Ground beetle (Coleoptera: Carabidae) population declines and phenological changes: Is there a connection?

Ecological Indicators 41 (2014) 15–24

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Ground beetle (Coleoptera: Carabidae) population declines and phenological changes: Is there a connection? Gabor Pozsgai ∗ , Nick A. Littlewood The James Hutton Institute, Craigiebuckler, Aberdeen AB15 8QH, United Kingdom

a r t i c l e

i n f o

Article history: Received 4 July 2013 Received in revised form 16 January 2014 Accepted 21 January 2014 Keywords: Climate change Population decline Environmental Change Network Phenology Upland

a b s t r a c t Long-term monitoring data were analyzed to reveal correlations between declining ground beetle (Coleoptera: Carabidae) populations and phenological changes at two Environmental Change Network sites in Scotland. The potential role of advancing phenology as an adaptation function in population stability was investigated. Analysis focussed on the 25 most abundant species over an 18 year sampling period. Pitfall trap catches were used to calculate mean activity-densities both for the whole sampling period and for dates limited to those within the activity period. Several phenological measurements were calculated (i.e. first day of appearance, peak activity date, median activity, length of activity and winter inactivity periods, and the last day of presence) for each species. Robust non-parametric estimation was used to model changes in both activity density and phenology. Eight species declined in activity density over time, three increased and fourteen showed no change. The mean rate of decline was greater than that of increase. Most of the species included in the analysis changed their phenology. Advancing onset of activity and earlier cessation were the most pronounced changes. However, a slow advancing trend in the peak activity was also shown. Only Nebria brevicollis, an autumn species with recorded winter activity, extended its activity period to later dates, suggesting that cessation of activity for the remaining species may be more closely linked to photoperiod. The earlier termination of activity shortened substantially the activity window for several species. Declines in activity density showed a strong relationship with a narrowing window of activity, mainly caused by earlier cessation of activity. Declining species were found more in bog or dry heather moorland habitats compared to grassland, emphasizing the vulnerability of these vegetation types, and associated insect assemblages depending on them, to environmental or climatic changes. The reciprocal relationship found between the trend of timing of initiation of activity and changes in activity-densities suggests that populations with a higher capacity to advance their phenology are less prone to decline. Since phenological changes may drive changes in populations, investigating phenological variables is encouraged in both research and conservation planning. © 2014 Elsevier Ltd. All rights reserved.

1. Introduction Changes in an environment can lead to a number of responses from plant and animal communities (e.g. McCarty, 2001; Parmesan, 2006; Walther et al., 2002). Such responses, especially to drivers such as climate change, include shifts in geographical distribution (e.g. Drees et al., 2011; Parmesan et al., 1999; Parmesan, 1996) or phenological changes such as being active at different times of the year (e.g. Graham-Taylor et al., 2009; Hassall et al., 2007; Penuelas

∗ Corresponding author. Tel.: +44 1224395223. E-mail address: [email protected] (G. Pozsgai). 1470-160X/$ – see front matter © 2014 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.ecolind.2014.01.029

and Filella, 2001; Pozsgai and Littlewood, 2011; Roy and Sparks, 2000). If a species’ response does not keep pace with environmental change its abundance and relative importance in a community can decrease, and species can go extinct from areas where the environment becomes unsuitable (e.g. McCarty, 2001; Parmesan, 2006). All of these processes can result in a mismatch in species’ interactions at all levels, with the potential to threaten ecosystem stability (Durant et al., 2005; Visser and Holleman, 2001; Walther et al., 2002). Numerous studies have been conducted on the impacts of environmental change, such as global warming, in particular on insect communities, focusing on specific taxonomic groups (e.g. Warren et al., 2001; Woods et al., 2008) or on trophic levels

