<figure><img src="image_28.png" /><figcaption>FIGURE 7</figcaption></figure>

Increase in total number of *Ophrys apifera* from 2013 to 2021 and mean horse feeding frequency from April–March from 2011–2021 on the grid cells in the different calcareous grassland types (n=121; xeric n=17, dry n=96, dry-mesic n=8).

<figure><img src="image_29.png" /><figcaption>FIGURE 8</figcaption></figure>

Effect of horse feeding frequency on inverse-transformed density of *Ophrys apifera* in the different calcareous grassland types in 2013, 2018 and 2021 (n=121; xeric n=17, dry n=96, dry-mesic n=8). Shaded area represents 95% confidence intervals.

Schwabe, 2007; Schwabe et al., 2013; Henning et al., 2017a) assuming that a sufficient, adapted stocking rate is guaranteed as undergrazing can lead to ruderalization or tall species dominance (Sýkora et al., 2009; Moinardeau et al., 2021). Concerning faunistic diversity, forb enhancement of grassland is usually beneficial to insects as forbs provide both nesting and foraging sites (Bonari et al., 2017), and in the consequence beneficial to higher trophic animal species like birds (Lovász et al., 2021; Schmidt et al., 2022).

Compared to other grazing animal species such as cattle, sheep or goats, equids can compensate low-quality forage through increased, nearly permanent feeding (Duncan et al., 1990; own observation data). As equids show a preference for grass species, grasses are disproportionately more frequently used compared to forbs. Grass species, almost indigestible for many herbivores due to silicate inclusion in the cells and lignifying fibers, can be digested by equids because of their digestive system geared to fiber utilization and high-crowned
teeth protecting from abrasion (coevolution of grasses and grazers beginning in the Tertiary; Stebbins, 1981; Bunzel-Drüke et al., 2008). These characteristics make horses well suited to reduce grass encroachment in dry grasslands (Süss and Schwabe, 2007; Schwabe et al., 2013).

We found that ruderals significantly increased only in the long-term observation period, probably due to enhanced disturbance and more bare soil patches created by the continuous reduction of herb and litter layers through grazing in the long run. The observed short-lived, small-growing, thermophilic ruderal species such as *Arenaria serpyllifolia*, *Cerastium pumilum* or *Draba verna* are seen as intrinsic part of pastures that enhance their biodiversity (Elias et al., 2018; Köhler et al., 2020) as well as nectar and pollen supply for insects, but do not affect the species assemblage of dry grasslands due to the low increase of cover. Extended nitrophilous ruderal patches that are often discussed as problematic in grazed areas, have not been observed in our study, most likely due to the generally low nutrient levels. Generally, the different types of zoochory leading to species dispersal inside the pasture are seen as the main driver of long-time establishment success of target species (Kiss et al., 2021), emphasizing the benefits of grazing compared to mowing.

## 4.2. Large-scale effects of year-round grazing on *Ophrys apifera*

We evaluated the long-term effects of year-round horse grazing on the target orchid species *O. apifera* in a dry calcareous grassland using data on habitat structures and horse feeding frequency. We found a continuous significant increase of the counted *O. apifera* individuals on most observed grid cells from 2013 to 2021. This important finding means that low-intensity horse grazing is generally suitable for orchid species maintenance and promotion on large-scale, plain dry calcareous grasslands. The orchid species' ability to cope with permanent grazing has already been demonstrated by Köhler et al. (2016) for a 5-year grazing period. In this study, we confirm these results for a 12-year period and clarified relevant mechanisms in the different calcareous grassland types. Across all three calcareous grassland types, horse feeding frequency had an overall positive effect on the density of *O. apifera* individuals. However, *O. apifera* abundance varied inside the pasture due to differently appropriate site conditions. Particularly, the increase of bare soil patches and the reduction of litter layers induced by feeding and trampling led to lower competition through the surrounding vegetation and to enhanced light availability on the ground (Borer et al., 2014) creating niches for orchid recruitment and seedling establishment (Shefferson et al., 2020). In addition, biomass and particularly grass cover reduction also diminished vegetation competition.

Sites characterized by shallow soils, a high share of bare soil patches and low vegetation cover that harbor xeric calcareous grassland species showed highest initial *O. apifera* abundance at the beginning of the study. Here, lowest horse feeding frequencies were recorded over the observation period, as grazers probably avoided grassland patches with both low fodder availability and quality (Gilhaus et al., 2014). However, under this condition of generally very low grazing intensity, the higher the feeding frequency, the more positive it affected *O. apifera* density due to the better reduction of litter layer and the dynamic creation of bare soil patches. Hence, these high-value xeric grasslands can be integrated into the grazing system contrary to often recommended fencing off (Calaciura and Spinelli, 2008; Catorci et al., 2013), at least in large-scaled

pastures (> 10, better > 50 ha; Bunzel-Drüke et al., 2008) where sites of higher productivity are preferentially grazed.

Sites, that were initially less suited for orchid occurrence due to heavy grass encroachment and thick litter layers, were characterized by higher vegetation cover of dry calcareous grasslands. These sites were significantly more frequented by the horses due to higher fodder availability and quality. On this grassland type, that had by far the largest area share (80.9%), *O. apifera* density was significantly lower at the beginning of grazing compared to xeric grasslands. But, surprisingly, it also increased significantly and contributed most to the increase in numbers of individuals due to enhanced habitat quality. The benefits of grazing for the development of *O. apifera* in this grassland type were most effective at a medium horse feeding frequency, since grazing frequency did not show an explicit effect on the individual density. On grid cells with a higher grazing frequency, positive grazing effects seemed to be able to outbalance negative effects such as trampling and feeding on *O. apifera*.

Grid cells of the dry-mesic calcareous grassland type were per-se sub-optimal sites for orchid occurrence as characterized by deeper soils, supporting more mesic, higher-competitive species. On these (rather small) sites of the pasture, horse feeding frequency showed significantly highest values probably due to higher protein forage compared to dry and xeric grasslands (Gilhaus et al., 2014). But even on these sites, *O. apifera* individual density was stabilized during the horse grazing experiment. Here, grid cells with lower or medium horse feeding frequency maintained *O. apifera* populations, while at higher horse feeding frequency a slightly negative effect was observed (Figure 8).

In contrast to other studies reporting direct negative grazing effects on orchids by feeding and trampling (Calaciura and Spinelli, 2008; Catorci et al., 2013; Cleavitt et al., 2016), we found positive effects of year-round grazing. Horses that intentionally graze orchid inflorescences or leaves were not observed in our study (field observation). *Vice-versa*, horse grazing was shown to maintain or create habitat structures that supported the strong population increase of the target orchid species. A crucial factor for this development was the increase of bare soil patches mainly induced by grazing and trampling. The establishment of many orchid species is not limited by seed but microsite availability (Shefferson et al., 2020). Hutchings (2010) reported site disturbance as a key driver of species recruitment of *O. sphegodes* and Gardiner and Vaughan (2009) outlined that site disturbance at road verges enhanced *O. apifera* abundance. The regular and dynamic creation of bare soil patches might be the main advantage of large herbivore grazing in comparison to mowing or small-scale rudimentary grazing. Intermediate disturbance often initiates positive processes such as species establishment and should not only be discussed as generally problematic in conservation and restoration ecology (Grubb, 1977).

In contrast to other reports of the failure of year-round horse grazing to counteract woody plant encroachment (e.g., Cosyns et al., 2001), our study revealed that woody plant cover only slightly increased under grazing impact in the long term, indicating the potential of horses to prevent significant increase of woody plant cover in low precipitation habitats with relatively low initial woody plant cover. In addition, only a few potentially invasive, stoloniferous woody species were present on the pasture (Köhler et al., 2016) and additional mineral supply was limited enhancing the intensity of browsing (especially bark peeling) to meet the mineral requirements particularly in the winter months (Putfarken et al., 2008). Moreover, only marginal areas needed additional restoration treatment such as mechanical shrub removal to achieve restoration targets compared to other similar pasture regimes (e.g., Henning et al., 2017a). For the
positive development of target bird communities, remaining individual shrubs as well as groups of shrubs and small forest patches should be considered as beneficial (Köhler et al., 2016; Lovász et al., 2021).

Besides grazing, there might be other factors influencing orchid density. *Ophrys apifera* shows a general increase and range extension probably due to climate warming in Saxony-Anhalt and Central Europe since the 1990s (Hutchings, 2010; Meysel, 2011; Osiadacz and Kręciala, 2014). Increases to such an extent have not been observed in other pastures in Saxony-Anhalt but the species often even decreased due to poor management. A not yet published dataset of the regional working group Arbeitskreis Heimische Orchideen Saxony-Anhalt e.V. shows clearly that in an area managed by mowing the number of *O. apifera* individuals remained stable over the same period, but did not increase. Yet, clarifying studies about climate impact on population growth and range expansion of *O. apifera* are missing. Fast population development in the pasture may also derive from relatively short reproduction cycles between germination and first appearance of *O. apifera*. Ziegenspeck (1936) suggested that the species can produce leaves after only 1 or 2 years.

## 4.3. Conclusion

Our results indicated that year-round horse grazing as a relatively new grassland restoration tool has the potential to enhance floristic biodiversity of dry calcareous grasslands not only in the short, but also in the long term. Particularly, the development of orchid species has to be evaluated differentially, when assessing the restoration success. Since pastures usually consist of heterogeneous vegetation types with varying suitability for orchid establishment, practitioners should not only rely on mean values concerning the entire pasture to regulate management but also on decrease and increase of individuals in individual vegetation types as the potential for local new establishment can strongly vary. In our study site, numbers of *O. apifera* increased most significantly in the prevailing medium dry calcareous grassland type and this type contributed most to the overall increase of individuals. Furthermore, our study clearly showed that even high-value xeric grasslands with very high orchid abundances can be integrated into the year-round grazing system also in the long term, and there is no need to the often recommended fencing off (exclosure) of such particularly valuable areas. Our study also revealed that short-term, especially one-year-only evaluation of restoration schemes can be misleading, particularly at slow-changing indicators such as dry calcareous vegetation types and at slow-growing and longer-lived species or those exhibiting high interannual population fluctuations such as many orchid species. Long-term monitoring schemes should be implemented to meet these requirements and to obtain found results on the effects of restoration and management measures.

## Data availability statement

The raw data supporting the conclusions of this article will be made available by the authors, without undue reservation.

## Ethics statement

Ethical review and approval was not required for the animal study because our work was a purely observational study and no ethical approval was needed. The GPS collar that provided our horse feeding

frequency data was fitted and regularly controlled on one female horse without using any external devices such as sedation. In order to access the telemetry data, no direct animal contact was needed. Written informed consent was obtained from the owners for the participation of their animals in this study.

## Author contributions

MK and ST developed the research questions and plot selection. MK collected the data and wrote the manuscript. MK and AS analyzed the data. AS, AB, NH, and ST contributed critically to the drafts and gave final approval for publication. All authors contributed to the article and approved the submitted version.

## Funding

Funding was provided by the European Agricultural Fund for Rural Development (EAFRD) and the state of Saxony-Anhalt (407.1.7-60128/323009000054, 323010000089, 323012000034, and 630116000006) and by the Graduate Scholarship Program of Saxony-Anhalt. We acknowledge support by the German Research Foundation (Deutsche Forschungsgemeinschaft DFG) -project number 491460386 - and the Open Access Publishing Fund of Anhalt University of Applied Sciences.

## Acknowledgments

The authors thank Georg Hiller and Frank Meysel as well as all students who helped with the data collection in the field, particularly horse observations, the Nature Conservation Authority Burgenlandkreis for implementing the financial support of the pasture, the Agrar GmbH Crawinkel for implementing the management, the owner of the site Naturstiftung David - Die Stiftung des BUND Thüringen and further cooperation partners.

## Conflict of interest

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

## Publisher's note

All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.

## Supplementary material

The Supplementary material for this article can be found online at:
https://www.frontiersin.org/articles/10.3389/fevo.2023.1107987/full#
supplementary-material
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<figure><img src="image_30.png" /></figure>

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EDITED BY

Péter Török,
University of Debrecen,
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Katherina Ng,
Australian National University,
Australia
Judit Sonkoly,
University of Debrecen,
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Sophie P. Ewert

Sophie.ewert@web.de

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CITATION

Ewert SP, Knörnschild M, Jung K and Frommolt
K-H (2023) Structurally rich dry grasslands –
Potential stepping stones for bats in open
farmland.

Front. Ecol. Evol. 11:995133.
doi: 10.3389/fevo.2023.995133

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Structurally rich dry grasslands –
Potential stepping stones for bats in
open farmland

Sophie P. Ewert¹,²*, Mirjam Knörnschild¹,³, Kirsten Jung⁴ and
Karl-Heinz Frommolt¹

¹Department of Evolutionary Diversity Dynamics, Museum für Naturkunde – Leibniz Institute for Evolution and Biodiversity Science, Berlin, Germany, ²Animal Behavior Lab, Department of Biology, Chemistry, and Pharmacy, Institute of Biology, Freie Universität Berlin, Berlin, Germany, ³Deutsche Fledermauswarte e.V., Berlin, Germany, ⁴Evolutionary Ecology and Conservation Genomics, Ulm University, Ulm, Baden-Württemberg, Germany

Agricultural intensification has caused decrease and fragmentation of European semi-natural dry grasslands. While a high biodiversity value of dry grasslands is acknowledged for plants and insects, locally and on landscape level, their relevance for mobile species, such as bats, is unknown. Here we investigate the use of dry grassland fragments by bats in an agriculturally intensified region in Germany and evaluate local and landscape factors influencing bat activity and assemblages. Specifically, we predicted that a combination of local dry grassland structural richness and landscape features as well as their interactions affect bat activity and foraging above dry grasslands. We also expected that these features influence compositions of local bat assemblages. We repeatedly sampled at 12 dry grassland plots with acoustic monitoring and assessed activity and foraging of bat species/sonotypes, which we grouped into guilds known for foraging in open land, at vegetation edges and in narrow spaces. We determined structural richness of the dry grassland plots in field and derived landscape features from digital landscape data. A relatively high proportion of bat species/sonotypes used dry grasslands regularly. The edge space foragers responded positively to higher local structural richness. Their dry grassland use increased when surrounding forests and woody features were less available, but they foraged more on dry grasslands closer to water bodies. Narrow space bat activity on dry grasslands decreased with less landscape connectivity. Open and narrow space foragers responded to local structural richness only in landscape context. For all bat guilds we found increased use of structurally richer dry grasslands when there was more open farmland in the surroundings. This was also the case for edge space foragers, when landscapes were more homogeneous. Lastly, with increasing structural richness, bat assemblages were more dominated by edge space foragers. We show the importance of European dry grassland fragments for the highly mobile group of bats under certain local structural and landscape compositional conditions. Our results underline the value of heterogeneous dry grassland fragments as potential stepping stones in intensively used farmland areas and contribute to evidence based decision making in dry grassland management and bat conservation.

