Insect Conservation and Australia’s Grasslands [1st ed. 2019] 978-3-030-22779-1, 978-3-030-22780-7

Table of contents :
Front Matter ....Pages i-xv
Introduction to Grasses and Grasslands (Tim R. New)....Pages 1-35
Australian Grasslands – Variety and Extent (Tim R. New)....Pages 37-57
Agents of Change – Management and Succession (Tim R. New)....Pages 59-69
Intricacies of Grassland Management for Conservation (Tim R. New)....Pages 71-88
Urban Grasslands (Tim R. New)....Pages 89-97
Insects in Grasslands: The Key Groups for Understanding (Tim R. New)....Pages 99-141
Flagship Insect Species in Australia’s Grasslands (Tim R. New)....Pages 143-151
Pasture Pests (Tim R. New)....Pages 153-165
Maintaining Ecological Integrity and Processes (Tim R. New)....Pages 167-178
Grassland Management for Insect Conservation: Grazing, Mowing, and Fire (Tim R. New)....Pages 179-234
Grassland Management for Insect Conservation: Restoration (Tim R. New)....Pages 235-256
Back Matter ....Pages 257-272

Citation preview

Tim R. New

Insect Conservation and Australia’s Grasslands

Insect Conservation and Australia’s Grasslands

Tim R. New

Insect Conservation and Australia’s Grasslands

Tim R. New Department of Ecology, Environment & Evolution La Trobe University Melbourne, VIC, Australia

ISBN 978-3-030-22779-1    ISBN 978-3-030-22780-7 (eBook) https://doi.org/10.1007/978-3-030-22780-7 © Springer Nature Switzerland AG 2019 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, express or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Switzerland AG. The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Preface

‘Grassland’ is much more than ‘land covered with grass’, as the name implies. In practice, the term covers a multitude of vegetated areas dominated by low-growing plants and largely free of woody vegetation, so that absence or paucity of shrubs and trees is perhaps their most unifying feature. As well as grasses, these ecosystems may contain a great variety of forbs, and sedges or other plants may even predominate over grasses in their composition. Many are among the most species-rich terrestrial ecosystems. Grassland ecosystems comprise three complex and interacting components, listed by Lemaire et al. (2005) as (1) one or more vegetation communities with a varied population of herbivores, (2) the physical and chemical components of soil and (3) a diverse soil microbial community and microflora. Their management must accommodate the various functions and sectoral priorities of value and use – environment, biodiversity, landscape ecology and agricultural production within a range of socioeconomic contexts – and sustain all of these in the face of increasing exploitation and change. The themes of this book focus largely on the first of the above components, with plants and insects the foundation of, often complex and diverse, communities influenced by anthropogenic activities that can fundamentally change grassland areas and their roles as support systems for biodiversity. The term ‘grassland’, although used universally, is thus a considerable oversimplification that embraces this vast variety of environmental complexities – but still conveys the general impression of the open landscapes that dominate many parts of the world. They are the areas on which most human settlement has occurred; on which agriculture and pastoral activities were founded, and are the basis for human food production through conversions to cropping and grazing by domestic stock, most notably cattle and sheep; and on which many undomesticated mammals also depend; the classic visages of vast hordes of grazers on African plains and historically abundant bison on North American prairies represent much wider uses. Indeed, it has been claimed that 25% of global land area is devoted to grazing (Asner et al.

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2004). Grasses include cereals  – which have been proclaimed (Tscharntke and Greiler 1995) to be consumed by people in greater amounts than all other human foodstuffs combined and, in providing about 49% of calorific intake of humankind, are the staple crops for ‘billions of people’ (Overholt and Franck 2017). Recent interests in biofuel represent a further dimension of grassland changes. Conversely, grasses also include many of the world’s worst weeds, as either natural invaders or unplanned spread from deliberate plantings elsewhere. The collective changes wrought to satisfy and protect continuing human needs remain one of the greatest transformative agents to native grasslands, whatever their more precise botanical designation, and anthropogenic changes amongst the major causes of natural grassland losses and ensuing conservation concerns for the increasingly fragmented and vulnerable remainder. In short, native grasslands are under threat from extensive, often permanent, changes, including alienation by predominance of introduced plants that benefit humanity – but often at the expense of native species and interactions. Many localised grassland ecosystems have disappeared completely or declined to small fractions of their former extents. Those changes continue as increasing human populations seek to assure food for the future and a place to live, both as formidable pressures on open land areas and transforming them to conditions far from their natural states. Harmonising those needs with conservation of little-appreciated biological complexity is a major challenge. The uses and ecological significance of wise management of grasslands have generated a vast literature on grassland ecology, collectively embracing all major grassland regions, values and categories. Much of the collective emphasis has been driven by economic needs – of livestock production and food crop protection – and amenity values (such as human recreation), as well as the consequences for biodiversity of the transformations of grasslands from their natural conditions. A comprehensive account of grassland ecology (Gibson 2009) provides much background but, other than noting the importance of insect herbivores (notably grasshoppers and termites) and wider roles of termites and ants, made little reference to insects, and the classic volume on grassland invertebrates by Curry (1994) remains an invaluable source of information. The natural floristic variety of grasslands, extending over large areas in many parts of the world, supports equivalent variety of insects and other invertebrates, many of them also localised, ecologically specialised and restricted and equally vulnerable to changes as their host grasslands decline or disappear. Increasing recognition of the diversity, susceptibility and functional roles of grassland insects, together with the consequences of their loss, leads to appreciation of the importance of conserving grasslands and the unique insect faunas they support. The open structure of grasslands and their rich complements of associated insects have considerable advantages for ecological investigations, beyond their fundamental importance. The simple architecture – compared to forest, for example – renders grassland occupants relatively easy to observe, sample and study. However, in the public eye, initiatives for conservation of grasslands are perhaps not as prominent as are calls for

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forest conservation, but the peculiarities of grasslands and the variety of insects they support form notable components of the world’s biological heritage, in systems that are familiar  – as meadows, home gardens, urban parks, pastures and cropping areas – to people in all walks of life. The conservation of insects in these grasslands is the major theme of this book, in which I hope to provide sufficient background information to the many non-entomologists concerned with grassland biology and charged with managing grasslands in Australia for them to appreciate the importance of conserving grassland-associated insects and the ammunition to achieve positive outcomes for this to occur. Whilst to many people ‘grass is grass’, to many insects ‘grass’ is a very specific subset or association of Poaceae, sedges and associated forbs, often accompanied by restricted structural conditions such as sward height and density and the prevailing microclimate. Many of those subsets are characteristic ecosystems that are both geographically and ecologically restricted. Poaceae (Gramineae), with around 1300 Australian species, are the third largest plant family in the country. They occur at all elevations from sea level to the highest mountain ranges, and the introduction to the family in the flora of Australia (Mallet and Orchard 2002) also noted that grasslands and grassy woodlands of temperate Australia are ‘now largely modified’ to fulfil and satisfy their ‘indispensable roles in human economies’. Those alienations continue to threaten sensitive native biota, including many insects about which very little is known. Far more attention has been paid to the relatively few insects that have become pests of pastures and cultivated grasslands, simply because of human need to counter their impacts. This book embraces a range of structurally similar ‘open habitats’, unified by absence or very low incidence of woody vegetation and dominated by grasses and associated forbs or by other low-growing vascular vegetation such as sedges. Despite their botanical variety, these systems have much in common when their management and conservation are considered, so that experiences from each overlap and inter-fertilise. They are subject to very similar disturbances and losses by anthropogenic land-use changes, and the threats to their resident biota are broadly similar. So, therefore, are the measures that may alleviate those impacts and the ambit needs of resident insects as their habitats change by successions or human interventions. The concept of ‘grassland’ adopted here largely follows the broad scope of Daubenmire (1968) in his classic overview of fire in grasslands, as ‘any herb-dominated vegetation, steppe to tundra to marsh’, but I have omitted much of his corollary ‘as well as herb-dominated layers of savanna or open forest’. The book is a companion to a parallel treatment of insect conservation and Australian forests (New 2018), and the fates of Australian grasslands are outlined in a wider global context of grassland insect ecology and management for their conservation. As for that earlier book, the requirement for each chapter to be read independently has led to some overlap in contents as different themes and contexts are given priority: cross-references to page numbers across chapters are given in paren-

