Biodiversity and Conservation (2019) 28:1343–1360 https://doi.org/10.1007/s10531-019-01741-8

REVIEW PAPER

Mitigating the precipitous decline of terrestrial European : Requirements for a new strategy

Jan Christian Habel1,2 · Michael J. Samways3 · Thomas Schmitt4,5

Received: 27 December 2018 / Revised: 8 March 2019 / Accepted: 11 March 2019 / Published online: 22 March 2019 © Springer Nature B.V. 2019

Abstract Severe decline in terrestrial species richness, abundance, fying biomass, and local extinctions across are cause for alarm. Here, we summarize this decline, and iden- tify species afected most. We then focus on the species that might respond best to mitiga- tion measures relative to their traits. We review apparent drivers of decline, and critically refect on strengths and weaknesses of existing studies, while emphasising their general signifcance. Generality of recent scientifc fndings on insect decline have shortcomings, as results have been based on irregular time series of insect inventories, and have been car- ried out on restricted species sets, or have been undertaken only in a particular geographi- cal area. Agricultural intensifcation is the main driver of recent terrestrial insect decline, through habitat loss, reduced functional connectivity, overly intense management, nitrogen infux, and use of other fertilisers, as well as application of harmful pesticides. However, there are also supplementary and adversely synergistic factors especially climate change, increasingly intense urbanisation, and associated increase in trafc volume, artifcial light- ing and environmental pollution. Despite these various synergistic impacts, there are miti- gating factors that can be implemented to stem the precipitous insect decline. Science can provide the fundamental information on potential synergistic and antagonistic mechanisms of multiple drivers of insect decline, while implementation research can help develop alter- native approaches to agriculture and forestry to mitigate impacts on insects. We argue for more nature-friendly land-use practices to re-establish Europe’s insect diversity.

Keywords Biodiversity crisis · Insect decline · Species richness · Abundance · Agricultural intensifcation · Habitat fragmentation · Habitat degradation · Pesticides · Climate change · Insect conservation

Communicated by P. Ponel.

* Jan Christian Habel [email protected] Extended author information available on the last page of the article

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Introduction

Europe’s insect diversity was recorded as declining in the ­19th century, when it was described as ‘heartlessly being swept away in this era of steam and telegraphy’ (Swin- ton 1880). Since then, the situation has worsened, with accelerating velocity of losses across Europe since the 1950s. This decline has been documented for various, mainly charismatic insect groups, such as lepidopterans (Maes et al. 2001; Conrad et al. 2004; Thomas et al. 2004; Thomas 2005; Fox 2013) and carabid beetles (Brooks et al. 2012). However, these well-studied groups are not unique, but refect a general decline in insects (Sánchez-Bayo and Wyckhuys, 2019; Samways 2019). Overall, Europe has seen a signifcant and on-going loss of general insect species richness (Thomas et al. 2004; Conrad et al. 2006; Müller et al. 2012; Habel et al. 2016; Simons et al. 2017). For example, the European Red List of Butterfies indicates a decline in these insects at the continental scale (Van Swaay et al. 2010). Similar trends have occurred at regional and local scales. Macro- and micro-lepidopterans of the fed- eral state of Bavaria in southern Germany have experienced losses of >13% since the year 2000 (Haslberger and Segerer 2016), and in south-western Germany, Luxembourg, and north-eastern France, butterfies and carabid beetles associated with calcareous grassland patches have declined precipitously in the last 50 years (Augenstein et al. 2012; Filz et al. 2013; Habel et al. 2016). In addition to decline in species richness, there have also been temporal shifts in species composition, as for example among but- terfies (Filz et al. 2013; Habel et al. 2016), wild bees (Potts et al. 2010), and carabid beetles (Augenstein et al. 2012). These species losses and assemblage shifts are often associated with declines in abundance of specifc taxa, and decreases in total biomass. Hallmann et al. (2017) showed a biomass reduction of 75% in fying insects at various sites in Germany over three decades, while similar negative trends have been recorded for other parts of Europe (Conrad et al. 2004; Shortall et al. 2009; Knowler et al. 2016; Storkey et al. 2016). Although these studies have temporal and spatial restrictions, they have the same conclusions: (1) reduction of insect species richness, (2) shift in species assemblages, with reduced species evenness, and (3) decline in many species’ abun- dance, i.e. losses of biomass. These conclusions lead to three general questions, which we explore in the following:

1. Which taxon traits are sufering the most, and which are not afected, or alternatively, beneft from the continuing environmental changes? 2. What are the main drivers leading to these negative trends? 3. How can science, policy, and management mitigate further insect decline, or preferably, reverse this negative trend?