KEYWORDS

dry grasslands, bats (Chiroptera), agricultural intensification, landscape, structural richness, acoustic monitoring, habitat fragments
# 1. Introduction

The agricultural intensification and industrialization, starting in the second half of the 20th century, have led to significant changes in land use and ecosystems worldwide. Both belong to the major suspected causes for global biodiversity loss (Díaz et al., 2019; Raven and Wagner, 2021) and population declines of various plant, insect and animal taxa (Burns et al., 2016; Young et al., 2016). Among the drivers of these detrimental effects are homogenization and simplification of landscapes, increased patch sizes of agricultural fields, reduced landscape structures (such as hedgerows) and loss of natural and semi-natural habitats (Matson et al., 1997; Stoate et al., 2001; Tilman et al., 2001; Jongman, 2002).

Semi-natural dry grasslands are open habitats, dominated by herbaceous (vascular) plants, mosses and lichens with occasional heterogeneous shrub and tree growth. In central Europe they have been significantly reduced during the 20th century, mainly due to land conversion, abandonment and forest encroachment, leaving small remnants fragmented in agricultural areas (Poschlod and Wallis de Vries, 2002; Habel et al., 2013; Vrahnakis et al., 2013; Leuschner and Ellenberg, 2017). Dry grasslands are known for their extremely high biodiversity in special plant communities, resulting from extremely nutrient poor, dry and sun exposed local conditions (Dengler et al., 2012; Wilson et al., 2012), and a rich insect fauna (Wallis de Vries and Van Swaay, 2009; Rasran et al., 2018; Schuldt et al., 2019). Therefore, they are now protected under the EU Habitats Directive (Council Directive 92/43/EEC). Conservation management targets these characteristic dry grassland plant species communities, by aiming to reduce nutrient inflow from agriculture (e.g., buffer zones) and by the application of grazing regimes, to reduce succession of shrubs and trees (Vrahnakis et al., 2013; Leuschner and Ellenberg, 2017). Especially smaller dry grassland fragments surrounded by intensively managed land are at greatest risk of further habitat degradation and loss of associated organisms due to external environmental influences (Butaye et al., 2005; Hodgson et al., 2005). However, exactly those small fragments have shown a great importance for species richness and communities of plants (Deák et al., 2020) and insects (Tscharntke et al., 2002; Rösch et al., 2015). Such positive effects on biodiversity are known to reach beyond their borders, affecting the species communities of the surrounding landscape. Lindborg et al. (2014) for example found increased plant richness on green islands in agricultural landscapes when dry grasslands were present in the vicinity. Their presence itself, as additional habitats in crop-dominated land, facilitates occurrence of various organisms and their exchange with adjacent landscape. This has been shown for insect communities (Tscharntke et al., 2002; Duelli and Obrist, 2003) and multiple taxa of insects and plants (Zulka et al., 2014). Dry grassland fragments can thus improve overall biodiversity of the landscape by providing local suitable conditions for specific species communities, which at the same time are influenced by the surrounding landscape. While research on plants and insects on dry grasslands has been plentiful so far, other more mobile organisms from higher trophic taxa have barely been taken into account. Therefore, researchers have recommended investigating the value of dry grasslands and their management regimes for other species and highlighted the importance of multi-scale approaches including local as well as surrounding landscape factors (Söderström et al., 2001; Zulka et al., 2014; Loos et al., 2021).

European bats are highly mobile insectivorous animals, protected by the EU Habitats Directive (Council Directive 92/43/EEC).

Consequences of agricultural intensification, such as habitat and connectivity losses, are among the major drivers of bat population declines (Browning et al., 2021). While mainly associated with forests and partly urban areas for roosting, bats utilize a variety of habitat types during their nightly foraging trips of several kilometers (Dietz and Kiefer, 2016). Their habitat use is thereby not only directed by insect food availability, but also by suitability of flight pathways (Kusch et al., 2004; Müller et al., 2012). The latter is defined for each bat species by their functional traits (e.g., echolocation system and wing morphology) which determine the abilities to navigate, maneuver and forage in different degrees of vegetation clutter (Aldridge and Rautenbach, 1987; Norberg and Rayner, 1987). Based on that, bat species can be categorized into three functional guilds related to their preferred commuting and foraging habitat of open space, edge structures (such as forest edges, hedgerows) or narrow space (for example inside forests; Fenton, 1990; Schnitzler and Kalko, 2001). Local clutter conditions, determined by the structural compositions of the surroundings, can thus facilitate or hinder movement of bats from functional guilds depending on their requirements. The composition of vegetation structures, such as shrubs and trees, has shown to be of influence on bat species occurrence and activity in studies in forest systems. These studies have found species-specific activity responses of bats to various horizontal and vertical structural forest stand features (Jung et al., 2012; Froidevaux et al., 2016; Blakey et al., 2017). Highest bat activity and species diversity have been recorded in heterogeneous forest plots with intermediate vegetation clutter on different height levels, offering various structural niches (Adams and Law, 2011; Hanspach et al., 2012; Jung et al., 2012; Starik et al., 2018). Hence, on dry grasslands, one would expect a positive influence of local heterogeneous shrub and tree encroachment on bat occurrence and activity. However, dry grassland fragments, especially in open and homogeneous agricultural areas, might often be isolated from other main habitats for roosting and foraging or from structural landscape features, such as hedgerows, that offer connectivity networks as sonar guidelines, with higher prey availability and as possible protection from predators (Verboom and Huitema, 1997). When commuting through these kind of landscapes, open habitat adapted bat species do not face major constraints, while others, which are adapted to commuting and hunting closer to edges or cluttered vegetation, are likely to reach a limit in how far they move through open areas (Frey-Ehrenbold et al., 2013). These sensory constrains and habitat associations underline the necessity of a multi-scale approach when investigating the use of dry grasslands by bats embedded in open agricultural landscapes. After all, while local dry grassland conditions can offer abundant resources for commuting and foraging, the features of the surrounding landscape predominantly determine the accessibility of dry grassland fragments (Boughey et al., 2011b; Kalda et al., 2015a,b; Froidevaux et al., 2022).

A landscape should enable dispersal and support of a diverse species pool by providing a variety of resources in non-crop habitats, which is however usually limited in agriculturally intensified regions (Tscharntke et al., 2005). It is therefore crucial to understand how remaining (semi-) natural habitat fragments in farmland, such as dry grassland systems, affect endangered species. The value of dry grassland fragments for highly mobile protected bats is however still unknown. Identifying which conditions at local and landscape scales are important for bats to occur and forage on dry grasslands, adds evidence to the biodiversity value of dry grassland fragments and contributes to informed decision making for dry grassland management and bat conservation efforts. Our study therefore focuses on investigating the use of dry grassland
fragments by bats in context of features of local and landscape scales as well as their interactions in an agriculturally rich region in Germany. We described the local conditions of dry grasslands with structural richness provided by shrubs and trees. In addition, we evaluated features describing openness, homogeneity and connectivity of the surrounding landscape as well as the availability of other main habitat types associated with foraging and roosting of bat species. We hypothesized that a combination of local and landscape effects as well as their interactions determines the use of dry grasslands by bats. Specifically, we expected that higher local structural richness on dry grassland fragments would increase bat activity and foraging [as in forests in Jung et al. (2012)]. Furthermore, we predicted effects of landscape features [as in Heim et al. (2015)] on use of dry grasslands by bats, and anticipated particularly the edge and narrow space adapted bats to show greater accessibility limitations (Frey-Ehrenbold et al., 2013). Most importantly however, we expected that landscape features would influence the correlation of bat activity with local structural richness [such as Froidevaux et al. (2022)]. Lastly, we predicted a dominance shift of functional bat guilds in bat assemblages above dry grasslands with changes in dry grassland and landscape features [given their different habitat preferences as in Schnitzler and Kalko (2001)].

# 2. Materials and methods

## 2.1. Study region

This study has been conducted on designated protected dry grassland fragments (Figure 1) in the study region North-East Brandenburg in Northeastern Germany. Brandenburg is characterized by a comparatively low human population density. It is a young glacial landscape with many wetlands. The district Uckermark (Statistical Office Berlin-Brandenburg, 2020), where the majority of dry grassland fragments from this study are located, consists of roughly 60 percent of agricultural land (mainly used for cereal grains and a low proportion for pastures). About a quarter of the landscape is covered by forest, predominantly coniferous forests, followed by deciduous and a few floodplain forests. Although there are some fragmented forest patches, the majority of forests in the study region are rather continuous and extensive. This also means that in-between, large areas are covered by open land. The dry grassland fragments are typically rather small (in this study: min. 2 ha, max. 41 ha, mean 16 ha) and sometimes long and narrow, as they have been pushed to margins of agriculturally unusable land (e.g., slopes). They are found on dry, nutrient poor and sun exposed locations, reaching sub-continental extreme microclimatic conditions. The prominent type of management to reduce shrub and tree encroachment is grazing by sheep for a short period once a year. Brandenburg carries great conservation responsibility, containing Germany's largest proportion of xeric sand calcareous and sub-pannonic steppic grassland (Landesamt für Umwelt Brandenburg, 2022). The latter type is dominating in the study region and is characterized by an open landscape with a high number of vascular plants, mosses and lichens. The dry grassland fragments chosen for this study showed a range of cover of woody structures (e.g., shrubs, small trees) from none up to 30%. These conditions are described as excellent to mediocre, respectively, in conservation value (Landesamt für Umwelt Brandenburg, 2022) and underline the heterogeneity of dry grasslands. This is especially relevant when seen in comparison to other grassland types (pastures and meadows) in the surroundings, which are mainly used for

cattle grazing and are usually very homogeneous and lacking structural richness. Our dry grassland fragments differed not only in structural compositions but also in embedment in the landscape (regarding for example distances to other habitats).

## 2.2. Acoustic monitoring

We used Batlogger (Batlogger A, Elekon AG, Luzern, Switzerland) for repeated passive acoustic monitoring to assess bat activity above dry grassland fragments. Batlogger microphones were mounted on poles in the center of 12 designated study plots, which were intended for vegetation assessment (see 2.4), two meters above the ground and at least two meters away from larger trees to avoid strong echoes. We mounted one Batlogger per dry grassland fragment due to their small sizes, except on two larger fragments, where we used two plots each (see Figure 1). The locations of the plots were chosen in field depending on local structural conditions to cover the wide range of structural compositions. Their centers, hence Batlogger locations, were at least 500 m away from each other to prevent spatial autocorrelation and at least 30 m away from adjacent habitat types (e.g., crop and forest) to reduce edge effects. Note, however, that effects of other habitats were accounted for in the following landscape analysis. Each Batlogger sampled four nights (except at one plot, which was sampled three times due to access restrictions) throughout the main bat activity season (five plots in 2019, seven plots in 2020). Two sampling nights took place between June – July and (one) two between August – October. This resulted in 47 sampling nights. Recording started around 30 min before sunset and continued until approximately 30 min after sunrise. For triggering of sound recordings, we used the Crest Advanced setting that ensured the recording of quiet bat calls and reduced sensitivity to disturbing noise (Supplementary Datasheet S1). Sampling was avoided in nights with prolonged rainfall, high wind speed (>BFT 4; Martin et al., 2017) and low temperatures <8°C. We also avoided recording during sheep and cattle grazing, as movement of cattle can potentially increase bat activity (Downs and Sanderson, 2010; Ancillotto et al., 2017).

## 2.3. Bat identification and activity

We performed semi-automatic call analysis in the program Batexplorer 2.1 (Elekon AG, Luzern, Switzerland) to assign recordings with bat echolocation calls to species level or to an identifiable sonotype (groups of bat species with very similar call structures). This was based on a self-created list of identification criteria (Supplementary Table S1) derived from various sources (Parsons and Jones, 2000; Russo and Jones, 2002; Pfalzer and Kusch, 2003; Obrist et al., 2004; Skiba, 2009; Barataud, 2020). All recordings were checked by hand to reduce misidentification (Russo and Voigt, 2016). We analyzed recordings with sequences containing two or more recognizable bat echolocation signals. Non-identifiable bat calls were omitted from further analysis. We also identified recordings with one or more feeding buzzes (distinct phases with rapid, steep calls with a high frequency bandwidth) as indications for insect prey encounter and therefore foraging activity. We grouped all bat species/sonotypes into three functional bat guilds by similarities in eco-morphological adaptations and in their preferred vegetation clutter conditions of habitats (Schnitzler and Kalko, 2001; Zeale et al., 2012; Dietz and Kiefer, 2016; Barataud, 2020): (i) open space guild (open
<figure><img src="image_31.png" /></figure>

Map created with QGIS 3.16 in Dec. 2022, by S.P.Ewert; Basemap: ESRI Satellite (ArcGIS/World_Imagery); Spatial data: ATKIS-Basis-DLM © GeoBasis-DE/LGB, dide/by-2.0, 2021; SWF 2015 © European Union, Copernicus Land monitoring Service, 2020

<caption>FIGURE 1</caption>

Map of study region in Brandenburg in Germany with the landscape types, locations of the dry grassland plots (n=12) and the landscape radii for analysis.

habitats, high and fast flying bats with low maneuverability, narrowband and low-frequency echolocation calls), (ii) edge space guild (uses background cluttered habitats, such as forest edges, but is flexible in open and clutter; bats with intermediate speed and maneuverability; medium frequency echolocation calls with steep broadband and narrowband components), and (iii) narrow space guild (uses cluttered habitats close to vegetation; bats with low flight speed and high maneuverability; steep broadband echolocation calls). For a measure of habitat use of each guild, we took bat activity, which was defined by the recorded number of minutes of one night per plot, in which we found at least one echolocation recording (with a minimum of two calls) of the respective bat guild (Miller, 2001). We did the same for recordings including feeding buzzes as a measure of foraging activity. Feeding buzzes were not sampled for bats from the narrow space guild, as they mainly use gleaning and passive listening to find insects, instead of feeding buzzes (Entwistle et al., 1996; Schnitzler and Kalko, 2001).

## 2.4. Local structural richness of dry grasslands

From forest systems it is known, that higher three-dimensional structural heterogeneity supports a higher number of bat species and their activity (Adams and Law, 2011; Hanspach et al., 2012; Jung et al., 2012; Starik et al., 2018). We thus used an index that describes vertical as well as horizontal structural richness of the dry grassland

plots. Therefore, we used the foliage height diversity index (FHD), which was originally proposed by MacArthur and MacArthur (1961) as a description of heterogeneity in vegetation layer coverage and to predict bird species richness. It was later adapted in other studies for birds (Wood et al., 2013) and for bat occurrence in forest systems (Froidevaux et al., 2016). For calculating the FHD as a measure of structural richness, we applied the Shannon-Wiener Index [in R 4.0.3. by R Core Team (2020); package: vegan, Oksanen et al., 2020; function: diversity], as proposed in Froidevaux et al., (2016). For that, we used cover values of foliage height classes: ground 0–1 m, shrub 1–5 m, tree low 5–10 m, tree high >10 m. To assess vegetation cover per height class, we did in-field vegetation sampling of the study plots within a 25 m radius around the Batlogger locations. We identified the positions and average heights of woody vegetation structures (shrubs and trees) that were taller than 0.5 m and drew the vegetation layout true to scale on a grid map. For height estimation of larger trees, we used a tree-height-measuring app (App: “Baumhöhenmesser,” Foresttools, Göttingen, Germany). In the following, looking at the drawn map and knowing the heights of the woody vegetation, we estimated the amount of square meters covered by woody vegetation in each defined height class and then calculated the percentage of the total plot covered by each height class. The FHD correlated positively with other descriptives of the dry grassland plots (cover values of height classes and total foliage cover) with Pearson coefficients from 0.8 and higher [in R 4.0.3. by R Core Team (2020), package: Ggally, by
Schloerke et al. (2021), function “ggcorr”]. We therefore considered the FHD as an appropriate proxy for the description of local structural richness.