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theses. And, also as for forests, much of the pioneering work on insect conservation in grasslands has been undertaken in other parts of the world, and examples and outcomes from these are directly relevant as the discipline develops in Australia. This information is based on scientific publications and more informal ‘grey literature’ available to me by late 2018. The variety and complexity of Australian grasslands and of the insects they support, together with the extent of losses and needs for management, reflect those conditions elsewhere but have received relatively little attention. In contrast to much of the northern hemisphere, intensive grassland ­management has moulded grassland character over a much shorter period. The first Australians understood and used fire to manage grasslands sympathetically and sustainably over far longer periods, and the modern era of pastoralisation, intensive cropping and rapid urbanisation contrasts markedly in its impacts. That extensive change largely involves direct loss of native grasslands or their replacement by alien species – such as in ‘pasture improvement’ that increases productivity of grazing stock and sustains key industries on which Australia’s economic prosperity rests – and provides environments far different from those they replace. Those changes also create energetic and controversial discourse over ‘conservation versus development’. The needs to harmonise different priorities for the greatest common good reflect the extent and rapidity of changes to Australia’s grasslands and the losses of numerous insect and other species about which very little is known. Several centuries of low-intensity grassland management have moulded many northern temperate region grasslands as ‘seminatural’, a condition which reflects that legacy of gradual imposed change and has allowed them to continue as havens for much local biodiversity. Those gradual changes have provided models for recent grassland management in seeking to emulate natural disturbance regimes to which resident insects and others have become adapted and in which they can persist and – where necessary – be managed. In contrast, a parallel condition is largely absent in the more heavily populated parts of Australia, where rapidity of transformations over only about two centuries has prevented any such more gradual and less abrupt transitions. Many native grasslands have simply been lost or reduced to small, often isolated, remnants whose biological significance is only now being acknowledged. Those remnants, again reflecting a relatively short period of severe imposed changes, can still harbour much of the ‘pre-European’ complement of plants and invertebrates: they are irreplaceable ‘snapshots’ of biota largely displaced (and in some cases wholly lost) from the wider landscape as suitable habitats disappeared. Others are undergoing invasions by alien grasses and other taxa and are in the process of transformation from natural to seminatural in character, together with reduction in the ecological integrity and the ecological processes they support. Concerns for Australian grassland conservation are now widespread, but the future for many of

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their endemic and unusual denizens, including numerous insect taxa that do not occur elsewhere, remains uncertain. I hope that the background perspective given in this book may contribute to a more informed awareness of their importance and how their chances for persistence may be improved. Melbourne, Australia

Tim R. New

References Asner GP, Elmore AJ, Olander LP, Martin RE, Harris AT (2004) Grazing systems, ecosystem responses, and global change. Annu Rev Environ Resour 29:261–299 Curry JP (1994) Grassland invertebrates. Chapman and Hall, London Daubenmire R (1968) Ecology of fire in grasslands. Adv Ecol Res 5:209–266 Gibson DJ (2009) Grasses and grassland ecology. Oxford University Press, Oxford Lemaire G, Wilkins R, Hodgson J (2005) Challenges for grassland science: managing research priorities. Agric Ecosyst Environ 108:99–108 Mallet K, Orchard AE (eds) (2002) Introduction. In: Flora of Australia 43, Poaceae 1: introduction and Atlas. Australian biological resources study. Canberra/CSIRO Publishing, Collingwood, pp x–xii McCusker A (2002) Structure and variation in the grass plant. In: Mallett K, Orchard AE (eds) Flora of Australia 43, Poaceae 1: introduction and Atlas. Australian biological resources study, Canberra/CSIRO Publishing, Collingwood pp 3–18 New TR (2018) Forests and insect conservation in Australia. Springer, Cham Overholt WA, Franck AR (2017) The invasive legacy of forage grass introductions into Florida. Nat Areas J 37:254–264 Tscharntke T, Greiler H-J (1995) Insect communities, grasses, and grasslands. Annu Rev Entomol 40:535–558

Acknowledgements

Permission to reproduce or modify material for which they hold copyright has been granted generously by the following publishers and organisations, whose courtesy is acknowledged gratefully. Every effort has been made to obtain permissions to use such previously published material, and the publishers would welcome advice on any inadvertent omissions or corrections that should be included in any future editions or imprints of this book. Most illustrations used have been redrawn to ensure standardisation of lettering, and some have been simplified from their original format, as acknowledged in individual figure legends. Sources of advice on needs and permissions are as follows: American Midland Naturalist; Commission for Environmental Cooperation, Montreal; CSIRO Publishing, Melbourne; New Zealand Ecological Society; Elsevier; European Journal of Entomology; John Wiley & Sons; Natural Areas Journal; Oxford University Press, Oxford. At Springer, I thank Zuzana Bernhardt as commissioning editor for her welcome encouragement. And, as with several earlier books, Mariska van der Stigchel’s prompt and thorough practical support, coupled with her constructive advice and continuing good humour in dealing with a variety of practical problems and concerns, has greatly eased the transition toward completing this work: again, my very grateful thanks. My Project Manager, Mr Pandurangan Krishna Kumar, has again taken great care with final preparation of the book for publication.

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Contents

1 Introduction to Grasses and Grasslands������������������������������������������������    1 1.1 Introduction��������������������������������������������������������������������������������������    1 1.2 Grasses����������������������������������������������������������������������������������������������    2 1.3 Grasslands ����������������������������������������������������������������������������������������    5 1.4 Grassland Remnants��������������������������������������������������������������������������   21 References��������������������������������������������������������������������������������������������������   30 2 Australian Grasslands – Variety and Extent ����������������������������������������   37 2.1 Introduction��������������������������������������������������������������������������������������   37 2.2 Australia’s Natural Grassland Estate������������������������������������������������   37 2.3 Alien Grasses in Australia����������������������������������������������������������������   43 2.4 Economic and Ecological Importance����������������������������������������������   54 References��������������������������������������������������������������������������������������������������   55 3 Agents of Change – Management and Succession��������������������������������   59 3.1 Introduction��������������������������������������������������������������������������������������   59 3.2 Succession����������������������������������������������������������������������������������������   63 3.3 Spillover Effects��������������������������������������������������������������������������������   65 References��������������������������������������������������������������������������������������������������   68 4 Intricacies of Grassland Management for Conservation ��������������������   71 4.1 Introduction: Learning from a Global Perspective ��������������������������   71 4.2 European Calcareous Grassland ������������������������������������������������������   72 4.3 North American Prairies ������������������������������������������������������������������   77 4.4 South Africa’s Grassland Biome������������������������������������������������������   80 4.5 South American Grasslands��������������������������������������������������������������   82 4.6 New Zealand Tussock Grasslands����������������������������������������������������   84 References��������������������������������������������������������������������������������������������������   86

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5 Urban Grasslands������������������������������������������������������������������������������������   89 5.1 Introduction: The Scope of Urban Grasslands����������������������������������   89 5.2 Turfgrass ������������������������������������������������������������������������������������������   94 5.3 Green Roofs��������������������������������������������������������������������������������������   96 References��������������������������������������������������������������������������������������������������   96 6 Insects in Grasslands: The Key Groups for Understanding����������������   99 6.1 Introduction��������������������������������������������������������������������������������������   99 6.2 The Key Grassland Insect Groups����������������������������������������������������  102 6.2.1 Orthoptera ����������������������������������������������������������������������������  104 6.2.2 Hemiptera������������������������������������������������������������������������������  112 6.2.3 Coleoptera ����������������������������������������������������������������������������  122 6.2.4 Lepidoptera ��������������������������������������������������������������������������  124 6.2.5 Hymenoptera������������������������������������������������������������������������  130 6.3 Insect Communities as Grassland Indicators������������������������������������  135 References��������������������������������������������������������������������������������������������������  136 7 Flagship Insect Species in Australia’s Grasslands��������������������������������  143 7.1 Introduction: Individual Species as Flagships for Grasslands����������  143 7.2 Insect Species Conservation on Australia’s Grasslands��������������������  145 7.2.1 The Perunga Grasshopper, Perunga ochracea����������������������  146 7.2.2 The Matchstick Grasshopper, Keyacris scurra ��������������������  146 7.2.3 The Ptunarra Brown Butterfly, Oreixenica ptunarra������������  147 7.2.4 The Black Grass-Dart Butterfly, Ocybadistes knightorum ��  148 7.2.5 The Golden Sun-Moth, Synemon plana��������������������������������  148 References��������������������������������������������������������������������������������������������������  150 8 Pasture Pests ��������������������������������������������������������������������������������������������  153 8.1 Introduction��������������������������������������������������������������������������������������  153 8.2 Key Pest Taxa������������������������������������������������������������������������������������  155 8.2.1 Lepidoptera ��������������������������������������������������������������������������  155 8.2.2 Pasture Scarabs ��������������������������������������������������������������������  156 8.2.3 Orthoptera ����������������������������������������������������������������������������  159 8.3 Nutrition and Grass Quality��������������������������������������������������������������  160 8.4 Pest Management������������������������������������������������������������������������������  163 References��������������������������������������������������������������������������������������������������  164 9 Maintaining Ecological Integrity and Processes ����������������������������������  167 9.1 Introduction��������������������������������������������������������������������������������������  167 9.2 Pollination ����������������������������������������������������������������������������������������  167 9.3 Nectar Supply������������������������������������������������������������������������������������  176 References��������������������������������������������������������������������������������������������������  176

