From degradation to diversity – Insights and methods for restoring semi-natural grasslands in Central Europe

From degradation to diversity Insights and methods for restoring semi-natural grasslands in Central Europe FELIX HELBING First supervisor: Prof. Dr. Thomas Fartmann Second supervisor: Prof. Dr. Stefan Zerbe Date of disputation: 2 February 2026 From degradation to diversity Insights and methods for restoring semi-natural grasslands in Central Europe A thesis submitted for the degree of Doctor rerum naturalium Osnabrück University School of Biology/Chemistry Division of Biodiversity and Landscape Ecology submitted by FELIX HELBING Osnabrück | 2025 Contact Felix Helbing Osnabrück University Division of Biodiversity and Landscape Ecology Barbarastraße 11, 49076 Osnabrück (Germany)  [email protected] Photos on front and back cover Front (from left to right): Red cattle grazing on a calcareous grassland under restoration, Stachys germanica and Doratura horvathi. Back: Restored montane grassland with Hypochaeris radicata and Arnica montana. Photos on chapter opening pages Chapter I: Phyteuma nigrum. Chapter II: Control of shrub encroachment by mulching in a calcareous grassland. Chapter III: Arthropod sampling with a suction sampler (photo: M. Drung). Chapter IV: Stenobothrus lineatus. Unless otherwise stated, all photos in this thesis are © Felix Helbing Contents Chapter I General introduction ............................................................................................................................................... 1 Chapter II Restoration of semi-natural grasslands from former shrub and forest stages ............................................. 19 Paper 1 Restoration of calcareous grasslands: The early successional stage promotes biodiversity Poniatowski, D., Stuhldreher, G., Helbing, F., Hamer, U., Fartmann, T. Ecological Engineering 151 (2020), 105858 ..................................................................................................... 20 Paper 2 Restoration measures foster biodiversity of important primary consumers within calcareous grasslands Helbing, F., Fartmann, T., Poniatowski, D. Biological Conservation 256 (2021), 109058 ................................................................................................... 22 Paper 3 Rapid response of vascular plants and insects to restoration of montane grasslands Helbing, F., Fartmann, T., Morkel, C., Poniatowski, D. Frontiers in Ecology and Evolution 11 (2023), 1148266................................................................................... 24 Chapter III Methods for arthropod translocation and monitoring ...................................................................................... 53 Paper 4 Suction samplers are a valuable tool to sample arthropod assemblages for conservation translocation Helbing, F., Fartmann, T., Poniatowski, D. Entomologia Experimentalis et Applicata 168 (2020), 688–694 .................................................................... 54 Paper 5 A box quadrat for standardised sampling of Orthoptera in open habitats: Design, handling, applications and baseline data Fartmann, T., Freienstein, F. M., Helbing, F., Scherer, G., Poniatowski, D. Global Ecology and Conservation 55 (2024), e03217 .................................................................................... 62 Chapter IV Synthesis and perspectives ................................................................................................................................. 81 Acknowledgements ............................................................................................................................................. 90 Curriculum vitae .................................................................................................................................................... 91 Publications and conference contributions ...................................................................................................... 92 Chapter I General introduction CHAPTER I Fartmann, 2006; leafhoppers: Helbing and Poniatowski, 2015; true bugs: Morkel et al., 2018). In addition to the Diemel Valley, the study area includes parts of the neighbouring Rothaar Mountains, a low mountain range at the northwestern edge of the German uplands. The cooler, more humid climate compared to the Diemel Valley, combined with oligotrophic soils on acidic bedrock and a long history of low-intensity land use, has fostered the development of species-rich montane grasslands (Fartmann et al., 2022a; Geologisches Landesamt Nordrhein-Westfalen, 1998). However, with land-use change, large areas of montane grasslands have been converted into Norway spruce forests (Picea abies), leading to a strong decrease in grassland species. Between 2011 and 2016, an EU LIFE project was implemented to restore speciesrich montane grasslands from former spruce forests in the study area. Outline of the thesis The thesis comprises five scientific papers, structured into Chapters II and III. Chapter II focuses on the evaluation of the restoration of calcareous (Papers 1 and 2) and montane grasslands (Paper 3) from former shrub and forest stages. In the first study (Paper 1), phytodiversity was examined in detail by comparing the early successional stage of restored calcareous grasslands with both extant, regularly managed sites and abandoned shrubby calcareous grasslands in the Diemel Valley. To capture a wide range of plant species with potentially high conservation value, species characteristic of calcareous grasslands, along with thermophilous fringe and ruderal species, were included in the study. In addition, the functional role of plants for insects was explored. The second study (Paper 2) provides a thorough analysis of the role of the early successional stage for leafhoppers. In Paper 3, vascular plants and three insect taxa (leafhoppers, true bugs, and grasshoppers) were used to evaluate the restoration of mountain hay meadows and Nardus grasslands on former spruce forests in the Rothaar Mountains. Species richness, diversity, and composition of the restoration sites were compared to nutrient-poor pastures with a long-term grazing history. Another aim of the three studies (Papers 1-3) was to apply multifactorial models to assess the effects of environmental factors on the biodiversity of the restored sites. Based on the results, recommendations are provided for the selection of future restoration sites and habitat management. Chapter III is dedicated to methods for arthropod translocation and monitoring. The suitability of suction samplers as a translocation tool was assessed in calcareous grasslands by comparing the mortality rates of sampled arthropods (leafhoppers, spiders, beetles, and true bugs) across different storage periods and conditions (Paper IV). The selected arthropod taxa are considered to be the most dominant when using a suction sampler and therefore have the greatest potential for successful translocation (Brook et al., 2008; Sanders and Entling, 2011). Paper V presents detailed instructions on the design of a box quadrat for sampling grasshoppers, along with a precise description of the workflow in the field. Additionally, baseline data on grasshopper species richness and abundance for open habitats in Germany are provided. These data can be used by biodiversity scientists and nature conservationists to aid in the interpretation of monitoring results obtained through box quadrat sampling. 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Conserv. 15, 275–294. doi: 10.1007/s10531-0048231-7 CHAPTER I Species-rich calcareous grassland in the Diemel Valley, with Galium verum, Helianthemum nummularium, Prunella grandiflora and Trifolium montanum. The Diemel Valley contains the largest area of semi-dry calcareous grasslands in the northern part of Germany. (Udorf, 06/2019) Semi-natural grasslands are of outstanding value for nature conservation. However, their extent has declined dramatically in recent decades, mainly due to agricultural intensification and the abandonment of traditional land use. Today, the remaining grasslands are largely protected under the EU Habitats Directive. The photo shows species-rich montane grasslands in the Rothaar Mountains, interspersed with spruce forests. (Winterberg, 05/2018) Species-rich calcareous grassland in the Diemel Valley, with Galium verum, Helianthemum nummularium, Prunella grandiflora and Trifolium montanum. The Diemel Valley contains the largest area of semi-dry calcareous grasslands in the northern part of Germany. (Udorf, 06/2019) GENERAL INTRODUCTION The abandonment of semi-natural grasslands initiates a gradual process of habitat change, with detrimental impacts on biodiversity in the long-term. With increasing succession, the vegetation becomes dominated by a few competitive grass and shrub species suppressing less-competitive, light-demanding plants and thermophilic animals. (‘Wulsenberg’ near Marsberg, 06/2019) At the time the photo was taken, the right rear part of this slope in the Diemel Valley (‘Opferberg’ near Lamerden) was completely overgrown with shrubs after decades of abandonment. In contrast, the left front part (‘Schwiemelkopf’) had already been restored in 2014 through shrub removal and the reintroduction of low-intensity goat grazing. (06/2020) CHAPTER II 1 Introduction For centuries, European landscapes were shaped by traditional land-use practices (Plieninger et al., 2006; Eriksson, 2021; Fartmann, 2023). Small-scale low-intensity farming created a wide variety of interconnected habitats, considerably fostering biodiversity (Plieninger et al., 2006; Halada et al., 2011). With the onset of industrialisation, however, drastic changes in land use were ushered in throughout Europe, accelerating from the mid-20th century onwards (Fartmann, 2023). A wide range of organizational and technological innovations, such as land consolidation instruments, progressive mechanisation and the increasing use of artificial fertilisers, enabled large-scale intensification of agriculture on productive soils (Stoate et al., 2001; Robinson and Sutherland, 2002; Kleijn et al., 2009). By contrast, marginal land that became increasingly unviable for agricultural use was often abandoned or afforested (Henle et al., 2008; Fartmann, 2023). The transformation from traditional to industrial agricultural landscapes is today considered the Paper 3 · Frontiers in Ecology and Evolution 11 (2023), 1148266, doi: 10.3389/fevo.2023.1148266 Rapid response of vascular plants and insects to restoration of montane grasslands Felix Helbing, Thomas Fartmann, Carsten Morkel, Dominik Poniatowski ABSTRACT Introduction: Industrialisation has ushered in massive changes in agriculture. Particularly in low mountain ranges, large-scale afforestation with Norway spruce on traditionally managed, semi-natural grasslands has caused a severe decline in biodiversity. Tree removal, hay transfer and resumption of grazing or mowing are typical measures to re-create species-rich grasslands. The aim of this study was to use vascular plants and three insect taxa (leafhoppers, true bugs, and grasshoppers) as bioindicators to evaluate the success of montane grassland restoration on former spruce forests in Central Europe. In addition, we intended to identify the drivers of species richness within the studied grasslands in order to derive suitable recommendations for habitat management. Methods: We analysed two different treatments: (i) grazed restoration sites where trees had been cut and species-rich green hay had been applied (n = 9) and (ii) target sites with a long continuity of low-intensity grazing (n = 9). Results and Discussion: Our study revealed that all studied taxa responded rapidly to the restoration measures. After a development period of three to five years, we found no differences in species richness and diversity of leafhoppers, true bugs and grasshoppers (all, target and threatened species). In addition, non-metric multidimensional scaling showed a large overlap in species composition between restoration and target grasslands. Among target and threatened species, vascular plants displayed the same pattern as insects and reached similar values when comparing the two treatments. However, total species richness and diversity of vascular plants were still higher on the target sites and species composition overlapped only partially. Grazing intensity was the predictor with the highest explanatory power in multivariable (Generalized) Linear Mixed-effects Models, being negatively related to species richness of leafhoppers and true bugs. We conclude that the measures implemented were effective in re-establishing target communities of different taxa. The transfer of seed-containing hay enabled or accelerated the development of the vegetation. Insects, on the other hand, were able to recolonise the restored grasslands on their own, given that these sites were embedded in a network of species-rich grasslands. With regard to insects (e.g., leafhoppers and true bugs), it should be ensured that grazing is applied at low intensity. KEYWORDS Auchenorrhyncha Green hay transfer Habitat quality Heteroptera Host plant Nature conservation Orthoptera Picea abies RESTORATION OF SEMI-NATURAL GRASSLANDS major driver of the dramatic decrease in farmland biodiversity across Europe (Plieninger et al., 2006; Pereira et al., 2012; Fartmann, 2023). Temperate, semi-natural grasslands, arising from traditional management, are very rich in plant and animal species and therefore have an outstanding value for nature conservation (Bonari et al., 2017; Feurdean et al., 2018; Fartmann et al., 2021). However, for decades, they have suffered particularly from persistent pressure exerted by land-use change, which has resulted in continuous habitat loss, deterioration and fragmentation (Isselstein et al., 2005; Fartmann et al., 2021; Fartmann, 2023). The contrasting drivers of this development have been unevenly distributed along elevation gradients: The conversion of semi-natural grassland to cropland and the intensification of management have been more pronounced in lowlands, whereas spontaneous (i.e., succession after abandonment) and artificial afforestation on former grasslands have especially taken place at higher elevations (Mottet et al., 2006; Fartmann, 2023). In Central European low mountain ranges, montane grasslands in particular have been widely planted with introduced Norway spruce (Picea abies; Hölzel and Tischew, 2019). This has led to rapid changes in habitat conditions. Following site preparation, trees severely alter light and soil conditions as they grow, hampering the development of ground vegetation (Strand et