## 2.5. Landscape features

For the landscape analysis, we considered landscape features that are of known relevance for bats and would aid in describing the landscape in terms of openness, homogeneity, connectivity and availability of resources for bats. We thus included forests and urban areas, as important roosting and foraging habitats (e.g., Dietz and Kiefer 2016), as well as water bodies, offering valuable foraging sites especially in open land (e.g., Heim et al., 2018). We also considered woody features (e.g., hedges, tree lines and small woodland patches), as important connectivity features for many bat species using them for sonar guidance, protection and foraging (e.g., Verboom and Huitema, 1997). In addition, we included open land, consisting of pastures and predominantly cropland, known to negatively affect bat activity (e.g., Put et al., 2019). Lastly, we considered patchiness of the landscape as a measure for homogeneity. A homogeneous landscape, with less patches, can be detrimental for some bat species (e.g., Mendes et al., 2017) as it provides less heterogeneous niches and edge habitats for commuting and hunting (Ethier and Fahrig, 2011). The landscape features and the used measurements can be found in Table 1. To derive measurements of landscape features we performed landscape analysis (Supplementary image S1) within a 200 m radius around the Batlogger position using the geographic information system QGIS 3.16.3 (QGIS Development Team, 2021). We adapted a 200 m radius, as this distance has already shown significant effects of landscape features on bat activity and species composition, for example in Heim et al. (2015). We used layers with detailed landscape data (spatial data: ATKIS-Basis-DLM, source: ®GeoBasis-DE/LGB, dl-de/by-2-0, 2021). The categories of the landscape

layers were then reclassified into broader categories: urban area, water body, forest and open land (pastures + cropland; Supplementary Table S2) and used for subsequent spatial analysis in QGIS. Furthermore, we used a high-resolution layer of woody features, which consisted of patchy, linear and additional woody structures, such as hedges and tree lines (spatial data: Small Woody Features 2015, source: ®European Union, Copernicus Land monitoring Service, 2020). Woody features directly connected to forest were reclassified as forest when they were protruding not more than 5 m from the forest. Further, we corrected major misclassification errors in landscape categorizations by checking in the field. We then calculated within the 200 m radius the cover of forest, open land (pastures + cropland) and woody features in percentage. Additionally, we counted the number of landscape patches regardless of their types within the 200 m radius. Separate patches could be for example landscape features distinct from their surroundings (e.g., forest or woody patches) or when divided by paths or roads (e.g., separate fields of crop or pastures). Lastly, we measured the linear distances from the Batlogger locations to the nearest urban areas, woody features and water bodies. We checked all variables for collinearity and found Pearson coefficients only between -0.6 and 0.6, indicating no notable correlations [in R 4.0.3. by R Core Team (2020), package: Ggally, by Schloerke et al. (2021), function “ggcorr”].

## 2.6. Statistical analysis

The following statistical analyses were performed in R (version 4.0.3, R Core Team, 2020).

### 2.6.1. Modeling responses to local and landscape features

To assess whether bat activity and foraging of each functional guild responded to increased structural richness (FHD) of dry grassland

<table><caption>TABLE 1 Final choice of local and landscape features and their abbreviations used for further analysis.</caption><thead><tr><th>Local & Landscape features</th><th>Abbreviation</th><th>Relevance for bats found in:</th></tr></thead><tbody><tr><td>Measurements</td><td></td><td></td></tr><tr><td></td><td colspan="2">Dry grassland (r = 25 m)</td></tr><tr><td>Foliage height diversity (structural richness)</td><td>FHD</td><td>Froidevaux et al. (2016)</td></tr><tr><td></td><td colspan="2">Landscape (r = 200 m)</td></tr><tr><td>Forest:<br>Percentage cover</td><td>For</td><td>Fuentes-Montemayor et al. (2013); Heim et al. (2015, 2018); Kalda et al. (2015b); Treitler et al. (2016)</td></tr><tr><td>Open land (pastures + cropland):<br>Percentage cover</td><td>Open</td><td>Put et al. (2019)</td></tr><tr><td>Patches:<br>Counted number</td><td>Patch</td><td>Mendes et al. (2017)</td></tr><tr><td>Urban area:<br>*Distance</td><td>DistUrb</td><td>Fuentes-Montemayor et al. (2013); Heim et al. (2015)</td></tr><tr><td>Water bodies:<br>*Distance</td><td>DistWat</td><td>Downs and Racey (2006); Stahlschmidt et al. (2012); Fuentes-Montemayor et al. (2013); Heim et al. (2015, 2018); Mendes et al. (2017)</td></tr><tr><td>Woody features:<br>Percentage cover<br>*Distance</td><td>Wood<br>DistWood</td><td>Verboom and Huitema (1997); Downs and Racey (2006); Boughey et al. (2011b); Frey-Ehrenbold et al. (2013); Kelm et al. (2014); Heim et al. (2015); Kalda et al. (2015a); Toffoli (2016)</td></tr></tbody></table>

Right column shows examples of literature where the landscape variable has shown a relevance for bats. *Distances were also measured when reaching outside of the 200 m radius.
<table><caption>TABLE 2 List of identified species/sonotypes in their associated functional guilds and the accumulated number of activity/foraging minutes based on 47 sampling nights.</caption><thead><tr><th rowspan="2">Species/sonotypes</th><th colspan="2">Total numbers of minutes</th></tr><tr><th>Activity</th><th>Foraging</th></tr></thead><tbody><tr><td colspan="3"><strong>Open space guild</strong></td></tr><tr><td><em>Eptesicus serotinus</em></td><td>177</td><td>3</td></tr><tr><td><em>Nyctaloid</em></td><td>424</td><td>23</td></tr><tr><td><em>Nyctalus leisleri</em></td><td>139</td><td>6</td></tr><tr><td><em>Nyctalus noctula</em></td><td>2,278</td><td>215</td></tr><tr><td colspan="3"><strong>Edge space guild</strong></td></tr><tr><td><em>Barbastella barbastellus</em></td><td>2</td><td>0</td></tr><tr><td><em>Pipistrellus high</em></td><td>177</td><td>15</td></tr><tr><td><em>Pipistrellus low</em></td><td>168</td><td>7</td></tr><tr><td><em>Pipistrellus nathusii</em></td><td>1,180</td><td>49</td></tr><tr><td><em>Pipistrellus pipistrellus</em></td><td>784</td><td>56</td></tr><tr><td><em>Pipistrellus pygmaeus</em></td><td>923</td><td>107</td></tr><tr><td colspan="3"><strong>Narrow space guild</strong></td></tr><tr><td><em>Myotis myotis</em></td><td>11</td><td>0</td></tr><tr><td><em>Myotis spp.</em></td><td>215</td><td>2</td></tr><tr><td><em>Plecotus spp.</em></td><td>39</td><td>0</td></tr></tbody></table>

<figure><img src="image_32.png" /><figcaption>FIGURE 2
Nightly activity minutes of bat guilds (open, edge, and narrow space) counted during 47 nights on 12 dry grassland plots.</figcaption></figure>

plots, as well as to landscape features and/or the interaction of both, we applied information-theoretic model selection with generalized linear mixed models (R package: glmmTMB, Brooks et al., 2017). This approach was chosen over an *a priori* set of candidate models, as we expected all predictors to have effects on bat activity and foraging to some extent (see “Landscape features”) and it allowed us to find the most explanatory set of predictors (Bolker et al., 2009; Frey-Ehrenbold et al., 2013; Kalda et al., 2015a; Mendes et al., 2017). We used a negative binomial link due to overdispersion. FHD and landscape features were taken as fixed effects and mean nightly temperatures and the IDs of the

dry grassland plots were taken as random effects. All predictors were scaled using the scale function (Becker et al., 1988). We conducted stepwise forward selection while keeping FHD in every model and in a stepwise manner adding landscape features as main effects and then as interaction terms with FHD, looking for the model with the best AIC value indicating the most parsimonious model (Supplementary Table S3). We chose forward selection for the following reasons: first, to reduce the risk of convergence issues due to overly complex models; secondly, to mirror our research questions approach, by initially testing for effects of dry grassland structural richness and then adding features from the broader landscape scale. Model selection was performed for all guilds and their activity as well as their foraging activity (except for the narrow space guild).

### 2.6.2. Functional composition of bat assemblages

From the cumulative total bat activity of all three functional guilds per night and plot, we calculated the percentages of nightly activity per guild. From there we derived the median activity per dry grassland plot for each guild. This made the dry grassland plots comparable in their proportions of functional guild activity counts. To investigate the proportional differences on dry grassland fragments we performed non-metric multidimensional scaling (NMDS in R package: vegan; function: metaNMDS, Oksanen et al., 2020). We then applied environmental fitting of the dry grassland and landscape features onto the ordination (R package: vegan, Oksanen et al., 2020; function: envfit) to investigate their potential to explain proportional changes of bat assemblages between dry grassland plots. Results were supported by using permutational multivariate analysis of variance [PERMANOVA in R package: vegan, Oksanen et al., 2020; function: adonis, based on Anderson (2001)].

## 3. Results

### 3.1. General results

At 12 dry grassland plots, during a total of 47 sampling nights, we found 13 bat species/sonotypes that we had grouped into the three functional guilds (Table 2) and 5,488 minutes with echolocation activity of one or more bat species/sonotypes. Bats were active during 19% of the total recording time (28,560 min) in which one or more bats could have been recorded. Per night at a plot, we counted on average almost 2 hours of bat activity (mean: 117 min). Eight percent of activity minutes contained feeding buzzes of one or more bat species/sonotypes from the open and edge space guild, indicating foraging activity. Nightly activity (Figure 2) mainly consisted of the open and the edge space guild. Activity of bats from the narrow space guild was generally much lower. The activity of the open space guild was dominated by *Nyctalus noctula* (75%). The edge space guild showed shared dominances of *Pipistrellus nathusii* (37%), *Pipistrellus pygmaeus* (29%) and *Pipistrellus pipistrellus* (23%). Lastly, the *Myotis* group (81%) formed the largest part of the narrow space guild. Each dry grassland plot was visited by an average of 11 bat species/sonotypes throughout the year, with a nightly mean of eight, indicating a regular use by a high proportion of the recorded bat species/sonotypes. Additionally, we recorded at least three species/sonotypes of higher conservation concerns according to the German Red List (Meinig et al., 2020): *Eptesicus serotinus* (threatened) on 11 plots, *Plecotus* spp. (threatened) on 9 plots and *Barbastella barbastellus* (highly threatened) on one plot.
## 3.2. Responses to local and landscape features

Model selections revealed that a combination of both local (plot level) and landscape scales, and especially their interactions, is important for explaining activity (Table 3) and foraging (Table 4) of bat functional guilds at the dry grasslands.

Bats from the edge space guild showed strong associations with local structural conditions. Their activity generally increased significantly with higher structural richness (FHD), regardless of landscape interactions. We also found significant additive effects of features on the landscape scale on edge space bats. They foraged more above dry grassland plots when water bodies were closer by, indicating associations with other main habitat types. However, they were more active on dry grasslands plots when there was less forest in the surrounding and their foraging activity increased with greater distances to woody features (Figure 3), suggesting an increased use of dry grasslands when other habitat sources are less available. Activity of bats from the narrow space guild decreased at dry grassland plots further away from woody features (Figure 3), showing a stronger accessibility limitation.

For all functional guilds, we found significant interaction effects of local structural richness in combination with landscape features describing openness, homogeneity and availability of main habitats. The activity of open and narrow space bats (Table 3) as well as the foraging activity of open and edge space bats (Table 4) increased above dry grassland plots with more local structural richness, when they were

surrounded by a higher percentage of open land (pastures and cropland). In less open landscapes, structural richness was less important for the open space guild (Figures 4A,B) and the narrow space guild (Figure 4D), while in the foraging activity of the edge space guild, the effect was still positive but less prominent (Figure 4C). Additionally, edge space bat foraging increased on structurally richer dry grassland plots, when the number of patches in the surrounding decreased (Figure 5). Lastly, structurally richer dry grassland plots were used more by open space bats (increase in activity and foraging) and foraged more by edge space bats, when urban areas were further away.

## 3.3. Functional compositions of bat assemblages

Non-metric multidimensional scaling separated dry grassland plots based on the median percentage of nightly activities of the function guilds (Figure 6; NMDS: k=2, stress=0.006, perm=999, Supplementary Table S4). Environmental fitting including local and landscape features showed that the structural richness (FHD) significantly explained separation of dry grassland plots along the y-axis ($r^2$=0.68, *p*=0.010). Especially the activity proportion of edge space guild increased with higher structural richness. This was supported by results of the PERMANOVA (FHD: $r^2$=0.54, *p*=0.022, Supplementary Table S5) and suggests that structural composition of dry grassland fragments influences assemblage composition of bats. Landscape features did not show significant effects.