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10 Grassland Management for Insect Conservation: Grazing, Mowing, and Fire ������������������������������������������������������������������������������������  179 10.1 Introduction������������������������������������������������������������������������������������  179 10.2 Grazing��������������������������������������������������������������������������������������������  196 10.3 Mowing ������������������������������������������������������������������������������������������  214 10.4 Fire��������������������������������������������������������������������������������������������������  219 References��������������������������������������������������������������������������������������������������  227 11 Grassland Management for Insect Conservation: Restoration ����������  235 11.1 Introduction������������������������������������������������������������������������������������  235 11.2 Grassland Restoration ��������������������������������������������������������������������  238 References��������������������������������������������������������������������������������������������������  253 �������������������������������������������������������������������������������������������������������� 257 ������������������������������������������������������������������������������������������������������������������ 267

Chapter 1

Introduction to Grasses and Grasslands

1.1  Introduction A broad perspective of the major constituent vegetation (‘grasses’), the areas they occupy (‘grasslands’), and the problems they face from historical and contemporary changes and losses helps to appreciate the scope, variety and urgency of their conservation needs, wider ecological importance, and benefits to humanity. Four major categories of ‘why grasslands matter’ were recognised by White et  al. (2000) (Table 1.1). The first two of these, and the interactions between them are the major focus here. This brief entrée to the massive literature on grassland variety places the Australian needs into a wider global context, with the broad parameters of grassland management treated in several later chapters. The major consequences of human interventions into grassland include the conversion of vast areas of previously natural grasslands for other uses, and their reduction to small fragments (‘remnants’) of their former extent. Those, commonly very small, areas are now refuges for many previously more widespread biota and – although often overlooked – are the last strongholds for many localised insects and other species. Threats to grasslands are varied and widespread. Paralleled elsewhere, for tropical grasslands, Bond and Parr (2010) listed the major threats as (1) land clearing, the major conservation concern and usually more extensive than for forests; (2) afforestation; (3) natural areas undergoing switches to closed forest, often from fire suppression; and (4) natural areas suffering invasions by alien species, especially when fire has been suppressed.

© Springer Nature Switzerland AG 2019 T. R. New, Insect Conservation and Australia’s Grasslands, https://doi.org/10.1007/978-3-030-22780-7_1

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1  Introduction to Grasses and Grasslands

Table 1.1  Major roles of grasslands: why grasslands matter 1. Provide food, forage and livestock, on which humanity largely depends throughout the world. Much grassland has been degraded or changed from natural conditions to support these human needs 2. Biodiversity – numerous species of all biotic groups depend on grasslands and the resources they provide. Many of those species are ecologically specialised and occur nowhere else, and their survival depends on maintenance of grasslands and the resources needed. Both tangible and less obvious benefits to humanity occur through maintenance of ecological processes. Almost half of global designated Centres of Plant Diversity include grassland habitats 3. Carbon storage – grasslands are major absorbers of carbon dioxide (CO2) and storage aids in reducing accumulation in the atmosphere and mitigating climate change impacts. Grasslands store about 34% of the global stock of carbon in terrestrial ecosystems, mostly in the soil 4. Tourism and recreation. Major sources of local or national incomes through activities such as safaris, hunting and wider visitations for recreation. Balance may be needed between revenue income and ecosystem degradation from ‘visitor pressures’ Based on White et al. (2000)

1.2  Grasses The grasses, Poaceae (formerly ‘Gramineae’) are annual or perennial herbs – with a few, such as the bamboos, sometimes large herbs – and are almost cosmopolitan in terrestrial ecosystems. Very broadly, the family comprises ‘grasses, reeds and bamboos’ (McCusker 2002). Collectively, they exploit virtually every habitat available to vascular plants. Various ‘life forms’ occur, and these may be very characteristic so that descriptors such as ‘tussock grasslands’ occur widely. Globally, McCusker (2002) noted that the family includes more than 700 genera distributed across 12 subfamilies and more than 40 tribes. Australia’s 230 or so grass genera include representatives of 10 subfamilies and 29 tribes (Table 1.2). Native grassland communities vary greatly in composition and diversity, and cover vast areas of Australia. They are augmented by grass-dominated understorey layers in many woodland communities, and the secondary grasslands resulting from clearing of land for grazing and crop production. Floristically, ‘grasslands’ are defined by domination of Poaceae but, in accord with several wider concepts of ‘grass’ as ‘herbaceous monocotyledons with narrow leaves’ the definition can be expanded to include Cyperaceae, Restionaceae and others with similar attributes. Many of those species are restricted geographically and edaphically. Many grasses are thus localised, and Linder et al. (2002) recognised five floristic regimes as useful for assessing grass distribution patterns (Fig.  1.1). All parts of Australia, however, support diverse grasses – most States/Territories have representatives of about seven subfamilies and numbers of tribes ranging from 12 to 21. Some of those distributions, and subsequent modifications by people – including species additions to improve pastoral lands – directly reflect grass metabolism patterns. Perennial grasses manifest two different photosynthetic pathways, known respectively as ‘C3’ and ‘C4’ pathways. All species exhibit the more primitive C3 pathway, but many grass species in the tropics have the additional C4. Several

1.2 Grasses Table 1.2  List of the subfamilies and tribes of Poaceae in Australia as adopted in the Flora of Australia (McCusker 2002, where attributed as compiled by E.A. Kellogg)

3 Subfamily Pharoideae Pooideae

Tribe Phareae Nardeae Stipeae Meliceae Brachypodieae Bromeae Triticeae Avenaceae Poeae Bambusoideae Bambuseae Ehrhartoideae Oryzeae Ehrharteae Centothecoideae Centroheceae ‘Cyperochloeae’ ‘Spartochloeae’ Arundinoideae Arundineae Amphipogoneae Danthoniodeae Danthonieae Aristoideae Aristideae Incertae sedis Micraireae Eriachneae Chloridoideae Pappophoreae Triodeae Cynondonteae Panicoideae Isachneae Paniceae Neurachneae Arundineleae Andropogonaceae

c­ orrelates of the two have been proposed in Australian pastoral management – both in seeking optimal native species and selecting the most suitable alien species for introductions. Thus, ‘C3 species’ are often more abundant in shady conditions, and C4 species predominate in sunny, warmer conditions. In pastures, the combination can provide a broader spread of productivity. Their distributions in Australia are thus broadly related to climate (Hattersley 1983). All but four of the 1025 Australian grass species (833 native, 292 naturalised) surveyed by Hattersley were known to have one or other pathway, providing a solid template for assessment. Findings endorsed earlier interpretations in that (1) C3 species were most numerous where spring was cool and wet; (2) C4 species were more numerous where summers were hot and wet; (3) both increased in abundance with increasing rainfall in their ‘preferred’ temperature regimes; (4) C4 species decline with decreasing

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1  Introduction to Grasses and Grasslands

Fig. 1.1  The five major regions (‘phytochorological regions’) denoting distinctive grassland floras in Australia (from Linder et al. 2002). Political divisions of Australia, bounded by dashed lines in figure, are shown as WA Western Australia, NT Northern Territory, SA South Australia, QLD Queensland, NSW New South Wales, ACT Australian Capital Territory, VIC Victoria, TAS Tasmania

temperature and/or decreasing summer rainfall; and (5) C3 species decline with increasing temperature and/or decreasing spring rainfall. Despite gaps in knowledge, these generalisations appeared valid. As noted by MacPhail and Hill (2002), Hattersley’s conclusion that about 65% of native Australian grasses (540 of the 833 species, above) have C4 photosynthesis, is a rather higher proportion than the approximately 50% of grasses worldwide. The transition from C3 to C4 pathways apparently reflected combinations of temperature and precipitation, with C4 considered an adaptation to low atmospheric CO2 levels, and may have been important in leading to increase of C4-dominated grasslands. Interpretation, though, is complex (Sinclair 2002). If C4 pathways indeed arose in response to falling atmospheric CO2 concentrations, the converse (rising CO2 concentrations) would then reduce their competitive advantage. Increasing atmospheric CO2 concentrations would thus favour C3 plants, with availability of particular nutrients (notably nitrogen and phosphorus) also influencing their roles in pasture productivity and, more broadly, grass ecology. The two pathways occur differently in the various major taxon segregates of Australian grasses: Linder et al. (2002) summarised these as (1) C3 pathway is the only one in Bambusoideae, Pooideae, Danthonioideae; (2) both C3 and C4 occur in Arundinoideae; and (3) C4 occurs in most Panicoideae. As Australia’s tropical areas are dominated by C4 grasses, and cooler areas by C3 species, with various intermediate mixes, the pathways are relevant to interpreting distributions of the various

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5

grass groups. C4 photosynthesis has developed at least 15 times as climates cooled and became drier and open areas expanded between 32 and 3 million years ago (Mya) (Vincentini et al. 2008). Thus, Australian tropical savannas are regimes of sparse or moderate tree cover over a ground layer dominated by C4 grasses, so that grass biology is fundamental to understanding savanna ecology (Williams et  al. 2017), with influences of fire regimes and land management integrated with their phenology in informed conservation.