al., 2021). Grassland plant and animal species not only lose habitat but are also hindered in dispersal between remaining habitat patches (Poniatowski et al., 2016). Consequently, the conversion of grasslands to shady high forests, primarily intended to produce firewood and timber, has caused a strong decline in montane species, and semi-natural montane grasslands are seriously threatened today (Finck et al., 2017; Hölzel and Tischew, 2019). Besides preserving and improving extant fragments of semi-natural montane grasslands, habitat restoration is a key component of nature conservation for increasing biodiversity. In the case of spruce forests, restoration measures include clear-cutting and subsequent tillage. However, the soil seed bank is often depleted after decades of forest use and spontaneous recolonisation by dispersing propagules is limited for many plant species (Bakker et al., 2002; Kiehl et al., 2010). Target species are therefore actively introduced by transferring green hay or brush-harvested seeds from nearby species-rich donor grasslands to restoration sites (Kiehl et al., 2010; Durbecq et al., 2022). However, the success of implemented measures varies between restoration projects as it depends on several factors, such as timing, species composition of donor sites, soil conditions and effectiveness of reintroduced habitat management (Kiehl et al., 2010; Wagner et al., 2020). A careful evaluation of habitat restoration is thus essential. Contrary to common practice, evaluation studies should ideally focus not only on plants but also include animals, as otherwise there is a risk of overestimating restoration success (Baur, 2014; McAlpine et al., 2016). Although plant recovery is an important prerequisite for animal recolonisation, sites may lack suitable habitat quality in terms of microclimate or vegetation structure (Jones and Davidson, 2016; Helbing et al., 2021). Moreover, dispersal-limited species often fail to reach restoration sites spontaneously, especially when dispersal corridors are limited (Baur, 2014; Poniatowski et al., 2016). This applies in particular to arthropods that are major drivers of ecosystem functioning in semi-natural grasslands (Soliveres et al., 2016; Poniatowski et al., 2018; Neff et al., 2020). Studies evaluating both the flora and arthropod fauna of restored montane grasslands, though, are still rare (but see Kurtogullari et al., 2020; Neff et al., 2020). In addition, research to date has tended to focus on the restoration of montane grassland habitats in open areas (e.g., Lencová and Prach, 2011; Sullivan et al., 2020; Durbecq et al., 2022), whereas restoration projects that involve clearing spruce forests are still understudied (but see Dassonville et al., 2013). The objective of this multi-taxon study was to use vascular plants and three insect taxa (leafhoppers [Hemiptera: Auchenorrhyncha], true bugs [Hemiptera: Heteroptera] and grasshoppers [Orthoptera]) as bioindicators to evaluate the success of montane grassland restoration on former spruce forests in a Central European low mountain range. We compared species richness, diversity and composition of grazed restoration sites and nutrient-poor pastures with long-term grazing history (target grassland). The studied insect taxa (i) play an important role both as consumers and as an abundant food source for insectivores in grassland ecosystems, (ii) occupy different ecological niches, (iii) respond quickly to habitat changes and (iv) allow standardised sampling (Kőrösi et al., 2012; Helbing et al., 2020; Fartmann et al., 2022c). With a development period of three to five years after the implemented measures, we studied the short-term effects of restoration. In particular, we addressed the following questions: (i) Do environmental conditions in restored grasslands differ from those in target grasslands? (ii) What are the effects of restoration measures on species richness and diversity of vascular plants CHAPTER II and insects (all, target and threatened species), and how do species compositions differ between both types of studied grasslands? (iii) Which environmental parameters explain species richness of vascular plants and insects within the studied grasslands, and, on the basis of the results, which recommendations for further habitat management can be given? 2 Material and methods 2.1 Study area The study area is located in the Rothaar Mountains, a low mountain range at the northwestern edge of the German uplands (51°11'17 N, 8°32'25 E; Fig. 1). It covers an elevation range from 450 to 842 m a.s.l. and is characterised by a cool (mean annual temperature: 5 °C) and humid (mean annual precipitation: 1,450 mm) montane climate with snowy winters (mean snowcover duration: 100 days/year) (Fartmann et al., 2022b). Oligotrophic cambisols on acidic bedrock are the prevalent soils (Geologisches Landesamt Nordrhein-Westfalen (NRW), 1998). Interspersed with grasslands, the landscape is dominated by Norway spruce forests (Picea abies), which are non-native to the study region. Although agriculture has been partially intensified here as well, montane grasslands that are grazed or mown at low intensity are still widespread. These comprise mountain hay meadows (order: Arrhenatheretalia, alliance: Polygono-Trisetion) and species-rich Nardus grasslands (order: Nardetalia strictae, alliance: Violion caninae), both of which are protected under the EU Habitats Directive (92/43/EEC, habitat types 6520 and 6230). Within the study area, a large-scale EU LIFE project (‘Mountain hay meadows near Winterberg’) was carried out from 2011 to 2016. It aimed to restore, improve and protect mountain hay meadows and species-rich Nardus grasslands. The project measures included the conversion of spruce forests into speciesrich grasslands (Schulte, 2017). After the conduction of the restoration measures, the grasslands have been either grazed or mown at low intensity. In this study, we only considered grazed grasslands that were restored on former spruce forests. 2.2 Restoration methods The restoration measures were conducted three to five years prior to our study (i.e., between 2012 and 2014) and are described in detail in Schulte (2017). Overall, the restoration comprised four steps: (i) deforestation of spruce forests, (ii) removal of remaining branches, (iii) forestry mulching, and (iv) transfer of autochthonous green hay from adjacent donor sites. Only in exceptional cases was the soil Figure 1 Location of the study area and study sites in Germany. RESTORATION OF SEMI-NATURAL GRASSLANDS additionally milled and harrowed. Long-existing Nardus grasslands were rarely accessible by heavy machinery due to topographic and edaphic constraints. Thus, oligotrophic hay meadows with a high number of typical Nardus grassland species were mostly chosen as donor sites. For the green hay transfer, these were mown once in early to mid-August of the respective year with a rotary disc mower. The hay was then transported and spread onto the recipient sites using a forage wagon (application rate of 1:1). Only on steeper slopes was spreading by hand necessary. 2.3 Study sites and plots The effectiveness of the restoration measures was evaluated by a comparative approach. We analysed two treatments: (i) Montane grasslands where restoration measures described above were carried out and which have since been managed as rough pastures (restored grassland, n = 9, mean size: 0.66 ± 0.11 ha, Fig. 2a, b). (ii) Species-rich pastures with a long continuity of low-intensity grazing. These include Nardus grasslands and closely related oligotrophic grassland communities. The pastures are characterised by a high phytodiversity and represent the target state of the restoration of grazed grasslands in the study area (target grassland, n = 9, mean size: 1.69 ± 0.39 ha, Fig. 2c, d). The majority of the sites were grazed by cattle (61%), followed by horses (28%) and goats (11%). In the centre of each site, we established a rectangular plot with a total size of 500 m² and a distance of at least 10 m to the site boundary (to avoid edge effects) (cf. Kurtogullari et al., 2020). The plots were marked with magnets and tent pegs, which were placed in the ground together. With the help of a magnetic detector, the locations were accurately retrieved during subsequent visits. In order to address potential issues with spatial autocorrelation, the study area was divided into five subareas, each comprising at least one restored grassland site (Fig. 1, see also Section 2.6: LMM and Figure 2 Impressions of the study area: (a, b) aerial views with restored grasslands framed in red (photos: F. Helbing), (c) species-rich pasture with long-term grazing history inhabited by Geranium sylvaticum (photo D. Poniatowski) and (d) long-term pasture grazed by red cattle (photo D. Poniatowski). CHAPTER II GLMM with random effects). We attempted to balance the number of sites of the two treatments within each subarea. However, due to the limited availability of suitable pastures, this was not feasible in subarea IV. We therefore increased the number of studied target grasslands in the adjacent subarea V. 2.4 Species sampling and classification We sampled vascular plants and insects in 2017. Insect sampling was always conducted under favorable weather conditions (dry and sunny days with temperatures > 15 °C). Depending on the phenology of each species group, the period and frequency of the surveys varied. In order to evaluate the effects of grassland restoration, we classified three groups of vascular plants and insects: all, target and threatened species. Criteria for defining target species were determined individually for each species group and are described below. If available, we obtained the threat status of the species from the latest red data books of the federal state North Rhine-Westphalia (vascular plants, grasshoppers), otherwise red data books of Germany (leafhoppers, true bugs) were used. Hereinafter, species that are listed as ‘critically endangered’, ‘endangered’, ‘vulnerable’ or ‘near threatened’ (red-list categories: 1, 2, 3 and V) are termed ‘threatened species’. Literature on species identification, scientific nomenclature and the threat status of the studied taxa is listed in the supplementary material. 2.4.1 Vascular plants We established a subplot covering 16 m² (4 m × 4 m) in the centre of each plot to record vascular plant species richness and cover. The main vegetation survey was carried out in early June. In early August, we revisited the subplots to search for late-flowering plants such as eyebright (Euphrasia spp.). Species cover was estimated according to the Wilmanns scale (Wilmanns, 1998). For statistical analyses, the categories were translated into percentage-cover values: r = 0.1%, + = 0.5%, 1 = 2.5%, 2m = 2.5%, 2a = 10%, 2b = 20.5%, 3 = 38%, 4 = 63% and 5 = 88% (Dierschke, 1994). In addition to the quantitative data gathered within the subplots, we searched the 500 m² plots in their entirety for additional plant species during both surveys. From the resulting species list, we determined the host plants of sampled leafhoppers (Nickel and Remane, 2002; Nickel, 2003) and true bugs (Wachmann et al., 2004, 2006, 2007, 2008) and used their counts as explanatory variables (see Section 2.6). Both the donor and target sites were characterised by plant species typical of Nardus grasslands as well as mountain hay meadows. Thus, we defined all characteristic species of the phytosociological class Nardo-Callunetea, the alliance Polygono-Trisetion and their respective subunits as target species (Dierschke, 1997; Oberdorfer, 2001; Peppler-Lisbach and Petersen, 2001) (Supplementary Table 1). 2.4.2 Leafhoppers and true bugs Leafhoppers and true bugs were surveyed together three times in early June, mid-July and early September. We combined a sweep net (30 cm diameter) and a G-Vac suction apparatus (Stihl SH 56, Waiblingen, Germany; 12.5 cm diameter suction tube; 710 m3 h-1 air flow rate) with a fine gauze collection bag (300 μm mesh size) on the inner side of the inlet nozzle to cover all available microhabitats and sample the complete range of herband ground-dwelling species (cf. Stewart, 2002; Helbing et al., 2020). We walked in loops over the whole plot and randomly performed 50 sweep net strokes and 50 suction samples. The latter was done by holding the nozzle onto the ground for 10 s (cf. Helbing et al., 2021). The catches were killed with ethyl acetate and identified in the laboratory using a digital microscope (leafhoppers: Keyence VHX-900F, Osaka, Japan) or binoculars (true bugs: Helmut Hund 7–90×, Wetzlar, Germany). Data from both sampling techniques were pooled for statistical analyses. For some genera, identification of females at the species level is not possible (e.g., Aphrodes, Psammotettix, Ribautodelphax). These females were identified at the genus level or, if present, assigned to the corresponding males. In the case of more than one species of a genus, the number of females was assumed to correspond to the number of males (cf. Helbing et al., 2021). We defined typical inhabitants of semi-natural grasslands as target species. For leafhoppers, we listed all sampled grassland species that avoid intensive to very intensive management, based on Nickel and Achtziger (1999). From this list, we then removed species that occur exclusively on unmanaged or wet grasslands, as well as pioneer species (Nickel and Achtziger, 1999; Nickel, 2003) (Supplementary Table 2). True bug target species were classified according to Wachmann et al. (2004, 2006, 2007, 2008). We included all species that primarily occur on grasslands with a low management intensity except for typical inhabitants of fringe and wetland habitats as well as pioneer species (Supplementary Table 3). 2.4.3 Grasshoppers Grasshopper sampling took place once per plot in early August. We used a box quadrat (1.41 × 1.41 m ≙ RESTORATION OF SEMI-NATURAL GRASSLANDS 2.0 m²), which is a well-approved and accurate tool for surveying species richness and abundance of grasshoppers (Gardiner and Hill, 2006). In each plot, we randomly dropped the box quadrat at 10 different points, covering a total area of 20 m² (cf. Fartmann et al., 2022a). Species were identified in the field and released afterwards. As with leafhoppers and true bugs, we intended to classify grasshopper species typical of semi-natural grasslands as target species. Poniatowski et al. (2020) presented a species farmland index (SFI) for German grasshopper species that measures the average availability of high nature value farmland (HNV) within a species' range. HNV is a term used to describe biodiversity-rich farming systems, usually resulting from low-intensity traditional landuse practices (Lomba et al., 2014). Habitat specialists that strongly depend on open, semi-natural habitats receive high SFI values, whereas values for habitat generalists are low. In this study, we classified all species with an SFI value of at least 17.0 as target species (Supplementary Table 4). 