<table><caption>TABLE 3 Results of bat guild activity predictions from the forward selections of generalized linear mixed models (GLMM) with negative binomial links based on selection of the best Akaike's information criterion (AIC).</caption><thead><tr><th>GLMM model</th><th>Parameter</th><th>Estimate</th><th>Standard error</th><th>Z value</th><th>Pr(&gt;|z|)</th></tr></thead><tbody><tr><td>Open space guild activity</td><td>(Intercept)</td><td>3.1427</td><td>0.2767</td><td>11.360</td><td>***</td></tr><tr><td rowspan="6">AIC: 468.1</td><td>FHD</td><td>-0.1365</td><td>0.1291</td><td>-1.057</td><td>n.s.</td></tr><tr><td>Open</td><td>0.1408</td><td>0.1343</td><td>1.048</td><td>n.s.</td></tr><tr><td>DistUrb</td><td>0.6843</td><td>0.1945</td><td>3.518</td><td>***</td></tr><tr><td>Wood</td><td>-0.3267</td><td>0.1907</td><td>-1.713</td><td>.</td></tr><tr><td>FHD*Open (Figure 4)</td><td>1.1413</td><td>0.2499</td><td>4.567</td><td>***</td></tr><tr><td>FHD*DistUrb</td><td>0.9694</td><td>0.2784</td><td>3.482</td><td>***</td></tr><tr><td rowspan="3">Edge space guild activity</td><td>(Intercept)</td><td>3.8142</td><td>0.2088</td><td>18.264</td><td>***</td></tr><tr><td>FHD</td><td>0.3939</td><td>0.1656</td><td>2.379</td><td>*</td></tr><tr><td>For</td><td>-0.5750</td><td>0.2398</td><td>-2.398</td><td>*</td></tr><tr><td>Narrow space guild activity</td><td>(Intercept)</td><td>0.88980</td><td>0.27875</td><td>3.192</td><td>**</td></tr><tr><td rowspan="6">AIC: 257.4</td><td>FHD</td><td>0.40062</td><td>0.28511</td><td>1.405</td><td>n.s.</td></tr><tr><td>Open</td><td>0.14072</td><td>0.15188</td><td>0.927</td><td>n.s.</td></tr><tr><td>DistWood (Figure 3)</td><td>-0.76303</td><td>0.28555</td><td>-2.672</td><td>**</td></tr><tr><td>Patch</td><td>0.05497</td><td>0.26329</td><td>0.209</td><td>n.s.</td></tr><tr><td>FHD*Patch (Figure 4)</td><td>1.10942</td><td>0.31057</td><td>3.572</td><td>***</td></tr><tr><td>FHD*Patch</td><td>0.64107</td><td>0.37177</td><td>1.724</td><td>.</td></tr></tbody></table>

Listed are the final models for activities of open, edge and narrow space guilds based on 47 nights at 12 grassland dry grassland plots and the habitat and landscape features (refer to Table 1). Positive estimate values (+) indicate positive effects and negative values (-) negative effects. n.s., non-significant; "*p"<0.1; "*p"<0.05; "**p"<0.01; "***p"<0.001.
<table><caption>TABLE 4 Results of bat guild foraging activity predictions from the forward selections of generalized linear mixed models (GLMM) with negative binomial links based on selection of the best Akaike's information criterion (AIC).</caption><thead><tr><th>GLMM model</th><th>Parameter</th><th>Estimate</th><th>Standard error</th><th>Z value</th><th>Pr(&gt;|z|)</th></tr></thead><tbody><tr><td>Open space guild foraging</td><td>(Intercept)</td><td>0.76786</td><td>0.42613</td><td>1.802</td><td>.</td></tr><tr><td>AIC:252.1</td><td>FHD</td><td>-0.04104</td><td>0.35316</td><td>-0.116</td><td>n.s.</td></tr><tr><td></td><td>Open</td><td>0.59091</td><td>0.29060</td><td>2.033</td><td>*</td></tr><tr><td></td><td>DistUrb</td><td>0.74269</td><td>0.35610</td><td>2.086</td><td>*</td></tr><tr><td></td><td>FHD*Open (Figure 4)</td><td>1.47709</td><td>0.60054</td><td>2.460</td><td>*</td></tr><tr><td></td><td>FHD*DistUrb</td><td>1.10715</td><td>0.55029</td><td>2.012</td><td>*</td></tr><tr><td>Edge space guild foraging</td><td>(Intercept)</td><td>-0.4091</td><td>2.1632</td><td>-0.189</td><td>n.s.</td></tr><tr><td>AIC: 221.8</td><td>FHD</td><td>-2.3540</td><td>2.3254</td><td>-1.012</td><td>n.s.</td></tr><tr><td></td><td>DistWat</td><td>-3.4008</td><td>0.9687</td><td>-3.511</td><td>***</td></tr><tr><td></td><td>DistUrb</td><td>0.9301</td><td>0.9068</td><td>1.026</td><td>n.s.</td></tr><tr><td></td><td>Open</td><td>1.2013</td><td>0.2157</td><td>5.568</td><td>***</td></tr><tr><td></td><td>Patch</td><td>-3.9534</td><td>0.7601</td><td>-5.201</td><td>***</td></tr><tr><td></td><td>DistWood (Figure 3)</td><td>2.3413</td><td>1.1500</td><td>2.036</td><td>*</td></tr><tr><td></td><td>FHD*DistUrb</td><td>3.4652</td><td>1.3265</td><td>2.612</td><td>**</td></tr><tr><td></td><td>FHD*Open (Figure 4)</td><td>4.0477</td><td>0.7864</td><td>5.147</td><td>***</td></tr><tr><td></td><td>FHD*Patch (Figure 5)</td><td>-8.9213</td><td>1.8411</td><td>-4.846</td><td>***</td></tr><tr><td></td><td>FHD*DistWood</td><td>1.9979</td><td>1.8752</td><td>1.065</td><td>n.s.</td></tr></tbody></table>

Listed are the final models for foraging activities of open, edge and narrow space guilds based on 47 nights at 12 grassland dry grassland plots and the habitat and landscape features (refer to Table 1). Positive estimate values (+) indicate positive effects and negative values (–) negative effects. n.s., non-significant; “:p<0.1; *:p<0.05; **:p<0.01; ***:p<0.001.

## 4. Discussion

Loss of (semi-) natural habitats and ecosystem changes are consequences of agricultural intensification, which is globally identified as the major cause of the current biodiversity crisis (Burns et al., 2016; Young et al., 2016; Díaz et al., 2019; Raven and Wagner, 2021). The presence of non-crop habitats in intensified agro-ecosystems is crucial to enable occurrence and dispersal of a variety of organisms that are necessary to sustain a resilient species pool and associated ecosystem services (Tscharntke et al., 2005; Fischer et al., 2006). In this study we focused on fragmented, protected European semi-natural dry grasslands, which are known for a very high plant biodiversity (Dengler et al., 2012; Wilson et al., 2012) and a rich insect fauna (Wallis de Vries and Van Swaay, 2009; Rasran et al., 2018; Schuldt et al., 2019). An increased value has been particularly proven for small dry grassland fragments in agro-ecosystems, locally (Rösch et al., 2015; Deák et al., 2020) and in exchange with the surrounding landscape (Tscharntke et al., 2002; Duelli and Obrist, 2003; Lindborg et al., 2014; Zulka et al., 2014). However, their importance for highly mobile species, such as bats, is not yet understood, even though many bat species face challenges of agricultural intensification, such as habitat and connectivity losses (Browning et al., 2021) and when moving through open land (Frey-Ehrenbold et al., 2013). Revealing which local structural conditions and landscape features as well as their interactions increase commuting and foraging of bats on dry grassland fragments, is of utmost importance for bat conservation and for evidence-based decision making in management and conservation efforts of dry grasslands. Our results shed light onto the importance of dry grassland fragments for bats in a farmland-dominated region in

Germany in relation to local vegetation structural richness and the availability of other important habitats. This is especially relevant in agriculturally rich regions, where other non-crop habitats are reduced (Burns et al., 2016) or more distant. We show regular visitations by a relatively high number of bat species/sonotypes throughout the year, also by species/sonotypes of higher conservation concern. Our results clearly show, that dry grassland use by bats is determined by local structural richness, the features of the surrounding landscape and the interaction of both.

### 4.1. Local and landscape features determine dry grassland use

As predicted, our results show additive effects of local and landscape features on dry grassland use by bats. This is in accordance with findings of other studies that highlighted the influence of landscape features at different scales on bat occurrence and their activities (Boughey et al., 2011a; Kalda et al., 2015b; Heim et al., 2017; Mendes et al., 2017).

Bats from the edge space guild appeared to be the most responsive to local conditions of dry grassland fragments. Their activity above dry grassland plots generally increased with a higher structural richness, hence more heterogeneous shrubs and trees. This is corroborated by findings of previous studies, showing that edge space bats prefer moving along structural features (Verboom and Huitema, 1997; Downs and Racey, 2006; Toffoli, 2016), specifically close to taller trees (Froidevaux et al., 2022) and it suggests that structurally richer dry grasslands have a higher value than very open ones, for edge space bats commuting in relatively open agricultural
A

<figure><img src="image_33.png" /></figure>

B

<figure><img src="image_34.png" /><figcaption>FIGURE 3
Effect plot for the predicted effects of distance to woody features (scaled) on (A) edge space guild foraging and (B) narrow space guild activity.</figcaption></figure>

landscapes settings. When looking at landscape features, edge space
bats were more active above dry grassland plots when less forest was
available in the surrounding, and they foraged more at dry grassland
plots when woody features (such as hedgerows) were further away.
This is opposite to findings of other studies that show higher activity
of the edge space bats (such as *Pipistrellus* species) close to forest
(Davidson-Watts et al., 2006), forest edges (Heim et al., 2018) and
woody features (Verboom and Huitema, 1997; Downs and Racey,
2006; Frey-Ehrenbold et al., 2013). However, the latter studies also
highlight that *Pipistrellus* species can cover greater distances in open
land away from favorite habitats to reach other favorable land
patches. According to Kelm et al. (2014), the biggest activity drop of
*Pipistrellus* species away from structural edges into open land is
usually within the first 50 m. Since we show increased activity at dry
grassland plots away from woody features (here average distance:
44 m) and with less surrounding forest cover, we propose that dry
grassland fragments are possibly favorable habitats for bats from the
edge space guild (see also Davidson-Watts et al., 2006 for *Pipistrellus*
species in unimproved grasslands) and potentially fulfill an
important function for commuting and foraging. These results are
not supporting our expectation of accessibility limitations for edge
space bats, although this could possibly become the case in regions
where dry grasslands fragments are too isolated from other
connecting landscape features (Frey-Ehrenbold et al., 2013). In
contrast, but as expected, the narrow space guild exhibited stronger
accessibility restrictions by showing reduced activity on dry
grassland plots further away from woody features. This is in
accordance with findings from Frey-Ehrenbold et al. (2013) showing
the greatest limitations in open land for the narrow space bats of the
*Myotis* and *Plecotus* groups. Therefore, isolated dry grasslands
fragments might eventually fall out of their reach, if they are beyond
their accessibility threshold.

Our results also highlight the importance of other habitat types in
the landscape. Edge space bats foraged more on dry grassland plots
closer to waterbodies. The high value of the latter for various bat species

as foraging ground has been supported by other studies before
(Stahlschmidt et al., 2012; Kalda et al., 2015a; Heim et al., 2018).
Wetlands are also among the most preferred habitats for the edge space
bat species *Pipistrellus pipistrellus* and *Pipistrellus pygmaeus* (Davidson-
Watts et al., 2006; Dietz and Kiefer, 2016).

We did not find independent effects of local structural richness
on bats from the open and the narrow space guild, but only in
interaction with landscape features, indicating that for these bats
surrounding conditions determine the importance of dry grasslands
and their structures (discussed below). We also did not find
significant effects of surrounding forest cover on bat activity and
foraging in the open and narrow space guilds, even though other
studies have shown positive associations of several bat species with
forest proximity (Frey-Ehrenbold et al., 2013; Treitler et al., 2016)
and cover (Boughey et al., 2011a; Heim et al., 2015; Mendes et al.,
2017). We suppose that changes in the overall low forest cover in the
200 m radius (mean: 7%) were irrelevant for the fast flying
open space bats and that the long distances (mean: 153 m) excluded
many woodland dwelling narrow space species, keeping their
activity low.

Our results here indicate the use of dry grasslands by bats in relation
to local conditions and to availability of other main bat habitats and
landscape features, which mainly affected bats from the edge and narrow
space guild. While dry grassland fragments appear beneficial for
commuting and foraging of bats from the edge space guild, especially
when containing heterogeneous vegetation structures, narrow space bats
show a dependency on higher connectivity to other structural landscape
features (such as hedgerows). Our findings therefore also underline the
necessity of connectivity networks and corridors in open farmland
(Rosenberg et al., 1997) and the availability of other important habitats
(Heim et al., 2015) to facilitate bat dispersal and provide
additional resources.

## 4.2. Landscape setting determines the importance of local conditions for bats

We furthermore confirmed our prediction that landscape
features would influence the correlations with local structural
conditions and therefore determine dry grassland use. Key findings
in this study are the impacts of the amount of open land (pastures
and cropland) and of the number of landscape patches on the bat
activity correlation with local structural richness. Bats used
structurally richer dry grassland plots more for commuting and/or
foraging, when the surroundings were more open (all guilds) and
less patchy, hence more homogeneous (edge space guild). Both
parameters, openness and homogeneity of landscapes, are of special
importance, since they usually are negative consequences of
agricultural intensification (Matson et al., 1997; Stoate et al., 2001;
Tilman et al., 2001; Jongman, 2002; Burns et al., 2016). They very
likely determine resource availability for bats in the landscape due
to the following reasons: A higher amount of open farmland means
the reduction of other main non-crop habitats, necessary for
commuting and foraging. Similarly, homogeneous landscapes with
less patches usually contain fewer crop types with reduced insect
diversity (Olimpi and Philpott, 2018), and fewer habitat edges
(Ethier and Fahrig, 2011). Consequently, previous studies have
found decreasing bat activities of several bat species in farmland
with greater amounts of open agricultural land (Put et al., 2019),
A

Open space guild: activity

<figure><img src="image_35.png" /></figure>

B

Open space guild: foraging

<figure><img src="image_36.png" /></figure>

C

Edge space guild: foraging

<figure><img src="image_37.png" /></figure>

D

Narrow space guild: activity

<figure><img src="image_38.png" /></figure>

FIGURE 4

Effect plots of the interaction of the foliage height diversity (FHD, scaled) and % open land (scaled). Boxes show the predicted effects of FHD in low and high % of open land on (A) open space guild activity and (B) foraging, (C) edge space guild foraging, and (D) narrow space guild activity. Low and high values are represented by the first and third quartile of the open land data (scaled) respectively.

above improved meadows and pastures when further away from other important habitats (Heim et al., 2015; Treitler et al., 2016), and with less landscape heterogeneity (Monck-Whipp et al., 2018; Froidevaux et al., 2022). The latter especially facilitates edge bats movement (Mendes et al., 2017), which could explain why patchiness only appeared in results for the edge guild. Despite the just named known negative effects of a high amount of open farmland and heterogeneity loss, we found locally increased bat activity and foraging above dry grassland plots, when these provided higher structural richness. This is also contrary to findings of Froidevaux et al. (2022), who showed that local features become more relevant for many bat species only when the landscape is more heterogeneous. We therefore argue that the value of dry grassland fragments with diverse shrub and tree vegetation in intensive farmland increases, when the surrounding landscape becomes more unfavorable. For that, we provide the following explanations: first,

the higher richness in vegetation structures can offer more niches, aiding a high variety of bat species with different sensory requirements and habitat adaptations, as it is known from forest systems (Adams and Law, 2011; Hanspach et al., 2012; Jung et al., 2012; Müller et al., 2013; Froidevaux et al., 2016; Blakey et al., 2017; Starik et al., 2018). Secondly, dry grasslands contain a rich insect fauna (Wallis de Vries and Van Swaay, 2009; Rasran et al., 2018) as potential food sources. The presence of more and denser vertical woody vegetation structures, such as trees, can further increase insect abundance (Müller et al., 2012; Fuentes-Montemayor et al., 2013), especially in agricultural regions [such as at trees and hedgerows in Grüebler et al. (2008)]. In the surrounding landscape, certain types of farmland, such as orchards or vineyards, could also provide diverse vegetation structures. Additionally, even open cropland can offer occasional plentiful insect food sources for bats. However, the latter is usually ephemeral and depending on crop
Edge space guild: foraging

<figure><img src="image_39.png" /></figure>

FIGURE 5
Effect plots of the interaction of the foliage height diversity (FHD, scaled) and number of patches (scaled). Boxes show the predicted effects of FHD in low and high amount of number of patches for foraging activity of the edge space guild. Low and high values are represented by the first and third quartile of the number of patches (scaled) respectively.

<figure><img src="image_40.png" /></figure>

FIGURE 6
NMDS ordination plot showing the 12 dry grassland plots in k=2 dimensions. Placement of plots is based on Bray-Curtis dissimilarity of the activity compositions of functional bat guilds (indicated by stars). Environmental fitting was used to find explanatory variables for differences in guild composition. The significant variable (FHD) is represented as a vector.

growth and season (Heim et al., 2016, 2017; Stahlschmidt et al., 2017). Moreover, in conventional farmland, feeding on insects will lead to direct contamination of bats with pesticides (Stahlschmidt and Brühl, 2012; Stahlschmidt et al., 2017). Dry grasslands on the other hand, as semi-natural habitats, can provide stable food supplies for bats throughout the growing season (Davidai et al., 2015) and thus buffer negative effects of intensive farming practices.