1.3  Grasslands Grasslands occur across a very wide range of climate regimes, in all of which they can contain numerous different plant types. Simply defining grasslands can become a very complex process, leading Blair et  al. (2014) to affirm that ‘A simple all-­ encompassing definition of grasslands is surprisingly difficult to come by’. Some ecologists opt to consider grasslands and savannas to be essentially a continuum, but arbitrary limits have been recommended to distinguish them, largely in relation to presence of woody vegetation. White et al. (2000), for example, adopted an earlier distinction used for Africa (Scholes and Hall 1996) of grasslands having 0–10%; II, >10–30%; III, >30–50%; IV, >50–70%; V, >70%

a

Table 6.6  Reasons why leafhoppers are considered an appropriate group for studying grassland insect biodiversity and conservation in Europe (after Biedermann et al. 2005) (cf. Table 6.4, p. 114) 1. Numerical abundance – often a significant proportion of above ground insect richness and biomass 2. Functionally important as consumers of plant assimilates or xylem sap contents; some of economic importance as direct pests or vectors of plant diseases 3. Taxonomically diverse, with stable nomenclature and good identification aids 4. Ecological diversity, with good background of ecological knowledge and understanding 5. Standardised and simple/cheap methods for sampling and monitoring, with appreciation of needs for different habitat components 6. Responsive to environmental disturbances and stresses

The roles of leafhoppers in studying grassland biodiversity and associated conservation needs are listed in Table 6.6, after Biedermann et al. (2005), who noted that leafhopper assemblages are influenced by several interacting factors. The arrangement shown in Fig. 6.7 has much wider application to include other insects, but leafhoppers display the influences of both physical structure and taxon composition of the vegetation, both in turn affected strongly by disturbances, management and various soil and microclimate regimes. Some leafhoppers are useful indicators of disturbance. However, even small changes in vegetation or microclimate may threaten local populations, and changes to land use and habitat condition represent the most widespread category of threat to them. A high proportion of species may be sensitive to decreased habitat extent, and resulting fragmentation.

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Fig. 6.7  Scheme of interacting habitat factors that directly (solid lines) or indirectly (broken lines) affect the numbers of species and assemblage composition of leafhoppers in grassland ecosystems. (Biedermann et al. 2005)

Long-term changes (over about 40 years) in Auchenorrhyncha of protected dry grasslands in Germany were assessed by comparing samples from the same reserved sites from 1964–1967 and 2008–2010 (Schuch et al. 2012a) to constitute a (rarely recorded over such a long interval) indication of differences in richness, species incidence, and abundance. Each series comprised samples over three years, and thorough description of sampling methods in the first accounts allowed precise repetition of the methods, using sweep-netting with a similar–sized net. It was anticipated that changes might help to assess the effectiveness of those reserves, as other dry grass areas of central Europe declined considerably over that interval. Richness was very similar: 146 species in the earlier survey and 152 in the more recent samples on the same 26 sites, and with a median site richness of 22 species. Collectively across both series, approximately 76,000 individuals were collected and identified. Abundance had declined considerably, however, from a total of 49,744 to only 14,466 individuals – the latter not including >13,000 individuals of Zyginidia scutellaris, which was absent from the earlier samples but predominated in the more recent series. Seven species declined over the interval, with some suggestion of frequency of generalist species increasing whilst specialists declined. It seemed that, although the protected status of the sites may have helped to prevent losses in richness, the changes in composition included declines of dry grassland specialist leafhoppers. The net losses in abundance, notwithstanding site preservation, might also have implications for insects feeding at higher trophic levels (Schuch et al. 2012a).

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In a parallel exercise, comparisons across nine grassland sites originally sampled in 1951 and re-sampled in 2009 (Schuch et  al. 2012b) revealed decreased abundance of Auchenorrhyncha (pool of 94 species) and Orthoptera (15 species), but increased numbers of Heteroptera (88 species). Richness of both hemipteran groups increased, but remained unchanged for Orthoptera. Human use of some of the sites changed markedly over the intervening period, and the sites  – in contrast to the above example – were not reserves. Changes in abundance indicated that homogenisation and increasing proportions of generalist species (Table  6.7) were possibly linked to decreases in plant richness as land-use changed. The finding of major differences between Auchenorrhyncha and Heteroptera, as subsets of the same order, again suggest that reliance on any single taxon for evaluating differences and changes to the whole communities may be unwise. Some of the 88 Heteroptera species increased over 1951 to 2009. These, and increasing Auchenorrhyncha species, were taxa that prefer disturbed sites, whilst intensively managed sites were dominated by widespread generalists. All Auchenorrhyncha that decreased significantly prefer low-productive habitats, vulnerable to loss through agricultural fertiliser applications. Leafhopper assemblages can thus be a useful indicator of habitat quality, and relevant to assessing grassland restoration and conservation management (Hollier et al. 2005). Assemblages respond rapidly to grazing or cutting, and also reflect the structure and composition of vegetation. Surveys across 95 sites of calcareous grassland in southern England suggested that leafhoppers responded less to site management than to site vegetation. Nevertheless, in restored prairies in Indiana (United States) leafhopper richness and density were correlated with plant richness. Restoration of tallgrass prairie to increase plant species richness should in principle lead to higher insect diversity there. Investigating this hypothesis for leafhoppers in Indiana, Rowe and Holland (2013) found that increased plant richness indeed led to three to sevenfold increases in leafhoppers and in prairie-dependent leafhopper richness on restored sites. Their restoration sites provided habitat comparable to that on remnant prairie, and supported more species than either restorations with lower plant richness or sites along prairie edges. However, composition of the leafhopper assemblages on both high and low plant richness restoration sites differed from those in prairie interiors. High richness restoration provided new habitat that enabled different leafhoppers to thrive, so that those sites increased the overall local Table 6.7  Changes in species richness, total number and medians of individuals of three insect groups, from sweep samples in 1951 and 2009, given as per plot data (Schuch et al. 2012b)

Auchenorrhyncha Heteroptera Orthoptera

1 1951 55 52 9

2009 75 63 11

2 1951 20 14 3

2009 25 19 3

3 1951 2.8 3.7 3.0

2009 3.0 3.3 3.7

4 1951 16,088 1426 335

2009 5799 1820 125

5 1951 1762 108 31

2009 710 172 5

1–5, in sequence, are total number of species, median number of species, beta species richness, total number of individuals, median number of individuals

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leafhopper diversity. In addition, because leafhoppers comprise a substantial fraction of insect biomass on plants, and in turn are exploited by many higher-level consumers, the high abundance from restoration activities may be important in re-­ establishing or reinforcing food webs (Rowe and Holland 2013). Similar outcomes were obtained, also for tallgrass prairie, by Nemec and Bragg (2008). More limited restoration, to achieve a lower level of plant richness, may not be sufficient to increase leafhopper diversity. Both the above studies infer that high diversity of insect herbivores is best explained by the diversity of the plant community, irrespective of the extent of restoration undertaken. Features of the abundance and diversity of Auchenorhyncha associated specifically with grasslands led (Wallner et  al. 2013) to devise the ‘Auchenorrhyncha Quality Index’, later discussed and somewhat modified by Primi et  al. (2016) to expand the variety of grasslands for which it is relevant. The basis of Wallner et al.’s index reflected features of Auchenorrhyncha that earlier workers had regarded as giving them useful indicator values: the variety of species includes (1) those that are habitat specific and do not tolerate degradation and (2) those that are generalists and tolerate degradation; they respond rapidly and distinctly to grazing; and after disturbance they recover very slowly compared with their host plants. Species richness in a given habitat is combined with an index that integrated features of voltinism, origin, overwintering microhabitat, wing length, habitat fidelity and host plant affinity. The taxa included in Primi et al.’s evaluation comprised 91 genera and 132 species. The inferences (notably of reduced diversity and especially richness of specialist species) supported earlier data (Nickel and Hildebrandt 2003) in Germany. Different grassland types support rather different leafhopper assemblages, related to vegetation structure and composition, but also responding regionally to management by cutting. Populations can change rapidly (Waloff 1980, 1994), and changes from unknown causes might mask values as indicators. Leafhopper assemblages on British chalk grasslands were sampled over two (non-consecutive) years on 95 sites managed as components of agrienvironment schemes, to examine the hypothesis that constant management should lead to between-year differences representing stochastic changes, and between-site differences reflect individual site suitability. Hollier et al. (2005) suggested that if the two kinds of differences were similar, leafhoppers would not be good indicators of habitat quality  – but if site quality alone was important, the case for their value becomes strong. The influence of sampling year was indeed small in this study, so that the assemblages were adjudged a good indicator of habitat quality. Sampling grassland leafhoppers generally involves use of sweep nets (with some bias toward species occurring higher on vegetation, or on taller plants) or vacuum samplers, and, especially beyond the relatively well-documented fauna of the northern hemisphere, identification of all the species in samples may be difficult.