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(e.g. Bale et al., 2002) and the implications for nature conservation (Arribas et al., 2012; Kotze and O’Hara, 2003; Samways, 2007; Schuldt and Assmann, 2010) of the observed trends. Much research has focussed on life-traits such as dispersal ability, thermal sensitivity (Biesmeijer, 2012; Butterfield, 1996; Charmantier et al., 2008; Harper and Peckarsky, 2006; Kotze and O’Hara, 2003; Poniatowski et al., 2012) ecophysiological changes of a species related to climate change (Bale and Hayward, 2010; Chown and Gaston, 1999) or species’ responses to various environmental conditions (e.g. changes in diversity, species richness or overall abundance) (Andrew and Hughes, 2005; Templer et al., 2012; Warren et al., 2001; Woods et al., 2008). Although an increasing body of knowledge suggests that phenological changes and species’ declines may be linked (Jones and Cresswell, 2010; McNamara et al., 2011), there are no studies that question whether the change in a species’ phenology can enable population stability or even growth in the face of drivers such global warming. This strategy could provide advantage to individuals adapted to the changed circumstances and even significantly improve the competitiveness of species over others lacking this ability. It would also enable populations of the species to maintain, or even increase, their importance in a community without taking the risk of colonizing new locations or habitats. As Pau et al. (2011) pointed out for species living in a temperate climate, changing phenology is a more important adaptive strategy than shifting geographical distribution. Ground beetles (Coleopera: Carabidae) have been the focus of phenological and climate change-related studies for almost two decades (Butterfield, 1996; Grechanichenko, 2001; Kotze and O’Hara, 2003; Morecroft et al., 2009; Scott and Anderson, 2003). Their well-known life-history, ecology and quick responses to environmental changes make them suitable for use as biological indicators (Lövei and Sunderland, 1996; Pearce and Venier, 2006; Rainio and Niemela, 2003; Thiele, 1977). Long-term datasets are rare though, especially those where data were collected using standardized sampling techniques (Walther et al., 2002). The United Kingdom Environmental Change Network (ECN) is a long-term monitoring project that was established in 1993 and now operates across 12 terrestrial sites in the UK. It uses standardized methods to trace the impacts of changes in climate and other environmental drivers on natural ecosystems (Sykes and Lane, 1996). Among the monitoring methods, the ECN Ground Predator Protocol, focussing primarily on ground beetles, has been running at two ECN sites in Scotland, Glensaugh and Sourhope, since 1994, giving 18 years of continuous data. A significant decrease in the abundance of ground beetles was reported at several ECN sites over a seven (Scott and Anderson, 2003) and over a 14 year period (Morecroft et al., 2009), particularly in upland habitats. Moreover, Pozsgai and Littlewood (2011) reported an advancing phenology of Pterostichus madidus from the same sites. Both phenomena are likely to be a caused by a changing climate, although the main drivers are not yet determined. Our research investigates temporal changes in abundance and phenology of ground beetle assemblages, aiming to gain a deeper insight into the adaptation aspects of phenological change, and describes the potential links between changing phenology and declining ground beetle populations. In this study we aimed to (1) describe the changes in activity-densities (a measure of relative frequencies of capture (Thiele, 1977) of the most abundant species in the assemblages we investigated, (2) describe the changes in these species’ phenology, (3) investigate whether phenological plasticity can influence the likelihood of population declines and (4) identify species and habitats that are most vulnerable to drivers of phenological change.

2. Materials and methods 2.1. Site locations and sampling protocol Ground beetle data from two Scottish ECN sites, Glensaugh and Sourhope, were used for the analysis. Glensaugh Research Station is located in North-East Scotland, 56 km south-west of Aberdeen (latitude/longitude: 56.895363, −2.541496, altitude: 165 m, mean annual temperature: 7.8 ◦ C, mean annual rainfall: 1048 mm). Sourhope lies 24 km south-east of Kelso (latitude/longitude: 55.481188, −2.245039, altitude: 231 m, mean annual temperature: 7.5 ◦ C, mean annual rainfall: 893 mm) (www.ecn.ac.uk). At both sites data were available from 1994 until 2011, from 3 pitfall trap transects, each. Transects consisted of 10 traps, 10 m apart from each other and positioned on an improved grassland, dry heather moorland and blanket bog vegetation at each site. Both dry heather moorland and blanket bog habitats were mostly dominated by heather (Calluna vulgaris) and grasslands were intensively grazed. The transect positioning and sampling procedure followed the ECN Ground Predator Protocol (Sykes and Lane, 1996). Sampling frequency was fortnightly from April or May until November each year. Material collected later than 2004 has been identified by the first author, using the works of Lindroth (1985, 1986), Freude et al. (2004) and Hurka (1996). Data from earlier samples are those held by the ECN database (http://www.ecn.ac.uk/data info.htm). 2.2. Statistical analysis To avoid biased results caused by insufficient data, only species with a combined abundance of greater than 100 captured individuals over the entire sampling period at both sites were included in the analysis. Years with fewer than three recording dates for each of these species were also excluded. Due to the difficulty in their identification, Patrobus assimilis and P. atrorufus were considered as one ‘pseudospecies’; Patrobus sp. The data from the ten traps in each transect were pooled to generate activity-densities (number of individuals captured per ten traps per sampling event, AD) and yearly means were calculated using these batches for each species. Means were calculated in two different ways; (a) including all sampling events, counting as zeroes when the traps were operating but the particular species was not present (all mean, AM) and (b) including only sampling events between the first and last date when the species was present (limited mean, LM). AM is more useful and the most commonly used measure in ecological studies when investigating the changes in abundances of populations from a general aspect. However, with AM being sensitive to the length of activity period, it would not be possible, using this measure alone, to determine whether changes were caused solely by a shortened activity period or by a general decrease in AD over the entire sampling period. LM, on the other hand, can detect changes in AD even when the length of the activity period is changing. In order to investigate trends in species’ abundance changes we used robust non-parametric Mann–Kendall tau, and estimated the slope and confidence limits (CL) by using Sen’s method (Sen, 1968; Sokal and Rohlf, 1995). Two-sided p-values were calculated to test significance. Although both the scale and quantity of activity-density data would have been sufficient for using parametric statistics, to aid comparability we preferred to use the same method that was used for the trend analysis of phenological data. Moreover, this method was shown to perform well when used for analyzing time-series data (Önöz and Bayazit, 2003). For the phenological analysis, when the given species was present, dates of sampling events were converted to ordinal dates (days of the year). Although, for several years, data from months