We focus here specifcally on above-ground terrestrial insects, rather than aquatic insects, which face diferent (e.g. alien fsh predation, channelization, damming of riv- ers, functionally signifcant reduced leaf litter), but sometimes overlapping (e.g. pesti- cide impact, increased nitrogen and fertilizer input, fragmentation of habitats) threats and stressors. In turn, soil insects are subject to some diferent (e.g. loss of leaf lit- ter, soil erosion, changes in soil fungi composition) and some similar (e.g. changing microhabitats, landscape fragmentation, impact of invasive alien vegetation) impacts. All three realms, terrestrial, freshwater and soil, are afected by the pervasive efect of climate change, directly, or indirectly through changes in interaction networks.

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Trait analyses to discriminate losers from winners

Application of morphological, life-history, and physiological traits is useful for ana- lysing changes in patterns and abundance of biodiversity (Cleary et al. 2007; Mayfeld et al. 2010; Gossner and Müller 2011; Gossner et al. 2013; Thorn et al. 2015; Simons et al. 2016, 2017), and decreases in evenness, and hence homogenisation, of insect assemblages (Gossner et al. 2016). Thus, identifying response traits (e.g. life-history and feeding-related traits) to environmental changes may help to explain why certain species become less (or more) abundant over time. An underlying assumption is that the drivers act as flters allowing certain species to survive or remain in high abundance, and others not. Furthermore, trait analyses are informative for understanding commu- nity assembly rules (Mouchet et al. 2010; Vandewalle et al. 2010). For example, when analysing the characteristics of communities in response to land-use intensity, average insect body size decreases, and fight ability increases, with higher land-use intensity in grasslands (Simons et al. 2016). Consequently, more rapidly developing and mobile taxa are less afected than others. Furthermore, an increase in mobile species, and a decrease in sedentary species, indicates a response to increasing habitat isolation through higher mobility safeguarding species survival in a fragmented landscape (Thomas 2016). Spe- cies relying on specifc microhabitat structures and/or resources can sufer greatly from their reduction or loss. Consequently, deterioration of habitat quality can be as rele- vant as decreasing habitat size and increasing habitat isolation (Dennis and Eales 1997; Thomas et al. 2001). This means that a mechanistic understanding of how land-use changes afect assemblages can be achieved by trait-based analyses of com- munity shifts (Birkhofer et al. 2015). As species equipped with specialised traits are the ones sufering the most, it is essential to identify the specifc drivers causing disappear- ance of specifc taxa, as well as shifts in species composition. Highest rate of regional extinction in Danish butterfies has been among sedentary habitat specialists overwin- tering as eggs, while the most severe local-scale declines have occurred among seden- tary host plant specialists overwintering as larvae. In contrast, mobile generalists with a mature overwintering larval stage, have been least afected (Eskildsen et al. 2015). Trait-based analyses identify the following four drivers, which have diverging efects on diferent taxa, depending on their respective ecological requirements (Fig. 1).

Habitat loss

Habitat losses, e.g. of semi-natural habitats like grasslands, hedgerows, small set-aside areas, which were once integral components of traditional low-intensity agriculture, are greatly reducing regional species diversity (Reidsma et al. 2006). Loss of these habitats has severely afected extreme specialists, whose combination of traits does not allow any shift to alternative habitats. An example is the great loss of monophagous species through having lost their specialized bog habitats (Ebert and Rennwald 1991; Gelbrecht et al. 2016). The severe decline, and even regional extinction, of the highly specialised Maculinea butterfies, with their extraordinarily complex life cycles, in semi-natural, grazed grasslands is another example (Thomas et al. 2009). This makes trait space a major feature of this sensitive group of species. Consequently, even small habitat changes, can see local loss of these extreme specialists.