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6.2.3  Coleoptera Ground beetles (Carabidae) are perhaps the most frequently studied group of putative indicators of habitat quality amongst all insects, with many studies examining their representation along some form of habitat gradient of land use or vegetation composition and structure (Stork 1987). Their significance as predators of crop pests has fostered investigations of their ecology, diversity and abundance in agroecosystems, and has also led to developments such as ‘beetle banks’ that can act as refuges and increase their access and movement into crop interiors from field margins, and that can have parallel importance in wider conservation studies. The value of a single sampling technique, pitfall trapping, has led to its almost universal use in carabid beetle assessments. Many carabids are generalist predators, and are thus an important component of the natural enemy complex potentially available to suppress crop pests. Reliance on pitfall traps for their evaluation (with ‘actvity-density’ patterns taken to represent diversity and relative abundance) emphasises that the physical openness of the habitat, influencing beetle movement on the ground, is likely to influence catches (Carcamo et al. 1995). Most studies, whether in croplands or more natural grasslands, have been in North America or Europe, many indicating clear assemblage differences across different forms of grassland (Eyre and Luff 1990). Carabidae have thus been investigated frequently for their values in indicating changes to European semi-natural grasslands, as a group that is both species-rich and ecologically varied  – so that assemblages in many vegetation systems have been characterised in terms of the species present and their habitat ‘preferences’ in the broader picture of sound ecological knowledge of dispersal ability, dietary habits and life cycle information. Surveys of carabids of semi-natural grasslands in north-western Spain, for example, yielded 112 species across a comparison of two kinds of grassland associated with forests. Taboada et al. (2011) compared the assemblages of (1) ‘interior’ or ‘gap’ grasslands, small and embedded within surrounding oak or beech forest and (2) ‘exterior’ grasslands, large areas connected to a variety of habitat types. One outcome anticipated was that the latter should support more species, because the areas would be reached by species from varied adjacent habitats but that could not traverse forest to reach the ‘gap’ grasslands within. Within each of the two forested landscapes, gap and exterior assemblages were distinct, and also distinct from the forest carabid assemblages. The most abundant species trapped were generalists, collected from both grassland categories. Slightly under half (47.3%) of the species from grasslands were also captured in the forests, and about 60% of open habitat species were exclusive to one or other grassland type. Both categories of the remnants thereby contribute to the high regional carabid diversity, and are important components of the managed forest landscapes. Most of the open habitat species may be conserved on small remnants, even if embedded in other vegetation types, so that management plans could usefully aim to maintain a diverse set of grassland patches to promote regional carabid diversity and allow specialised species to persist, together with measures to develop appropriate traditional farming regimes to assure long-term persistence of the grasslands.

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Comparison of carabids in mown and grazed grassland plots in Switzerland revealed rather different assemblages (Grandchamp et al. 2005). Altogether, 74 species were represented, by 9379 individuals across the total 41 plots sampled. Total richness in the treatments was about the same (62, 63 species), but mown plots yielded almost twice the number of individuals (5980) as the grazed plots (3399), largely reflecting abundance of a few common species (such as Poecilus cupreus and Pterostichus madidus) but many species showed abundance skewed to one or other treatment. Likewise, management intensity, as the number of cuts, cattle or grazing intensity, and fertiliser intensity, all influenced the carabid assemblages. The lesson that even intensively managed meadows can support high carabid beetle diversity and abundance was augmented by both regimes having distinct indicator species, with these more numerous in mown than in grazed plots and most being species characteristic of open habitats. However, in these montane grasslands, cattle grazing is traditionally low, and further intensification might have negative impacts on carabid diversity, as Grandchamp et al. suggested. As another representative beetle group, Steiner et al. (2016) focused on weevils (Curculionoidea) to detect assemblage changes across abandoned, restored and continuously grazed semi-natural grasslands in Sweden. The weevil fauna (104 species) include two red-listed taxa, but many were found in only low numbers – only five species (including Sitona spp., treated as a single species because of identification difficulties) made up more than half the >3000 individuals captured. Unexpectedly, no significant differences in richness or abundance were found across treatments. However, increased cover of woody vegetation was associated with lower richness, and such structural influences depended on grassland management. Restoration (Chap. 11) by removing woody vegetation and re-introducing grazing here enabled natural weevil assemblages to occur. Weevil assemblage composition on abandoned pasture differed from the other two regimes in being dominated by polyphagous (generalist) species, compared to higher proportions of monophagous or oligophagous taxa elsewhere. Vegetation height and connectivity of patches also led to increased similarity and richness. The relative richness and abundance of the three predominant groups of beetles on paired extensively and intensively grazed pastures in Hungary showed that habitat generalists were more affected, and affected more negatively, by reduced heterogeneity than were habitat specialists (Batary et  al. 2007). All three groups were diverse: Carabidae included 98 species (34 specialists, 64 generalists), Chrysomelidae comprised 93 species (33, 60), and Curculionoidaea contained 99 species (28, 71). Generalists might benefit more than specialists from heterogeneity in larger grasslands. Batary et al.’s study implied that grassland coverage may strongly influence the composition of beetle assemblages, with trends toward those assemblages becoming dominated by generalist species as specialists disappear from fragments and are not replaced. The limited surveys available on the beetles of remnant grassland sites in Australia (Victoria) suggest that many of the species are not common or widely distributed, and that considerable seasonal changes in species composition and abundance occur (Yen and Kobelt 2009). That survey, across 12 sites and yielding

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114 beetle morphospecies also revealed about two-thirds of the species trapped at only one site, most of them in very low numbers. This trend reflected findings from broader surveys involving more sites in the region (Yen et al. 1994, 1995).

6.2.4  Lepidoptera As in many aspects of insect conservation, the fate of butterflies as environments change is an important theme in grassland ecology. Their relevance and practical value flows from considerable awareness of their declines, and of the major causes of those losses. Whilst butterflies are by far the most popular, best documented and most frequently assessed grassland insects, some diurnal moths are also included commonly. Many European surveys, for example, include the colourful Burnet moths (Zygaenidae, sometimes designated as ‘honorary butterflies’), and some studies also include Sesiidae in the ‘diurnal category’. Many nocturnal moths also depend on near-natural grasslands for their sole or predominant resources. Many grassland Lepidoptera occur only in those biomes, are highly characteristic of grassland environments, and their responses to changes and loss have received considerable attention. Erhardt and Thomas (1991) showed that diurnal Lepidoptera can be sensitive indicators of even apparently minor changes to grassland structures, such as those occurring through abandonment (Erhardt 1985). The latter study, in central Switzerland, showed that butterfly species richness was (1) highest in grasslands that had been abandoned recently (50 larvae/m2 in winter) in Tasmania as a keystone taxon. Larvae sever stems of tussock grasses, and the fallen grasses constitute a loose mat suitable for seed germination, and a strong component in tussock regeneration and sustainability. New Zealand’s extensive tussock grasslands (p. 84) also harbour a very diverse moth fauna and, despite extensive light trap surveys, many species remain poorly known and appear to be scarce as being known from few specimens. White (2002) reported 446 tussock grassland moth species – about a quarter of the New Zealand moth fauna – and noted that many species known from early records have been seen rarely (or not at all) in recent surveys, and ‘their inclusion might no longer be appropriate to the modified fauna of today’s grasslands’. Patrick (2004) noted that few of the diverse New Zealand tussock grassland moths are known to have become extinct, although (following White 1991, 2002) substantial declines in abundance and richness have occurred in some key areas. Many of these moths seem to be resilient, and continue to flourish in habitats that have been substantially altered. However, the perhaps more tenuous status of some species is exemplified by Patrick’s comments on selected taxa. For example, Orocrambus fugivitellus