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Table 1 Number of individuals of the 25 species caught in each habitat. Percentages given in brackets reflect the relative abundances of the species taking into account all collected species from all samples. Species name abbreviations given are those used in Fig. 1 and the graphical abstract. Species

Abbreviation

Bog

Dry heather

Grassland

Agonum fuliginosum Amara communis Amara lunicollis Calathus fuscipes Calathus melanocephalus Calathus micropterus Carabus arvensis Carabus glabratus Carabus problematicus Carabus violaceus Cychrus caraboides Leistus terminatus Loricera pilicorni Nebria salina Nebria brevicollis Notiophilus aquaticus Notiophilus biguttatus Patrobus sp. Pterostichus adstrictus Pterostichus diligens Pterostichus madidus Pterostichus melanarius Pterostichus niger Pterostichus strenuus Trechus obtusus 

Agon.ful Amar.com Amar.lun Cala.fus Cala.mel Cala.mic Cara.arv Cara.gla Cara.pro Cara.vio Cych.car Leis.ter Lori.pil Nebr.sal Nebr.bre Noti.aqu Noti.big Patr.sp Pter.ads Pter.dil Pter.mad Pter.mel Pter.nig Pter.str Trec.obt

235 (4.64%) 3 (0.06%) 73 (1.44%) 41 (0.81%) 159 (3.14%) 907 (17.91%) 313 (6.18%) 93 (1.84%) 342 (6.75%) 452 (8.93%) 89 (1.76%) 777 (15.34%) 293 (5.79%) 8 (0.16%) 3 (0.06%) 50 (0.99%) 40 (0.79%) 161 (3.18%) 407 (8.04%) 107 (2.11%) 37 (0.73%) 9 (0.18%) 9 (0.18%) 61 (1.2%) 395 (7.8%) 5064 (21.09%)

3 (0.09%) 68 (1.97%) 398 (11.55%) 34 (0.99%) 366 (10.62%) 295 (8.56%) 12 (0.35%) 233 (6.76%) 141 (4.09%) 249 (7.23%) 54 (1.57%) 286 (8.3%) 155 (4.5%) 14 (0.41%) 12 (0.35%) 89 (2.58%) 48 (1.39%) 37 (1.07%) 200 (5.8%) 303 (8.79%) 194 (5.63%) 1 (0.03%) 46 (1.33%) 118 (3.42%) 90 (2.61%) 3446 (14.35%)

4 (0.03%) 47 (0.3%) 47 (0.3%) 6044 (38.99%) 2719 (17.54%) 26 (0.17%) 4 (0.03%) 5 (0.03%) 123 (0.79%) 52 (0.34%) 0 (0%) 12 (0.08%) 96 (0.62%) 495 (3.19%) 91 (0.59%) 6 (0.04%) 19 (0.12%) 3 (0.02%) 21 (0.14%) 18 (0.12%) 5048 (32.57%) 321 (2.07%) 125 (0.81%) 156 (1.01%) 19 (0.12%) 15501 (64.56%)

other than these were available, in our analysis sampling days were restricted from May to November each year so as to avoid potential artefacts caused by changes of the first and last sampling dates. Minimum date value (MV, day when the third individual was captured), median values weighted with activity-densities of capture dates (WM), peak activity (PA, median of the ordinal dates with higher activity density than the yearly average) and maximum of the ordinal dates of capture events (LP, the day of last recorded presence) were calculated. Activity period (AP) was defined as the difference between maximum and minimum values, and winter inactivity period (WI) as the difference between MV and LP of the previous year. Trend-like changes were investigated using the same method described above. Slopes of all abundance-related and phenological measures (see above) were calculated, using the method described above, for each species in each vegetation type. Slopes with connected p-values greater than 0.25 were considered to represent no change. Although this is an arbitrary value, the 75% probability in support of the alternative hypothesis (trend-like changes) helps to provide large enough sample sizes for statistical comparison. Differences in the magnitude of beetle species abundance changes in the three vegetation types were then compared for each abundance-related and phenological measure, using Kruskall–Wallis-tests and two-sided pairwise Wilcoxon rank sum tests with Holm–Bonferroni method adjusted p-values. To investigate whether changes in phenology and ADs are correlated we used Spearman’s rank correlation test. Since LM is independent of the length of activity period, and therefore of all phenological measures, it was tested against each of those measures. Since these tests were not independent (see Blend, 2000) instead of using Bonferroni correction to control false positive test results we adjusted p-values according to the Benjamini–Hochberg procedure (Benjamini and Hochberg, 1995). Significant correlations between the changes of mean activity-density (LM) and phenological variables may suggest that changing phenology is influential in carabid population dynamics. All statistical analyses were carried out using the R software (R Core Team, 2013), with the help of “date” (Therneau et al., 2012)