We additionally found prominent interaction effects of distance to urban areas and local dry grassland conditions in the open space (activity and foraging) and edge space (foraging) guild, as their use of dry grassland plots increased with higher structural richness and with greater distance to urban areas. It is known that some species especially from open and edge space guilds regularly utilize urban areas for roosting and foraging (Jung and Threlfall, 2018) and that the proximity to anthropogenic areas can positively affect bat species occurrence and activity in grasslands nearby (Heim et al., 2015). Our study region contained mainly rural, low-density settlements, which have great potential as roosting or feeding sites (Gehrt and Chelsvig, 2004; Dixon, 2012). We therefore argue, that structurally richer dry grasslands increase in value when further away from such urban sites.

Our results here could only be obtained by investigating interaction terms of landscape and local features. As other studies suggested, effects of landscape conditions on bats at a broader scale are usually dominant over local conditions (Kalda et al., 2015a,b) but interaction terms can reveal scenarios where local features become relevant. Hence, the availability of resources for commuting and foraging provided by the surroundings determine the importance of local features. We thus conclude that in simpler and more open agricultural regions, structurally richer dry grasslands are favored by bats over less structured ones.

4.3. Functional guild assemblages shift with local features

In accordance to our prediction, we found that local structural conditions on dry grasslands change bat guild activity compositions. Our results show a proportional shift from open space to edge space guild activity with higher structural richness of plots. This again highlights the previously discussed positive affinity of edge space bats with structurally richer dry grasslands, which makes them the dominating guild, when diverse shrubs and trees are available. This corroborates the results of Davidson-Watts et al. (2006), who found that (scrubby) grasslanda are among favored habitat types for *Pipistrellus pipistrellus* and *P. pygmaeus*. We did not clearly see dominance changes along the structural richness gradient in the narrow space guild, which seems counterintuitive given their affinity to vegetation clutter. The activity of the narrow space guild however was generally low, which might have been caused by the overall open character of the landscape, and proportional changes were likely to be minor, compared to the other dominant guilds.

Heim et al. (2015) found shifts in species composition above pastures and meadows in correlation with cover of and distances to landscape features in a 200 m radius. In contrast, we did not find any explanation for shifts of bat assemblages by any landscape features. Here the limited sample size of dry grassland plots embedded in not extremely varying landscapes (given that most dry grassland fragments are located in farmland), might have obscured possible evidences.

4.4. Conclusion and further applications

To our knowledge, we show for the first time the importance of European dry grassland fragments for bats under certain local and landscape compositional conditions in an agriculturally rich region in Germany. While we confirm a constant use of dry grasslands by a high number of bat species/sonotypes throughout the season, we also
show that effects of local and surrounding landscape features on bat activity and foraging vary between functional bat guilds. Nonetheless, our study emphasizes for all bat guilds the significant value of dry grasslands with increased local structural richness in unfavorable, open and homogeneous farmland. We thus argue that dry grassland fragments especially with heterogeneous shrub and tree vegetation have a high potential as stepping stones for bats in intensive agricultural regions. By that means, we underline the necessity of heterogeneity on different spatial scales for biodiversity in such landscapes (Benton et al., 2003). Our findings are of high ecological importance and conservation concern, as many bat species are facing increased challenges in agriculturally intensified regions through habitat and connectivity loss (Browning et al., 2021), but are providing important services as insect pest control (Cleveland et al., 2006; Kalka and Kalko, 2006; Kalka et al., 2008; Williams-Guillén et al., 2008). With our results, we add evidence to the high biodiversity value of European dry grassland fragments in agricultural regions, which was previously mainly shown for plants (Lindborg et al., 2014; Deák et al., 2020) and insects (Tscharntke et al., 2002; Duelli and Obrist, 2003; Rösch et al., 2015). Lastly, our findings contribute to informed decision making for dry grassland management and bat conservation.

For further applications, we highlight the importance of careful considerations before removal of heterogeneous shrub and tree vegetation on dry grassland fragments. This is contradicting the current management goals of dry grasslands, aiming to reduce shrub and tree encroachment. However, in this study we showed positive correlations of bat activity with increased structural richness, while the maximum structural richness corresponded with a rather low foliage cover of 25% (summed up cover of shrub and tree foliage), which is still equivalent to a mediocre management state for the dry grassland types found in Brandenburg, Germany (Landesamt für Umwelt Brandenburg, 2022). We thus propose considering leaving heterogeneous shrubs and trees on dry grassland fragments particularly in open, homogenous and intensively used agricultural regions, while this appears less important in heterogeneous landscapes. In line with previous studies, we highlight the need of connectivity networks (such as hedgerows) in landscapes to allow accessibility of edge and narrow space bat species (Frey-Ehrenbold et al., 2013; Heim et al., 2015). More importantly, we strongly advise to co-evaluate interactions of local and landscape features, as this leads to a better understanding of the landscape settings that determine the importance of dry grassland conditions for bats. Finally, we encourage more research on the relevance of these habitat types for single bat species, in different regions and subsequently also for other non-target organisms.

## Data availability statement

The raw data supporting the conclusions of this article will be made available by the authors, without undue reservation.

## Ethics statement

Ethical review and approval was not required for the animal study because acoustic monitoring of bats is non-invasive and does not involve any animal contact. However, fieldwork and methodological approaches were approved by the State Office for the Environment and access was

permitted by the nature conservation authorities of Barnim and Uckermark (districts in Brandenburg, Germany).

## Author contributions

SE conceptualized and designed the study. Data collection in the field was performed by SE and K-HF. SE did the data processing and statistical analysis. The manuscript was written by SE. MK, KJ, and K-HF contributed to all previous steps through guidance and supervision. All authors further contributed to manuscript revision and read and approved the submitted version.

## Funding

This research was funded as a part of the project INPEDIV (SAW-2019-ZFMK-3-INPEDIV). Open Access Funding was provided by the Freie Universität Berlin.

## Acknowledgments

We thank Livia Schäffler (Zoological Research Museum Alexander Koenig, Bonn, Germany) for her contribution as project lead and coordination in INPEDIV and in acquisition of fieldwork permissions for dry grassland sites. We greatly appreciate the assistance and efforts of Artur and Anna Ewert in fieldwork. We thank the various landowners and managers as well as cattle herders and farmers for their cooperation as well as their great interests and openness toward our work. We would like to thank participants of INPEDIV for their exchange and specifically Carsten Brühl, Bodo Möseler, and Benito Schoepke for their support and input. We greatly appreciate Neil Middleton for his exchange about bat acoustics. Lastly, we acknowledge support by the Open Access Publication Initiative of Freie Universität Berlin.

## Conflict of interest

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

## Publisher's note

All claims expressed in this article are solely those of the authors and do not necessarily represent those of their affiliated organizations, or those of the publisher, the editors and the reviewers. Any product that may be evaluated in this article, or claim that may be made by its manufacturer, is not guaranteed or endorsed by the publisher.

## Supplementary material

The Supplementary material for this article can be found online at:
https://www.frontiersin.org/articles/10.3389/fevo.2023.995133/
full#supplementary-material
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Frontiers Production Office (2023) Erratum: Structurally rich dry grasslands – Potential stepping stones for bats in open farmland.
*Front. Ecol. Evol.* **11**:1168225.
doi: 10.3389/fevo.2023.1168225

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# Erratum: Structurally rich dry grasslands – Potential stepping stones for bats in open farmland

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KEYWORDS

dry grasslands, bats (Chiroptera), agricultural intensification, landscape, structural richness, acoustic monitoring, habitat fragments

An Erratum on

Structurally rich dry grasslands – Potential stepping stones for bats in
open farmland

by Ewert, S. P., Knörnschild, M., Jung, K., and Frommolt, K.-H. (2023). *Front. Ecol. Evol.* **11**:995133.
doi: 10.3389/fevo.2023.995133

Due to a production error, there was a mistake in **Table 1** as published. There were several formatting issues. The corrected **Table 1** appears below.

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References

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Kalda, R., Kalda, O., Lõhmus, K., and Liira, J. (2015b). Multi-scale ecology of woodland bat the role of species pool, landscape complexity and stand structure. *Biodivers. Conserv.* 24, 337–353. doi: 10.1007/s10531-014-0811-6

Kelm, D. H., Lenski, J., Kelm, V., Toelch, U., and Dziock, F. (2014). Seasonal bat activity in relation to distance to hedgerows in an agricultural landscape in Central Europe and implications for wind energy development. *Acta Chiropterologica* 16, 65–73. doi: 10.3161/150811014x683273

Mendes, E. S., Fonseca, C., Marques, S. F., Maia, D., and Pereira, M. J. R. (2017). Bat richness and activity in heterogeneous landscapes: guild-specific and scale-dependent? *Landsc. Ecol.* 32, 295–311. doi: 10.1007/s10980-016-0444-0

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<table><caption>TABLE 1 Final choice of local and landscape features and their abbreviations used for further analysis.</caption><thead><tr><th>Local & Landscape features<br>Measurements</th><th>Abbreviation</th><th>Relevance for bats found in:</th></tr></thead><tbody><tr><td colspan="3" style="text-align: center; font-weight: bold;">Dry grassland (r = 25 m)</td></tr><tr><td>Foliage height diversity (structural richness)</td><td>FHD</td><td>Froidevaux et al. (2016)</td></tr><tr><td colspan="3" style="text-align: center; font-weight: bold;">Landscape (r = 200 m)</td></tr><tr><td>Forest:<br>Percentage cover</td><td>For</td><td>Fuentes-Montemayor et al. (2013), Heim et al. (2015, 2018), Kalda et al. (2015b), Treitler et al. (2016)</td></tr><tr><td>Open land (pastures + cropland):<br>Percentage cover</td><td>Open</td><td>Put et al. (2019)</td></tr><tr><td>Patches:<br>Counted number</td><td>Patch</td><td>Mendes et al. (2017)</td></tr><tr><td>Urban area:<br>*Distance</td><td>DistUrb</td><td>Fuentes-Montemayor et al. (2013), Heim et al. (2015)</td></tr><tr><td>Water bodies:<br>*Distance</td><td>DistWat</td><td>Downs and Racey (2006), Stahlschmidt et al. (2012), Fuentes-Montemayor et al. (2013), Heim et al. (2015, 2018), Mendes et al. (2017)</td></tr><tr><td>Woody features:<br>Percentage cover<br>*Distance</td><td>Wood<br>DistWood</td><td>Verboom and Huitema (1997), Downs and Racey (2006), Boughey et al. (2011b), Frey-Ehrenbold et al. (2013), Kelm et al. (2014), Heim et al. (2015), Kalda et al. (2015a), Toffoli (2016)</td></tr></tbody></table>

Right column shows examples of literature where the landscape variable has shown a relevance for bats. *Distances were also measured when reaching outside of the 200 m radius.
<figure><img src="image_43.png" /></figure>

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EDITED BY

Grégory Mahy,
University of Liège,
Belgium

REVIEWED BY

Leonie K. Fischer,
University of Stuttgart,
Germany
Réka Kiss,
Hungarian Academy of Sciences,
Hungary

*CORRESPONDENCE

Mariana P. Fernandes

<figure><img src="image_44.png" /></figure>

mpfernandes@uevora.pt

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Frontiers in Ecology and Evolution

RECEIVED 30 November 2022

ACCEPTED 31 January 2023

PUBLISHED 02 March 2023

CITATION

Fernandes MP, Matono P, Almeida E, Pinto-Cruz C and Belo ADF (2023) Sowing wildflower meadows in Mediterranean peri-urban green areas to promote grassland diversity.

Front. Ecol. Evol. 11:1112596.
doi: 10.3389/fevo.2023.1112596

COPYRIGHT

© 2023 Fernandes, Matono, Almeida, Pinto-Cruz and Belo. This is an open-access article distributed under the terms of the [Creative Commons Attribution License (CC BY)](https://creativecommons.org/licenses/by/4.0/). The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Sowing wildflower meadows in Mediterranean peri-urban green areas to promote grassland diversity

Mariana P. Fernandes¹,²*, Paula Matono¹, Erika Almeida¹, Carla Pinto-Cruz³ and Anabela D. F. Belo³

¹MED – Mediterranean Institute for Agriculture, Environment and Development & CHANGE – Global Change and Sustainability Institute, Instituto de Investigação e Formação Avançada, Universidade de Évora, Pólo da Mitra, Évora, Portugal, ²cE3c – Center for Ecology, Evolution and Environmental Changes & CHANGE – Global Change and Sustainability Institute, Faculdade de Ciências, Universidade de Lisboa, Campo Grande, Portugal, ³MED – Mediterranean Institute for Agriculture, Environment and Development & CHANGE – Global Change and Sustainability Institute, Departamento de Biologia, Universidade de Évora, Pólo da Mitra, Évora, Portugal

**Introduction:** The increase of urban areas and their infrastructure network is homogenizing the landscape and threatening biodiversity and ecosystems functions and services. Wildflower meadows have a high biodiversity value and can prosper in degraded areas dominated by nitrophilous species, making them suitable to be used in peri-urban and urban areas to promote local flora, create habitat for pollinators and other small fauna, and increase overall biodiversity. Moreover, the application of wildflowers seed mixes suitable for rehabilitating anthropized environments should be restricted to native species of regional origin, and the results properly monitored. However, thorough monitoring of seed mixes evolution is uncommon. This study evaluates the effectiveness of a seed mix of wild native species developed to promote grassland diversity in Mediterranean peri-urban areas.

**Methods:** The study was divided into two sequential phases. Firstly, a preparatory phase consisted in developing two seed mixes and sowing them (autumn 2016) in *ex-situ* plots (three plots of 5x2m² per mix) at an experimental field to choose the one with the best performance. The second phase consisted of the *in-situ* application (autumn 2018) of the chosen seed mix by sowing 14 plots (10x2m²) in pocket parks distributed along pedestrian trails of South Portugal. All plots were monitored through floristic surveys for two springs (*ex-situ* trials: 2017 and 2018; *in-situ* trials: 2019 and 2020).

**Results:** All sowed species germinated in the *in-situ* plots over the first 2 years. The seed mix application positively contributed to the floristic community, generating a significant increase in the total species richness, diversity, evenness, and vegetation cover. The seed mix establishment did not require watering nor soil fertilizing and the mowing frequency was low (once in late spring), contributing to sustainable and low-cost management of these green areas.

**Discussion:** The tested seed mix promoted native flora diversity rapidly and seems suitable for use in peri-urban context under identical climate conditions. Given the small number of native seed mixes tested in the Mediterranean, this study represents a contribution toward improved management standards of native flora diversity in Mediterranean green urban and peri-urban areas.