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(Crambidae) was reported to be confined to 70%) were identified as indicators of the control sites, but none for the rehabilitated sites. Two-thirds of the species were found in both treatments, 10 were unique to the control sites, and only two occurred only in the rehabilitated sites. The clear assemblage differences probably reflected a combination of factors: Jamison et al. suggested the presence of a particular dominant ant species, and differences in features such as bare ground and soil characteristics – and such a list of ‘possibilities’ helps to emphasise the difficulty of attributing faunal differences to any simple or single cause or combination of causes. A general pattern of disturbance impacts on ants is elusive, but Australian studies (Andersen 1995a) have suggested that direct disturbances, such as grazing and fire, may have little direct impact but act indirectly through changing habitat structure, competitive interactions, and food availability. This scenario was contrasted with active grassland restoration through ploughing and seeding, disturbances that in Moranz et al.’s (2013) words ‘may be so intense so as to directly reduce ant abundance’, after which some ant species may take years to recover. Those trajectories are an important basis for appreciating ants as indicators of grassland restoration. As for other endangered grassland insects, the ants of North American tallgrass prairies have assumed greater importance as restoration efforts proliferate, in part through use of seed mixes (p. 236) on former croplands, where notice of invertebrates now complements earlier focus on plants alone as the major index of restoration progress. The values of ants in this context, reviewed by Nemec (2014), include leading to better understanding of the ecological conditions on restored grasslands. Most ant species have been recorded in both native and restored prairie – but a few have been found in only one or other category. Likewise, reports on ant responses to disturbance vary. About 100 ant species occur on tallgrass prairie, and about 60 of these are common. Any single substantial prairie remnant may harbour 25–35 species (Trager 1998). Reduced fragmentation of those patches is likely to increase ant diversity and abundance, with some benefits likely from conservation efforts towards single large patches or grouped small sites sufficiently close to each other for connectivity to be effective. Moranz et al. (2013) compared ant responses to three tallgrass prairie management treatments on remnant and restored grasslands in Iowa and Missouri, namely

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(1) ‘patch burn-graze’, rotational burning and growing-season cattle grazing; (2) ‘graze and burn’, entire area burned every three years and growing-season cattle grazing; and (3) ‘burn only’, burning every three years, but no cattle grazing. The assemblages included 14 ant species (sampled by sweep-netting) and were dominated by Formica montana, which comprised about 81% of the total 5794 individuals obtained. Ants were allocated among four functional groups, and analysis suggested that functional group abundance is a better measure of disturbance than either total abundance or species richness. The four groups differed in responses to fire, grazing and restoration – but alteration of competitive interactions was believed to be the most important influence. No groups were eliminated by prescribed fires, in part probably reflecting the long history of association with plants that have co-­ evolved with fires. The dominant species (here, only F. montana) may mediate the impacts of disturbance on ant community structure, but any such responses are likely to be context-specific and vary according to the type and intensity of disturbance. The high dominance of F. montana in the prairie ant community might be maintained in part by frequent fires – but these, in turn, might keep generalist ants at low abundance levels, perhaps over many years, and continuing use of frequent fires might induce extirpations of these. Moranz et al. speculated that moderate levels of cattle grazing might enhance prairie suitability for generalist ants. A similar suggestion for open-habitat specialist ants following grassland restoration in Sweden by removing woody vegetation and introducing grazing was made by Dahms et al. (2010), in concluding that regular moderate grazing would be beneficial. In that relatively cool northern climate, the proportion of ground with exposed bare rock was also correlated with higher ant richness, possibly associated with warmer ground conditions. Nevertheless, continuing management of the restored grasslands appeared to be necessary to support the highest abundance of open-habitat ant species, whilst retention of some trees and shrubs increased overall richness by providing for forest ant species. Within the pool of 27 ant species in the Swedish survey, 11 were characteristic of open habitats and seven were forest specialists. Arid grazing land in inland Australia supports many ant taxa. From one area of South Australia that had been used almost exclusively for sheep grazing for many decades, Hoffmann and James (2011) collected 171 species from 25 genera, and examined the patterns of distribution in that area. Globally, ants show four major patterns on grazing lands (Table 6.9), and Hoffmann and James explored their representation by sampling ants along ‘grazing gradients’, as varying distances from watering points. Overall ant abundance and richness did not differ significantly with distance from water – but the abundance of 10 of the 29 most common species did change: three were each ‘increasers’ or ‘decreasers’, and the remaining four showed mixed responses influenced by soil and vegetation characteristics. Monitoring, using ants or other tools, is needed in a variety of contexts, so that the objectives of any given study must be formulated clearly. Five major contexts (Table 6.10) are each relevant, but may involve different time scales and sampling approaches. Ants are useful in most of these exercises involving grasslands, on which impacts of fire, grazing, land transformation and fragmentation are important

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Table 6.9  The four major patterns exhibited by ants in grazing lands 1. Vegetation and soil type are the primary determinants of ant community composition, and have far greater effects on ant community composition than grazing 2. Grazing induces species compositional change, but does not necessarily affect species richness or abundance 3. A species’ response to grazing is not necessarily consistent across different habitats 4. Approximately one quarter to one half of species that are sufficiently common for statistical analysis have significant responses to grazing After Hoffmann and James (2011) Table 6.10  The five contexts in which ants can provide useful information for monitoring of management in ecosystems

1. Detecting the presence of invasive species 2. Detecting trends among endangered or threatened species or other focal taxa 3. Detecting trends among keystone species 4. Evaluating land management actions or processes 5. Assessing long-term ecosystem changes After Underwood and Fisher (2006)

in Australia, together with recording the presence and spread of alien or ‘tramp’ ant species as ecosystem invaders. However, in reviewing around 60 studies, Underwood and Fisher (2006) remarked on the difficulty of drawing any general conclusions on the outcomes – and also that variations in scale, conditions, duration, and purpose of the studies often rendered objective comparisons difficult. Habitat heterogeneity has strong effects on ant diversity patterns, but comparative surveys of ant richness along an extensive (500 Km) longitudinal gradient in the Argentine pampas also highlighted the influence of temperature as a driver of that representation (Ramos et al. 2018). An important lesson for future conservation was recommendation that restoration of this highly degraded land (now used predominantly for agriculture and grazing) to more natural grassland should represent that extensive environmental gradient rather than give priority to any single site with high species richness. The ant communities of prairie and savanna in Illinois grasslands are distinct, but respond in similar ways to restoration, with age of the restoration sites leading to recognition of distinct ant assemblages (from pitfall trap surveys), as (1) prairie: sites 5 years, remnant prairie; and (2) savanna: sites 15 years old, remnant savanna (Menke et  al. 2015). In essence, ant assemblages changed along predictable trajectories with time-since-restoration, so that they reflect the success of that activity. That study also revealed several reasonably common ant species that can indicate those changes in prairie and, less clearly, in savanna. The responses of ants to grazing largely followed four key patterns, and were elucidated from Hoffmann’s (2010) exploration of studies and extrapolation from

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findings in Australia to a global scope. They help to clarify reasons for the perceived values of ants as monitors or indicators of rangeland condition. Those patterns are findings from one of the most intensively and frequently investigated insect groups in grasslands in relation to land use, and are those listed in Table 6.9. Nevertheless, the principles are a useful guide to how other insect groups might respond, and clues to themes that merit more detailed investigation to determine the variety or generality of those responses by ants when applied to other taxa. Clearly, lack of biological information on many grassland insects effectively precludes their valid use as indicators of habitat quality at present. Grazing disturbances to ants are indirect  – through changing vegetation composition and structure, competitive interactions between species, and food supply, so that many responses by ants and other insects are context-dependent and unpredictable. Wider insect diversity in grasslands may be mirrored by the parasitoid Hymenoptera, a variety of mostly very small wasps, present. Their abundance and richness were more closely related than any other group studied to overall arthropod diversity in Irish grasslands (Anderson et al. 2011). These Hymenoptera furnished ‘a simple and practicable monitoring tool for tracking changes in wider diversity in agroecosystems’. Anderson et al.’s survey spanned samples from agricultural grassland in 48 small farms. They showed that simple parasitoid abundance was the strongest predictor of arthropod taxon richness. Importantly, this correlation overcomes the needs for taxonomic expertise to identify and enumerate wasp taxa. Even if only to family level, this task can be daunting, and in Australia many small wasps can not be identified reliably beyond that broad level. ‘Taxon richness’ among such groups is thereby highly uncertain, and the practical implications from Sanderson et al.’s findings deserve further investigations in grasslands.

