterfly faunas. Based on this it could be argued that species that went extinct earlier elsewhere probably could have persisted for a longer time in the studied sites,

9 572 BIOLOGICAL CONSERVATION 128 (2006) but finally went extinct also in these sites in the time period between our two surveys. On the other hand, since no major changes in habitat quality have occurred, it seems more likely that our studied sites still contain a relatively large number of species compared to other sites. A regular monitoring program at sites representing several different habitat types, such as the programs in Great Britain (Pollard and Yates, 1993) and The Netherlands (van Swaay et al., 1997) could give indications of general, large-scale population trends. In addition, more focused monitoring of the state of populations within protected areas is necessary in order to optimize reserve management. Acknowledgements Magnus Sylvén participated in both the planning and the field work of the initial study , which was financed by the Swedish Environmental Protection Agency. The re-survey and the analyses were financed by the Swedish Environmental Protection Agency through the research program The Conservation Chain to H.G.S. S.G.N. was supported by FOR- MAS. Markus Franzén, Daniel Rogers and three anonymous reviewers gave valuable comments on the manuscript. REFERENCES Anonymus, Ängs- och hagmarker i Sverige. Rapport 4819, Naturvårdsverket, Stockholm. Asher, J., Warren, M., Fox, R., Harding, P., Jeffcoate, G., Jeffcoate, S., The Millennium Atlas of Butterflies in Britain and Ireland. Oxford University Press, Oxford. Balmer, O., Erhardt, A., Consequences of succession on extensively grazed grasslands for Central European butterfly communities: rethinking conservation practices. Conservation Biology 14, Battin, J., When good animals love bad habitats? Ecological traps and the conservation of animal populations. Conservation Biology 18, Bergman, K.-O., Askling, J., Ekberg, O., Ignell, H., Wahlman, H., Milberg, P., Landscape effects on butterfly assemblages in an agricultural region. Ecography 27, Bernes, C., Grundsten, C., The Environment. National Atlas of Sweden, Stockholm. Bourn, N.A.D., Thomas, J.A., The challenge of conserving grassland insects at the margins of their range in Europe. Biological Conservation 104, Burke, I.C., Lauenroth, W.K., Vinton, M.A., Hook, P.B., Kelly, R.H., Epstein, H.E., Aguiar, M.R., Robles, M.D., Aguilera, M.O., Murphy, K.L., Gill, R.A., Plant soil interactions in temperate grasslands. Biogeochemistry 42, Cabeza, M., Moilanen, A., Possingham, H.P., Metapopulaton dynamics and reserve network design. In: Hanski, I., Gaggiotti, O.E. (Eds.), Metapopulation Ecology, Genetics, and Evolution. Academic Press, Amsterdam, pp Dennis, R.H.L., Hodgson, J.G., Grenyer, R., Shreeve, T.G., Roy, D.B., Host plants and butterfly biology. Do host-plant strategies drive butterfly status? Ecological Entomology 29, Dias, P.C., Sources and sinks in population biology. Trends in Ecology & Evolution 11, Ellenberg, H., Weber, H.E., Düll, R., Wirth, V., Werner, W., Paulissen, D., Indicator values of plants in Central Europe. Scripta Geobotanica 18. Erhardt, A., Diurnal Lepidoptera: sensitive indicators of cultivated and abandoned grasslands. Journal of Applied Ecology 22, Erhardt, A., Thomas, J.A., Lepidoptera as indicators of changes in the semi-natural grasslands of lowland and upland Europe. In: Collins, N.M., Thomas, J.A. (Eds.), The Conservation of Insects and Their Habitats. Academic Press, London, pp Franzén, M., Ranius, T., Occurence patterns of butterflies (Rhopalocera) in semi-natural pastures in southeastern Sweden. Journal for Nature Conservation 12, Gärdenfors, U Rödlistade arter i Sverige 2000, Artdatabanken, SLU, Uppsala. Greatorex-Davies, J.N., Sparks, T.H., Hall, M.L., Marrs, R.H., The influence of shade on butterflies in rides of coniferised lowland woods in southern England and implications for conservation management. Biological Conservation 63, Hammarstedt, O Miljöövervakning av och med dagfjärilar, Länet i utveckling 1996:21, Länsstyrelsen i Malmöhus län, Malmö. Hanski, I., Metapopulation Ecology. Oxford University Press, Oxford. Henriksen, H.J., Kreutzer, I., The butterflies of Scandinavia in nature. Skandinavisk Bogforlag, Odense. Hodgson, J.G., Commonness and rarity in British butterflies. Journal of Applied Ecology 30, Huldén, L. (Ed.), Finlands Storfjärilsatlas. Lepidoperologiska Sällskapet i Finland & Naturhistoriska centralmuseet, Helsinki. Karsholt, O., Razowski, J., The Lepidoptera of Europe: A Distributional Checklist. Apollo Books, Stenstrup. Kerr, J.T., Sugar, A., Packer, L., Indicator taxa, rapid biodiversity assessment, and nestedness in an endangered ecosystem. Conservation Biology 14, Konvicka, M., Maradova, M., Benes, J., Fric, Z., Kepka, P., Uphill shifts in distribution of butterflies in the Czech Republic: effects of changing climate detected on a regional scale. Global Ecology & Biogeography 12, Kotiaho, J.S., Kaitala, V., Komonen, A., Päivinen, J., Predicting the risk of extinction from shared ecological