242 (1.01%) 118 (0.49%) 518 (2.16%) 6119 (25.48%) 3244 (13.51%) 1228 (5.11%) 329 (1.37%) 331 (1.38%) 606 (2.52%) 753 (3.14%) 143 (0.6%) 1075 (4.48%) 544 (2.27%) 517 (2.15%) 106 (0.44%) 145 (0.6%) 107 (0.45%) 201 (0.84%) 628 (2.62%) 428 (1.78%) 5279 (21.99%) 331 (1.38%) 180 (0.75%) 335 (1.4%) 504 (2.1%) 24011 (100%)

“kendall” (McLeod, 2011), “BiodiversityR” (Kindt and Coe, 2005) and “zyp” (Bronaugh and Werner, 2012) packages. 3. Results There were 24011 individuals, of 25 species, caught over the 18 years of sampling at Glensaugh and Sourhope ECN sites, that met the criteria for inclusion of this analysis. This amounts to 94.25% of all Carabids caught from both sites over the entire sampling period. Most species were found in more than one habitat. Several species were characteristic of one habitat and only represented by a few individuals occurring in other habitats (e.g. Leistus terminatus was abundant in both of the heather-dominated vegetations and very rare in grasslands) though only Cychrus caraboides did not occur in all habitats at least once (Table 1). 3.1. Changes in abundance-related variables The AM of seven of the twenty-five species involved in the analysis decreased significantly over the study period and one decreased marginally significantly (0.05 < p < 0.1). Calathus fuscipes suffered the most severe decline in AM (Sen = −1.1 individual year−1 ; CL = −2.67, −0.09; p = 0.045), but the extent of decrease was major in the case of Calathus melanocephalus and Calathus micropterus (Sen = −0.88 individual year−1 , CL = −1.66, −0.06, p = 0.034; Sen = −0.5 individual year−1 , CL = −0.86, −0.33, p < 0.01 respectively). Pterostichus adstrictus, from a high population activity-density in the early years with a clear declining trend (−0.4 individual year−1 , CL = −0.73, −0.11, p < 0.01), almost completely disappeared from our samples by the end of the sampling period (one specimen was caught in the final year). The LM of seven species decreased significantly. Patrobus sp. did not show a change in LM despite of its marginally significant decrease in AM value. (Fig. 1, Supplementary material 1). The populations of thirteen species can be regarded as being stable due to the lack of significant changes in their activity-densities over the sampling period.

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Fig. 1. Trend-like changes in abundance-related variables. Arrows pointing down indicate declining trends (p < 0.1), arrows pointing up indicate increasing trends (p < 0.1), circles indicate no changes (p > 0.1). Widths of arrows correspond with the magnitude of the change.

C. caraboides, Nebria brevicollis, Pterostichus niger and Pterostichus melanarius became more abundant (Fig. 1, Supplementary material 1). The latter three species increased both in AM and LM whilst C. caraboides increased in LM only. The greatest increase both in AM and LM was recorded for N. brevicollis (Sen = 0.3; CL = 0.15, 0.63, p < 0.01; Sen = 0.3; 0.04, 0.80; p = 0.018, respectively. Slopes represent individual sampling date−1 year−1 changes). In summary, 32% of the species included in this analysis showed a decreasing trend in activity density, 16% increased and 52% showed no change in activity-density. Whilst the mean rate of change among those that decreased was 0.32 ± 0.31 individuals year−1 the mean rate of change among those that increased was only 0.09 ± 0.08 individuals year−1 (one-sided Wilcoxon rank sum test p = 0.02), demonstrating that the magnitude of decline was greater than that of the growth.

3.2. Changes in phenology There were significant (p < 0.05) changes in some of the phenological measure of 12 species with five additional species showing marginally-significant changes (p ≤ 0.1). In eight species no significant changes were detectable (Fig. 1, Supplementary material 2). The first day of appearance advanced for seven species and became later for two species over the observed period. N. brevicollis displayed the greatest advance (Sen = −7.0; CL = −8.89, −2.12; p < 0.01) whilst the greatest delay in activity was shown for Leistus terminatus, though this was only at a marginally significant level (Sen = 2.7; CL = −0.21, 5.50; p = 0.081). Four species ceased their activity (LP) progressively earlier over the monitored period with the greatest extent show by P. adstrictus (Sen = −9.0; CL = −15.0, −3.0; p < 0.01). Only N. brevicollis showed a trend of extending its activity to later dates. Six species showed an advancing trend in the WM values and five in PAs, whilst the PA of Carabus problematicus shifted to later dates (Fig. 2). The length of AP decreased significantly in four species and near significantly in two further species. In contrast P. madidus (Sen = 0.8; CL = 0.00, 1.50; p < 0.01) and N. brevicollis (Sen = 8.7; CL = 3.11, 10.62; p < 0.01) extended their activity (Fig. 1, Supplementary material 2). The calculated winter inactivity period (WI) became significantly longer in five species and shortened in three species (Fig. 1, Supplementary material 2).