2.5 Environmental conditions 2.5.1 Local climate We determined the mean elevation (m a.s.l.) of the study sites from topographic maps using the geographical information system ArcGIS 10.6. Moreover, long-term averages of annual temperature (°C) and precipitation (mm) (period 1981–2010) were derived from grid maps with a spatial resolution of 1 km2 (German Meteorological Service, 2022). If sites covered more than one grid cell, we calculated mean values. During fieldwork, we measured aspect and slope of the plots. In combination with the latitude of the study area (51°), these data were used to calculate the potential amount of direct incident radiation (in MJ cm−2 yr−1) as an approximate measure of the local climate (McCune and Keon, 2002). 2.5.2 Habitat characteristics We calculated the size of the study sites (ha) using orthophotos in ArcGIS 10.6. The intensity of grazing was determined on a metric scale. Combining own observations and information provided by farmers, we recorded the composition of grazing livestock and the duration of grazing (with an accuracy of 0.5 months) for each site during the 2017 season. Each type of animal (species and age) was converted into a livestock unit (LSU) using a conversion key for North RhineWestphalia (Table 1; LANUV, 2022). The grazing intensity of a site was calculated as the product of the density of livestock (summed LSU per hectare) and the months of grazing (MoG). In the course of the main vegetation survey in June 2017, we recorded parameters of horizontal and vertical habitat structures within each subplot. We estimated the cover of grasses, herbs, cryptogams, bare soil and litter with an accuracy of 5%. If values were above 95% or below 5%, we used steps of 2.5% (Helbing et al., 2021). The vegetation height of the field layer was measured with an accuracy of 2.5 cm. To characterise soil conditions, we also calculated Ellenberg’s indicator values for vascular plants (Ellenberg et al., 2001) (mean values for moisture, nitrogen and soil reaction). 2.6 Statistical analysis 2.6.1 Comparison of habitat types All statistical analyses were performed using R 4.1.2 (R Core Team, 2022). We quantified the taxonomic richness and diversity of vascular plants and insects using Hill numbers, which provide a mathematically unified set of diversity indices in units of species (Hill, 1973; Jost, 2006; Chao et al., 2014). The sensitivity of these indices to species relative abundances is controlled by the order q (Chao et al., 2014). We considered Hill numbers at q = 0, which represents species richness, and at q = 1, which is equivalent to the exponential of Shannon entropy (hereinafter termed ‘diversity‘) (Jost, 2006; Chao et al., 2014). We tested for significant differences in environmental conditions, species richness and diversity between restored and target grasslands. For the former, we used the Mann-Whitney U test since the majority of environmental variables showed highly skewed distributions and could thus not be analysed by mixed-effects models (cf. Helbing et al., 2021). By contrast, species richness and diversity of vascular plants, leafhoppers, true bugs and grasshoppers (all, target and threatened species) were compared via Linear Mixed-effects Models (LMM) or Generalized Linear Mixedeffects Models (GLMM) with habitat type as a nominal predictor and subarea (n = 5; see Section 2.3) as a random intercept (R package ‘lme4’; Bates et al., 2022). For normally distributed and log-transformed variables with a normal distribution, we applied LMMs, Table 1 Conversion key for the calculation of livestock units per site (LANUV [Landesamt für Natur, Umwelt und Verbraucherschutz Nordrhein-Westfalen], 2022). Animal type Livestock unit Cattle < 6 months of age 0.4 Cattle 6 months − 2 years of age 0.6 Cattle > 2 years of age 1 Horse > 6 months of age 1 Goat 0.15 CHAPTER II otherwise we calculated GLMMs with a negative-binomial error structure to account for overdispersion (Richards, 2008). Species compositions of restored and target grasslands were compared via non-metric multidimensional scaling (NMDS; R package ‘vegan’, metaMDS function; Oksanen et al., 2022) with the Bray-Curtis distance as a distance measure and a maximum number of 200 random starts in search of a stable solution. The four studied species groups were analysed separately. We included only those species in the analyses that were sampled on at least three plots. For each species group, we performed a Permutational Multivariate Analysis of Variance with 999 permutations (PERMANOVA; Bray-Curtis distance; R package ‘vegan’, ‘adonis2’ function; Oksanen et al., 2022) to test for significant differences between restored and target grasslands. 2.6.2 Effects of environmental conditions on species richness of vascular plants and insects We fitted (G)LMMs with the same random effects and error structure as described above to analyse the effects of environmental parameters on species richness of vascular plants and insects (target and threatened species). For each of the studied species groups, we included a specific and ecologically reasonable set of variables into the models (e.g., the number of leafhopper host plant species was used only to analyse leafhopper species richness, whereas Ellenberg indicator values characterising soil conditions were included only in the modelling of vascular plant species richness). A detailed assignment of the parameters to the species groups is given in Table 2. Prior to the analyses, we standardised all fixed effects to achieve comparability of effect sizes (cf. Fumy and Fartmann, 2021). Furthermore, we graphically examined all combinations of response and predictor variables before computing the (G)LMMs. If we detected unimodal rather than linear relationships between two variables, we also entered centred and squared values of the predictor into the full model in addition to the untransformed values (cf. Kämpfer and Fartmann, 2022). This was the case for the Ellenberg nitrogen value and threatened vascular plant species. To avoid multicollinearity, we also calculated Spearman’s rank correlation (rS) for all combinations of predictor variables (Dormann et al., 2013). In cases where two or more variables were strongly intercorrelated (|rS| ≥ 0.6, P < 0.05), only one was included in statistical modelling (Supplementary Table 5). We applied model averaging based on an information-theoretic approach for all multivariable analyses (Grueber et al., 2011). In a first step, all possible combinations of environmental parameters were tested, resulting in a set of different candidate models. These candidate models were then ranked based on Akaike’s information criterion (AICC) using the ‘dredge’ function (R package ‘MuMIn’; Bartoń, 2022). Finally, the top-ranked models within ΔAICC < 2 were used for model averaging (Grueber et al., 2011). We evaluated the explanatory power of the models by calculating marginal R² (variance explained by fixed effects) and conditional R² (variance explained by both fixed and random effects) (Nakagawa et al., 2017). The P values for LMMs were determined with the R package ‘lmerTest’ (Kuznetsova et al., 2022). 3 Results 3.1 Environmental conditions Restored and target grasslands differed only slightly in environmental conditions (Table 2). Local climate, grazing intensity, host plants of leafhoppers and true bugs as well as most parameters of habitat structure did not differ. Restored grasslands were merely smaller, had a lower cover of herbs and lower Ellenberg indicator values for nitrogen and soil reaction. 3.2 Species richness, diversity, and composition of vascular plants and insects We sampled a total of 267 species in the four species groups (Table 3, Supplementary Tables 1–4). Vascular plants and leafhoppers were the most species-rich taxa, followed by true bugs and grasshoppers. On average, 16% of the species per group were classified as threatened (ranging between 8% in true bugs and 29% in grasshoppers) and 34% as target species (ranging between 15% in vascular plants and 58% in leafhoppers). With a frequency of 100% in both studied habitat types, Agrostis capillaris, Festuca rubra agg., Luzula campestris and Rumex acetosa were the most widespread vascular plant species in our study. The most abundant species of leafhoppers were Deltocephalus pulicaris (nindividuals = 4,051; 22%) and Verdanus abdominalis (nindividuals = 3,451; 18%), those of true bugs were Pachytomella parallela (nindividuals = 955; 46%) and Nabis flavomarginatus (nindividuals = 385; 19%), and those of grasshoppers were Chorthippus biguttulus (nindividuals = 512; 45%) and Pseudochorthippus parallelus (nindividuals = 414; 36%). RESTORATION OF SEMI-NATURAL GRASSLANDS When comparing restored and target grasslands, no differences in species richness and diversity of leafhoppers, true bugs and grasshoppers (all, target and threatened species) were found (Fig. 3c–h). This was also true for target and threatened species of vascular plants (Fig. 3a, b). It was only the total species richness and the total species diversity of vascular plants that were higher in target than in restored grasslands (Fig. 3a, b). NMDS revealed similar patterns in terms of species composition. There was a large overlap in leafhoppers (F1,16 = 1.39, P = 0.20) and grasshoppers (F1,16 = 1.13, P = 0.34) between restored and target grasslands (Fig. 4b, d). True bugs also exhibited similarity between the two grassland types, but only for the centrally located target grasslands, while other target grasslands were more dispersed (F1,16 = 4.20, P < 0.01) (Fig. 4c). Vascular plants showed some overlap between restored and target grasslands as well, but overall, they were more segregated from one another (F1,16 = 4.28, P < 0.01) (Fig. 4a). 3.3 Response of vascular plants and insects to environmental conditions The (G)LMM analyses identified effects of environmental parameters on species richness of vascular Table 2 Metric environmental parameters (mean ± standard error [SE] as well as minimum [Min] and maximum [Max] values) in restored (n = 9) and target (n = 9) grasslands. Parameter Restored grassland Target grassland P Used for (G)LMM Mean ± SE Min–Max Mean ± SE Min– Max analyses ofa Local climate Elevation (m a.s.l.) 698 ± 29 (518–775) 660 ± 23 (531–757) n.s. P, L, T, G Average annual temperature (C°) 6.4 ± 0.1 (5.8–7.2) 6.6 ± 0.1 (5.8–7.2) n.s. P, L, T, G Annual precipitation (mm) 1,282 ± 21 (1,147–1,351) 1,261 ± 21 (1,147–1,351) n.s. P, L, T, G Direct incident radiationb (MJ cm-2 yr-1) 0.78 ± 0.04 (0.63–0.91) 0.84 ± 0.04 (0.61–0.94) n.s. P, L, T, G Habitat characteristics Site size (ha) 0.66 ± 0.11 (0.22–1.25) 1.69 ± 0.39 (0.21–3.94) * P, L, T, G Grazing intensityc (LSU MoG ha−1) 3.9 ± 0.8 (0.6–7.8) 5.3 ± 1.1 (0.6–8.9) n.s. P, L, T, G Habitat structure Cover (%) Grasses 74.4 ± 4.5 (55–95) 63.9 ± 4.1 (50–90) n.s. L, T, G Herbs 21.4 ± 4.2 (2.5–40) 40.0 ± 3.8 (15–50) ** L, T, G Cryptogams 58.3 ± 7.8 (20–90) 34.2 ± 11.9 (2.5–90) n.s. L, T, G Bare ground 2.5 ± 1.2 (0–10) 3.6 ± 1.0 (0–10) n.s. L, T, G Litter 40.0 ± 9.0 (15–85) 35.0 ± 7.8 (5–85) n.s. L, T, G Vegetation height (cm) 14.9 ± 2.1 (8–27) 14.4 ± 1.8 (8–22) n.s. L, T, G No. vascular plant species Leafhopper host plants 18.0 ± 1.7 (12–30) 21.6 ± 1.5 (14–29) n.s. L True bug host plants 16.1 ± 1.5 (12–27) 17.2 ± 1.2 (13–23) n.s. T Ellenberg indicator value (unitless) Moisture 5.0 ± 0.1 (4.6–5.8) 5.0 ± 0.1 (4.6–5.3) n.s. P Nitrogen 3.8 ± 0.2 (3.1–4.8) 4.5 ± 0.2 (3.4–5.2) * P Soil reaction 4.0 ± 0.2 (3.1–4.9) 4.8 ± 0.1 (4.3–5.3) * P Differences between the two habitat types were analysed using the Mann-Whitney U test. Statistical significances are indicated as follows: n.s. (not significant), P ≥ 0.05, *P < 0.05, **P < 0.01. a P = Plants, L = Leafhoppers, T = True bugs, G = Grasshoppers. a P = Plants; L = Leafhoppers; T = True bugs; G = Grasshoppers. b Direct incident radiation was calculated according to McCune and Keon (2002). c LSU = Livestock units, MoG = Months of grazing (accuracy: 0.5 months) (for details see Section 2.5.2). Table 3 Overview of the studied taxa with numbers of species and individuals (total, mean ± standard error [SE] as well as minimum [Min] and maximum [Max] values) per taxon (nsites = 18). Taxon Number of species Number of individuals Scientific Aliasa Total Mean ± SE Min–Max Total Mean ± SE Min–Max Tracheophyta Vascular plants 129 30.3 ± 2.2 14–50 – – – Hemiptera: Auchenorrhyncha Leafhoppers 73 25.9 ± 0.9 19–32 18,723 1,042.1 ± 93.9 378–1,811 Hemiptera: Heteroptera True bugs 51 11.3 ± 1.0 4–21 2,073 115.8 ± 14.3 38–255 Orthoptera Grasshoppers 14 4.9 ± 0.3 4–8 1,151 63.6 ± 10.5 10–145 a This term is used throughout the paper. CHAPTER II Figure 3 Mean values (± standard error) of species richness (Hill numbers of order q = 0) and diversity (expressed as exponential Shannon diversity; Hill numbers of order q = 1) for all, target and threatened species of (a, b) vascular plants, (c, d) leafhoppers, (e, f) true bugs and (g, h) grasshoppers in restored (n = 9) and target grasslands (n = 9). Asterisks indicate a significant difference between the two habitat types (tested via mixed-effects models with subarea [n = 5] as a random intercept; *P < 0.05). RESTORATION OF SEMI-NATURAL GRASSLANDS plants, leafhoppers and true bugs. The number of vascular plant target species increased with annual precipitation and decreased with Ellenberg nitrogen values (Table 4a, Fig. 5a, b). By contrast, the number of threatened vascular plant species was highest at intermediate Ellenberg nitrogen values (Table 4b, Fig. 5c). Grazing intensity was the parameter with the highest explanatory power in the multivariable analyses for two of the three insect groups: it negatively affected the species number of leafhoppers (target and threatened species) and true bugs (target species) (Table 4c– e, Fig. 5d, e, h). 