characteristics. Proceedings of the National Academy of Sciences of the United States of America 102, Krauss, J., Steffan-Dewenter, I., Tscharntke, T., Local species immigration, extinction, and turnover of butterflies in relation to habitat area and habitat isolation. Oecologia 137, Kudrna, O., The Distribution Atlas of European Butterflies. Apollo Books, Stenstrup. Littell, R.C., Milliken, G.A., Stroup, W.W., Wolfinger, R.D., SAS System for Mixed Models. SAS Institute Inc., Cary, NC. MacArthur, R.H., Wilson, E.O., The Theory of Island Biogeography. Princeton University Press, Princeton, NJ. Maes, D., van Dyck, H., Butterfly diversity loss in Flandres (north Belgium): Europes worst case scenario? Biological Conservation 99, Magurran, A.E., Measuring Biological Diversity. Blackwell Science, Oxford. Naumann, C.M., Tarmann, G.M., Tremewan, W.G., The Western Palaearctic Zygaenidae. Apollo Books, Stenstrup. New, T.R., Are Lepidoptera an effective umbrella group for biodiversity conservation? Journal of Insect Conservation 1, Oostermeijer, J.G.B., van Swaay, C.A.M., The relationship between butterflies and environmental indicator values: a tool for conservation in a changing landscape. Biological Conservation 86, Palmqvist, G., Intressanta fynd av Macrolepidoptera i Sverige Entomologisk Tidskrift 104,

10 BIOLOGICAL CONSERVATION 128 (2006) Parmesan, C., Ryrholm, N., Stefanescu, C., Hill, J.K., Thomas, C.D., Descimon, H., Huntley, B., Kaila, L., Kullberg, J., Tammaru, T., Tennent, W.J., Thomas, J.A., Warren, M., Poleward shift in geographical ranges of butterfly species associated with regional warming. Nature 399, Pollard, E., Yates, T.J., Monitoring Butterflies for Ecology and Conservation. Chapman & Hall, London. Pollard, E., Moss, D., Yates, T.J., Population trends of common British butterflies at monitored sites. Journal of Applied Ecology 32, Pollard, E., Woiwod, I.P., Greatorex-Davies, J.N., Yates, T.J., Welch, R.C., The spread of coarse grasses and changes in numbers of Lepidoptera in a woodland nature reserve. Biological Conservation 84, Pywell, R.F., Warman, E.A., Sparks, T.H., Greatorex-Davies, J.N., Walker, K.J., Meek, W.R., Caewell, C., Petit, S., Firbank, L.G., Assessing habitat quality for butterflies on intensively managed arable farmland. Biological Conservation 118, Reynolds, J.D., Life histories and extinction risk. In: Blackburn, T.M., Gaston, K.J. (Eds.), Macroecology: Concepts and Consequences. Cambridge University Press, Cambridge, pp Saarinen, K., Lahti, T., Marttila, O., Population trends of Finnish butterflies (Lepidoptera: Hesperioidea, Papilionoidea) in Biodiversity and Conservation 12, Smith, R.S., Shiel, R.S., Millward, D., Corkhill, P., The interactive effects of management on the productivity. Journal of Applied Ecology 37, Stoltze, M., Danske dagsommerfugle. Gyldendal, Copenhagen. Sutcliffe, O.L., Thomas, C.D., Moss, D., Spatial synchrony and asynchrony in butterfly population dynamics. Journal of Animal Ecology 65, Thomas, J.A., Holocene climate changes and warm man-made refugia may explain why a sixth of British butterflies possess unnatural early-successional habitats. Ecography 16, Thomas, C.D., Dispersal and extinction in fragmented landscapes. Proceedings of the Royal Society of London, Series B Biological Sciences 267, Thomas, J.A., Thomas, C.D., Simcox, D.J., Clarke, R.T., Ecology and declining status of the silver-spotted skipper butterfly (Hesperia comma) in Britain. Journal of Applied Ecology 23, Thomas, J.A., Rose, R.J., Clarke, R.T., Thomas, C.D., Webb, N.R., Intraspecific variation in habitat availability among ectothermic animals near their climatic limits and their centres of range. Functional Ecology 13 (Suppl. 1), s55 s64. Thomas, J.A., Bourn, N.A.D., Clarke, R.T., Stewart, K.E., Simcox, D.J., Pearman, G.S., Curtis, R., Goodger, B., The quality and isolation of habitat patches both determine where butterflies persist in fragmented landscapes. Proceedings of the Royal Society of London, Series B Biological Sciences 268, Thomas, J.A., Telfer, M.G., Roy, D.B., Preston, C.D., Greenwood, J.J.D., Asher, J., Fox, R., Clarke, R.T., Lawton, J.H., Comparative losses of British butterflies, birds, and plants and the global extinction crisis. Science 303, Tyler, T., Olsson, K.A., Förändringar i Skånes flora under perioden statistisk analys av resultat från två inventeringar. Svensk Botanisk Tidskrift 91, van Swaay, C.A.M., An assessment of the changes in butterfly abundance in The Netherlands during the 20th century. Biological Conservation 52, van Swaay, C., Warren, M., Red Data Book of European Butterflies (Rhopalocera). Council of Europe Publishing, Strasbourg. van Swaay, C.A.M., Maes, D., Plate, C., Monitoring butterflies in The Netherlands and Flanders: the first results. Journal of Insect Conservation 1, Warren, M.S., The conservation of British butterflies. In: Dennis, R.H.L. (Ed.), The Ecology of Butterflies in Britain. Oxford University Press, Oxford. Warren, M.S., A review of butterfly conservation in central southern Britain: I. Protection, evaluation and extinction on prime sites. Biological Conservation 64,