3.3. Relationship of trends to habitats Kruskall–Wallis tests showed no significant differences in the mean change of abundance-related measures between different habitats. However, with relatively low p-values, the possibility of a difference in AM between both heather-dominated habitats and grassland cannot be disregarded (Table 2). The same test revealed significant differences in the mean slopes of phenological measures between the three habitats. In pairwise comparisons, bog and dry heather habitats did not differ from each other but both proved to be different from grassland sites in the magnitude of changes in LP, AP and WI. The difference in the change of MV was significant between bog and grassland habitats, whilst the 11.2% probability for these values being equal between the dry heather and grassland habitats is low enough to give cautious support for there being a difference in this measure (Table 2). 3.4. Relationship between activity-density and phenology Since values of LM do not depend on any phenological measures, correlations between them could be analyzed. Spearman Rank Correlation revealed a strong negative relationship between LM and the day of appearance of the third individual (MV) (rho = −0.54, p = 0.016) suggesting that species that advanced in phenology declined less. LM and both the LP and AP were positively correlated (rho = 0.59, p < 0.01; rho = 0.58, p < 0.01, respectively), indicating a relationship between the narrowing window of activity, and population declines. LM and WI were also strongly but reciprocally correlated (rho = −0.55, p = 0.012). No significant correlations were found between WM or PA and any other measures. 4. Discussion Our research is focused on temporal changes in abundance and phenology of ground beetle assemblages, aiming at gaining a deeper insight into evolutionary aspects of phenological change and at investigating potential links between changing phenology and declining ground beetle populations. Numerous trends have been revealed, both in phenology and abundance-related variables, in ground beetle assemblages at two Scottish ECN sites, Glensaugh and Sourhope. However, the presence of trends and the magnitude and directionality of changes varied greatly among the Carabidae species involved in the study.

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Fig. 2. Weekly activities of the six species that showed significant change in their PA value, with the dashed line showing the linear trend. X and Y axes represent the weeks in a year and the sampled years, respectively. Radius of circles is proportional to the number of individuals caught. Darkened circles indicate the peak activity week for each year. First and last week of capture events also can be seen in the figure.

4.1. Abundance trends Of the twenty-five species, eight have undergone a significant decline. In contrast only four species became more frequent. Moreover, the mean declining trend was significantly greater than that of growth. The general decline of ground beetles at our ECN sites is in line with the findings of Morecroft et al. (2009) and Brooks

et al. (2012). Calathus fuscipes, C. melanocephalus and C. micropterus, which between them showed the three greatest declines, are wingless species, breeding in the autumn. Thus, higher late summer temperatures and droughts in autumn could have narrowed their activity window significantly, decreasing the possibility of successful mating and egg laying. Moreover, the lack of wings of these species decreases their mobility, hence their likelihood of finding

Table 2 Adjusted p-values of the two-sided pairwise Wilcoxon-tests between each habitat combination calculated for each phenological and abundance related variable. The bottom row contains the p-values of the Kruskall–Wallis tests.