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Fig. 1 Summary of drivers causing insect decline, and actions needed to stop it

Increasing habitat isolation

Ongoing landscape modifcation leads to optimal habitats becoming increasingly spa- tially isolated from each other over time. This results in reduction of individuals, and hence gene fow, between habitat patches. As a consequence, metapopulation structures (cf. Hanski 1999) are altered, and in the case of the Glanville fritillary (Melitaea cinxia) in southern Finland, complete destruction of its metapopulation dynamics (Hanski et al. 1994). Metapopulation dynamics loss is leading to a decline and even local disappear- ance of sedentary species, as this trait specialisation is strongly dependent on dynamic population networks (Thomas 2016). Even species with moderately specialised traits, that in the past were widespread across the landscape matrix, are also sufering a decline in metapopulation dynamics, as in the case of butterfies with intermediate mobility (Thomas et al. 2001). Such species, not adapted for survival in small isolated popula- tions by their genetic composition, often sufer from inbreeding, and can become extinct for that reason (Habel and Schmitt 2018). This genetic trait constraint (where at least some mobility is required to ensure metapopulation dynamics) might also explain the apparent random loss of species from protected areas, in particular for species, which in former times, occurred as metapopulations (Augenstein et al. 2012; Filz et al. 2013; Habel et al. 2016). This has occurred in the hermit butterfy ( briseis) in north- ern Bohemia, where population genetic structure and genetic bottlenecks, along with the random (and hence maladapted) shift of alleles, or even the fxation of (weakly) delete- rious genes at some loci, have contributed to this species’ decline (Kadlec et al. 2010).

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Decreasing habitat quality

Several, partly independent factors, are responsible for habitat-quality reduction, and which afect diferent traits independently, or in combination. We focus here on the four most important aspects.

Fertilisers and atmospheric nitrogen deposition

Nitrogen accumulation and other fertilisers afect habitat quality, particularly in nitrogen- limited ecosystems (such as semi-natural grasslands, bogs, and heathlands) through dis- placement of plants that are outcompeted by species beneftting from increased nitrogen availability. In particular, fowering herbs are replaced by grasses (Stevens et al. 2004). These shifts reduce nectar availability for fower-visiting insects and specifc (and often rare) larval food plants for many phytophagous species. Consequently, pollinators over- all are more severely afected than non-pollinators, and species demanding specifc larval food plants more than oligo- or polyphagous species, leading, for example, to a decline in pollinators associated with decrease in insect-pollinated plants in parts of Europe (Bies- meijer et al. 2006). Most European butterfy species, for example, are particularly sensitive to increasingly intensive fertilizer regimes. Thus, butterfies adapted to living in nitrogen- limited ecosystems, such as semi-natural calcareous grasslands, show decreases in relative abundance (Habel and Schmitt 2018), and even local extinctions in some regions, from increased fertilizer use (Van Swaay et al. 2010). Species beneftting are those with high mobility in their trait space, and are multivoltine and/or polyphagous, with high repro- ductive capacity, rapid larval development, and that hibernate as pupae or adults (Wal- lisDeVries and van Swaay 2017). Only rarely can species with specialized traits beneft from increased stressors, for example, when a species is specialized on a larval food plant beneftting from eutrophication. Thus the peacock butterfy (Inachis io), which feeds exclu- sively as a larva on nitrogen-loving nettles ( Urtica), and the meadow brown butterfy (Maniola jurtina) that uses many grass species for development, are overall stable (IUCN 2018).

Pesticides

Insecticides act directly as mortal agents, while herbicides reduce host plant availability and negatively impact insect population size (Theiling and Croft 1988; Brittain et al. 2010; Geiger et al. 2010; Henry et al. 2012). Herbicides diminish area of occupancy of wild herbs, and are important nectar sources and larval food plants for many insects, includ- ing wild bees, hoverfies, and butterfies (Freemark and Boutin 1995). Furthermore, aeri- ally applied insecticides often drift, having an impact often far beyond where they were sprayed. These impacts are often stochastic, as they depend on various factors such as wind velocity and method of application (Frost and Ware 1970), as well as thermic uplift, often leading to the deposition of pesticides on slopes at higher elevations (De Jong et al. 2008). The point here is that drift results in sub-lethal doses of various pesticides, which can afect aspects of behaviour and neurophysiology, including learning performance of various insects (Desneux et al. 2007). Neonicotinoid exposure (from seed dressings) reduces wild bee density, solitary bee nesting density, and bumble bee colony growth (Rundlöf et al. 1 3 1348 Biodiversity and Conservation (2019) 28:1343–1360

2015), and exposure to sulfoxafor, a potential replacement for neonicotinoids, results in fewer bumble bee workers and reproductive ofspring, and leads to an overall decline in ft- ness (Siviter et al. 2018).