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Rothmaler –Exkursionsflora von Deutschland, Gefäßpflanzen: Atlasband, 13th ed. Springer Spektrum, Berlin, Germany. Oberdorfer, E., 2001. Pflanzensoziologische Exkursionsflora für Deutschland und angrenzende Gebiete, 8th ed. Eugen Ulmer, Stuttgart, Germany. Leafhoppers Biedermann, R., Niedringhaus, R., 2004. Die Zikaden Deutschlands – Bestimmungstafeln für alle Arten.: WABV Fründ, Scheeßel, Germany. Holzinger, W. E., Kammerlander, I., Nickel, H., 2003. Die Zikaden Mitteleuropas. Vol. 1: Fulgoromorpha, Cicadomorpha (excl. Cicadellidae). Brill, Leiden, The Netherlands. Kunz, G., Nickel, H., Niedringhaus, R., 2011. Fotoatlas der Zikaden Deutschlands – A photographic atlas of the planthoppers and leafhoppers of Germany. WABV Fründ, Scheeßel, Germany. True bugs Bozdĕchová, J., 1973. Diagnostische Merkmale der Arten Trigonotylus ruficornis und T. coelestialium (Heteroptera, Miridae). Fol. Mus. Rer. Nat. Bohem. Occ. Plze., Zoologica 3: 1–18. Rieger, C., 1985. Zur Systematik und Faunistik der Weichwanzen Orthops kalmi Linné und Orthops basalis Costa (Heteroptera, Miridae). Veröffentl. Natursch. Landschaftspfl. BaWü. 59/60: 457–465. Rinne, V., 1989. Review of the European Polymerus subgenus Poeciloscytus (Heteroptera, Miridae), with two new species and special reference to the Finnish fauna. Ann. Entomol. Fenn. 55: 89–101. Wagner, E., 1952. Blindwanzen oder Miriden. In: Dahl, F., Dahl, M., Bischoff, H. (Eds.), Die Tierwelt Deutschlands und der angrenzenden Meeresteile 41. Gustav Fischer, Jena, Germany, pp. 1–218. Wagner, E., 1966. Wanzen oder Heteropteren - I Pentatomorpha. In: Dahl, F., Dahl, M., Peus, F. (Eds.), Die Tierwelt Deutschlands und der angrenzenden Meeresteile 54, Gustav Fischer, Jena, Germany, pp. 1–235. Wagner, E., 1967. Wanzen oder Heteropteren - II Cimicomorpha. In: Dahl, F., Dahl, M., Peus, F. (Eds.), Die Tierwelt Deutschlands und der angrenzenden Meeresteile 55, Gustav Fischer, Jena, Germany, pp. 1–179. Grasshoppers Fischer, J., Steinlechner, D., Zehm, A., Poniatowski, D., Fartmann, T., Beckmann, A., Stettmer, C., 2016. Die Heuschrecken Deutschlands und Nordtirols: Bestimmen – Beobachten – Schützen. Quelle & Meyer, Wiebelsheim, Germany. CHAPTER II Scientific nomenclature Vascular plants Hand, R., Thieme, M., 2022. Florenliste von Deutschland (Gefäßpflanzen). Version 12. https://www.kpbuttler.de/florenliste/index.htm (accessed 24/01/2022) Leafhoppers Nickel, H., Achtziger, R., Biedermann, R., Bückle, C., Deutschmann, U., Niedringhaus, R., Remane, R., Walter, S., Witsack, W., 2016. Rote Liste und Gesamtartenliste der Zikaden (Hemiptera: Auchenorrhyncha) Deutschlands. Nat. Schutz. Biol. Vielfalt. 70, 247–298. True bugs Aukema, B., Rieger, C. (Eds.), 1995. Catalogue of the Heteroptera of the Palaearctic Region. Enicocephalomorpha, Dipsocoromorpha, Nepomorpha, Gerromorpha and Leptopodomorpha. Neth. Entomol. Soc. 1, 1–222. Aukema, B., Rieger, C. (Eds.), 1996. Catalogue of the Heteroptera of the Palaearctic Region. Cimicomorpha I. Neth. Entomol. Soc. 2, 1–361. Aukema, B., Rieger, C., (Eds.), 1999. Catalogue of the Heteroptera of the Palaearctic Region. Cimicomorpha II. Neth. Entomol. Soc. 3, 1–577. Aukema, B., Rieger, C. (Eds.), 2001. Catalogue of the Heteroptera of the Palaearctic Region. Pentatomomorpha I. Neth. Entomol. Soc. 4, 1–346. Aukema, B., Rieger, C. (Eds.), 2006. Catalogue of the Heteroptera of the Palaearctic Region. Pentatomomorpha II. Neth. Entomol. Soc. 5, 1–550. Aukema, B., Rieger, C., Rabitsch, W. (Eds.), 2013. Catalogue of the Heteroptera of the Palaearctic Region. Supplement. Neth. Entomol. Soc. 6, 1–629. Grasshoppers Fischer, J., Steinlechner, D., Zehm, A., Poniatowski, D., Fartmann, T., Beckmann, A., Stettmer, C., 2020. Die Heuschrecken Deutschlands und Nordtirols: Bestimmen – Beobachten – Schützen, 2nd ed. Quelle & Meyer, Wiebelsheim, Germany. Threat status Vascular plants Verbücheln, G., Götte, R., Hövelmann, T., Itjeshorst, W., Keil, P., Kulbrock, P., Kulbrock, G., Luwe, M., Mause, R., Neikes, N., Schubert, W., Schumacher, W., Schwartze, P., van de Weye, K., 2021. Rote Liste der Farnund Blütenpflanzen – Pteridophyta et Spermatophyta – in Nordrhein-Westfalen. 5th ed., as of October 2020. LANUV-Fachbericht 118, Recklinghausen, Germany. Leafhoppers Nickel, H., Achtziger, R., Biedermann, R., Bückle, C., Deutschmann, U., Niedringhaus, R., Remane, R., Walter, S., Witsack, W., 2016. Rote Liste und Gesamtartenliste der Zikaden (Hemiptera: Auchenorrhyncha) Deutschlands. Nat. Schutz. Biol. Vielfalt. 70, 247–298. True bugs Simon, H., Achtziger, R., Bräu, M., Dorow, W. H. O., Göricke, P., Gossner, M. M., Göricke, P., Gruschwitz, W., Heckmann, R., Hoffmann, H.-J., Kallenborn, H., Kleinsteuber, W., Martschei, T., Melber, A., Morkel, C., Münch, M., Nawratil, J., Remane, R., Rieger, C., Voigt, K., Winkelmann, H., 2021. Rote Liste und Gesamtartenliste der Wanzen (Heteroptera) Deutschlands. Nat. Schutz. Biol. Vielfalt 70, 465–624. Grasshoppers Volpers, M., Vaut, L., 2010. Rote Liste und Artenverzeichnis der Heuschrecken – Saltatoria – in NordrheinWestfalen. LANUV-Fachbericht 36, 489–501. RESTORATION OF SEMI-NATURAL GRASSLANDS Supplementary Table 1 Species list of vascular plants, their threat status according to Verbücheln et al. (2021) (VU = vulnerable, NT = near threatened, * = least concern), classification as target species, suitability as host plants for leafhoppers and true bugs and frequency on plots of restored (n = 9) and target grassland (n = 9). No. Species Threat Target Leafhopper True bug Frequency (%) status species host plant host plant Restored grassland Target grassland 1 Achillea millefolium * . ✓ ✓ . 88.9 2 Agrostis capillaris * . ✓ ✓ 100.0 100.0 3 Ajuga reptans * . . . 22.2 55.6 4 Alchemilla glabra * . . . . 11.1 5 Alchemilla monticola * . . . 22.2 55.6 6 Alchemilla vulgaris * . . . 22.2 55.6 7 Alopecurus pratensis * . ✓ ✓ . 11.1 8 Anemone nemorosa * . . . 11.1 11.1 9 Anthoxanthum odoratum * . ✓ . 100.0 88.9 10 Anthriscus sylvestris * . ✓ . . 11.1 11 Arnica montana VU ✓ . . 11.1 . 12 Arrhenatherum elatius * . ✓ ✓ 22.2 11.1 13 Bellis perennis * . . . . 22.2 14 Betula pendula juv. * . . . 11.1 . 15 Bistorta officinalis * . . . 11.1 22.2 16 Calluna vulgaris * ✓ . ✓ 33.3 . 17 Campanula rapunculoides * . . ✓ . 11.1 18 Campanula rotundifolia * . . ✓ 66.7 88.9 19 Cardamine pratensis * . ✓ . 22.2 44.4 20 Carex leporina * ✓ ✓ ✓ 11.1 11.1 21 Carex nigra * . ✓ ✓ . 11.1 22 Carex pallescens * ✓ ✓ ✓ 22.2 . 23 Carex panicea VU . ✓ ✓ 11.1 . 24 Carex pilulifera * ✓ ✓ ✓ 55.6 33.3 25 Carex remota * . ✓ ✓ 11.1 . 26 Carex sylvatica * . ✓ ✓ 11.1 . 27 Centaurea jacea * . . . 33.3 44.4 28 Centaurea montana * . . . . 11.1 29 Cerastium glomeratum * . . . . 11.1 30 Cerastium holosteoides * . . . 66.7 66.7 31 Cirsium arvense * . ✓ ✓ 11.1 11.1 32 Cirsium palustre * . ✓ ✓ 33.3 33.3 33 Cirsium vulgare * . ✓ ✓ 11.1 . 34 Colchicum autumnale VU . . . . 22.2 35 Crataegus sp. juv. * . . ✓ 11.1 . 36 Crepis mollis VU ✓ . . . 11.1 37 Cynosurus cristatus VU . . . . 33.3 38 Cytisus scoparius juv. * . . ✓ 11.1 11.1 39 Dactylis glomerata * . ✓ ✓ 55.6 100.0 40 Deschampsia cespitosa * . ✓ ✓ 44.4 22.2 41 Deschampsia flexuosa * ✓ ✓ ✓ 33.3 22.2 42 Digitalis purpurea * . . . 11.1 . 43 Dryopteris carthusiana * . . . 22.2 . 44 Epilobium angustifolium * . . . 33.3 . 45 Euphrasia officinalis VU . . . 22.2 11.1 46 Festuca ovina agg. * . ✓ ✓ 55.6 44.4 47 Festuca pratensis * . ✓ ✓ . 11.1 48 Festuca rubra agg. * . ✓ ✓ 100.0 100.0 49 Galeopsis tetrahit * . . . 11.1 33.3 50 Galium mollugo agg. * . ✓ ✓ 22.2 55.6 51 Galium pumilum VU . ✓ ✓ . 11.1 52 Galium saxatile * ✓ ✓ ✓ 77.8 22.2 53 Galium verum * . ✓ ✓ . 22.2 54 Genista tinctoria VU . . . 11.1 11.1 CHAPTER II Supplementary Table 1 continued. No. Species Threat Target Leafhopper True bug Frequency (%) status species host plant host plant Restored grassland Target grassland 55 Geranium sylvaticum * ✓ . . . 11.1 56 Helictotrichon pubescens * . ✓ . . 11.1 57 Heracleum sphondylium * . . . 11.1 44.4 58 Hieracium aurantiacum * . . ✓ 33.3 . 59 Hieracium caespitosum * . . ✓ . 11.1 60 Hieracium lachenalii * . . ✓ 55.6 11.1 61 Hieracium pilosella * . ✓ ✓ . 33.3 62 Holcus lanatus * . ✓ ✓ 22.2 44.4 63 Holcus mollis * . ✓ ✓ 44.4 55.6 64 Hypericum maculatum * ✓ . . 33.3 66.7 65 Hypericum perforatum * . . . 11.1 11.1 66 Hypochaeris radicata * . ✓ . 77.8 55.6 67 Juncus articulatus * . ✓ ✓ 11.1 . 68 Juncus effusus * . ✓ ✓ 33.3 11.1 69 Knautia arvensis * . . . 33.3 77.8 70 Lapsana communis * . . . 11.1 . 71 Lathyrus linifolius VU ✓ . . 55.6 66.7 72 Lathyrus pratensis * . . . . 33.3 73 Leontodon hispidus * . ✓ . 33.3 11.1 74 Leucanthemum vulgare agg. * . . . 55.6 100.0 75 Lolium perenne * . ✓ ✓ . 33.3 76 Lotus corniculatus * . ✓ . 55.6 77.8 77 Luzula campestris * ✓ ✓ ✓ 100.0 100.0 78 Luzula luzuloides * . ✓ ✓ 33.3 . 79 Lysimachia nemorum * . . . 11.1 . 80 Malva moschata * . . . . 22.2 81 Moehringia trinervia * . . . 11.1 . 82 Nardus stricta VU ✓ ✓ ✓ 22.2 11.1 83 Phleum pratense * . ✓ ✓ 44.4 66.7 84 Phyteuma nigrum * ✓ . . 11.1 11.1 85 Phyteuma spicatum * ✓ . . . 11.1 86 Picea abies juv. * . . . 22.2 . 87 Pimpinella saxifraga * . . . 22.2 55.6 88 Plantago lanceolata * . ✓ . 55.6 88.9 89 Poa pratensis * . ✓ ✓ . 88.9 90 Poa trivialis * . ✓ ✓ 11.1 33.3 91 Polygala vulgaris agg. VU ✓ . . . 22.2 92 Potentilla erecta NT ✓ . . 44.4 66.7 93 Prunella vulgaris * . ✓ . . 33.3 94 Ranunculus acris * . . . 66.7 100.0 95 Ranunculus bulbosus * . . . . 11.1 96 Ranunculus flammula NT . . . 11.1 . 97 Ranunculus repens * . ✓ . 11.1 33.3 98 Rhinanthus minor VU . . . 66.7 33.3 99 Rhinanthus serotinus VU . . . 11.1 11.1 100 Rubus idaeus juv. * . . . 22.2 . 101 Rumex acetosa * . ✓ ✓ 100.0 100.0 102 Rumex acetosella * . ✓ ✓ 77.8 11.1 103 Rumex obtusifolius * . ✓ ✓ 22.2 44.4 104 Sanguisorba officinalis VU . . . 33.3 77.8 105 Scorzoneroides autumnalis * . . . 33.3 33.3 106 Senecio jacobaea * . . ✓ . 33.3 107 Senecio ovatus * . . ✓ 33.3 . 108 Silene vulgaris * . . . 11.1 11.1 109 Sorbus aucuparia juv. * . . . 11.1 11.1 110 Stachys sylvatica * . . . 11.1 . RESTORATION OF SEMI-NATURAL GRASSLANDS Supplementary Table 1 continued. No. Species Threat Target Leafhopper True bug Frequency (%) status species host plant host plant Restored grassland Target grassland 111 Stellaria graminea * . . . 66.7 88.9 112 Taraxacum sect. Ruderalia * . ✓ . 44.4 77.8 113 Teucrium scorodonia * . ✓ . 11.1 11.1 114 Thymus pulegioides VU . ✓ . 11.1 33.3 115 Tragopogon pratensis * . . . . 22.2 116 Trientalis europaea * . . . 22.2 . 117 Trifolium dubium * . ✓ ✓ 11.1 11.1 118 Trifolium medium * . ✓ ✓ 11.1 22.2 119 Trifolium pratense * . ✓ ✓ 66.7 77.8 120 Trifolium repens * . ✓ ✓ 66.7 100.0 121 Trisetum flavescens * . ✓ . 11.1 44.4 122 Vaccinium myrtillus * ✓ ✓ . 22.2 22.2 123 Veronica arvensis * . . . . 33.3 124 Veronica chamaedrys * . . . 66.7 100.0 125 Veronica officinalis * ✓ . . 55.6 44.4 126 Vicia cracca * . . . 66.7 88.9 127 Vicia sepium * . . . 33.3 55.6 128 Viola riviniana * . . . 11.1 22.2 129 Viola tricolor ssp. tricolor VU . . . 11.1 . Total number 18 19 58 53 . . Supplementary Table 2 Species list of leafhoppers, their threat status according to Nickel et al. (2016) (CR = critically endangered, EN = endangered, VU = vulnerable, NT = near threatened, * = least concern), classification as target species, frequency on plots of restored (n = 9) and target grassland (n = 9) and number of individuals. No. Species Threat Target Frequency (%) No. status species Restored grassland Target grassland individuals 1 Acanthodelphax spinosa * ✓ 88.9 88.9 724 2 Agallia brachyptera NT ✓ 11.1 44.4 27 3 Allygus mixtus * . 11.1 . 1 4 Anoscopus albifrons * ✓ 44.4 33.3 72 5 Anoscopus flavostriatus * ✓ 44.4 66.7 50 6 Anoscopus serratulae * . 11.1 22.2 10 7 Aphrodes diminuta NT ✓ . 11.1 12 8 Aphrodes makarovi * . 11.1 . 6 9 Aphrophora alni * . . 11.1 1 10 Arthaldeus pascuellus * . 88.9 100.0 1,481 11 Athysanus argentarius * ✓ 33.3 22.2 7 12 Balclutha punctata * . 88.9 88.9 404 13 Cercopis vulnerata * . . 33.3 7 14 Chlorita paolii * ✓ . 11.1 1 15 Cicadella viridis * ✓ 11.1 11.1 5 16 Cicadula persimilis * ✓ 22.2 77.8 28 17 Cixius nervosus * . . 11.1 1 18 Conomelus anceps * ✓ 22.2 . 54 19 Criomorphus albomarginatus * ✓ 55.6 22.2 59 20 Delphacinus mesomelas EN ✓ 88.9 88.9 979 21 Delphacodes venosus * ✓ 55.6 22.2 286 22 Deltocephalus pulicaris * . 100.0 100.0 4,051 23 Dicranotropis divergens VU ✓ 77.8 100.0 560 24 Dicranotropis hamata * ✓ 11.1 . 3 25 Dikraneura variata * . 11.1 . 1 26 Doliotettix lunulatus * . 22.2 22.2 512 27 Doratura stylata * ✓ 77.8 66.7 67 CHAPTER II Supplementary Table 2 continued. No. Species Threat Target Frequency (%) No. status species Restored grassland Target grassland individuals 28 Elymana sulphurella * ✓ 100.0 100.0 765 29 Emelyanoviana mollicula * ✓ . 22.2 3 30 Errastunus ocellaris * . 22.2 55.6 59 31 Eupelix cuspidata NT ✓ 44.4 33.3 19 32 Eupteryx notata * ✓ 44.4 66.7 154 33 Euscelis incisus * . 33.3 22.2 5 34 Fagocyba cruenta * . . 11.1 1 35 Forcipata citrinella NT ✓ 11.1 11.1 4 36 Graphocraerus ventralis * ✓ 88.9 55.6 185 37 Hesium domino NT . . 11.1 9 38 Hyledelphax elegantula * . 11.1 . 8 39 Jassargus pseudocellaris * ✓ 55.6 . 9 40 Javesella discolor * . 22.2 11.1 53 41 Javesella dubia * . 88.9 55.6 881 42 Javesella forcipata NT ✓ . 11.1 90 43 Javesella pellucida * . 44.4 55.6 99 44 Macrosteles laevis * . 44.4 44.4 214 45 Macrosteles septemnotatus * . . 11.1 1 46 Macrosteles sexnotatus * . . 22.2 8 47 Macrosteles viridigriseus NT . 11.1 22.2 110 48 Macustus grisescens * . 22.2 . 4 49 Megophthalmus scanicus * ✓ 88.9 100.0 376 50 Muellerianella brevipennis * ✓ 22.2 33.3 7 51 Muellerianella fairmairei * . 11.1 11.1 7 52 Neophilaenus lineatus * ✓ 77.8 33.3 21 53 Philaenus spumarius * . 33.3 66.7 37 54 Planaphrodes bifasciata NT ✓ 33.3 33.3 25 55 Planaphrodes nigrita NT ✓ 11.1 11.1 3 56 Psammotettix confinis * . 100.0 100.0 1,771 57 Psammotettix helvolus * . 33.3 44.4 117 58 Psammotettix nodosus NT . 22.2 . 19 59 Rhopalopyx adumbrata NT ✓ 11.1 11.1 30 60 Rhopalopyx vitripennis NT ✓ . 11.1 1 61 Ribautodelphax albostriata * ✓ . 11.1 1 62 Ribautodelphax angulosa EN ✓ 11.1 . 5 63 Ribautodelphax collina VU ✓ 22.2 . 12 64 Stenocranus minutus * ✓ . 11.1 1 65 Stiroma affinis * . 11.1 . 1 66 Stiroma bicarinata * ✓ 66.7 11.1 72 67 Streptanus aemulans * ✓ 11.1 44.4 41 68 Streptanus marginatus * ✓ 33.3 . 7 69 Streptanus sordidus * ✓ 66.7 55.6 321 70 Verdanus abdominalis * ✓ 88.9 88.9 3,451 71 Xanthodelphax flaveola CR ✓ 44.4 55.6 266 72 Xanthodelphax straminea NT ✓ 66.7 33.3 27 73 Zyginidia scutellaris * . 33.3 22.2 14 Total number 18 42 . . 18,723 RESTORATION OF SEMI-NATURAL GRASSLANDS Supplementary Table 3 Species list of true bugs, their threat status according to Simon et al. (2021) (NT = near threatened, * = least concern), classification as target species, frequency on plots of restored (n = 9) and target grassland (n = 9) and number of individuals. No. Species Threat Target Frequency (%) No. status species Restored grassland Target grassland individuals 1 Aelia klugii NT ✓ 11.1 . 1 2 Anthocoris nemorum * . . 11.1 1 3 Anthocoris sarothamni * . . 11.1 1 4 Berytinus minor * ✓ 22.2 44.4 45 5 Capsus ater * . 77.8 77.8 58 6 Carpocoris fuscispinus * . . 11.1 1 7 Charagochilus gyllenhalii * ✓ 11.1 . 1 8 Chlamydatus pulicarius * ✓ 11.1 44.4 11 9 Chlamydatus pullus * ✓ 11.1 . 1 10 Dolycoris baccarum * . 