Bog–dry heather Bog–grassland Dry heather–grassland Kruskall p-value

MV

LP

WM

PA

AP

AM

LM

WI

0.523 0.032 0.112 0.031

0.664 0.029 0.029 0.016

1 1 1 0.644

0.662 0.462 0.462 0.282

0.703 0.001 0.001 <0.001

0.458 0.111 0.124 0.17

0.915 0.223 0.249 0.427

0.5 0.03 0.027 0.007

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suitable microhabitats for reproduction before the onset of adverse late autumn weather. Besides the widely used metric of average number per trap sampling date (AM) we introduced another measure which is independent of the length of the sampling period (LM). The reason for doing so was to investigate whether detected declines in carabid populations were only the result of a shortened period of activity or whether ADs are lower overall. Since almost all changes in abundance-related variables were detectable with both measures and of a similar magnitude, we can conclude that for most of the species, when lower numbers were captured this was not only caused by a shortened activity period but also by a general decline in population size during this active period. Only Patrobus sp. decreased in AM marginally significantly but without showing such a decline in the LM values suggesting that shrinking AP is likely to be responsible for the detected decrease in ADs of this species. In contrast Cychrus caraboides increased in LM without any trend-like changes in AMs. This may be the first sign of an increasing trend of this species, albeit below the threshold of statistical significance. There is a slight indication that species of blanket bog and dry heather moorland habitats are affected more severely than those of grassland, although differences between these habitats in the mean slopes of AM or LM were not statistically significant (Kruskal–Wallis test p = 0.17, 0.427, respectively). Any such trend could be masked by the high rate of decline in abundance in Calathus fuscipes, the only grassland species that decreasing in AD (both AM and MV values, see Fig. 1, Supplementary material 2) values. However, with bog and dry heather moorland habitats pooled to ‘heathland’ there was a significantly greater decrease in AM values than was the case for grassland. The remaining seven of the eight declining species have high fidelity to heather-dominated habitats and the three species that become more frequent are all typical of grasslands. Our results are in line with several other studies emphasizing the vulnerability of these habitat types together with the specialized species associated with them (Essl et al., 2012; Morecroft et al., 2002; Thompson et al., 1995). In our study, there were no differences found in changes of any abundance-related or phenological measures between bog and dry heather moorland communities. Hence, hereafter we refer to them jointly as ‘heathlands’ or ‘heath-dominated habitats’. ADs of Carabid species increased only in grassland although the biggest single species decline was also in this habitat. This could be a reflection of a trend-like change in ground beetle assemblages as a reaction to changing climate, causing the replacement of a species in the habitat by one or more other species (McCarty, 2001; Parmesan and Yohe, 2003; Parmesan, 2006). 4.2. Phenology Phenological changes in several species were revealed by this study. Ten species from twenty-five advanced their activity at least in one of the phenological measures calculated. Similar advances have been reported for plants (Cleland et al., 2007), fish (Ahas, 1999) and birds (Crick et al., 1997; Cotton, 2003). Phenological changes in insect life-histories have been explored with hoverflies (Diptera: Syrphidae, Graham-Taylor et al., 2009), craneflies (Diptera: Tipulidae, Pearce-Higgins et al., 2005) and in Odonata by Hassall et al. (2007). Nonetheless, despite their well-studied ecology, only Pozsgai and Littlewood (2011) report an advance in the phenology of a carabid species. The potential impacts of changing phenologies are summarized in Penuelas and Filella (2001). In our study advances in the onset of activity (MV) were more pronounced than other timing measures. However median and peak activity dates also advanced markedly. Since global warming has tended to produce greater increases in winter temperatures than in temperatures during the remainder of the year (Easterling

et al., 1997), changes in early phenological stages are more likely. Nonetheless early emergence might be seen as a risky strategy, and slowly changing (Fig. 2) or static PA and WM values (Fig. 1) indicate that a significant part of the population continues to follow the established seasonality of activity. However, if advanced phenology is significantly increasing the fitness and fecundity of populations, and if this behavioural adaptation is hereditable, it is likely that the ‘genes of early emergence’ will ultimately drive an advancement in main activity peaks to earlier dates. In contrast to advancing early activities, a trend of later emergence was only demonstrated in two cases. Both species also experienced a high rate of decline. Thus, a decreased probability of being recorded may be the cause of them appearing later in our samples. This would be consistent with the findings of Ellwood et al. (2012) who suggested that changes in abundances influence the frequency of capture events, and thus the recorded phenology. Additionally, in species such as Patrobus sp. and Trechus obtusus exhibiting similar, marginally significant, trends increasing MVs coincided with decreasing LP values leading to a strong decrease in AP. In our study, all species showing this phenomenon declined. This may lead to the prospect that even if long-term abundance data are not available, phenological data may provide warning signs for species threatened by climate-driven changes in the environment. Only N. brevicollis extended its activity to later dates, supporting the findings of Butterfield (1996), Pozsgai and Littlewood (2011) and Shintani and Numata (2010) that the timing of hibernation of most ground beetles is more likely to depend on photoperiodic cues whilst emergence and spring activity are driven by temperature related factors. N. brevicollis is typically active in autumn and has been reported several times to be active over the winter period (Jaskula and Soszynska-Maj, 2011), so can remain active later with increasing autumn temperatures. PA only shifted significantly to later dates for one species, C. problematicus. Unfavourable weather conditions early in the activity period and lack of the main food resource may be the drivers in this observation. Changing voltinism may also explain the delay in activity peaks, although no significant indications were found in any abundance-related variables to support this hypothesis. Some species split they activity period by going to a summer aestivation (Matalin, 2008; Thiele, 1969). With changing environment the timing and length of this aestivation may also change, resulting in merging the dual activity peaks into one, separating them by a longer time period or the disappearance of any of the peaks. In our results we visually checked the activity curves of every species and only N. brevicollis showed an annual activity with two obvious peaks. For this species the same curves were checked yearly and no apparent changes in the number of activity peaks were visible. Although we have not tested whether any of these peaks had shifted, it is likely that at least one of our measures would have picked up on any significant changes. Furthermore, according to Thiele (1969), the first peak is more influenced by temperaturerelated environmental factors whilst the second is more influenced by photoperiod. Hence the first peak is more susceptible to change and since, in our dataset, this always was higher than the second (probably because of summer mortality), the introduction of a new measure was unnecessary. For many species the length of winter hibernation may be crucial for development. Between our calculated measures WI reflected how much time the studied species was inactive during the winter period. Very similarly to the AP, this measure was also strongly dependent on the ADs, as low ADs decrease the probability of capturing an animal and thus can strongly influence the post-peak records. Therefore, it is not surprising that all five species showing a significant increase in this measure also suffered a significant loss in at least one of the abundance related measures. Parallel to this, from two from those three species that shortened their winter