Inappropriate habitat management

Protected areas are just a small fraction of the landscape, and they are not always efec- tive in protecting insect species. The European Natura 2000 network for example, while high in butterfy species richness, has seen an overall decline of 10% in only 11 years. Although butterfy declines are also taking place outside the protected areas, the important point is that these geographically explicit areas are not necessarily efective for protecting insect targets (Rada et al. 2019). Lack of efectiveness is due in part to poor management of special and unique habitats. Good management of these small areas is a demanding task, especially as management that is benefcial for some insect groups is detrimental to others (Habel et al. 2016). Furthermore, most protected areas were not chosen for their high insect diversity, so the subsequent management plans usually frst target plants (often orchids) and/or vertebrates. Resulting management plans therefore, are often unsuitable for insects, especially as conservation measures are usually performed at large spatial scales for eco- nomic reasons, hence destroying many essential insect microhabitats.

Intensifcation and abandonment

Continuing mechanisation of agriculture is leading to more intensive, economically viable, and efcient agriculture, leaving insect species with fewer places for survival (Perović et al. 2015). Within felds, intensive use of herbicides is erasing plants apart from the desired crop, so reducing survival of insects, apart from crop pests. Highly diverse mowed or grazed meadows of former times, which constituted much of the agricultural land, have now largely been converted to high productivity grasslands dominated by just one or a few fast growing grass species (e.g. perenne) (Paniatowski et al. 2018). Also, grass- lands have often been overgrazed and fragmented. Fragments are vulnerable to loss of their original heterogeneity and ecological integrity through isolation, matrix efects from conversion of the surrounding area to arable land, intensive livestock farming, plantation establishment, and replacement by alien grasses. All these factors change grassland phe- nology, and adversely afect indigenous biodiversity (Wilsey et al. 2018). Depletion of traditional rural habitats is a major driver of insect diversity loss in Europe, greatly afecting many species with at least some specialisation in their trait space (Kuus- saari et al. 2007). However, where intensifcation is not economically feasible, felds and meadows are set aside and become fallows, which after some years, lose much of their ecological value for species in need of open habitats, as demonstrated for butterfies (Baur et al. 2006). In this respect, precision farming can lead to loss of essential small habitats (e.g. grassland strips along feld margins, and fallows), which are additional suitable habi- tats (Ranjha and Irmler 2014; Toivonen et al. 2016). The vanishing insects at the landscape level, and even in habitats where we are not able to identify any signifcant changes of eco- system quality, suggests a tipping point has been reached, with entire population networks of many insect populations at risk of collapse. This is manifested by the huge declines seen among many insect species across Europe. Intensifcation has also occurred in urban habi- tats, with drivers of insect decline including loss of adjacent rural areas and subsequent loss of greenspace (Guenat et al. 2019), increased trafc intensity (Martin et al. 2018), and 1 3 Biodiversity and Conservation (2019) 28:1343–1360 1349 artifcial lighting (Wilson et al. 2018). These declines are refective of the traits of the vari- ous insect species (Iserhard et al. 2019), as in the case of agricultural intensifcation.