11.1 . 1 11 Drymus brunneus * . . 11.1 1 12 Drymus sylvaticus * . 11.1 22.2 9 13 Eurygaster testudinaria * . 11.1 . 1 14 Legnotus picipes * ✓ . 11.1 1 15 Leptopterna dolabrata * . 33.3 44.4 11 16 Leptopterna ferrugata * ✓ 11.1 22.2 8 17 Lopus decolor * ✓ 33.3 . 19 18 Lygocoris pabulinus * . . 11.1 1 19 Lygus pratensis * . 11.1 11.1 5 20 Lygus rugulipennis * . . 22.2 3 21 Lygus wagneri * . . 11.1 3 22 Mecomma ambulans NT . . 11.1 16 23 Megaloceroea recticornis * . 33.3 44.4 25 24 Nabis brevis * ✓ 22.2 11.1 8 25 Nabis flavomarginatus * ✓ 100.0 88.9 385 26 Nabis limbatus * ✓ 11.1 11.1 2 27 Nabis rugosus * . 44.4 22.2 23 28 Neottiglossa pusilla * ✓ 11.1 . 1 29 Nithecus jacobaeae * ✓ 22.2 11.1 25 30 Notostira elongata * . 22.2 22.2 6 31 Notostira erratica * . . 22.2 11 32 Orthocephalus saltator * ✓ . 11.1 1 33 Orthops basalis * . . 11.1 1 34 Orthotylus virescens * . . 22.2 2 35 Pachytomella parallela NT ✓ 100.0 100.0 955 36 Palomena prasina * . 11.1 11.1 2 37 Peribalus strictus * . . 11.1 1 38 Peritrechus geniculatus * ✓ 55.6 11.1 10 39 Pithanus maerkelii * ✓ 33.3 55.6 31 40 Plagiognathus chrysanthemi * ✓ 33.3 44.4 218 41 Polymerus unifasciatus * ✓ 11.1 . 2 42 Rhopalus sp. * . . 11.1 1 43 Stenodema calcarata * . 88.9 88.9 71 44 Stenodema holsata * ✓ 66.7 44.4 58 45 Stenodema laevigata * . 11.1 22.2 3 46 Stenotus binotatus * . 11.1 11.1 2 47 Strongylocoris steganoides * ✓ . 22.2 4 48 Stygnocoris rusticus * ✓ . 33.3 7 49 Stygnocoris sabulosus * ✓ 44.4 . 17 50 Trigonotylus caelestialium * . 11.1 . 1 51 Tytthus pygmaeus NT . 11.1 . 1 Total number 4 23 . . 2,073 CHAPTER III our samples consisted of the majority of transferable arthropods when using a suction sampler (Brook et al., 2008; Sanders and Entling, 2011). Throughout the duration of sample storage, we measured temperatures adjacent to the collection bags outside and inside the cool box using data loggers (iButton, Maxim Integrated, San José, CA, USA) with a measurement resolution of 1 min and an accuracy of 0.5 °C. To prevent the logger outside the cool box from being affected by direct sunlight, it was attached to a radiation shield at a height of 28 cm above the ground. 2.2 Statistical analysis Prior to analyses, we visually checked histograms for normal distribution and homoscedasticity (Quinn and Keough, 2002). Normally distributed data or data that could be normalised by transformation were analysed via parametric tests, otherwise we applied non-parametric approaches. Differences between the four taxonomic groups (leafhoppers, spiders, beetles, and true bugs) and between the three storage periods (1, 2, and 3 h) were compared using repeated measures ANOVA (parametric: rANOVA; non-parametric: Friedman test). As post-hoc test, we applied the HolmSidak test and Dunn’s test, respectively. Storage times of 3 h outside the cool box and 3 h inside the cool box were compared with a paired t-test (parametric) or paired Wilcoxon-test (non-parametric). We used the paired t-test to test for significant differences in the temperatures outside and inside the cool box. All tests were conducted with SigmaPlot v.14 (Systat Software, San José, CA, USA). 3 Results We counted in total 26,357 individuals in the suction samples of the 21 study patches. The four taxonomic groups were found in different abundance (Friedman test: χ² = 53.97, d.f. = 3, P < 0.001). Leafhoppers (12,782 specimens) and spiders (10,551) were most abundant, beetles (2,223) and true bugs (801) were also sampled in all patches, but in much lower abundance (Fig. 2). Temperature was significantly lower inside than outside the cool box (Fig. 3); however, the cooling capacity of the box was poor, as the median temperature inside the box over a period of 3 h was 22 °C. This is also reflected by the minimum values, which ranged between 13 and 26 °C (median 21 °C). Mortality rate was generally low with an overall Figure 1 (a) Suction sampler. (b) Fine gauze collection bag (300 μm mesh size; ca. 6 l volume) on the suction tube. (c) Passive cool box and suction exhauster. ARTHROPOD TRANSLOCATION AND MONITORING number of 2,437 dead individuals (9% of all sampled individuals). After a storage period of 1 h without cooling, about 7% of the leafhoppers had died (Fig. 4a). Extending the storage time significantly increased mortality (maximum of 19% after 3 h; rANOVA: F2,40 = 23.87, P < 0.001). Comparing the storage time of 3 h outside vs. inside the cool box also revealed a significant difference: using the cool box reduced the mortality to 14%. Overall, the mortality of spiders was lower than that of leafhoppers but the pattern was similar: storage period affected mortality, with lowest values after 1 h of storage and highest values after 3 h (F2,40 = 3.38, P < 0.05; Fig. 4b). Mortality after 2 h of storage was not significantly different from the two other storage periods. Similar to leafhoppers, the use of the cool box for 3 h significantly reduced the mortality rate of spiders compared to storage for 3 h outside. In beetles and true bugs, the mortality was generally low and did not differ among the storage periods or the cool box (Friedman test, beetles: χ² = 0.22, P = 0.90; true bugs: χ² = 0.62, P = 0.74, both d.f. = 2; Figure 4c, d). 4 Discussion Our study revealed that arthropod mortality after suction sampling was generally low (9% of all sampled individuals). Overall, the survival rate was affected by (i) storage time, (ii) storage conditions, and (iii) arthropod group. The mortality of beetles and true bugs was low and not influenced by storage time or storage conditions. In contrast, leafhoppers and spiders had a higher mortality; mortality increased with storage time and decreased by the use of a cool box. Arthropod mortality in our study may be caused by effects of the suction or the storage. Suction sampling may lead to lethal damage through collision of the individuals with the walls of the suction sampler or with swirling debris. In addition, storing arthropods within collection bags may result in death due to predation, lethal damages through jumping, or Leafhoppers Spiders Beetles True bugs No. individuals per collection bag 0 20 40 60 80 100 120 140 160 180 a a b b Figure 2 Mean (± SE; n = 21 plots) number of individuals per collection bag (i.e., 50 suction samples per bag) of leafhoppers, spiders, beetles, and true bugs. Means capped with different letters are significantly different (Dunn's test: P < 0.05). Uncooled Cooled Temperature (°C) 0 5 10 15 20 25 30 *** Figure 3 Temperature outside and inside the cool box (based on mean values generated from measurements recorded every minute for 3 h). Box plots show 10th and 90th percentiles (whiskers), 25th and 75th percentiles (top and bottom of the box), and median (line within the box). The asterisks indicate a significant difference between treatments (paired t-test: P < 0.001). CHAPTER III suffocation and dehydration of individuals covered by debris inside the collection bags (Stewart, 2002; Ramires et al., 2007). Predation rates and jumping behaviour are strongly dependent on the activity of the individuals (Speight et al., 2008). Hence, both should be higher under the warmer and lighter conditions in storage bags outside the cool box than in those inside the cool box. The two arthropod groups with on average larger specimens – beetles and true bugs – were relatively resistant to the effects of suction and storing as the low mortality rates in our study indicate. Beetles are characterised by a heavily chitinised body with hardshelled forewings (Dettner and Peters, 2003) and thus seem to be well protected against mechanical damage and against most predators in the bags. True bugs suffered higher mortality rates than beetles. Nevertheless, they were also robust and even soft-bodied species such as plant bugs (Miridae) mostly survived the storage. Leafhoppers are small, hardly chitinised, and actively jumping arthropods (Dettner and Peters, 2003; Speight et al., 2008). These characteristics make them more sensitive to mechanical damage and arthropod predation (Stewart, 2002; Ramires et al., 2007). Additionally, when buried by debris inside the collection bags they may be too weak to escape, risking suffocation and dehydration (cf. Ramires et al., 2007). Spiders are also known to be sensitive to mechanical damage (Parry and Brown, 1959; Wilson, 1970; Anderson and Prestwich, 1975; Ramires et al., 2007; Kropf, 2013) and especially the smaller ones should also suffer from predation by other arthropods in the collection bags. In line with this, we regularly observed spiders and some predatory true bugs preying on other spiders and, more frequently, leafhoppers. Storage in the cool box reduced the mortality of both leafhoppers and spiders. The lower temperature may have lowered the activity of the individuals and, hence, activity-dependent mortality may have been (c) Beetles 1 h 2 h 3 h 3 h % mortality 0 5 10 15 20 (a) Leafhoppers 1 h 2 h 3 h 3 h % mortality 0 5 10 15 20 a b c ** (d) True bugs 1 h 2 h 3 h 3 h 0 5 10 15 20 (b) Spiders 1 h 2 h 3 h 3 h 0 5 10 15 20 aab * b Cooled Cooled Cooled Cooled Uncooled Uncooled Uncooled Uncooled % mortality % mortality Figure 4 Mean (± SE; n = 21 plots) mortality (%) of (a) leafhoppers (ntotal = 12,782), (b) spiders (ntotal = 10,551), (c) beetles (ntotal = 2,223), and (d) true bugs (ntotal = 801) after storage for 1-3 h outside the cool box or 3 h inside the cool box. Means within a panel capped with different letters are significantly different (Holm-Sidak test: P < 0.05). The asterisks indicate a difference between cooled and uncooled (3 h storage) samples (a, c, d: non-parametric test for paired samples; b: paired t-test; *0.01 < P < 0.05, **P < 0.01). ARTHROPOD TRANSLOCATION AND MONITORING lower. Despite a significant cooling effect, the conditions inside the cool box were still quite warm (median temperature ca. 22 °C). Hence, we speculate that the darkness inside the box was also important for a reduced activity and higher survival rates. However, a device with a higher cooling capacity may lead to a lower mortality than the passive cool box we used. 5 Implications for conservation translocations using a suction sampler Suction samplers are well suited to sample arthropods (Stewart, 2002) and to translocate plant seeds in restoration projects (Zerbe and Wiegleb, 2009; Zerbe, 2019). According to our results, they are also a valuable tool to sample arthropod assemblages for conservation translocation. A large proportion of transported individuals survived and the mortality of the more sensitive taxa (leafhoppers and spiders) could be reduced by decreasing storage time and by using a cool box. Low loss rates are mandatory in translocation projects for reasons of ethics, economics, and ecological sustainability. The current guidelines for translocations underline that it is important to preserve the source populations and to minimise stress or suffering of translocated individuals (JCCBI, 2010; IUCN/SSC, 2013). Translocation projects are often expensive and time consuming (Fischer and Lindenmayer, 2000; Carter et al., 2017; Berger-Tal et al., 2019). An advantage of using a suction sampler for translocation is the little material requirements – no heavy machinery is needed for transportation and handling can be done by a single operator. Assessing the scale of expected benefits of a conservation translocation is part of its planning stage (IUCN/SSC, 2013). The dispersal abilities of arthropods vary greatly among species (Carter et al., 2017). Some are able to move long distances (Kisimoto and Rosenberg, 1994; Reynolds et al., 2017) or can be transported passively by hay transfer (Kiehl and Wagner, 2006), sheep (Fischer et al., 1996) or vacuum-harvested seeds (Kiehl et al., 2010). For such good dispersers, active translocations are of subordinate importance because they are able to recolonise restored sites on their own (Carter et al., 2017). Furthermore, the capture efficiency of a suction sampler differs between taxa. For example, it is low for soil-dwelling beetles or hidden nocturnal beetles and spiders, whereas vegetation-dwelling species are caught easily (Sanders and Entling, 2011). But if source populations of catchable, rare and dispersal-limited species exist, a suction sampler should work well for conservation translocation (Baur, 2014; Carter et al., 2017) and has the potential to complement the reintroduction of plant species. Suction samplers can be a valuable tool to translocate species assemblages. 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Zerbe, S., Wiegleb, G. (Eds.), 2009. Renaturierung von Ökosystemen in Mitteleuropa. Springer Spektrum, Berlin, Heidelberg, Germany. CHAPTER III 1 Introduction Globally, insects are the most species-rich branch of the tree of life (Stork, 2018). However, they are declining rapidly (Seibold et al., 2019; van Klink et al., 2020) and their loss rates are much higher than those of many other taxa (Thomas et al., 2005; Cardoso et al., 2020; Wagner, 2020). Insects provide a large variety of key ecosystem services (Ripple et al., 2017; IPBES, 2019; Cardoso et al., 2020). Therefore, their decrease has cascading effects on a wide range of other species and threatens human well-being on our planet. Consequently, to conduct suitable countermeasures timely, there is an urgent need for systematic long-term monitoring that uses standardised methods to detect changes in insect populations and their drivers at an early stage (Streitberger et al., 2024). Indeed, the development of monitoring schemes and funding of monitoring has gained much more attention in recent years (Streitberger et al., 2024; van Klink et al., 2024) Orthoptera play a pivotal role in grassland ecosystems. They are often the major arthropod consumers (Samways, 2005) and a vital food resource for insectivorous vertebrates, particularly birds (Ingrisch and Köhler, 1998; González del Portillo et al., 2021). Moreover, they are highly sensitive to environmental change, such as alterations in land use or climate (Bazelet and Samways, 2011; Uchida and Ushimaru, 2014; Kenyeres et al., 2020; Theron et al., 2022). Thus far, the long-term effects of environmental alterations on Orthoptera have mainly been assessed by modelling approaches using opportunistic distribution data (Poniatowski et al., 2020; Engelhardt et al., 2022; Neff et al., 2022) or re-surveys that compared historical and recent distributions (Schuch et al., 2011; Löffler et al., 2019; Fartmann et al., 2022a; Thorn et al., 2022). For assessments of Orthoptera species richness, abundance and community composition in grasslands, different sampling techniques are regularly applied (Ingrisch and Köhler, 1998; Gardiner et al., 2005). Paper 5 · Global