KEYWORDS

native flora conservation, native seed mixes, pedestrian path management, rehabilitation, semi-natural pastures
Introduction

Human activities have consistently affected natural and semi-
natural habitats. Land use and land cover changes, particularly land
taken by agriculture and urbanization, are transforming the terrestrial
environment at unprecedented rates and scales, homogenizing the
landscape and threatening biodiversity and ecosystems functions and
services (Stoate et al., 2009; Seto et al., 2011). With the expansion of
the urban network and the decline of natural ecosystems, urban areas
management has a crucial role in preserving and enhancing
biodiversity and presents unparalleled opportunities to enhance the
urban systems' resilience and ecological functioning (Farinha-
Marques et al., 2011; Elmqvist et al., 2015). Conservation of urban
grasslands can reduce cities' ecological footprints and ecological debts
while enhancing their inhabitants' resilience, health, and quality of life
(Gómez-Baggethun and Barton, 2013). One way of diminishing the
negative impact of cities is to use their green spaces to conserve
biodiversity and improve these places' ecosystem value and function.

Natural and semi-natural grasslands are among the world's most
threatened biomes (Janssen et al., 2016), making the conservation of
their floristic assemblages, individual species, and ecological functions
essential (Wesche et al., 2012). Urban areas usually have many patches
of unused land (e.g., road and trail verges, roundabouts) and amenity
areas (e.g., public parks, pocket parks, and lawns) frequently managed
by regular mowing, whose potential for promoting biodiversity is
often ignored (Blackmore and Goulson, 2014). These areas can
be considered urban grasslands (Klaus, 2013), which are herbaceous,
perennial landscapes of primarily turf-forming grass species adapted
to semi-regular mowing and kept for aesthetic or recreational value
(Thompson and Kao-Kniffin, 2017, 2019). Urban grasslands are
considered novel habitats with a species assemblage partially altered
due to an intense modification of biotic and abiotic conditions caused
by the high level of human disturbance (Kowarik, 2011). Nevertheless,
rehabilitated urban grasslands can maintain characteristic species of
semi-natural grasslands and even rare species (Fischer et al., 2013a).

The urban green areas, including trail verges and pocket parks, are usually mown and, thus, kept in an early successional stage (Norton et al., 2019), promoting a vegetation composition and structure similar to natural and semi-natural grasslands. Due to their linear format, the trail verges can represent vegetation corridors between larger areas, such as parks or community gardens, promoting ecological connectivity. These linear infrastructures are, therefore, useful for grassland rehabilitation, i.e., for applying management actions that aim to reinstate a certain level of ecosystem functioning without the necessity to match up the biodiversity or integrity of the reference habitat with accuracy (Gann et al., 2019). Alter the management of these green areas to combine nature conservation with socio-economic aspects and landscape planning could be a win-win strategy for biodiversity conservation and improvement of socio-economic conditions (Gómez-Baggethun and Barton, 2013; Blackmore and Goulson, 2014; Bretzel et al., 2016).

Native wildflowers seed mixes sown in pocket parks and trail verges can enhance biodiversity and create an ecological continuity between urban and rural landscapes with low management costs (Blackmore and Goulson, 2014; Bretzel et al., 2016). Species-rich herbaceous communities, like grasslands, wildflower meadows and pastures, have a high biodiversity value and are adapted to soils with structural degradation and low organic matter content (Bretzel et al.,

2016). Additionally, these floristic communities benefit from a regular mowing regime, which is the standard practice in urban green areas (Kelemen et al., 2014; Tälle et al., 2018). Thus, these herbaceous communities can constitute models to be adapted for landscape management of urban areas to increase biodiversity, create habitat for pollinators or small fauna, and conserve the local flora (Aldrich, 2002; Norton et al., 2019).

The effectiveness of wildflower meadows as a biodiversity promotion measure is related to the composition of the seed mix and its adaptation to local conditions, since this affects the number of species, flower abundance, plant species diversity, and vegetation structure in the sown areas (Haaland and Gyllin, 2011). Promoting a wide range of species in the urban grasslands increases life forms and survival strategies, favoring biodiversity promotion, and improving the establishment capacities of the floristic community (Bretzel et al., 2012). Although the species of semi-natural grasslands can adapt to urban environments, their choice should be judicious to guarantee an effective establishment: it is recommended to select native species, preferably of regional provenance, since they are well adapted to specific edaphoclimatic conditions, have fewer invasion risks, and help to conserve genetic diversity (Hufford and Mazer, 2003; Gann et al., 2019). The species selection should also consider the species' ability to survive in semi-arid and/or disturbed environments, the existence of an extended flowering period assured by the set of the species, the use of species with different life types (Bretzel et al., 2012), their germination behavior, particularly when stored seeds are used (Bhatt et al., 2019), and the species establishment features (Scheper et al., 2021). Using seed mixes containing annual, biennial, and perennial species is a holistic approach that makes it possible to combine the advantages of different life strategies. Annual species enhance the aesthetic value and may persist by self-seeding (Dunnett, 2011; Hoyle, 2016). In contrast, biennial and perennial species persist and flower over multiple years, reducing the need for re-sowing (Hoyle, 2016). The Mediterranean grassland communities naturally include patches of biennial and perennial species (Chust et al., 2006), and promoting these species in urban grasslands, along with annual species, increases the cover of a functional group already present in native ecosystems while maintaining the local biogeographic species pool (Filibeck et al., 2016). Furthermore, using species with different life strategies extends the flowering period: annual species flower in early spring and biennial and perennial species flower in late spring or summer (Bretzel et al., 2009, 2012). Promoting species with different flowering periods is beneficial because it increases the period of attractiveness to people and of resources available for pollinators (Hoyle et al., 2018).

In novel habitats, species assemblages can change the competitive
dynamics between species (sown or spontaneous) compared to the
habitats they usually inhabit (Hobbs et al., 2006). Consequently, the
new floristic community will not have the same composition as in a
non-anthropogenic environment (Fischer et al., 2013a). Therefore,
species selection must consider the potential competition between all
species (sown or spontaneous) One key question for the successful
rehabilitation of urban grassland is how species cope with competitive
pressure in an environment dominated by ruderal and exotic species
(Fischer et al., 2013b), that are often competitive and tolerant to high
stress and disturbance (Craine, 2005). The human disturbances in
urban grasslands tend to favor the colonizing species with traits that
increase a plant's ability to cope with stress and ruderal conditions
(Fischer et al., 2013b). Usually, the competitive strategy of perennial
species is less favored and increases the proportion of annual species (Fischer et al., 2013b).

Besides the ecological aspect, promoting semi-natural grasslands and wildflowers in urban areas also has economic, health promotion, and aesthetic value. To succeed, wildflowers should be sustainable, not require supplementary watering nor soil fertilizing, and have a low mowing frequency, and are, therefore, a low maintenance cost strategy (Bretzel et al., 2009; Köppler et al., 2014). Wildflowers can also provide ecosystem services like climate regulation, pollination, or soil and air quality improvement, being a link between urban environments and rural areas (Aldrich, 2002; Bretzel et al., 2016). This access to a wild natural environment positively impacts human health and quality of life and increments urban green spaces' aesthetic interest (Lindemann-Matthies and Brieger, 2016; Norton et al., 2019). The long-term rehabilitation of urban grasslands requires residents' acceptance and local authorities' willingness to maintain them (Norton et al., 2019). Recent studies (e.g., Southon et al., 2018; Norton et al., 2019) suggested that replacing some of the amenity grasslands with a range of different types of meadow vegetation is well received by the general population and can generate positive outcomes for both biodiversity and people. For instance, using seed mixes with an extended flowering period meets the aesthetic expectations of the population (Hoyle et al., 2017a; Colombo et al., 2021) and promotes pollinators (Hoyle et al., 2018; Angelella et al., 2019), favoring flora and fauna diversity and benefiting the human population. However, some aspects concerning the actual use of wildflowers in urban settings remain – the use of tall herbaceous and the vegetation's wild appearance close to residential areas is not always easily accepted (Hoyle et al., 2017b; Colombo et al., 2021). In the Mediterranean, the citizens' displeasure is even more significant after the flowering season when the vegetation dries up (Filibeck et al., 2016; Colombo et al., 2021). However, mowing the vegetation in late spring removes the dry biomass, minimizing the negative effect on the aesthetic appearance and reducing fire risk during the dry season. Informing the citizens about the benefits to biodiversity of rehabilitated urban grasslands and a wise selection of areas seems to be decisive for the population acceptance of these areas (Filibeck et al., 2016; Southon et al., 2018; Norton et al., 2019).

Using wildflowers or herbaceous seed mixes in urban environments is becoming widespread (Norton et al., 2019; Vega and Küffer, 2021); however, many seed mixes contain exotic species, use species not adapted to the specific edaphoclimatic conditions or do not consider the species' establishment and development capacity (Bretzel et al., 2016). In the Mediterranean region, native seeds' availability and applicability for urban environment rehabilitation is still limited (Bochet et al., 2010; Bretzel et al., 2012; de Vitis et al., 2017). Furthermore, this type of rehabilitation often reveals limited effectiveness, particularly in the Mediterranean region (Nunes et al., 2016), what highlights a still low predictive ability concerning rehabilitation trajectories and outcomes (Suding et al., 2015).

Applied research is needed, so that ecological rehabilitation of urban and peri-urban green areas, such as grasslands, can fulfil its potential and meet its goals as a conservation tool. We developed a seed mix of wild native species to promote floristic diversity in peri-urban areas and improve conditions for promoting small fauna (e.g., butterflies and small mammals). The seed mix was applied in pocket parks distributed throughout pedestrian trails of Southern Portugal. The study aims to (a) promote plant diversity, including endemic

species, and favor habitat for small fauna, (b) evaluate the effect of a seed mix developed for Mediterranean areas on floristic diversity, evenness, and vegetation cover, and (c) evaluate the potential of peri-urban green areas as a tool for biodiversity conservation in the South of Portugal.

# Materials and methods

## Study area, species selection and seed collection

The study was carried out in Évora and Montemor-o-Novo (Central Alentejo - Southern Portugal) between 2016 and 2020 in a peri-urban context. The study area presents the typical land use pattern of the South of Portugal, with urban centers surrounded by cork oak forests (montado) and, nowadays, also by areas of intensive agriculture (e.g., vineyards and olive groves). Évora and Montemor-o-Novo are small cities with an urban area of 1,307 km² and 1,233 km² and a resident population of 53,577 and 15,799 inhabitants, respectively (FFMS, 2021). The area has a typical winter-wet and summer-dry pattern of the Mediterranean-type climate with a mean annual rainfall of 609 mm; the mean annual temperature is 15.9°C, and the warmest month is August (23.3°C) while the coldest is January (9.3°C; IPMA, 2019).

The study was divided into two subsequent phases: ex-situ (autumn 2016 to spring 2018) and in-situ trials (autumn 2018 to spring 2020; Figure 1). Firstly, ex-situ trials of two seed mixes were accomplished in an experimental field at Herdade da Mitra (University of Évora). Afterwards, in-situ application of the more appropriate seed mix was made in pocket parks distributed throughout pedestrian trails of Évora and Montemor-o-Novo.

Forty-one native species were selected to be used in the seed mixes, considering:

1. Their suitability to pedestrian trails and pocket parks, namely by being

* resistant to drought and intense solar radiation;

* annual plants that flower and set seed in early spring, ensuring their permanence in pedestrian trails and pocket parks (that are usually mown in late spring);

* non-annual plants with subterranean organs that allow re-sprouting each year after being cut;

* plants with no known toxicity.

2. Their conservation interest, namely by using endemic species.

3. The existence of well-established donor populations to collect seeds in the study area.

The species ecological characteristics were compiled from the literature (e.g., Castroviejo, 1986-2012; Flora-On, 2014; Flora Vascular, 2016) and benefited also from the authors' expert knowledge. The nomenclature of the species followed Flora Iberica (Castroviejo, 1986-2012). Seeds from the species selected were collected from plants growing in montado areas or in the intervened linear infrastructures surroundings during the springs and summers of
<figure><img src="image_45.png" /><figcaption>FIGURE 1</figcaption></figure>

*Ex-situ* and *in-situ* plots installed on the experimental field of the University of Évora and pocket parks distributed throughout pedestrian trails of Évora (six plots) and Montemor-o-Novo (eight plots), respectively.

2016–2018, and their geographical coordinates were recorded with GPS technology (Garmin Montana 680 t). In all collecting campaigns, care was taken to never collect more than 20% of the seeds available at each site to avoid significant damage to the donor populations and assure natural regeneration (ENSCONET, 2009). In the seed bank of the University of Évora, the seeds were then cleaned, dehumidified at 15% RH (FITOCLIMA S600PDH – Aralab 1681; Portugal), stored in glass containers, and maintained in the dark at room temperature until being used in the *ex-situ* or the *in-situ* trials (autumn 2016 and 2018, respectively).

Viability trials

The selected species were submitted to viability trials in a germination chamber before each study phase to ensure that their viability was higher than 50% (Bellairs and Caswell, 2016), and only the seed lots - i.e., seeds belonging to the same batch of seeds, with a single reference number, origin, and history - meeting this criterion were considered suitable to be used in the trials. When using the same seed lot of a given species was not possible throughout the experiment (*ex-situ* and *in-situ* trials), viability trials were performed separately for each lot.
Seeds were germinated, without pre-germination treatments, in plastic Petri dishes (9 cm diameter) with agar (1%; VWR Chemicals) in a germination chamber (FITOCLIMA S600PLH - Aralab 1680; Portugal). For each species, four 25-seeds replicates were incubated with a photoperiod of 8 h of light and 16 h of darkness with an alternating temperature of 15/10°C. These incubation conditions simulate autumn conditions, which is the optimum germination season for most Mediterranean species, according to Copete et al. (2009) and Fernandes et al. (2021). Petri dishes were re-randomized daily to ensure that there were no systematic effects due to position within the germination chamber (Cerabolini et al., 2004).

The counting of germinated seeds was carried out daily over 30 days, and germinated seeds were removed from the Petri dishes. Though germination is considered to occur when the embryonic tissue emerges (Bewley et al., 2013), a more conservative approach was used by considering germination only when a radicle with at least half of the seed size emerged. Non-germinated seeds at the end of the trial were cut, and those with a well-developed, white, and firm embryo were considered viable (Gosling, 2003).

## Ex-situ trials

The *ex-situ* trials occurred between November 2016 and June 2018 in an experimental field at Herdade da Mitra. Beside the experimental field there is a weather station managed by the University of Évora,<fnref n="1" /> what allowed accurate recording of meteorological conditions. During the *ex-situ* trial, the maximum monthly rainfall was in March 2018 (207 mm). Rainfall was higher in the autumn/winter of 2017/2018 than in the remaining study periods. The mean monthly temperature ranged between 26.3 ± 4.3°C in August 2018 and 8.4 ± 2.1°C in January 2017 (ICT, 2021).

The 41 selected species were combined in two seed mixes with 23 species each (annex 1). Both seed mixes met the following criteria:

1. have a proportion of 30% Fabaceae, 30% Poaceae, 10% Asteraceae, and 30% other families. The chosen proportions intended to maximize floristic diversity by creating a mix with more than 15 species as recommended by Lepš et al. (2007) and Carter and Blair (2012), and to promote the existence of different life strategies (annual, biennial, and perennial species) as recommend by Bretzel et al. (2012) and Filibeck et al. (2016). Regular mowing of vegetation promotes grass dominance (Simões et al., 2014). As such, the balance between Poaceae, Fabaceae, and Asteraceae aimed to minimize grass dominance pace.

2. have at least one endemic species.

3. have at least one species for small fauna promotion, i.e., ensure the inclusion of at least one flora species that provides food or shelter for small animals, namely butterflies and small mammals.