Climate change

Higher temperatures are resulting in higher individual growth rates, and in poleward geo- graphical range shifts. Consequently, cold-adapted and continental species in general are declining across Europe, as predicted by climate envelope models in the Climate Risk Atlas of European Butterfies with its detailed single-species studies (Settele et al. 2008). The boreo-montane butterfy Lycaena helle, for example, is not only limited by its ecologi- cal traits to a highly complex and difcult-to-preserve habitat type, but is also sensitive to global warming (Habel et al. 2011). Increasing frequencies of extreme weather events (e.g. droughts, heavy rainfalls, extremely high but also low temperatures and rapid temperature changes) taking place in the wake of climate change also have additional negative efects on local populations of many species (McLaughlin 2002). Furthermore, a warmer climate in general, triggers faster and denser growth of the vegetation, detrimental for many insect species, as the microclimatic conditions at the specifc places of larval development are deteriorating by becoming less sunny and more humid (Filz et al. 2013). This denser vege- tation is subject to drought, overall leading to increasingly hot and frequent fres, which can have detrimental efects on many insect taxa, especially those not well adapted to such fres (Yekwayo et al. 2018). However, in contrast to much current opinion that climate change is the main driver of insect decline, other factors related to agricultural intensifcation appear overall to be of much higher relevance, notwithstanding the increasingly synergistic impacts of a whole range of stressors and pressures. The relative importance of these various drivers vary from one geographical region to another, and for the diferent land-use systems, habitats, and ecological species groups (Hallmann et al. 2017). The diferential variation among drivers is inevitably interac- tive with the various traits among species. For example, species depending on particular resources (e.g. specifc food plants for larval development or specifc microhabitats) are greatly afected by deteriorating habitat quality. In contrast, sedentary species sufer more severely under the driver of increasing habitat isolation (Habel et al. 2016). Furthermore, some traits among certain species enable survival under one driver, while other traits among other species enable to survival under a diferent driver. In short, any one driver at one time depends on the specifc driver at a particular time. As it is only the most generalist and widely-adapted species that can survive all the impacts at the same time, and the whole complex of drivers can interact in an adversely synergistic way, these two facets (traits vs. drivers contemporaneously) inevitably are likely to exacerbate the decline in insect diversity in a rapidly changing world. This occurs when climate change and habitat loss act synergistically with each other, leading to ‘a deadly anthropogenic cocktail’, which is more ‘deadly’ when there is an increase in intensity at the same time (Travis 2003). Thus, diferent divers might have similar efects, and consequently show synergistic reinforcement, as with nitrogen deposi- tion and climate warming. This results in denser vegetation cover being detrimental for many insect species, making it difcult to identify the individual impact of each of the drivers. Alternatively, interaction of diferent drivers might have counterbalanc- ing efects. Thus, the disadvantage conferred on an insect by a cooler and more humid microclimate, resulting from the enhanced vegetation growth, could be counterbalanced 1 3 1350 Biodiversity and Conservation (2019) 28:1343–1360 by rising temperature from climate change. However, detailed knowledge on mode and consequences of interactions between diferent drivers is still largely unresolved (Hall- mann et al. 2017).

Trait analyses of time‑series help to understand insect decline

Insect populations naturally fuctuate over time, and can vary over at least three orders of magnitude over a few years (Den Boer 1985), as with some forest , driven in part by the relative impact of the host’s natural enemies from one year to the next (Myers 2018). Insect population peaks also change in local position around the landscape in response to optimal conditions, as illustrated by the classical studies on the common Cinnabar (Tyria jacobaeae) (Dempster 1975). Therefore, comparisons of single years, or limited numbers of time cohorts, are often problematic, and they may even indicate misleading trends. As a consequence, there is an urgent need for data from long-term insect assess- ments (including species assemblages, abundances, and biomass), and to relate the trait space of the respective temporal assemblages at specifc localities to the prevailing envi- ronmental conditions, abiotic and biotic, natural and anthropogenic (i.e. the above-men- tioned four main drivers of biodiversity loss). Use of large data sets and sophisticated sta- tistical analyses can then allow us to disentangle natural from anthropogenic drivers, as well as subsequent discrimination between the diferent anthropogenic divers. Consequently, the challenge for long-term studies is that high-quality data sets, well suited for analyses, are currently very few. Those available have frequently not been assessed within the framework of a standardized data collection, meaning that they have various shortcomings (Saunders 2017), and sampling has: (1) been intensively conducted but restricted to a single species (e.g. Reichholf 2005, 2006, 2008), thus not allowing anal- yses of changes in the trait space of communities, (2) considered only few time steps (e.g. Wenzel et al. 2006; Augenstein et al. 2012; Filz et al. 2013; Hallmann et al. 2017), mean- ing that the obtained changes in the trait composition might be blurred by natural fuctua- tions, (3) been done in geographically restricted regions (e.g. Habel et al. 2016), implying that local phenomena might be overriding large-scale efects, which are more relevant for biodiversity conservation as a whole, and/or (4) covered geographically large areas, but show many data gaps (across time and/or space) (e.g. Desender et al. 2010), hence combin- ing the analytical problems of points 2 and 3 (McGill et al. 2015). Considering these shortcomings in the temporal analyses of trait spaces of assem- blages or communities allows disclosure, and partly disentangles, the efects responsible for insect decline in the recent past, and provides pointers for future conservation activi- ties. The diferences seen, in part, are due to the shortcomings of time-gaps in data sets from voluntary and/or temporally random collection of data, which nevertheless have been mostly overcome by development of appropriate statistical techniques (Desender et al. 2010; Szabo et al. 2010; Isaac et al. 2014; Johnston et al. 2018). However, the explanatory power and generality of these data are still controversial, and so it is impor- tant that all projects identify and verify congruent, negative trends in insect diversity, abundance, and/or biomass. This congruence would then provide the highly informative data from time-series in general, and especially when they are linked to the trait space of the respective insect assemblages or communities. Against this background, we now discuss potential solutions on how to address these challenges and shortcomings, and suggest protocols for overcoming them.