Ecology and Conservation 55 (2024), e03217, doi: 10.1016/j.gecco.2024.e03217 A box quadrat for standardised sampling of Orthoptera in open habitats: Design, handling, applications and baseline data Thomas Fartmann, Felix Maximilian Freienstein, Felix Helbing, Gwydion Scherer, Dominik Poniatowski ABSTRACT Globally, insects are declining at an alarming rate. Therefore, there is an urgent need for standardised monitoring methods to detect changes in insect populations at an early stage. Box quadrat sampling enables highly precise assessments of Orthoptera community composition within a specific area. However, a detailed design of a box quadrat that guarantees an easy workflow, precise descriptions of its handling and a compilation of baseline data for future comparisons have yet been lacking. Based on thirty years of our own experience in box quadrat sampling, we present here: (i) the detailed design of a box quadrat, (ii) its standardised handling in the field (including a video as a supplement), (iii) its applications and (iv) Orthoptera species richness and abundance baseline data for open habitats. The low construction costs (~300 €) and simple design allow for a wide distribution of this type of box quadrat in future. Moreover, due to the plug system and the light aluminium construction, the quadrat can easily be transported, assembled, handled and dismantled by one person. For comparisons of Orthoptera species richness, abundance and community composition between habitats, we recommend using a quadrat of 2 m² size and 0.8 m height and a minimum sampling area of 20 m² per plot (10 touchdowns per plot). For long-term monitoring or when species are sampled that generally occur in low density, a sampling area of 30 m² (15 replicates) is suggested. By contrast, for the surveys of tiny species (e.g., groundhoppers) and developmental stages (nymphs), smaller box quadrats can also be used and the sampled area per plot can be below 20 m². Box quadrat sampling is less dependent on favourable weather conditions and vegetation structure than many other insect survey techniques. Overall, it allows rapid and highly precise assessments of Orthoptera community composition. KEYWORDS Construction manual Grasshopper monitoring scheme Grassland ecosystem Insect abundance Population density Species richness Survey technique ARTHROPOD TRANSLOCATION AND MONITORING Particularly, sweep netting but also transect counts are very popular (Gardiner et al., 2005). However, popularity does not necessarily mean that the applied method is always the most suitable one (Samways et al., 2010). For example, Orthoptera data obtained by sweep netting and acoustic or visual counts along transects strongly vary depending on weather conditions and vegetation structure (e.g., Gardiner et al., 2005). By contrast, box quadrat sampling is less impacted by both and enables highly precise assessments of Orthoptera species richness, abundance and community composition within a specific area (Fartmann, 1997; Ingrisch and Köhler, 1998; Gardiner et al., 2005; Gardiner and Hill, 2006; Badenhausser et al., 2009). Consequently, it has been proposed as the standard method for systematic long-term monitoring of Orthoptera communities in grasslands as part of the recently established German insect monitoring scheme (Streitberger et al., 2024). However, yet, a detailed design of a box quadrat that guarantees an easy workflow and precise descriptions of its handling in the field have not been available. Based on thirty years of own experience in box quadrat sampling, we present here: (i) the detailed design of a box quadrat, (ii) its standardised handling in the field and (iii) Orthoptera species richness and abundance baseline data for open habitats in Germany. The first version of the box quadrat introduced here was used in 1994 for assessing Orthoptera community composition in dry and wet grasslands (Fartmann, 1997). Since then, it has been refined to allow a quick and simple assembly and disassembly as well as easy handling in the field. As a supplement to the guideline for the handling in the field, we provide a video that shows the precise workflow. Published data on Orthoptera species richness and abundance are often hardly comparable due to different survey techniques or differences in the standardisation of sampling (e.g., Ingrisch and Köhler, 1998). Consequently, our assessment of Orthoptera species richness and abundance baselines in open habitats rests upon data that were gathered by standardised box quadrat sampling. These baseline data can serve as a first basis for comparisons with surveys in the future. So far, the basics on the applications of box quadrat sampling for assessing Orthoptera species richness, abundance and community composition were scattered over several publications (Fartmann, 1997; Ingrisch and Köhler, 1998; Gardiner et al., 2005; Gardiner and Hill, 2006; Badenhausser et al., 2009). Accordingly, we here discuss (i) the main findings of these publications, including the advantages but also the limitations of the survey method, (ii) the most suitable size of the quadrat and (iii) the favourable size of the sampling area depending on the aim of the study. Moreover, we evaluate the presented baseline data on the basis of the current knowledge on Orthoptera species richness and abundance in open habitats across Central Europe. 2 Material and methods 2.1 Design of the box quadrat and its handling in the field Based on own long-term experience in constructing the box quadrat and its handling in the field, we here provide detailed instructions on the design and dimensions of the box quadrat (see Section 3.1). Moreover, we explain the workflow in the field (see Section 3.2). As a supplement, a detailed video on the setup of the box quadrat and its handling is made available. 2.2 Orthoptera species richness and abundance baselines in open habitats across Germany Overall, we analysed data from 2,478 sampled plots that had been surveyed between 1994 and 2022 across the five biogeographical regions in Germany (for data sources see Appendix A). All data had been sampled: (i) by using the box quadrat proposed here (see Section 3.1), (ii) in accordance with the handling guidelines described in this paper (see Section 3.2), (iii) on an area of 20 m² and 30 m² per plot, respectively, (see also Section 4.1) and (iv) during July or August, which is the phenology peak of the vast majority of Central European Orthoptera species (Ingrisch and Köhler, 1998; Schlumprecht and Waeber, 2003). To avoid any area effects on species richness, only data that had been assessed on 20 m² per plot were considered. Mean species richness and abundance were analysed for the eight most intensively sampled open habitat types (n > 70 plots per type; five types of grasslands [acidic, calcareous, mesic, tussock and wet grasslands], heathlands, mires and light woodlands). Acidic grasslands included semi-dry grasslands on sandy soils and acidic bedrock, tussock grasslands comprised dry, sparsely vegetated grasslands on sandy soils, where tussock grasses (Corynephorus canescens, Festuca polesia, Festuca psammophila, Koeleria glauca or Stipa capillata) dominated, and mires covered transition mires and raised bogs. CHAPTER III 2.3 Statistical analysis All statistical analyses were performed using R 4.3.1 (R Development Core Team, 2024). Differences in species richness (species/10 m²) and abundance (individuals/10 m²) between habitat types or species were analysed by using generalised linear mixed-effects models (GLMM) with gamma error structure (R package ‘lme4’; Bates et al., 2015). Habitat type and species, respectively, served as a categorical predictor, year and biogeographical region as random factors. For the analysis of abundance data, we included sampling effort (20 m² vs. 30 m² of sampled area) as a further random factor. Pairwise comparisons between habitat types were conducted using Tukey’s contrasts (R package ‘multcomp’; Hothorn et al., 2008). Differences in Ensifera and Caelifera abundance were analysed by using the Mann-Whitney U test. 3 Results 3.1 Design of the box quadrat To guarantee a quick and simple assembly, disassembly and easy handling in the field, the quadrat is based on a plug system. The rods and connector blocks consist of light aluminium (Fig. 1 and 2). Depending on the aim of the study, we regularly use box quadrates of three different sizes: 0.71 m × 0.71 m (0.5 m²), 1 m × 1 m (1 m²) and 1.41 m × 1.41 m (2 m²) (see Section 4.1). All these quadrates have a height of 0.8 m and differ only in the length of the horizontal rods (see below). Here, we present the design for the 2 m² box quadrat since it is recommended for most cases of sampling of Orthoptera species richness, abundance and community composition (see Section 4.1). The box quadrat consists of five components: (i) fabric sheet, (ii) corner rods, (iii) connector blocks, (iv) horizontal rods and (v) bungee cords (Fig. 2). The fabric sheet for the cover of the sidewalls of the box quadrat is sewn from a fabric of 6 m × 1 m size and a thickness of ~ 320 g/m² (Fig. 3a). The upper and lower 10 centimetres of the fabric sheet are sewn into a loop, so that the sheet can later be fixed between the horizontal rods. Where the connector blocks are to be placed, small notches 8 cm in length and 5 cm in height are cut out of the fabric. The distance from notch to notch is 1.37 m, and the fabric sheet for the cover of the sidewalls is sewn as a loop with a 5.80 m circumference. All rods and connector blocks consist of aluminium. Each of the four corner rods has a height of 0.8 m, a diameter of 1.7 cm and a material thickness of 1.5 mm (Fig. 3b). The length and width of each of the four connector blocks is 4 cm, and their height is 3.2 cm (Fig. 3c). On two adjacent lateral sides of the block, there are boreholes, one on each side, 1.5 cm in depth and 1.7 cm in diameter. The centre of the boreholes is situated 1.45 cm from the outer and 1.15 cm from the upper edge of the block. Moreover, a continuous borehole 1.7 cm in diameter is bored from the bottom to the top of the two adjacent sides. Its centre is located 1.2 cm from the lateral sides without boreholes. At the top and bottom of the box quadrat, a corner rod is inserted into the continuous borehole of a connector block; corner rod and block are then welded or bolted together. In total, eight horizontal rods are needed, two for each side of the quadrat (Fig. 3d). Each rod has a length of 1.42 m and the same diameter and material thickness as the corner rods. To plug the horizontal rods into the connector blocks, the rods have to be narrowed to a diameter of 1.6 cm over a length of 1.5 cm at both ends of the rod. The bungee cords for stabilising the sidewalls are 1 m long (Fig. 3e). The complete box quadrat has a weight of ~6 kg. Hence, it can easily be transported and handled by one person (see Section 3.2). The total costs for the construction amount to ~300 €. Aluminium for the rods and connector blocks cost ~50 €, the fabric sheet for the sidewalls ~36 € (6 €/m) and the bungee cords ~12 €. For aluminium processing and sewing, further costs of ~200 € arise. 3.2 Handling of the box quadrat in the field For the assembly of the box quadrat, the horizontal rods have to be inserted into the hems of the fabric (Video 1, see Appendix B). Afterwards, the fabric sheet has to be unfolded and spread out on the ground. For quick assembly, we recommend placing all remaining components in the centre of the quadrat. To set up the sidewalls, the corner rods have to be raised on the inside of the fabric loop and the horizontal rods must be inserted into the connector blocks. To prevent the Figure 1 Box quadrat in a semi-dry calcareous grassland (photo: Dominik Poniatowski). ARTHROPOD TRANSLOCATION AND MONITORING rods from slipping out of the connector blocks while sampling, each sidewall has to be stabilised with a bungee cord at the top and bottom. Overall, the assembly and disassembly (see below) of the box quadrat, respectively, takes less than five minutes. For sampling Orthoptera in the field, the box quadrat has to be thrown forward and down with momentum on the soil surface (Video 1). To avoid chasing Orthoptera away with the shadow of the researcher on sunny days, the movement has to be towards the sun during this procedure. To prevent it from bouncing off the ground, the touchdown has to be executed in a controlled movement without letting go of the quadrat (Video 1). In case of dense swards, the vegetation below the sidewalls has to be pressed down with the box quadrat. By ensuring the full contact between the sidewalls and the ground surface, any escape of individuals from the area inside the quadrat is avoided. After each touchdown, the Orthoptera usually jump on the sidewalls, where they can easily be collected by hand or with a collecting jar. For the remaining individuals, a thorough search in the vegetation is necessary. Adults of all Central European species can be identified in the field, some with assistance of a loupe or calliper (Fischer et al., 2020). This is also true for the vast majority of older nymphs (Thommen, 2021). Once identified, the individuals can be released directly outside the quadrat. Disassembling the box quadrat is as simple as its assembly (Video 1). First, all bungee cords have to be removed and the horizontal rods must be detached from the connector blocks without pulling them out of the fabric hems. Then, the fabric is folded and the four corner rods and six of the eight bungee cords are wrapped in it. Tightening of both ends of the rolled-up fabric loop with the two remaining bungee cords allows for easy transportation and storage. 