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hibernation, an increase in both abundance-related measures was detected. The decreased WI was caused by the advancing spring activity in the case of N. brevicollis and P. madidus. P. niger shows similar trends in WI in spite the lack of a significantly advancing trend in MVs. The adult body size of individuals spending less time developing over the winter may decrease, and this may significantly influence mating success (Babin-Fenske et al., 2008; Sheridan and Bickford, 2011). Hence, further research should be conducted on the effect of decreasing WI and carabid body size. Although it is beyond the scope of this paper to investigate climatic drivers of phenological changes, we propose that dry late summer and early autumn periods may have played an important role in the earlier cessation of carabid activity and may also drive the shortening activity period and general decline of populations of species as was suggested by Morecroft et al. (2002). False positive trends caused by generally decreasing ADs (species becoming generally scarce with the possibility of a decreasing AD being suggested) are possible, but unlikely in view of the minimum abundance threshold that was applied for a species to be included in this research. The capacity to change may depend on the ecology of a species and can broadly vary among taxa. Several signs, including declines and drastic shortening in activity periods, may indicate that species living in heathland habitats are more vulnerable to climate change than those in grasslands. The smaller advance of MVs in heathland than in grassland (one-sided Wilcoxon signed rank test, bog < grassland, dry heather < grassland, p = 0.016, 0.056, respectively) and the greater magnitude of decrease in LP (one-sided Wilcoxon signed rank test, bog > grassland, dry heather > grassland, p = 0.014) resulted in a significantly shortened AP (one-sided Wilcoxon signed rank test, bog > grassland, dry heather > grassland, p < 0.01) over the years and, in consequence, this led to a greater decline in overall beetle activity. Moreover, most of the species strictly living in heathland habitats experienced a shrinkage in their activity period that was significantly greater than for those living in grasslands or that are habitat generalists. With a decline in specialist species associated with this vegetation type the ecological links between species can disappear (Clavel et al., 2010), through cascading effects threatening the assemblage typical to this habitat (Parmesan, 2006). 4.3. Relationship between phenology and abundance trends Changing phenology can be a sign of populations adapting to new environmental conditions. Our prediction was that species with greater capacity to change their phenology may be more successful than those that have more restricted emergence times. Although the activity-density measure we used (LM) was statistically independent from all phenological measures in our study, declining trends were shown to be reciprocally correlated with advancing initiation of activity and the length of winter inactivity period. Moreover, declining trends correlated with earlier cessation of activity, giving a significantly narrowed temporal window of activity. Insects can benefit from advancing phenology in several ways. If other environmental conditions allow them to take advantage of the earlier warm days they can start mating and ovipositing earlier in the year, thus extending this most important period. With a longer season a higher proportion of the population may be able to reproduce, and there may be a greater chance to take on sufficient resources to survive the winter and hibernation. Better physical condition before hibernation may substantially increase the likelihood of a successful mating season in the following year. This is particularly important for those species living for longer than one year. A percentage of the population of some carabid species