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Diferences in sampling intensity

Traditionally, sampling intensity has varied greatly among years and geographical regions. To correct for unequal sampling intensity, bivariate comparisons of sampling years, in combination with random sampling approaches, need to be applied, so that despite difer- ent sample sizes, relative numbers of records can be compared among years. One approach is to focus on shifts in relative abundance in time compared to those expected by randomly assembled communities. These techniques are those used by Hallmann et al. (2017) and Lyons et al. (2016).

Data gaps across time and space

In addition to variable sampling intensity, sampling among almost all data sets has not been conducted continuously throughout time. To address this challenge, various studies have used species-occurrence time matrices as a basis for subsequent analysis. One solu- tion is to frst assess minimal sample sizes for reliable richness estimation using species accumulation curves (Mehr et al. 2011). If necessary, it is possible to pool subsequent study years to sample time windows (e.g. 3, 5 or 10 year frames) with appropriate sam- ple sizes. Similar approaches can cluster sample sites in combination with subsequent spe- cies-area curves. When data are treated as random samples from the whole assemblage or community (assuming no systematic sampling bias), these samples can be compared using bootstrap and random placement techniques to estimate degree of error for each temporal comparison. Randomisations that retain local numbers of incidences and species occur- rence are able to interpret such incomplete data sets for species co-occurrence and beta diversity analyses (Ulrich et al. 2017a, b, 2018). In cases of large sample sizes, standard richness estimators, such as those contained in INext, can be applied (Kunin et al. 2018).

Combining data collected with diferent sampling techniques

Sampling insects is conducted in diferent ways, such as use of diferent trapping meth- ods, including light trapping, baiting, pitfall trapping, transect counts, and random sam- pling. Comparing such data is difcult, as they include method-specifc sampling biases (Samways et al. 2010). As far as possible, only data obtained by the same sampling tech- nique, methodologically and spatially, should be combined. In cases where diferent sam- pling techniques have been used within the same habitat and time window, is it possible to compare relative abundance distributions obtained from each technique. Ideally, these should be identical. Species that deviate greatly between the sampling approaches have to be excluded from further analysis. Where data from diferent sampling techniques are included in the same analysis, they mostly have to be reduced to a presence-only analysis applied to raster grit cells.

Diferent spatial resolution

Data acquisition has often been performed at diferent spatial resolutions, from point data to Universal Transverse Mercator grid resolution. Where possible, the best solution is always to compare only data with similar resolution. If not possible, analysis has to refer

1 3 1352 Biodiversity and Conservation (2019) 28:1343–1360 to the largest spatial resolution within a data set. Prior to this, it is best to exclude low-res- olution data to improve general quality of the entire data set. Necessarily, this also applies to the respective environmental data. For a series of sites with diferent spatial extent, we suggest constructing species-accumulation and species-area curves (Kunin et al. 2018), as was also identifed above for data gaps across time and space (Mehr et al. 2011). In cases of temporal data, sample size-time curves also account for diferent sample sizes (Ulrich et al. 2013). Parameter values can be compared among studies and regions, and they also provide information on regional species turnover and diferences in total richness.