6.3  Insect Communities as Grassland Indicators Rather than focusing solely on particular taxonomic groups, wider insect communities can also be pursued as possible indicators of grassland status and condition. They have been investigated as components of the South African Grassland Scoring System (SAGraSS), with a number of procedural changes suggested to render the initial method more effective (Kaiser et al. 2008). A valid biomonitoring procedure for the threatened Grassland Biome (p. 80) was considered necessary as a component of monitoring biological integrity. Sweep net samples of insects from grassland sites on a military training site were compared from areas with different times since they were last burned (8 years, in sequence, sites 1–4) and separated to order level (n = 16), families (n = 80) and inferred trophic level, as the SAGraSS method employs functional feeding groups as quasi-taxonomic units. The taxa were given a weighting (equivalent to a ‘quality score’) related to the probability of their occurrence in a ‘healthy grassland’, and the sum of those values was the SAGraSS score for the site, and considered together with relative richness and diversity. Two patterns were found: (1) increased numbers of individuals, families and orders

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with >200 individuals from site 1 to site 3, and decrease at site 4; and (2) scores increased gradually from site 1 to site 4. The method was still laborious, even without detailed taxonomic resolution, and the variety of trophic habits within many insect families also necessitates more detailed examination towards species-level identification. The approach was not considered suitable for use by non-specialists, and had severe failings as a ‘rapid’ method – but Kaiser et al. were optimistic that further refinement could counter this. Concentration on Coleoptera alone, for example, might be a valid focus but the difficulty of validly sampling a variety of taxa through any single technique is also problematical. Another group selection criterion is therefore their amenability to sampling by the same method. A complementary approach is to seek ‘surrogate values’, whereby changes in one focal insect group can be taken validly as representing those in others without need for direct measurements. Thus, the proclaimed values of butterflies as grassland flagships include that trends they show may mirror those of selected other taxa. Thus, species richness of grassland butterflies and grasshoppers in ecological networks in KwaZulu Natal, South Africa, were correlated strongly (Bazelet and Samways 2012), leading to inference that the grasshopper species and guild richness represented the butterfly assemblage. Both groups could be conserved adequately in the fragmented grasslands within the networks, as long as suitable host plants were present. In Greece, grasshoppers and vascular plants showed considerable congruence, and might be viable as ‘mutual surrogates’ in humid grasslands (Kati et al. 2012). Butterflies were similar to the grasshoppers in benefiting from management that involved only occasional disturbance from mowing or grazing – with more intensive disturbances likely to be harmful to both taxa.

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Chapter 7

Flagship Insect Species in Australia’s Grasslands

7.1  I ntroduction: Individual Species as Flagships for Grasslands Grassland insect diversity is a key concern, but its conservation is aided greatly by efforts for individual selected species, whose well-being can garner widespread interest, focus wider concern (including that from groups of citizen scientists), and demonstrate the intricacies of the biology of those species. The focal species, usually those selected for priority treatment by being signalled in some way as ‘threatened’, are a major avenue toward communicating the importance and vulnerability of grassland insects, and their wider interactions. Almost by definition, these ‘flagship species’ are ecological specialists with particular limited resource needs, occurring in small and few populations, restricted in distribution, and showing signs of historical and/or current declines from a variety of threats. Some are the focus of continuing conservation campaigns, many necessarily dealing primarily with conservation of the species’ particular habitat and its wider enveloping grassland. Many studies, at least initially, may be driven by fates of individual sites, such as by remnant grasslands threatened with development, and in which optimal management measures may be highly individualistic to accommodate site features. Many conservation measures for rare insects on grasslands must necessarily focus on the single or few sites in which the focal species occurs. Programmes such as that devised for conservation of the grasshopper Stenobothrus stigmaticus at its only known site in Britain (on the Isle of Man) are highly instructive cases for assessing parallel needs elsewhere. Selman and Cherrill (2018) emphasised the breadth of consultative and scientific components needed to assure the credibility and acceptance of such a campaign, and cooperation among all interested parties in seeking the ‘best’ compromises for conservation success. In this case, these were based on formal measures for species and site protection, distribution mapping and grazing management aided by fencing (through agri-environment schemes) and weed clearing. The lessons from studies of S. stigmaticus included (1) © Springer Nature Switzerland AG 2019 T. R. New, Insect Conservation and Australia’s Grasslands, https://doi.org/10.1007/978-3-030-22780-7_7

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minimising restrictions on grazing, emphasising the habitat structure needed rather than prescriptive management techniques – in this case that harm from undergrazing was generally greater than from overgrazing, so flexibility in treatment was necessary; (2) attempt the simplest agreement/consultation route between the grazier and the conservation manager/agency; (3) if landowners agree, consider a public forum for management discussions, with wider publicity and consultations (in this case on matters such as recreational access – here involving a golf course and dog-walking); (4) consider winter cutting that avoids harm to grasshoppers but might extend management into areas where grazing is not possible; and (5) continue to explore opportunities for management such as targeted scrub control as a precursor to grazing. As in many related contexts, focus on individual species can contribute to conservation of wider diversity. ‘Umbrella effects’ of grassland flagship insect species have received more attention in North America and Europe than in Australia, and some classic examples contain valuable lessons for consideration and possible emulation. One of the pioneering, and classic, North American studies is of the Bay checkerspot butterfly (Euphydryas editha bayensis) on remnant serpentine grassland patches in California (Launer and Murphy 1994). At that time, the butterfly was the only animal species restricted to that environment that was also listed under the United States Endangered Species Act, so had considerable significance as an ‘ambassador’. The Bay checkerspot exemplifies the widespread conservation dilemma of whether efforts to conserve single species can effectively also conserve other deserving and ecologically specialised taxa on the same sites, as ‘umbrellas’ for entire communities or ecosystems, or whether such interests should, rather, be subsumed in a transformation from single species to multi-species management focus and emphasis. The dilemma was illustrated well by outcomes from Launer and Murphy’s study. Preservation of all the sites occupied by the butterfly would ensure that a high proportion of native plant species would be protected, at least in part. Conversely if only the sites with the largest butterfly populations were preserved, or if the patches regarded as having only ‘marginal value’ for E. editha were lost, that proportion dropped considerably. The study embraced 27 sites, 14 of which were occupied by the checkerspot, but no single plant species had distribution identical to that of the butterfly. Butterfly occupancy was favoured by both patch area and the number of native spring-flowering non-grass species present (Fig. 7.1). Of the 133 plant species assessed, 98 occurred on at least one site with the butterfly. Conservation efforts for E. e. bayensis were not effective for all other taxa – Launer and Murphy noted that these offer only ‘a rather “leaky” umbrella of protection …’, as an inference since echoed elsewhere for other ecologically specialised insects. In this example, values as an umbrella species may apply more effectively to locally unique and ecologically isolated ecosystems in which stronger commonality of needs of different species might occur. For grasslands, the variety of composition and conditions related to succession mitigate against persistent umbrella value of any species restricted to portions of such gradients.

7.2  Insect Species Conservation on Australia’s Grasslands

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Fig. 7.1 Patch characteristics for occupancy by the Bay checkerspot butterfly, Euphydryas editha bayensis: occupied (black, n = 14) and unoccupied (open, n = 13) patches categorised by area (ha) and number of native spring-flowering non-grass plant species. (Launer and Murphy 1994)

7.2  Insect Species Conservation on Australia’s Grasslands The examples of Australian grassland insects discussed below demonstrate some current foci and concerns. The amount of information available to support conservation plans for most is far less than that from more intensive longer surveys and study of threatened grassland species in Europe or North America. Even for those species listed under the Commonwealth’s Environment Protection and Biodiversity Conservation Act 1999, assessment of threat status has sometimes been made on relatively little information, couched in rather general terms and acknowledging the wisdom of precautionary measures for their protection. The examples demonstrate these general needs. The species noted are only slightly fewer than those grassland insects signaled as of conservation significance by their formal listing under Commonwealth or State/ Territory legislations. Those taxa are dicussed in Appendix 1, as candidates for future flagship species. Some have been listed for many years but, despite various formal obligations flowing from this recognition, many remain poorly known and practical conservation measures limited or non-existent. The general lack of logistic support for insect conservation in Australia, coupled with limited expertise and interest, seems likely to persist, notwithstanding recent call (Taylor et al. 2018) for increased interest in insect flagships selected to represent all key ecosystems in the country. Likewise, the formal listing of threatened ecological communities is possible under some of the relevant legislations. Those listed include a number of localised and biologically distinctive native grasslands, most of which have also received very little entomological attention. These are also listed in Appendix 1 and the two lists together furnish strong evidence of needs to recognise the dependence of some insects on restricted unique endemic grassland formations.