4. provide a sequential and extended flowering period, offering aesthetic recreation for citizens and habitat and food for pollinators and small mammals.

Each seed mix was tested *ex-situ* in three plots (5 × 2 m² each) installed in November 2016 and one control plot (not sown) was also established. The plots had a slight slope (7°) and were exposed to full sun. The plots were previously harrowed and sowed with a density of 2,000 seeds/m². Seed density followed Hitchmough et al. (2008), according to which sowing with high densities (>1,200 seeds/m²) positively contributes to promoting diverse floristic communities in grassland rehabilitation actions. The seeds did not receive any pre-germination treatment. The soil was then lightly raked to provide better contact between the seeds and the soil particles. No watering, fertilizing or pesticides were applied.

Floristic inventories in 50 × 50 cm² squares (three quadrats/plot evenly distributed throughout the center of the plot) were used to assess the total and species cover percentage (of both sown and spontaneous species) and the number of individuals of the sown species. Monitoring was done in 2017 and 2018 springs, to evaluate the mix performance over time. The plots were mowed at the end of the flowering season and after monitoring, to simulate the management regime usually carried out in marginal areas of pedestrian trails (June 2017 and 2018).

## In-situ trials

The *in-situ* trial occurred between October 2018 and June 2020 in pocket parks distributed throughout pedestrian trails of Évora and Montemor-o-Novo. The pedestrian trails expand from the urban area to each city's surroundings. The pocket parks used in the study are in the marginal areas of the pedestrian trails and occur throughout the entire trail. The temperature and precipitation conditions of Évora and Montemor-o-Novo plots were considered separately, and the records were obtained from the nearest meteorological station. During the *in-situ* trial (Figure 2), rainfall was higher between November and December 2019 and the maximum monthly temperature was in August 2019 (ICT, 2021).

The selected seed mix underwent minor composition or proportion adjustments to improve the less positive aspects identified during *ex-situ* trials. In autumn 2018, the final seed mix (annex 2) was sown in 14 *in-situ* plots (10 × 2 m²) in pocket parks distributed throughout pedestrian trails of Évora (six plots) and Montemor-o-Novo (eight plots). The sowing process was identical to that carried out in the *ex-situ* plots, with a density of 2,000 seeds/m² and the seeds used did not receive any pre-germination treatment. Twelve control plots (not sown) were also established in areas close to, and identical to, the sowing plots. All plots were in flat sites and exposed to full sun. No watering, fertilizing, or pesticides were applied. All the plots were mowed in late spring (June 2019 and 2020) as part of the current management regime of the pedestrian trails and respective pocket parks.

Citizens were informed of the ongoing conservation actions through signage placed at the beginning of pedestrian trails and in some of the intervened pocket parks. All the sown plots were marked with stakes.

The plots were monitored and sampled during the springs of 2017–2018 (before intervention: control and pre-sown plots) and 2019–2020 (after intervention: control and sown plots) through floristic inventories in 50 × 50 cm² squares (three quadrats/plot evenly distributed throughout the center of the plot). The following

<footnote n="1">1 https://www.icterra.pt/g1/index.php/meteo-data/</footnote>
<figure><img src="image_46.png" /><figcaption>FIGURE 2
Total monthly rainfall (R) and mean monthly temperature (T) recorded accordingly the location of the *in-situ* plots (Évora or Montemor-o-Novo).</figcaption></figure>

parameters were measured: total and species cover percentage (sown and spontaneous species), and the number of individuals of the sown species. The presence/absence of reproductive elements (flowers and/or fruits) of the sown species was also registered.

## Data analysis

At the end of the viability trials and to ensure the seeds' quality, the viability percentage (VI) was calculated for each species lot considering the formula $VI) = (V \times 100)/NT$, where V is the number of viable seeds (germinated seeds plus viable seeds that did not germinate) and NT is the total number of seeds.

In the *ex-situ* trials, the floristic community of the plots sown with the seed mixes was evaluated based on taxonomic terms and on the following parameters: sown species present in the floristic community (%), sown species that germinated (% and No.), global germination (%) of the seed mix (i.e., the total rate of sown seeds that germinated, regardless of species), cover of sown species (%), cover of spontaneous species (%), diversity of sown species (Shannon-Wiener index), and cover of sown species (evenness). The global germination of the mix was only evaluated in the first year after sowing (spring 2017).

The selection of the most appropriate mix of species to be applied in the *in-situ* plots was based on the best results obtained for these parameters. To support this process, the seed mixes under analysis, and respective control plot, were compared for each of these parameters using the Mann-Whitney non-parametric test.

For the purposes of comparing the results obtained in the *ex-situ* and *in-situ* trials, the initial floristic community of sites (control and pre-sown plots) was analyzed in terms of families, using the Wilcoxon non-parametric test. With the same purpose, this test was also used to compare the average annual values of air temperature (°C) and rainfall (mm) between the locations and years of the respective trials.

In the *in-situ* trials, both the seed mix's performance and the floristic community's evolution were evaluated. The global seed mix and species germination percentages were calculated only for the first spring after sowing (2019) and took into consideration a viability factor (number of non-viable seeds), ensuring a valid comparison between species and a correct results interpretation. The viability factor was applied for each species in a sown plot, considering the proportion of each species lot used in the experimental plot. The following formulas were applied:

1. Non-viable seeds (NS) = $(NT-N) \times (1-VI)$, where NT is the total number of seeds, N is the number of germinated seeds, and VI is the viability number obtained at the viability trial;

2. Germination percentage (%) = $(N \times 100)/(NT-NS)$, where N is the number of individuals (germinated seeds), NT is the number of total seeds, and NS is the number of non-viable seeds.

The germination percentage of the seed mix species was tested with the Mann-Whitney non-parametric test.

The floristic community of the plots was analyzed on taxonomic terms, as well as for sown richness (mean number of species), diversity (Shannon-Wiener Index), evenness (Pielou Index) and vegetation cover (%), considering sown species, spontaneous species (other native and ruderal) and the species total. Mann-Whitney non-parametric test was used to compare these parameters between years and between control and sown plots. To evaluate the efficiency of the seed mix, for each of the *in-situ* plots, the percentage increment of each of these metrics was calculated, according to the formula: $Increase(\%) = (final-initial)/|initial| \times 100$, where final corresponds to the value obtained in the plot after sowing and initial corresponds to the value obtained in the control plot/pre-sowing plot. The increment of each metric between control and sown plots was tested using the Mann-Whitney non-parametric test.

Data analysis was performed with the software Statistica 10, and the significance level was set at 0.05.
# Results

The seeds' viability was high, with a mean percentage of 93±9%. The viability percentages ranged between 57 and 100%. Sixty-eight percent of the lots registered a viability percentage above 95%, and only 7% registered a viability percentage lower than 65%. These results ensured the necessary seeds' quality for rehabilitation actions, such as the sowings carried out in this study.

The initial floristic community registered in the control and pre-sown plots of the *ex-situ* and *in-situ* trials did not show significant differences in the families' composition, allowing the comparability of the results, as well as the joint evaluation of post-sown data for the studied pedestrian trails. Also, the average air temperature and rainfall conditions did not reveal significant differences between the sites of *ex-situ* and *in-situ* trials, ensuring the comparability of the results without the need to consider the possible influence of these factors in the analysis of the results.

## Ex-situ trials

In the *ex-situ* trials, both mixes registered significant differences and higher values in the parameters analyzed between control and sown plots (*p* < 0.001), therefore suggesting the effectiveness of the mixes, and supporting the restriction of the analysis to the direct comparison between the mixes under test to select the most suitable for the defined objectives. Mix 1 tended to have consistently higher values in most of the parameters analyzed (Table 1). However, despite this trend, for most parameters there were no significant differences between the two mixes in 2017 and 2018.

The floristic community of the plots sown with mix 1 registered a significantly higher percentage of the sown species and lower spontaneous species cover. In fact, although the percentage of sown species that germinated in 2017 was identical between the two mixes,

in 2018 there was a significant decrease in the germination of the species in mix 2 compared to mix 1. Nevertheless, between years, the cover of sown species significantly increased for both mixes (*p* < 0.05).

The diversity and evenness of the germinated species were identical between the two seed mixes in 2017, and both registered a decreasing tendency regarding these parameters between 2017 and 2018.

## In-situ trials

All the species of the final seed mix germinated in the *in-situ* plots in the first year, with global mean germination of 11% and maximum germination of 43% (*Briza maxima* L.). The germination percentage registered a high variability between species (*p* < 0.001). Seven species obtained germination values above 15% and, among them, *Briza maxima*, *Cynosurus echinatus* L., and *Silene gallica* L. obtained germination values above 30% (Figure 3).

There were no significant differences in the number of species and percentage of vegetation cover between the control and the pre-sown plots in the first monitoring (springs 2017 and 2018). The mean number of species increased in 2019 and 2020 (*p* < 0.001), following the germination of sown species, though other native and ruderal species also increased after the intervention (Figure 4). Therefore, both spontaneous and sown species increased the total of native flora species in both years (*p* < 0.001).

The introduction of the seed mix increased the vegetation cover of the sown plots (*p* < 0.001) and enhanced the existence of herbaceous vegetation layers, especially in 2020 (Figure 5). All the species of the seed mix persisted between 2019 and 2020, with a seed mix cover of 33% in both years. In 2019, the cover of spontaneous native species diminished in both control and sown plots. However, in 2020, the cover of ruderal species remained the same in the sown plots, and the cover of other spontaneous native species increased.

<table><caption>TABLE 1 Comparative analysis of the results obtained in 2017 and 2018 for different parameters of the floristic community (mean±standard error) in the two mixes sowed in the *ex-situ* plots at Herdade da Mitra.</caption><thead><tr><th rowspan="2">Parameter</th><th colspan="3">2017</th><th colspan="3">2018</th></tr><tr><th>Mix 1</th><th>Mix 2</th><th>SD</th><th>Mix 1</th><th>Mix 2</th><th>SD</th></tr></thead><tbody><tr><td>Sown species in the floristic community (%)</td><td>56.8 ± 3.2</td><td>43.5 ± 3.1</td><td>**</td><td>48.7 ± 3.0</td><td>38.6 ± 2.4</td><td>*</td></tr><tr><td>Sown species that germinated (%)</td><td>43.0 ± 1.5</td><td>38.2 ± 3.1</td><td></td><td>43.5 ± 4.2</td><td>29.5 ± 3.1</td><td>*</td></tr><tr><td>Sown species that germinated (no.)</td><td>9.9 ± 0.4</td><td>8.8 ± 0.7</td><td></td><td>10.0 ± 1.0</td><td>6.8 ± 0.7</td><td>*</td></tr><tr><td>Global germination of the seed mix (%)</td><td>8.3 ± 0.77</td><td>6.2 ± 0.9</td><td></td><td></td><td></td><td></td></tr><tr><td>Cover of sown species (%)</td><td>48.8 ± 2.7</td><td>43.2 ± 4.6</td><td></td><td>68.6 ± 6.7</td><td>64.7 ± 8.7</td><td></td></tr><tr><td>Cover of spontaneous species (%)</td><td>36.4 ± 3.8</td><td>62.8 ± 6.5</td><td>**</td><td>41.9 ± 8.2</td><td>47.0 ± 5.4</td><td></td></tr><tr><td>Diversity of sown species (Shannon-Wiener index)</td><td>1.9 ± 0.06</td><td>1.9 ± 0.10</td><td></td><td>1.6 ± 0.09</td><td>1.1 ± 0.16</td><td></td></tr><tr><td>Cover of sown species (evenness)</td><td>0.8 ± 0.02</td><td>0.9 ± 0.02</td><td></td><td>0.7 ± 0.03</td><td>0.6 ± 0.06</td><td></td></tr></tbody></table>

The significant differences between each mix in each year are distinguished in column SD (significant differences): * ≤ 0.05; ** ≤ 0.01.
<figure><img src="image_47.png" /><figcaption>FIGURE 3
Germination percentage (mean±standard error) in 2019 of the seed mix species sown in the *in-situ* plots.</figcaption></figure>

<figure><img src="image_48.png" /><figcaption>FIGURE 4
Number of species (mean±standard error) in control and *in-situ* sown plots, considering the seed mix species and spontaneous species (other native and ruderal species), and total richness.</figcaption></figure>

Additionally, 96% of the sown species produced flowers and seeds in both monitoring years. The overall flowering period of the seed mix took place between March and June. The number of flora species that provide food or shelter for butterflies and small mammals increased in the sown plots through the germination and growth of species such as *Brachypodium distachyon* (L.) P.Beauv. and *Mantisalca salmantica* (L.) Briq&Cavill. The seed mix also contributed to the increase of endemic species in the sown plots through the germination and

persistence of *Pterocephalidium diandrum* (Lag.) G.López, *Sanguisorba hybrida* (Link ex G.Don) Ces., and *Silene scabriflora* Brot.

Diversity and evenness were similar for control and pre-sown plots before the intervention (springs 2017 and 2018). The application of the final seed mix resulted in higher (*p* < 0.001) species diversity and evenness in the sown plots (Figure 6).

The application of the final seed mix generated a significant increase (*p* < 0.001) of the evaluated flora community parameters
<figure><img src="image_49.png" /><figcaption>FIGURE 5</figcaption></figure>

Species cover percentage (mean±standard error) in control and *in-situ* sown plots, considering the seed mix species and spontaneous species (other native and ruderal species), and vegetation overlap.

<figure><img src="image_50.png" /><figcaption>FIGURE 6</figcaption></figure>

Species diversity (Shannon-Wiener index) and Evenness (Pielou Index) of the floristic community (mean±standard error) in control and *in-situ* sown plots.

(Table 2), highlighting the increase in total species richness (125%) and diversity (45%).

Discussion

In the *ex-situ* trials, the two mixes under analysis showed identical behavior for most parameters of the floristic community. This similarity between mixes can be explained by their composition, which includes some shared species or species with identical functional characteristics. In either mix, only some of the sown species managed to establish (mix 1: 43% in 2017 and 43.5% in 2018; mix 2: 38.2% in 2017 and 29% in 2018), which agrees with other studies carried out in semi-arid conditions (e.g., Tormo et al., 2006; Bochet et al., 2010), and the global germination percentage was less than 10%. This difference between the establishment performance of the sown species evidences the need to readjust the mixes' composition to ensure the success of the *in-situ* trials. Also, both mixes increased the vegetation cover in the second year despite a decrease in diversity and evenness of the sown species, reflecting a short-term persistence

ability for some species and an increase in some species' dominance, namely grasses. The increase in the vegetation cover could be related to the precipitation variability between years, since the rainfall was higher in the autumn/winter of 2017/2018. The decrease in diversity and evenness of sown species is expected over time (e.g., Haaland and Gyllin, 2011; Angelella et al., 2019); however, in these seed mixes, it started shortly after the intervention (2nd spring), which may indicate the need to readjust the seed mix proportions. Despite the similarities between mixes, in the 2 years monitored, plots sown with mix 1 registered: (a) a higher percentage of sown species present in the floristic community, (b) a lower cover of spontaneous species, and (c) a steady percentage of sown species that germinated (about 43%) in both springs. The performance of mix 1 in these parameters may indicate better suitability to the edaphoclimatic conditions and increased capacity to fulfil the study aims of promoting native flora in the pedestrian trails' surrounding. Thus, mix 1 was considered most suitable for implementation in *in-situ* trials. Nevertheless, to ensure better performance, minor adjustments were made to the composition and proportion of mix 1 to incorporate the positive aspects of both mixes. These adjustments included (a) the redistribution of some
<table><caption>TABLE 2 Proportional increase of flora community parameters in the *in-situ* plots sown with the final seed mix.</caption><thead><tr><th colspan="2">Increase of flora community parameters</th></tr></thead><tbody><tr><td>Species richness</td><td>+125%</td></tr><tr><td>Species diversity (Shannon-Wiener index)</td><td>+45%</td></tr><tr><td>Species evenness (Pielou Index)</td><td>+14%</td></tr><tr><td>Vegetation cover</td><td>+17%</td></tr></tbody></table>

species proportion (e.g., *Medicago doliata* Carmign. was excluded from the mix, and the proportions of the remaining Fabaceae species were therefore slightly adjusted), (b) the addition of one species from the unselected mix (*Aegilops triuncialis* L.), and (c) the replacement of species by others from the unselected mix that obtained a better germination percentage (namely *Salvia verbenaca* L. and *Stipa capensis* Thunb. were replaced by *Petrorhagia nanteuilii* (Burnat) P.W.Ball&Heywood and *Lamarckia aurea* (L.) Moench, respectively) or have higher conservation interest (namely *Cerinthe major* L. and *Sanguisorba verrucosa* (Link ex G.Don) Ces. were replaced by *Pterocephalidium diandrum* and *Sanguisorba hybrida*, respectively).