Future steps in conservation strategies and political discourse

Insects are of high ecological and economic importance, and play key roles in species interactions. They are major components in all terrestrial food webs, and provide resources for many organisms at higher trophic levels, especially for many vertebrates (Schowalter 2011). Insects also provide important ecosystem services in biological pest control (preda- tors, parasitoids), soil formation, and, most prominently, as pollinators (Losey and Vaughan 2006). Decline in pollinators is also associated with the local extinction of plants relying on the declining pollinators (Biesmeijer et al. 2006). This in turn, may lead to destabilized ecosystems, due to lower species richness causing higher instability across trophic levels (Gossner et al. 2016). Pollination in general has high economic value, with signifcant impact on crop yields, and so plays a pivotal role in agro-economy. Its annual value for agricultural plants is estimated to be 200–600 billion US$ (Klein et al. 2007). Furthermore, natural pest regulation strongly depends on the presence of intact insect assemblages and their trophic interactions (Bianchi et al. 2006). Consequently, stopping insect decline is not a solely ethical question, but is highly important for the well-being of humanity, so we now address the necessary conservation activities needed to counteract the four main drivers of insect decline. We also focus on policy decisions in these environmentally rapidly chang- ing times.

Conservation of high quality habitats

Protection of high-quality insect habitats is a frst priority. Laws prohibiting destruction of these valuable habitats have been implemented at the European Union (EU) level, and are mandatory for all member states. Although the implementation of the Natura 2000 network (N2000) aims to protect species and ecosystems, and covers 18% of the land surface of the European Union (European Commission 2015), it is the Habitats Directive that focuses on threatened species (European Economic Community 1992). However, many of the targeted species listed on the annexes of the Habitats Directive represent marginal or even relict populations, and thus the EU often does not cover the core (and thus the most relevant parts for conservation) of their global distribution range (Habel et al. 2015; Habel and Schmitt 2018; Gómez-Rodriguez and Baselga 2018). Furthermore, the Habitats Directive largely focuses on birds and plants, and now requires greater inclusion of invertebrates (Hochkirch et al. 2013). For achieving this conservation goal, high-quality habitats in the agricultural matrix have to be re-established, and extended in size, as small and isolated nature reserves may not be able to guarantee long-lasting preservation of insect species (Habel et al. 2016). In this context, we suggest a new approach, where public interest has priority over indi- vidual ownership rights. However, adequate compensations have to be paid. 1 3 Biodiversity and Conservation (2019) 28:1343–1360 1353

Increasing landscape permeability

Long-term persistence of many species depends on healthy population networks with high functional connectivity (Hewitt et al. 2016). Taking biodiversity-friendly actions to ensure such dynamics continue, and ensuring heterogeneity rather than homogeneity across the landscape, as well as ensuring landscape permeability, requires economic incentives. In this context, subsidies paid by the European Commission to support agriculture conse- quently must be re-thought, and only the most environment-friendly land-use practices should be economically supported. This would convert subsidies into payments for ecosys- tem conservation by land-users, a strategy that would also help to stem landscape fragmen- tation, through additional high-quality habitats being created, and existing habitats of low quality improved (e.g. reversion of monotonous, high productivity grasslands into diverse and fower-rich meadows). Furthermore, a general ecological intensifcation (Garibaldi et al. 2019) of agriculture, by using, for example, feld margin extension and roadside eco- logical landscaping, will further increase habitat connectivity, and will increase the area of fowering plants, so improving landscapes towards more insect-friendly conditions. For example, grassy strips encourage ground beetles to move into adjacent felds (Ranjha and Irmler 2014). Even small temporal fallows of arable felds improve conditions for bumble bees and butterfies, as for example in Finland, with long-term fallows being especially favorable for specialist species (Toivonen et al. 2016). However, these strategies do not imply any reduction of money transfer into agriculture, but rather a completely diferent distribution of the existing resources, which ideally, should be increased. This approach is, in efect, a long-term insurance policy for future delivery of irreplaceable and essential insect services.