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can live for up to four years as adults (Bérces and Elek, 2013; Luff, 1973) thus increasing an individual’s lifetime fecundity. However, whilst very little is known about the ageing of beetles, it is plausible to expect that individuals with more time to feed before hibernation have lower mortality rates and are thus more likely to reproduce again. Also, there are several reported examples of insects changing their voltinism with warming climate, specifically from univoltine to bivoltine (Altermatt, 2010; Jonsson et al., 2009, 2011). An extra generation per year can be highly advantageous for a population of a species and can mean a significant increase in its importance within an assemblage (e.g. Altermatt, 2010; Kingsolver et al., 2011). Altered phenology can also have a negative effect on populations. Emerging on early warm days could considerably increase the risk of a later cold stress if the weather changes (Bale and Hayward, 2010). Fluctuating extreme spring weather conditions, as are forecast in climate change scenarios (Easterling et al., 2000; Beniston et al., 2007), can cause high mortality, thus seriously lowering the fitness of populations where a high proportion of individuals choose to become active earlier (Frederiksen et al., 2008; Parmesan et al., 2000). Availability of food also can be crucial for survival. Asynchrony in timing with main prey populations may result in earlier emerging individuals being unable to feed (Parmesan, 2006; Singer and Parmesan, 2010). However, where possible, changing to another prey species (or group) may compensate the negative effect and benefits may thus outweigh the costs (Ikeda et al., 2010; Loreau, 1989). Highly specialized species are clearly in a disadvantaged position in this case. Due to these reasons, beetles that start to become active earlier in order to benefit from an early mating season may be in a disadvantaged position compared to those that did not become active until later. Thus, in a population, a balance between individuals taking and not taking the risk of early emergence is likely to be sustained by evolutionary processes for a longer period with only a slight advancing trend, as was suggested by the changes in PA and WM values in our study. From another perspective, ‘normal’ or later emergence with earlier cessation of activity shortens the temporal window within which species feed, mate, oviposit or prepare for hibernation. Hence, a shortened activity period can cause a significant decrease in the populations’ fitness, thus leading to an overall decline in population size. Yet, major drivers, either abiotic or biotic, causing the decrease in the length of activity are still unclear. There are obvious environmental and physiological limits to advancement of activity as adaption to a changing environment. Limits in phenotypic plasticity can limit behavioural responses to climatic change (Visser, 2008). If individuals of a species can reach the emerging threshold in accumulated day degrees they have less time for feeding and growing. Hence beetles whose larvae develop for a shorter period will be smaller in body size (Kingsolver and Huey, 2008; Kingsolver et al., 2011). The correlation between LM and WI in our study suggests that shortened winter inactivity can be beneficial to carabid populations, but the effect on body size should be further investigated. It is still open to question as to whether phenological changes described here are merely examples of phenotypic plasticity or results of microevolutionary processes. It is likely that factors that fluctuate with temperature, even if the change is gradual, will not rapidly drive hereditable changes in genotypes of species with a relatively long generation time. However, Reale et al. (2003) reported evolutionary responses of red squirrels to climate change in as short as a ten year period whilst Anderson et al. (2008) also revealed phenotypic changes over 30 years in the wing morphology of Arctia caja (Lepidoptera) which are likely to be driven by microevolutionary processes.

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4.4. Implications for research and conservation Several insect taxa, in particular those from within some of the more conspicuous groups, including carabids, are used as ecological indicators in conservation planning and assessing the impacts of climate change (Henle et al., 2013; Pearce and Venier, 2006). Although phenology is greatly neglected in these studies, we strongly believe that changing phenology can be just as important as abundance trends in explaining stability or change in the status of such species. A better understanding of the underlying processes of such impacts is necessary for generating more precise estimations of likely impacts as a response to future scenarios. Furthermore, phenological changes can be assessed using only inventory data, even when detailed records on abundances are not available. Hence, more research needs to be done on changing phenologies of plants and animals to assess their adaptive capabilities and the possible role of such capabilities in maintaining population densities of species in a changing environment. Moreover, McClure et al. (2011) have already raised the problem that advancing phenology can influence the results of year by year comparison of abundances if sampling was carried out at the same time each year (normally, when the main activity of the investigated taxa is expected). The direct and indirect effects of changing phenology are still unclear and whether such responses increase or decrease the fitness and reproduction of the individual may depend on several factors. Hence, awareness of whether particular species are prone to phenological changes is crucial in designing ecological studies. In future analyses, in particular when designing long-term monitoring projects, it may also be worthwhile including abundance measures that are independent of the length of period when a species is present. The issue of whether a population of a species is able to adapt to changing conditions locally or has to shift its geographical distribution is also important in nature conservation (McCarty, 2001). At our sites we found reciprocal correlations between LM values and MVs suggesting that species starting their activity earlier in a year reproduced successfully, thus helping them to maintain a sustainable population level. Attention should be paid to species showing no changes either in activity-densities or phenological measures. In our study there were also signs of trend-like changes in various phenological measures at a lower significance level (0.1 < p < 0.3, Supplementary material 2).The lack of mathematically-significant change does not necessarily mean that those populations are less sensitive to climate change. A static phenology can be caused by the inability to adapt to the fast changing environment. For eurytopic species it is possible that the changes have not yet reached the tolerance threshold, that they are still in the lag stage after the environmental change (Menendez et al., 2006) or that they are already in decline but fluctuations of previous years mask the process, hence the lack of statistical significance. Although Pozsgai and Littlewood (2011) investigated the interactions between phenological change and environmental factors, this question need further investigation to produce a deeper insight into the processes driving phenological adaptations and their correlation with climate change. 5. Conclusions In our multi-species approach, we used the most abundant species in upland communities of ground beetles to investigate the presence and correlation of trends in abundances and phenology. Several species declined over the 18 year study period. Heathland specialists exhibited a higher rate of decline than those typical of grassland habitats. Declines showed a strong relationship with a narrowing window of activity, mainly caused by earlier cessation of activity. In contrast, earlier initiation of activity seemed to be advantageous for numerous species. Since phenological changes

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