In the *in-situ* plots, all sown species germinated in the first and second years after the intervention, reflecting a better performance of the final seed mix than the original seed mixes tested in the *ex-situ* trials and ensuring a more suitable selection of the species composition and proportion. The seed mix germination percentage in the *in-situ* trial was not high (11%) but follows what is usually registered in rehabilitation actions (Ceccon et al., 2016).

Applying seed mixes can be challenging, and even in places with low floristic diversity, such as many urban areas, sowing a wildflower mix does not automatically result in vegetation with the same composition as the seed mix (Bretzel et al., 2009; Fischer et al., 2013a; Schmidt et al., 2020). These differences in the species assemblage may occur due to the seed bank response or changes in the species' competitive dynamic (Hobbs et al., 2006; Fischer et al., 2013b). In this study, the application of the seed mix increased the species number in the *in-situ* plots due to the germination of sown species but also of spontaneous species (other native and ruderal species), most surely because these benefited from the disturbance of the soil seed bank produced at sowing (Hölzel et al., 2012). Therefore, this seed mix complemented and improved the floristic richness of the sown areas, as intended, without replacing the existing flora.

Developing a seed mix must consider the characteristics of the selected species and its applicability to the target place, considering factors such as soil type, soil seed bank, original land use, sowing method and weather conditions (Bretzel et al., 2012; Scheper et al., 2021). Despite considering other environmental factors (sun exposure and weather conditions), it was not possible to incorporate soil analyses that might help to explain the species' different behavior in *ex-situ* and *in-situ* trials. All tested seed mixes (*ex-situ* and *in-situ* trials) incorporated a sufficiently broad range of species to successfully establish the seed mix; although, testing the seed mix in different soil types would benefit the study by evaluating its adaptive ability.

The vegetation cover and the overlap of herbaceous species layers increased in the sown plots, ensuring a seed mix representation of 33% in the floristic community, and densifying

the vegetation cover. The implemented seed mix is intended to promote grassland diversity and create favorable conditions for the permanency of small fauna such as butterflies and small mammals, what seems to be accomplished by increasing the number of these flora species, vegetation cover and vegetation layers. Among the sown species, the germination of *Brachypodium distachyon* and *Mantisalca salmantica*, for instance, stands out, because they contribute to the promotion of refuge and feeding areas for butterflies like *Melanargia ines* (Hoffmannsegg, 1804), *Thymelicus action* (Rottemburg, 1775), *Thymelicus lineola* (Ochsenheimer, 1808), and *Vanessa cardui* (Linnaeus, 1758) (Stefanescu, 1997; Maravalhas, 2003; Obregón and Prunier, 2014).

The final seed mix combined species from several families with different life forms (annual, biannual, and perennial), what increases the floristic community resiliency by allowing an extended flowering period and different survival strategies (Hoyle, 2016). Additionally, most species produced flowers and seeds, which is a good indicator of sustainability. The flowering period occurred throughout the spring (March to June), ensuring food availability for pollinators and an aesthetically pleasing appearance for pedestrians, an important feature in these circumstances, as noted by other studies (Lindemann-Matthies and Brieger, 2016; Hoyle et al., 2018; Scheper et al., 2021). An extended flowering period, without mowing, provides more opportunities for the self-reseeding of annual species (Hoyle, 2016), as the plants have more time to produce viable seeds, decreasing the need to reseed. The tested seed mix also allows a wide range of flower shapes and colors, contributing to more feeding opportunities for pollinators and improving the citizens' acceptance (Hoyle et al., 2018; Colombo et al., 2021; Scheper et al., 2021). Therefore, a seed mix composition that allows the longest and more diverse flowering season helps to increase biodiversity, minimizes reseeding costs, and improves the relationship between citizens and the urban rehabilitated grasslands. The long-term sustainability of urban rehabilitated grasslands depends on citizens' acceptance and local authorities' commitment (Norton et al., 2019); thus, ensuring a diverse and extended flowering season is a key factor when planning a seed mix composition.

The management timing and frequency of rehabilitated urban grasslands also play a decisive role in their long-term sustainability. Although mowing is essential to preserve grassland diversity (Tälle et al., 2018), recurrent mowing during the flowering season can make seed production unfeasible (e.g., Blažek and Lepš, 2015; Nakahama et al., 2016), decreasing the diversity of wildflower meadows over time. The studied pedestrian trails, and respective pocket parks, were managed by mowing the vegetation in late spring, allowing the sown species to complete their reproductive cycle, and contributing to a favorable evolution of the floristic community over the studied period. Based on these results and in other studies in the Mediterranean (Filibeck et al., 2016), we advise planning the mowing just after the flowering peak, i.e., between the middle and the end of June, in the case of the Mediterranean basin.

The three vegetation groups (other native, ruderal, and seed mix) under analysis were more balanced in the sown plots than in the control plots, with a decrease in the dominance of ruderal species and the fostering of less common spontaneous species. This balance between species follows other studies in urban areas (e.g., Fischer et al., 2013a,b; Norton et al., 2019), which verified a decrease in
ruderal species in rehabilitated areas and the ability of meadow species to adapt to ruderal environments. The increment in the species diversity and evenness in the sown plots also corroborates the improvement of the flora community balance. Besides, the application of the seed mix increased the presence of species with conservation interest, like endemic species such as *Pterocephalidium diandrum*, *Sanguisorba hybrida*, and *Silene scabriflora*. As pointed out by Hobbs *et al.* (2006) and Fischer *et al.* (2013a), the coexistence of sown and spontaneous species results in a new floristic community that is probably more ruderal than natural or semi-natural grasslands. However, the establishment of less common species highlights the green urban areas' potential to preserve rare species of natural and semi-natural grasslands.

The application of the final seed mix in pocket parks distributed throughout pedestrian trails of South Portugal largely contributed to ameliorate the indicators of the floristic community, generating a quick and significant increase of total species richness in 125%, diversity in 45%, evenness in 14%, and cover in 17%. In addition, the establishment of this seed mix did not require watering nor soil fertilizing and the mowing frequency was low (once in late spring), following the requirements for implementing an ecologically sustainable and low-cost wildflower meadow (Bretzel *et al.*, 2009; Köppler *et al.*, 2014). The low requirements for maintaining rehabilitated urban grasslands are one of the significant advantages of their implementation from the point of view of policymakers and local authorities (Hoyle *et al.*, 2017b). Thus, given these areas' economic and ecological advantages, ensuring that the population recognizes their ecological and social value can be a decisive factor in expanding their implementation.

The acceptance of native herbaceous vegetation in peri-urban or urban areas in the Mediterranean can be difficult, especially because of their appearance in summer (Filibeck *et al.*, 2016; Colombo *et al.*, 2021). Most native meadows and grassland species dry during summer, and the rehabilitated green areas can look unattractive and careless (Filibeck *et al.*, 2016; Colombo *et al.*, 2021). Mowing these areas in late spring can benefit the grassland species diversity, ensuring that they remain in an initial succession phase (Filibeck *et al.*, 2016), and allowing a more acceptable aesthetic appearance during the dry season.

This study did not include acceptance population surveys; thus, we have no information about the population reaction in the intervened cities. To our knowledge, during the monitoring period, the sown plots were not vandalized, and the citizens did not react negatively to the urban grasslands' rehabilitation. Informational signs were placed at the beginning of pedestrian trails and some pocket parks, and the sown plots were marked with stakes. The signage of the rehabilitated areas allows citizens to be informed, what improves the ecological culture and increases the citizens' respect (Filibeck *et al.*, 2016; Norton *et al.*, 2019). Signaling the plots also allows, if necessary, to adapt the vegetation management and ensure that the rehabilitated areas are only mowed after the flowering season, contributing to long-term preservation (Norton *et al.*, 2019). As referenced in other studies (e.g., Filibeck *et al.*, 2016; Hoyle *et al.*, 2017b; Southon *et al.*, 2018; Norton *et al.*, 2019), the size, location, distribution, and density of the biodiversity promotion plots in urbanized contexts can be critical factors in preserving the intervened areas

and their acceptance by the citizens. Thus, the plots' small size, sparse distribution through pedestrian trails/pocket parks, and localization in peri-urban areas may have also contributed to their preservation. Furthermore, during the summer, the plots and the remaining vegetation were mowed, and the sown plots appearance was similar to the surrounding matrix.

Our study shows that the application of seed mixes with regional native species allows the promotion of native flora in Mediterranean peri-urban green areas, enhancing the potential of these spaces as tools for biodiversity conservation. Furthermore, the development and commercialization of native seed mixes could prevent the use of seed mixes imported from different biogeographical regions and, thus, diminish the risk of introducing exotic species with invasive ability.

The final seed mix seems suitable for replicating actions in a peri-urban context under identical climate conditions, namely in the south of the Iberian Peninsula. In the Mediterranean context and given the small number of native seed mixes tested, our study represents an innovation and provides practical insights into grassland rehabilitation, contributing to the improvement of plant diversity' management in peri-urban and urban areas. Nevertheless, further monitoring of the seed mix dynamics and evolution will be necessary since some changes may occur over the following years, and adjustments to species composition or proportions may be required. Additionally, the sustainability of the seed mix must be assessed in the long term, including an evaluation of the need for seeding reinforcement.

## Data availability statement

The datasets presented in this article are not readily available because raw data were generated at the University of Évora, under the LIFE LINES project (LIFE14 NAT/PT/001081). Restrictions apply to the availability of these data, which were used under license for this study. Derived data supporting the findings of this study are available from the corresponding author MF on request. Requests to access the datasets should be directed to MF, mpfernandes@uevora.pt.

## Author contributions

MF, CP-C, and AB conceived and designed the study. All authors participated in field work and contributed to data collection. PM analyzed data in collaboration with MF. MF, PM, and AB wrote and edited the manuscript, with inputs from the other authors. All authors contributed to the article and approved the submitted version.

## Funding

This work is funded by National Funds through FCT – Foundation for Science and Technology under the Project UIDB/05183/2020.
Acknowledgments

MF was supported by a Ph.D. grant from FCT - Foundation for Science and Technology (2020.06339.BD). We thank the support of the project "LIFE LINES – Linear Infrastructure Networks with Ecological Solutions" (LIFE14 NAT/PT/001081), from the European Union LIFE Program. We also recognize the collaboration of Marlene Emídio during fieldwork and data collection.

Conflict of interest

The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.

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Supplementary material

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at: https://www.frontiersin.org/articles/10.3389/fevo.2023.1112596/
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OPEN ACCESS

EDITED BY
Péter Török,
University of Debrecen, Hungary

REVIEWED BY
Jana Rúžičková,
ELKH-ELTE-MTM Integrative Ecology Research
Group, Hungary
Béla Tóthmérész,
University of Debrecen, Hungary

*CORRESPONDENCE
Olivier Blight
olivier.blight@imbe.fr

†These authors have contributed equally to this work

SPECIALTY SECTION
This article was submitted to Conservation and Restoration Ecology, a section of the journal Frontiers in Ecology and Evolution

RECEIVED 09 September 2022

ACCEPTED 15 March 2023

PUBLISHED 17 April 2023

CITATION
Blight O, Geslin B, Mottet L and Albert CH (2023) Potential of RFID telemetry for monitoring ground-dwelling beetle movements: A Mediterranean dry grassland study.
Front. Ecol. Evol. 11:1040931.
doi: 10.3389/fevo.2023.1040931

COPYRIGHT
© 2023 Blight, Geslin, Mottet and Albert. This is an open-access article distributed under the terms of the [Creative Commons Attribution License (CC BY)](https://creativecommons.org/licenses/by/4.0/). The use, distribution or reproduction in other forums is permitted, provided the original author(s) and the copyright owner(s) are credited and that the original publication in this journal is cited, in accordance with accepted academic practice. No use, distribution or reproduction is permitted which does not comply with these terms.

# Potential of RFID telemetry for monitoring ground-dwelling beetle movements: A Mediterranean dry grassland study

Olivier Blight*†, Benoît Geslin†, Lola Mottet† and Cécile H. Albert†

Avignon University, Aix-Marseille University, CNRS, IRD, IMBE, Avignon, France

Better understanding insects' movements could help preserve and restore the insect communities that are key to the functioning of grasslands. Recent technological advances have led to spectacular achievements in movement ecology, making it possible to track the individual movements of a wide variety of organisms, including the smallest. However, monitoring systems such as RFID tags may negatively impact an organism's life history, with potential consequences on the reliability of data and conclusions. This study explored the potential of passive RFID tags to track the movements of three small ground-dwelling beetle species, a predator (*Poecilus sericeus*, Carabidae), a detritivore (*Asida sericea*, Tenebrionidae) and a granivore (*Acinopus picipes*, Carabidae), in a Mediterranean dry grassland degraded by years of cultivation. First, we tested whether carrying tags might impact individuals' behaviour, using a before-and-after design under laboratory conditions. Despite a trend toward shorter displacements, we found no significant short-term effect of the tags on individuals' movements. Second, we tracked a total of 25 tagged beetles in their natural environment every 4 h for 48 h. We highlight the principal limitation of using passive tags with small terrestrial beetles: the antenna has to pass over the tags to detect them, which restricts tracking to a few consecutive days after which the probability of locating an individual is low. However, the data obtained sheds light on the biological rhythms and daily movement capabilities of our target species: *A. sericea* is more mobile and *P. sericeus* less mobile than expected. Such knowledge could help predict the species' ability to recolonise degraded areas, enabling appropriate restoration actions to be designed based on landscape ecology principles.

KEYWORDS

movement ecology, conservation, grassland, insects, radio-tracking

## 1. Introduction

The global decline of grasslands is making conservation and restoration actions a priority (Buisson et al., 2022). Usually, it is plant species that are targetted in restoration actions, under the general assumption that insect populations will naturally follow the restoration