Safeguarding habitat quality

Constant enrichment of the habitats with nitrogen and other fertilising agents is not solv- able within agriculture alone. Therefore, other anthropogenic sources of nitric oxides and ammonia (e.g. trafc, industry, households) have to be assessed for a maximal possible reduction. Furthermore, the detrimental efects of pesticides have to be further reduced, both in the agricultural and urban arenas. This is particularly relevant for neonicotinoids, which cause, among other impacts, reduced capacity of bee species to establish new popu- lations in the year following exposure to these pesticides (Woodcock et al. 2017). In this context, more intense moratoria at the EU level in the case of chemicals contributing to insect decline are necessary, with implementation having started. Furthermore, substances known to strongly harm insect diversity, even in sub-lethal doses, have to be taken of the market. In general, an orientation towards organic farming with generous support for the transitional phase might contribute to stopping or even reversing insect decline. Using organic farming along with in-feld plant diversifcation (Lichtenberg et al. 2017) greatly benefts pollinators, suggesting that organic production schemes can support species deliv- ering ecosystem services (Kremen and Miles 2012), while using natural enemy services along with cultural control for containing pest populations i.e. promote multi-service deliv- ery (Isbell et al. 2011). As management practices vary among many types of crops, it is essential to have focal systems, and investigate each in their own merit within a particular area or in a particular landscape context (Ponisio et al. 2015). We must also consider dif- ferent aspects of crop management practices, such as crop rotation, intercrop combinations,

1 3 1354 Biodiversity and Conservation (2019) 28:1343–1360 composting, crop variety, biological control, and reliance on natural pollinators, among others. Instigating a suite of these management practices would close the yield gap between organic and conventional systems, especially when tailored to local conditions (Cunning- ham et al. 2013). For example, to be efective at reasonable cost, rotational grazing for maintaining insect diversity must be practicable, while also achieving conservation objec- tives (Morris et al. 2005), while also using mowing, especially alongside the leaving veg- etation refuges rich in fowering plants (Noordijk et al. 2010). For example, conservation areas around Regensburg (south-eastern Germany) have lost many red-listed species, which still occur in a neighbouring highly disturbed quarry, which provides many diferent habitats, including diferent successional stages (Segerer et al. 1987). In sum, points 2 and 3 underscore that a move to organic farming, away from conven- tional agriculture, more insightful biologically diverse and sensitive approaches are avail- able. Besides maintaining high yields, this approach invests greatly in a more sustain- able future. Various studies have shown that organic farming supports biodiversity, and strengthens persistence of local insect populations (Rundlöf et al. 2008; Crowder et al. 2012; Gabriel et al. 2013; Puech et al. 2014; Schneider et al. 2014; Tuck et al. 2014). However, still much more applied research is required to improve yields in the context of organic farming (Seufert et al. 2012).

Insect conservation, society, and citizen science

Insect decline and conservation have to be understood as a societal and economic chal- lenge, as well as scientifc, and requires (1) philosophy (establishing the ethical founda- tion), (2) research (the fnding out), (3) policy (the framework for action, (4) psychology (understanding how to engage humans in insect conservation action), (5) practice (imple- mentation of action) and (6) validation (establishing how well we are doing at conserving insects), all on an economically viable platform (Samways 2018), and with a wide range of expertise working intensively together to mitigate the four major drivers responsible for insect decline as outlined above. Furthermore, a more general understanding of the value of insects needs to be achieved, in terms of ecosystem services, not just in economic terms, but also for human wellbeing. Here, communication regarding the irreplaceability of many insect species is crucial, against current erroneous assumptions that ecosystem services can easily be substituted by technical solutions (Simaika and Samways 2018). There is also an urgent need for environmental education, especially of young people and teachers in urban areas to increase insect conservation awareness (Samways 2007), while at the same time accruing scientifc evidence alongside improvements in scientifc literacy and skills (Saun- ders et al. 2018). Furthermore, there is a pressing need to move towards more ecological expertise among entomologists, with the concerning fact that insects now feature less often in biology textbooks (Gangwani and Landin 2018), and which has to be urgently addressed so as to have an insect-rich and healthier future that benefts all.

Acknowledgements We thank two anonymous referees for helpful comments to improve our article.

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Afliations

Jan Christian Habel1,2 · Michael J. Samways3 · Thomas Schmitt4,5

1 Evolutionary Zoology Group, Department of Biosciences, University of Salzburg, Hellbrunner Str. 34, 5020 Salzburg, Austria 2 Terrestrial Ecology Research Group, Department of Ecology and Ecosystem Management, School of Life Science Weihenstephan, Technische Universität München, 85354 Freising, Germany 3 Department of Conservation Ecology and Entomology, Stellenbosch University, Stellenbosch 7602, South Africa 4 Senckenberg Deutsches Entomologisches Institut, 15374 Müncheberg, Germany 5 Department of Zoology, Institute of Biology, Faculty of Natural Sciences I, Martin-Luther- University Halle-Wittenberg, 06099 Halle (Saale), Germany

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