3.3 Orthoptera species richness and abundance baselines in open habitats across Germany The number of plots per biogeographical region varied from 235 to 698 (Fig. 4). Overall species richness decreased in the following order: calcareous grasslands (mean ± SE: 2.8 ± 0.1 species/10 m²), acidic grasslands (2.5 ± 0.1 species/10 m²), wet grasslands (2.2 ± 0.1 species/10 m²), heathlands (2.1 ± 0.1 species/10 m²), Figure 2 The five components of the box quadrat. CHAPTER III monitoring schemes aim to detect even small changes at a very early stage. Therefore, in the newly established German insect monitoring scheme, Orthoptera are sampled on 30 m² per grassland plot (15 touchdowns per plot) (Streitberger et al., 2024). Moreover, a specific monitoring for single species that usually occur in low abundance may also require larger sampling areas. In line with this, the Federal State of Lower Saxony has founded a monitoring of the large bush-cricket Gampsocleis glabra, which usually exhibits low population densities, on a sampled area of 30 m² per heathland plot (own observation). By contrast, for tiny species, the sampling area per plot can be much smaller than 20 m² to produce reliable data (e.g., Tetrix spp. or Xya spp.; see Schulte, 2003; Münsch et al., 2013). If besides the species also the sex of each individual is to be determined, biomass assessments are possible. Fartmann and Poniatowski (2025) provide sex-specific live weights for 37 common Central European Orthoptera species. These data have for example already been used to assess biomass loss of Orthoptera through mowing in hay meadows (Schwarz et al., 2023) or to compare Orthoptera biomass inside and outside nature reserves (Poniatowski et al., 2024b). To identify the drivers of Orthoptera species community composition and its changes, it is also important to record a set of suitable environmental parameters. Since assemblage composition strongly depends on vegetation structure and the interrelated microclimate (Gardiner and Dover, 2008; Ingrisch and Köhler, 1998), we recommend to gather at least the cover of relevant layers (e.g., gravel/stones, bare ground, grasses/herbs and shrubs), vegetation height and type of land use (Fartmann et al., 2012; Löffler and Fartmann, 2017; Streitberger et al., 2024). To compare assemblage composition and to analyse changes, species traits, such as mobility (low vs. high), habitat specificity (generalist vs. specialist), and indicator values such as the species farmland or species temperature index have been proven as very useful (Poniatowski et al., 2020; Fartmann and Poniatowski, 2025). 4.2 Orthoptera species richness and abundance baselines in open habitats Thus far, a compilation of Orthoptera species richness and abundance data that had been gathered by standardised box quadrat sampling was missing. Our analysis revealed that particularly calcareous and acidic grasslands exhibited a high number of species, whereas especially mires and mesic grasslands had a low species richness. By contrast, abundance was highest in wet, acidic, calcareous and mesic grasslands, and lowest in tussock grasslands, heathlands, mires and woodlands. Mean abundance of the ten most common Caelifera species was nearly three times and maximum abundance even almost five times higher than those of the ten most common Ensifera species. Among the most common Caelifera species, particularly Chorthippus mollis, but also Chorthippus biguttulus, Pseudochorthippus montanus, Pseudochorthippus parallelus and Chorthippus albomarginatus had a high mean abundance. Among the most common Ensifera species, however, Conocephalus dorsalis was the only one that had on average a relatively high abundance. At the habitat level, the heterogeneity of a habitat is known to have the strongest impact on species richness of Orthoptera (Löffler and Fartmann, 2017; Fartmann, 2024). A high habitat heterogeneity fosters species richness in three different ways: (i) For many Orthoptera species, structurally diverse habitats are mandatory, since they undergo microhabitat shifts during their ontogenetic development. A typical example is Decticus verrucivorus (Cherrill and Brown, 1992; Schirmel et al., 2010; Wünsch et al., 2012): The species requires bare soil for egg-laying and the young nymphs also stay there due to the warm microclimate. Larger, older nymphs and later especially adults, however, also depend on taller vegetation that provides shelter against predators and too high temperatures. (ii) Heterogeneous habitats exhibit a higher variety of microhabitats and therefore enable the coexistence of many grasshopper species with sometimes very different habitat requirements (Joern, 2005; Jerrentrup et al., 2014; Löffler and Fartmann, 2017). (iii) Moreover, heterogeneous habitats are significantly more resistant and resilient against the impacts of climate change and generally against strong environmental variation than homogeneous ones (Fartmann, 2024). This is also true for grasshopper populations (Kindvall, 1996; Fartmann et al., 2022b). Overall, habitat heterogeneity decisively contributes to the long-term viability of grasshopper populations, even under strongly fluctuating environmental conditions. In line with this, calcareous and acidic grasslands exhibited the highest species richness. Both habitat types are well-known for their high habitat ARTHROPOD TRANSLOCATION AND MONITORING heterogeneity (Diacon-Bolli et al., 2012; Ellenberg and Leuschner, 2010; Fumy et al., 2021; Fumy and Fartmann, 2023). Usually, they consist of heterogeneous mosaics of bare ground, short turf and taller vegetation. By contrast, those habitat types that had the lowest species richness, especially mesic grasslands but also mires, are usually characterised by a homogeneous habitat structure (Fumy et al., 2021; Fumy and Fartmann, 2023). High Orthoptera abundance, however, is often a trade-off between sufficient food and an ample warm microclimate (Gardiner and Dover, 2008; Schirmel et al., 2011; Helbing et al., 2014; Kenyeres et al., 2024). In fact, all four habitat types that were characterised by high Orthoptera densities (wet, acidic, calcareous and mesic grasslands) are normally characterised by a well-established grass/herbaceous layer, which provides sufficient plant and insect food resources (Fartmann, 2024), as well as sunlit conditions that result in a warm microclimate (Ellenberg and Leuschner, 2010). By contrast, in sparsely vegetated tussock grasslands, heathlands and mires, the low cover of the grass/herbaceous layer and hence, a limited amount of food for grass-feeding species (most Caelifera species; Ingrisch and Köhler, 1998; Schlumprecht and Waeber, 2003), but also for insectivorous or omnivorous ones, was very likely responsible for the low abundance. However, in woodlands, which had the lowest abundance, shade is known to limit Orthoptera densities (Helbing et al., 2014). Overall, Caelifera are known to be the dominant group of Orthoptera in open Central European habitats and usually only a few of them, such as C. mollis, C. biguttulus, P. montanus, P. parallelus and C. albomarginatus, account for most of the total abundance of the Orthoptera community (Ingrisch and Köhler, 1998; Schlumprecht and Waeber, 2003). This is also confirmed by our analysis. Both mean abundance and maximum abundance of the most common Caelifera species were much higher than those of the most common Ensifera species. 5 Conclusions and outlook Here, we presented the design and handling of a box quadrat for standardised assessments of Orthoptera species richness, abundance, biomass and community composition in open habitats. The low construction costs (~300 €) and simple design allow for a wide application of this type of box quadrat in future. Moreover, due to the plug system and the light aluminium construction, the quadrat can easily be transported, assembled, handled and disassembled by one person. For comparisons of Orthoptera communities between habitats and to work efficiently, we recommend using the largest version of the quadrat (size: 2 m², height: 0.8 m) and a minimum sampling area of 20 m² per plot. The specific handling of the quadrat during sampling should follow the guidelines in this paper. For longterm monitoring or when species are sampled that generally occur in low density, we suggest a sampling area of 30 m² (15 replicates). For the sampling of tiny species (e.g., groundhoppers) and developmental stages (nymphs), smaller box quadrats can also be used (size: 0.5 m² or 1 m²) and the sampled area per plot can be much smaller than 20 m². If besides the species also the sex of each individual within the quadrat is to be determined, biomass assessments are possible by using available sex-specific live weights. Overall, box quadrat sampling allows rapid and highly precise assessments of Orthoptera community composition, since it is less dependent on favourable weather conditions and vegetation structure than many other insect survey techniques. Moreover, this paper presents Orthoptera species richness and abundance baseline data in the most frequently studied open habitats and for the most common species within these habitats across Germany. These data are an important basis for comparisons with further box quadrat data from open habitats in Central Europe, for example those that will be generated by the German insect monitoring scheme (Streitberger et al., 2024). Besides Orthoptera, the described box quadrat sampling can also be applied to assess the species richness and abundance of Mantodea (cf. Streitberger et al., 2024) or, in combination with a suction sampler, other smaller arthropods of the grass/herbaceous layer (e.g., Auchenorrhyncha or Heteroptera) (Cherrill, 2015; Helbing et al., 2023). When using suction samplers, it is difficult to define a concrete sampling area since not only individuals directly below but also in the vicinity of the suction tube are soaked in. Therefore, Cherrill (2015) recommended using suction samplers to sample the complete area within small enclosures. To do so, small box quadrats (size: 0.5 m²) seem to be as good as the cylindrical enclosures (size: 0.17 m²) that were used by Cherrill (2015). Chisté et al. (2018) even sampled plantand leafhopper communities of grasslands within box quadrats of 1 m² size. Acknowledgements We are grateful to Verena Hartmann, Laura Hebling, Christopher Poschmann and Nadine Zeuner for providing unpublished Orthoptera abundance data. 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Schuch, S., Bock, J., Leuschner, C., Schaefer, M., Wesche, K., 2011. Minor changes in orthopteran assemblages of Central European protected dry grasslands during the last 40 years. J. Insect Conserv. 15, 811–822. doi: 10.1007/s10841011-9379-6 Schulte, A. M., 2003. Taxonomie, Verbreitung und Ökologie von Tetrix bipunctata (Linnaeus 1758) und Tetrix tenuicornis (Sahlberg 1893) (Saltatoria: Tetrigidae). Articulata Beih. 10, 1–226. ARTHROPOD TRANSLOCATION AND MONITORING Schwarz, C., Fartmann. T., 2022. Traditional grazing management creates heterogeneous swards and fosters grasshopper densities. Insect Sci. 29, 1805–1818. doi: 10.1111/ 1744-7917.13041 Schwarz, C., Fumy, F., Drung, M., Fartmann, T., 2023. Insectfriendly harvest in hay meadows – Uncut refuges are of vital importance for conservation management. Glob. Ecol. Conserv. 48, 02731. doi: 10.1016/j.gecco.2023.e02731 Stork, N. E., 2018. How many species of insects and other terrestrial arthropods are there on Earth? Ann. Rev. 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Temperature preferences drive additive biotic homogenization of Orthoptera assemblages. Biol. Lett. 18, 20220055. doi: 10.1098/rsbl.2022.0055 Uchida, K., Ushimaru, A., 2014. Biodiversity declines due to abandonment and intensification of agricultural lands: patterns and mechanisms. Ecol. Monogr. 84, 637–658. doi: 10.1890/13-2170.1 Wagner, D. L., 2020. Insect declines in the Anthropocene. Ann. Rev. Entomol. 65, 457–480. doi: 10.1146/annurevento-011019-025151 Wünsch, Y., Schirmel, J., Fartmann, T., 2012. Conservation management of coastal dunes for Orthoptera has to consider oviposition and nymphal preferences. J. Insect Cons. 16, 501–510. doi: 10.1007/s10841-011-9436-1 Appendix A Data sources: Behrens and Fartmann (2004a,b), Fartmann et al. (2022b), Fleischer (2011), Graser (2016), Helbing et al. (2014), (2023), Kämpf (2011), Lemke et al. (2010), Löffler and Fartmann (2017), Ludwig (2015), own unpublished data, Poniatowski and Fartmann (2005, 2007, 2008), Poniatowski et al. (2024b), Poschmann (2011), Poschmann et al. (2009), Schwarz and Fartmann (2022), Schwarz et al. (2023). Appendix B Video 1 Setup and workflow of the box quadrat in the field. The video can be be accessed by scanning the QR code above or directly via the following link: https://youtu.be/Pbl0LmKJ03I CHAPTER III The colonisation probability of target species depends on both species-specific mobility and the connectivity of restoration sites to donor populations. For strongly isolated sites, such as those in intensive agricultural landscapes (top) or within forests (bottom), the active introduction of target animal species in addition to plants can be a reasonable approach. (‘Offenberg’, 06/2021; ‘Weinberg’, 06/2020; both near Hofgeismar) ARTHROPOD TRANSLOCATION AND MONITORING Suction samplers (top) are well suited for collecting arthropods such as leafhoppers (bottom), spiders, beetles, and true bugs, and they can also serve as a useful tool for conservation translocation. Mortality during sampling and translocation is generally low but varies among taxa. It can be further reduced by using a cool box for transporting the samples and by minimising the time between collection and release at the restoration site. (Restored montane grassland near Winterberg, 05/2018 [photo: D. Poniatowski]; Turrutus socialis in the Diemel Valley, 06/2020) Chapter IV Synthesis and perspectives CHAPTER IV Witzenberger, K. A., Hochkirch, A., 2008. Genetic consequences of animal translocations: A case study using the field cricket, Gryllus campestris L. Biol. Conserv. 141, 3059–3068. doi: 10.1016/j.biocon.2008.09.017 Wünsch, Y., Schirmel, J., Fartmann, T., 2012. Conservation management of coastal dunes for Orthoptera has to consider oviposition and nymphal preferences. J. Insect Cons. 16, 501–510. doi: 10.1007/s10841-011-9436-1 Zerbe, S. (Ed.), 2019. Renaturierung von Ökosystemen im Spannungsfeld von Mensch und Umwelt. Springer Spektrum, Berlin, Heidelberg, Germany. Zurbrügg, C., Frank, T. (2006). Factors influencing bug diversity (Insecta: Heteroptera) in semi-natural habitats. Biodivers. Conserv. 15, 275–294. doi: 10.1007/ s10531-004-8231-7 SYNTHESIS AND PERSPECTIVES Semi-natural grasslands are biodiversity hotspots and serve as important refuges for numerous specialised plant and animal species, such as (in rows from left to right): 1st row – Satyrium spini, Xylocopa violacea; 2nd row – Gentiana cruciata, Lacerta agilis; 3rd row – Polistes nimpha, Melitaea aurelia; 4th row – Dactylorhiza maculata, Decticus verrucivorus. ACKNOWLEDGEMENTS ACKNOWLEDGEMENTS I would like to take this opportunity to express my sincere gratitude to all the people and institutions that have greatly supported me in the course of my PhD. First of all, I would like to thank Prof. Dr. Thomas Fartmann for his outstanding scientific supervision. His deep ecological knowledge, his constructive way of discussing my studies, and his motivating guidance greatly contributed to the development of my work. Beyond this project, he has encouraged me since the early stages of my academic journey and has played a vital role in shaping my scientific interests – particularly in the ecology of insects. Furthermore, I would like to thank Prof. Dr. Stefan Zerbe, who kindly agreed to be the second supervisor of my thesis. Most of the studies included in this thesis were funded by the German Federal Agency for Nature Conservation (Bundesamt für Naturschutz) (Grant number 3516892017), which I gratefully acknowledge. Further valuable contributions to this work came from the team of the Division of Biodiversity and Landscape Ecology at Osnabrück University. I am particularly grateful to Dr. Dominik Poniatowski for his tireless support throughout my PhD and beyond. His sympathetic ear and the countless conversations we had – both on a professional and personal level – were truly enriching. Special thanks also go to Max Freienstein, Dr. Steffen Kämpfer, Dr. Marcel Kettermann, Dr. Franz Löffler, Thorsten Münsch, Gwydion Scherer, Dr. Merle Streitberger, and Dr. Gregor Stuhldreher, who co-authored papers, provided statistical advice, and offered helpful suggestions on study designs and manuscript drafts. I would like to extend my gratitude to Dr. Carsten Morkel for identifying the true bug species, to Dr. Herbert Nickel for his helpful advice on the identification of challenging leafhopper taxa, and to Dr. Axel M. Schulte (Biologische Station Hochsauerlandkreis e.V.), who generously shared his knowledge on restoration methods and site characteristics relevant to the study of montane grasslands. Permissions to conduct fieldwork within protected areas were kindly granted by the Landscape Authority Höxter (Untere Naturschutzbehörde Höxter) and the Regional Council of Kassel (Regierungspräsidium Kassel). Finally, I am deeply grateful to my partner Raïna, my children Joscha and Thilo, and my parents Kirsten and Ulrich for their constant support and patience throughout the writing of this thesis. Raïna, in particular, continually encouraged me and enabled me to keep moving forward — often at the expense of her own time and energy. CURRICULUM VITAE CURRICULUM VITAE Felix Helbing *9th September 1987 in Wermelskirchen (Germany) Professional experience since 07/2016 Research assistant, Osnabrück University, Division of Biodiversity and Landscape Ecology 04–06/2016 Graduate assistant, Osnabrück University, Division of Biodiversity and Landscape Ecology 06–09/2015 Freelance work for the Bavarian Academy for Nature Conservation and Landscape Management 06–08/2013 Visiting lecturer, Osnabrück University, Division of Ecology 2012–2015 Student assistant, University of Münster, Institute of Landscape Ecology, Community Ecology Group Education 04/2017-present PhD studies, Osnabrück University 10/2012–06/2016 Master studies in Landscape Ecology, University of Münster 10/2009–08/2012 Bachelor studies in Landscape Ecology, University of Münster 06/2007 Abitur, Städtisches Gymnasium Wermelskirchen PUBLICATIONS AND CONFERENCE CONTRIBUTIONS PUBLICATIONS AND CONFERENCE CONTRIBUTIONS International publications (peer-reviewed) Fartmann, T., Freienstein, F. M., Helbing, F., Scherer, G., Poniatowski, D., 2024. A box quadrat for standardised sampling of Orthoptera in open habitats: design, handling, applications and baseline data. Global Ecol. Conserv. 55, e03217. doi: 10.1016/j.gecco.2024.e03217 Helbing, F., Fartmann, T., Morkel, C., Poniatowski, D., 2023. Rapid response of vascular plants and insects to restoration of montane grasslands. Front. Ecol. Evol. 11, 1148266. doi: 10.3389/fevo.2023.1148266 Löffler, F., Grüneberg, C., Drung, M., Freienstein, F.M., Helbing, F., Holtmann, L., Kämpfer, S., Kettermann, M., Münsch, T., Poniatowski, D., Streitberger, M., Stuhldreher, G., Fartmann, T., 2023. Different environmental conditions in lowlands and uplands highlight challenges for butterfly conservation in Central Europe. Biol. Conserv. 281, 110034. doi: 10.1016/j.biocon.2023.110034 Helbing, F., Fartmann, T., Poniatowski, D., 2021. Restoration measures foster biodiversity of important primary consumers within calcareous grasslands. Biol. Conserv. 256, 109058. doi: 10.1016/j.biocon.2021.109058 Poniatowski, D., Beckmann, C., Löffler, F., Münsch, T., Helbing, F., Samways, M.J., Fartmann, T., 2020. Relative impacts of land-use and climate change on grasshopper range shifts have changed over time. Global Ecol. Biogeogr. 29, 2190–2202. doi: 10.1111/geb.13188 Helbing, F., Fartmann, T., Poniatowski, D., 2020. Suction samplers are a valuable tool to sample arthropod assemblages for conservation translocation. Entomol. Exp. Appl. 168, 688–694. doi: 10.1111/eea.12952 Poniatowski, D., Stuhldreher, G., Helbing, F., Hamer, U., Fartmann, T., 2020. Restoration of calcareous grasslands: The early successional stage promotes biodiversity. Ecol. Eng. 151, 105858. doi: 10.1016/j.ecoleng.2020.105858 Münsch, T., Helbing, F., Fartmann, T., 2019. Habitat quality determines patch occupancy of two specialist Lepidoptera species in well-connected grasslands. J. Insect Conserv. 23, 247–258. doi: 10.1007/s10841-018-0109-1 Helbing, F., Fartmann, T., Löffler, F., Poniatowski, D., 2017. Effects of local climate, landscape structure and habitat quality on leafhopper assemblages of acidic grasslands. Agric. Ecosyst. Environ. 246, 94–101. doi: 10.1016/j.agee.2017.05.024. Helbing, F., Cornils, N., Stuhldreher, G., Fartmann, T., 2015. Populations of a shrub-feeding butterfly thrive after introduction of restorative shrub cutting on formerly abandoned calcareous grassland. J. Insect Conserv. 19, 457– 464. doi: 10.1007/s10841-015-9766-5 Borchard, F., Buchholz, S., Helbing, F., Fartmann, T., 2014. Carabid beetles and spiders as bioindicators for the evaluation of montane heathland restoration on former spruce forests. Biol. Conserv. 178, 185–192. doi: 10.1016/j.biocon.2014.08.006 Helbing, F., Blaeser, T.P., Löffler, F., Fartmann, T., 2014. Response of Orthoptera communities to succession in alluvial pine woodlands. J. Insect Conserv. 18, 215–224. doi: 10.1007/s10841-014-9632-x Other publications Helbing, F., Poniatowski, D., Fuhrmann, K., Grein, G., Dense, C., Klugkist, H., Schuhmacher, O., Fartmann, T., 2025. Rote Liste und Gesamtartenliste der Heuschrecken (Orthoptera) in Niedersachsen und Bremen, 4. Fassung – Stand 2024. Inform.d. Naturschutz Niedersachs. 44 (2): 81–120. Poniatowski, D., Detzel, P., Drews, A., Hochkirch, A., Hundertmark, I., Husemann, M., Klatt, R., Klugkist, H., Köhler, G., Kronshage, A., Maas, S., Moritz, R., Pfeifer, M. A., Stübing, S., Voith, J., Winkler, C., Wranik, W., Helbing, F., Fartmann, T., 2024. Rote Liste und Gesamtartenliste der Heuschrecken und Fangschrecken (Orthoptera et Mantodea) Deutschlands. Nat. Schutz. Biol. Vielfalt. 170 (7), 1–88. PUBLICATIONS AND CONFERENCE CONTRIBUTIONS Rösch, V., Biedermann, R., Entling, M. H., Gjonov, I., Helbing, F., Jakovljević, M., van Klink, R., Marinković, S., Sára, A., Sára, H., Schuch, S., Achtziger, R., 2023. Leafhopper diversity in home gardens – results of a survey in four countries across Europe (Hemiptera, Auchenorrhyncha). Cicadina 22, 39–57. doi: 10.25673/111331 Poniatowski, D., Helbing, F., Stuhldreher, G., Grawe, F., Stowitz-Lohne, M., Düster, J., Fartmann, T., 2022. Abschluss des länderübergreifenden Renaturierungsprojektes im Diemeltal (Nordhessen/Ostwestfalen). Jb. Naturschutz Hessen 21, 153. Poniatowski, D., Helbing, F., Stuhldreher, G., Düster, J., Fartmann, T., 2021. Renaturierung von Kalkmagerrasen im Diemeltal Nordhessen / Ostwestfalen). Jb. Naturschutz Hessen 20, 110–113. Fartmann, T., Helbing, F., Streitberger, M., Stuhldreher, G., Poniatowski, D., 2021. Kalkmagerrasenrenaturierung. In: Fartmann, T., Jedicke, E., Streitberger, M., Stuhldreher, G. (eds.). Insektensterben in Mitteleuropa: Ursachen und Gegenmaßnahmen. Eugen Ulmer, Stuttgart, pp. 156–158. Fartmann, T., Stuhldreher, G., Streitberger, M., Helbing, F., 2021. Die Bedeutung der Habitatqualität für den Schutz der Insektendiversität – Mikroklima, Phytodiversität, Habitatheterogenität und Totholz sind Schlüsselfaktoren für artenreiche Insektengemeinschaften. Naturschutz Landschaftspl. 53 (7), 2–7. Fartmann, T., Helbing, F., Streitberger, M., Stuhldreher, G., Poniatowski, D., 2021. Kalkmagerrasen: Biodiversität, Management und Renaturierung. Nat. Schutz. Biol. Vielfalt. 171: 293–310. Helbing, F., Cornils, N., Stuhldreher, G., Fartmann, T., 2020. Habitatmanagement für einen strauchbewohnenden Tagfalter. In: Trautner, J. (ed.). Artenschutz – Rechtliche Pflichten, fachliche Konzepte, Umsetzung in der Praxis. Eugen Ulmer, Stuttgart, pp. 242–246. Fartmann, T., Helbing, F., Streitberger, M., Stuhldreher, G., Poniatowski, D., 2019. Auswirkungen des globalen Wandels auf die Biodiversität von Trockenrasen. Nationalpark-Jahrb. Unteres Odertal 16, 59–73. van Klink, R., den Bieman, K., Funke, L., de Haas, M., Helbing, F., Herwig, S., Huber, E., Malenovský, I., Nickel, H., Niedringhaus, R., Poniatowski, D., Rösch, V., Seyen, F., Witsack, W., 2019. Auchenorrhyncha and Psylloidea collected during the 25th Central European Auchenorrhyncha meeting, Arnhem, The Netherlands (Hemiptera: Auchenorrhyncha and Psylloidea). Cicadina 18, 43–58. doi: 10.25673/92161 Poniatowski, D., Münsch, T., Helbing, F., Fartmann, T., 2018. Arealveränderungen mitteleuropäischer Heuschrecken als Folge des Klimawandels. Nat. Landsch. 93 (12), 553–561. Helbing F., Fartmann, T., Poniatowski, D., 2017. Die Zikadenfauna isolierter Schieferkuppen der Medebacher Bucht (Südwestfalen/Nordhessen) (Hemiptera, Auchenorrhyncha). Cicadina 17: 1–18. doi: 10.25673/92249 Helbing F., Poniatowski, D., 2015. Neue Funde seltener Zikadenarten auf den Kalkmagerrasen des Diemeltals (Ostwestfalen/Nordhessen) (Hemiptera, Auchenorrhyncha). Cicadina 15, 43–57. doi: 10.25673/92241 Helbing, F., Cornils, N., Stuhldreher, G., Fartmann, T., 2015. Renaturierung von Kalkmagerrasen zur Förderung einer Tagfalterzielart. Natur in NRW 2/2015, 35–39. Helbing, F., Löffler F., Thoms, A., Fartmann, T., 2014. Erstfund der Lauchschrecke (Mecostethus parapleurus, Hagenbach, 1822) in den mittleren Bayerischen Kalkalpen. Articulata 29 (1), 75–78. Conference contributions Poniatowski, D., Stuhldreher, G., Helbing, F., Fartmann, T.: Evaluation der Renaturierung von Kalkmagerrasen. Ökosystem Trockenrasen – Bestand, Projekte, Perspektiven, Brandenburgische Akademie Schloss Criewen, Schwedt/Oder (Germany) (06/2019). Poniatowski, D., Münsch, T., Helbing, F., Fartmann, T.: Arealveränderungen mitteleuropäischer Heuschrecken als Folge des Klimawandels. Tagung ‘Biodiversität und Klima – Vernetzung der Akteure in Deutschland XV’, German Federal Agency for Nature Conservation, International Nature Conservation Academy Isle of Vilm (Germany) (09/2018). PUBLICATIONS AND CONFERENCE CONTRIBUTIONS Helbing, F., Fartmann, T., Poniatowski, D.: Effects of local climate, landscape and habitat quality on leafhoppers of acidic grasslands. 24. Mitteleuropäische Zikadentagung, Lugano (Swiss) (06/2017). Helbing, F., Löffler, F., Fartmann, T.: Auswirkungen von Waldweide und Auendynamik auf Heuschreckengemeinschaften in Schneeheide-Kiefernwäldern der Bayerischen Alpen. 13. Jahrestagung der Deutschen Gesellschaft für Orthopterologie (DGfO), Salzburg (Austria) (03/2014). Helbing, F., Poniatowski, D., Fartmann, T.: Weit verbreitet, aber wählerisch: Eiablagepräferenzen des Blauen Eichenzipfelfalters (Neozephyrus quercus). Symposium für Schmetterlingsschutz und 15. UFZ-Workshop zur Populationsbiologie von Tagfaltern und Widderchen, Leipzig (Germany) (03/2013). Erklärung über die Eigenständigkeit der erbrachten wissenschaftlichen Leistung Ich erkläre hiermit, dass ich die vorliegende Arbeit ohne unzulässige Hilfe Dritter und ohne Benutzung anderer als der angegebenen Hilfsmittel angefertigt habe. Die aus anderen Quellen direkt oder indirekt übernommenen Daten und Konzepte sind unter Angabe der Quelle gekennzeichnet. Bei der Auswahl und Auswertung folgenden Materials haben mir die nachstehend aufgeführten Personen in der jeweils beschriebenen Weise unentgeltlich geholfen. Artikel 1: Poniatowski, D., Stuhldreher, G., Helbing, F., Hamer, U., Fartmann, T., 2020. Restoration of calcareous grasslands: The early successional stage promotes biodiversity. Ecol. Eng. 151, 105858. doi: 10.1016/j.ecoleng.2020.105858 • Dominik Poniatowski: Datenerhebung, Verfassen der Erstfassung des Manuskripts. Anteil: 35% • Gregor Stuhldreher: Datenerhebung, Datenanalyse, Redigieren des Manuskripts. Anteil: 25% • Ute Hamer: Redigieren des Manuskripts. Anteil: 5% • Thomas Fartmann: Redigieren des Manuskripts. Anteil: 10% • Zusätzlich wurden die Bodenproben von Gwydion Scherer und Antje Möhlmeyer analysiert. Artikel 2: Helbing, F., Fartmann, T., Poniatowski, D., 2021. Restoration measures foster biodiversity of important primary consumers within calcareous grasslands. Biol. Conserv. 256, 109058. doi: 10.1016/j.biocon.2021.109058 • Thomas Fartmann: Redigieren des Manuskripts. Anteil: 10% • Dominik Poniatowski: Redigieren des Manuskripts. Anteil: 10% Artikel 3: Helbing, F., Fartmann, T., Morkel, C., Poniatowski, D., 2023. Rapid response of vascular plants and insects to restoration of montane grasslands. Front. Ecol. Evol. 11, 1148266. doi: 10.3389/fevo.2023.1148266 • Thomas Fartmann: Redigieren des Manuskripts. Anteil: 10% • Carsten Morkel: Bestimmung der Wanzen, Redigieren des Manuskripts. Anteil: 10% • Dominik Poniatowski: Redigieren des Manuskripts. Anteil: 10% Artikel 4: Helbing, F., Fartmann, T., Poniatowski, D., 2020. Suction samplers are a valuable tool to sample arthropod assemblages for conservation translocation. Entomol. Exp. Appl. 168, 688–694. doi: 10.1111/eea.12952 • Thomas Fartmann: Redigieren des Manuskripts. Anteil: 10% • Dominik Poniatowski: Redigieren des Manuskripts. Anteil: 10% Artikel 5: Fartmann, T., Freienstein, F. M., Helbing, F., Scherer, G., Poniatowski, D., 2024. A box quadrat for standardised sampling of Orthoptera in open habitats: Design, handling, applications and baseline data. Global Ecol. Conserv. 55, e03217. doi: 10.1016/j.gecco.2024.e03217 • Thomas Fartmann: Verfassen der Erstfassung des Manuskripts. Anteil: 40% • Felix Max Freienstein: Unterstützung bei der Erstellung des Videos, Redigieren des Manuskripts. Anteil: 15% • Gwydion Scherer: Redigieren des Manuskripts. Anteil: 5% • Dominik Poniatowski: Unterstützung bei der Datenanalyse, Redigieren des Manuskripts. Anteil: 15% Weitere Personen oder Organisationen waren an der inhaltlichen materiellen Erstellung der vorliegenden Arbeit nicht beteiligt. Insbesondere habe ich hierfür nicht die entgeltliche Hilfe von Vermittlungsbzw. Beratungsdiensten, Promotionsberaterinnen oder Promotionsberatern oder anderen Personen in Anspruch genommen. Die Arbeit wurde bisher weder im Innoch im Ausland in gleicher oder ähnlicher Form einer anderen Prüfungsbehörde vorgelegt. ………………………………………... ……………………………………………………… (Ort, Datum) (Unterschrift)