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7.2.1  The Perunga Grasshopper, Perunga ochracea This flightless grasshopper occurs only in the Australian Capital Territory (ACT) and nearby parts of New South Wales, where it is found in native grasslands or grassy woodland. The major threat to Perunga is loss and degradation of this habitat, which is under continuing pressures for development and susceptible to weed invasions and agricultural changes. In the ACT Perunga is thought to occur only on some larger grassland remnants, and populations are progressively isolated because adults cannot disperse easily. A recently revised Action Plan (Anon 2017) noted that the grasshopper occurs usually in very low densities: most survey records are of single individuals, and most other records are simply opportunistic rather than based on extensive systematic surveys. Data on population sizes are not available, and the area of occupancy (within a range of about 180 km north-south and 150 km east-­west) is likely to be low. Protection of the remaining habitat patches is the primary conservation need, and the Action Plan recommends that all sites where the grasshopper is known within the ACT should be protected from undesirable impacts, and the larger native grassland patches protected formally. Management to promote heterogeneity through mosaics of short, moderate and long grass is considered desirable. The Action Plan listed these objectives, accompanied by needs to increase fundamental understanding of the species’ ecology, promote wider connectivity between remnants, and promote wider awareness and engagement in its conservation.

7.2.2  The Matchstick Grasshopper, Keyacris scurra This representative of Australia’s flightless morabine grasshoppers, an endemic radiation of around 250 species, was once common on grasslands and grassy woodlands on the south east mainland, but has been eliminated from most of its former range. It now occurs only in a few small patches of grassland in which the foodplant persists amidst tall Themeda grasses. K. scurra feeds only on two species of Helichrysum, but cover of the Kangaroo grass is also needed (Rentz 1994); joint occurrence of these plants is a critical determinant of habitat suitability. Although the south eastern temperate grassland is a nationally protected ecological community (Anon 2016), few sites are predominantly free of disturbance in the Australian Capital Territory (ACT) or Victoria. The grasshopper has gained some notoriety as conserved largely within fenced enclaves such as country cemeteries, from which grazing was effectively excluded over historical time, and Keyacris is absent from most unprotected grazing land beyond those fences. It was confined to a few cemetery sites by the 1950s and is now extremely rare. Several previous cemetery populations have become extinct, the losses attributed to changed management, notably mowing cemeteries close to the ground. Recent observations in the ACT (Mulvaney 2012) confirmed the grasshopper’s presence at sites under current conservation management. K. scurra has recently (February 2018) been listed under Victoria’s Flora and Fauna Guarantee Act 1988, but until very recently had not been confirmed to occur in the state for 40 years.

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7.2.3  The Ptunarra Brown Butterfly, Oreixenica ptunarra This Tasmanian endemic member of Australia’s Satyrinae has several relatives associated with grasslands on the mainland, but has become a notable flagship for conservation of Poa native tussock grasslands across parts of central Tasmania. It is taxonomically complex, with two or three localised subspecies generally recognised, despite some ambiguity over their relationships and distinctiveness. O. ptunarra has declined markedly from loss of its grassland habitat, and Neyland (1993) suggested that all remaining populations – most of them then already small – are threatened by continued stock grazing. In general, O. ptunarra occurs only above about 400 m, in areas with significant Poa cover. Bell (1999) suggested that over-burning and over-grazing may have eradicated the butterfly from parts of central Tasmania, but currently occupied habitat ranges from tussock grassland to open grassy eucalypt woodland. A recovery plan (Bell 1999) was based on survey of almost all actual and potential habitats to show that the two included subspecies recognised by McQuillan and Ek (1997) occurred in about 120 populations (O. p. ptunarra) and about 30 populations (O. p. ‘north-west’) respectively. Larvae feed on a variety of Poa species, but the butterfly is absent from areas converted to pasture, and many of the populations occur on the fringes of areas that are likely to have supported far larger populations in the past. Exclusion of grazing stock led to weed and introduced grass invasions of fenced areas, and declines of butterfly numbers. Observations in the Midlands area showed that grazed tussock areas were preferred over ungrazed areas, and that weed-infested roadside tussocks were less suitable than well-grazed pasture areas. Whilst frequent fires were detrimental, absence of burning led to succession of some sites, with shrub and tree invasion. Moderate burning and grazing were thought likely to be beneficial. Few butterflies are found on heavily grazed sites, but sites with large overgrown tussocks also seem to support only low numbers. Because O. ptunarra flies only weakly, recolonising other than very close sites was considered unlikely. It is wholly protected under state legislation, and the butterfly’s conservation need had been recognised by the 1980s, and Bell (1999) noted that many of the actions proposed in an initial recovery plan (1991) had been implemented, or were continuing. O. ptunarra can withstand moderate grazing by sheep, and can use nectar from weeds (such as the alien Hypochoeris radicata) if native nectar sources are not available (McQuillan and Ek 1997). Protection of specific populations on private land, together with recommending sympathetic management regimes emphasise the importance of that private land in the butterfly’s long-term security. The advice provided for listing of O. ptunarra under the Commonwealth’s Environment Protection and Biodiversity Conservation Act 1999 (as ‘endangered’, from 2014) noted that Bell’s (1999) state recovery plan, as the origin of the approved conservation advice for the species, ‘provided sufficient direction to implement priority actions and mitigate against key threats’, so that no additional plan was then considered to be needed.

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7.2.4  The Black Grass-Dart Butterfly, Ocybadistes knightorum This small skipper has one of the most localised distributions among Australia’s butterflies, occurring only in two localities in northern coastal New South Wales, where most habitat patches are very small. Andren and Cameron (2012, 2014) examined the entire coastal distribution of the butterfly to estimate the total habitat area of 32.4 hectares, with most of the 293 patches identified being small and largely unsuitable. Many patches were considered too small or too isolated to support viable populations, or poor quality, and O. knightorum was not found on 86 of the 100 smallest patches surveyed, and never on the very smallest (200/m2 in three year-old pastures in the North Island), and sward composition changes led to long-term losses in yield and quality. Major costs flowed from reduced stock carrying capacity and renovation costs for the affected pastures. One recurring problem of assessing impacts of pest insects in pastures is simply that several key pest species may occur together, and their effects may be additive or otherwise interact (such as through competition between them), or differing relative abundances and impacts on different co-occurring plant species. Zydenbos et al. (2011) quoted earlier New Zealand examples that (1) in a pasture species trial, Black beetle (Heteronychus arator) decreased ryegrass and White-fringed weevil (Graphognathus leucocoloma) reduced clover yield; (2) stress responses of plants to root loss from subterranean consumption can improve quality of food for above-­ ground feeders such as aphids; and (3) larval feeding by Clover root weevil (the alien Sitona lepidus) can induce production of defensive chemical compounds by Red clover. A comprehensive overview of insect pests of Australian pastures and field crops (Bailey 2007) includes much diagnostic and biological information on the great variety of species involved. The most significant and widespread pests of pasture grasslands are a number of native Orthoptera, Coleoptera and Lepidoptera, mostly with close relatives that are innocuous but with which they may easily be confused, and sometimes occurring in the same areas. Many are polyphagous, and some (such as the noctuoid moths noted below) feed mainly on forbs within grassy areas. Some of those non-pest relative species may inadvertently become non-target victims of control measures applied against the pests, and could need conservation if such trends continued. Representatives of many other insect taxa are also involved, and their roles in native grasslands are not always clear. Several alien aphids (for examples, Corn aphid [Rhopalosiphum maidis], Oat aphid [R. padi], and Rose-grain aphid [Metapolophium dirhodum]) are significant pests of cereal crops and their host plant range includes a variety of pasture grass species. Likewise, the 19 species of ‘cane grubs’ whose larvae attack sugarcane in Queensland are all native species and may feed on a wide range of Poaceae. Whilst the Greyback cane grub (Dermolepida albohirtum, p. 156) is a major sugarcane pest, some of the other cane grubs are relatively minor and sporadic in incidence. Early important alien pests include the African black beetle (Heteronychus ­arator, Scarabaeidae), in which both larvae and adults cause grass mortality by

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severing the roots, and several species of weevils (Curculionidae) that feed mainly on legumes in pasture environments. The main concerns for these are from Sitona weevil (Sitona discoideus) and White-fringed weevil (Naupactus [formerly Graphognathus] leucoloma). However, the grass-feeding South American Argentine stem weevil, Listronotus bonariensis, can be a major pest of Poa-based pastures and turfgrasses in southern Australia and New Zealand.

8.2  Key Pest Taxa Some of the species noted by Wallace half a century ago are still of concern, but his account also summarised the early needs for their suppression to be based on ecological understanding.

8.2.1  Lepidoptera The native pest taxa include Hednota spp. (webworm moths, Crambidae), mostly occurring amongst grass tussocks in temperate woodland environments but with the advent of fast-growing annual grasses leading to massively increased abundance. Wheat crops in Western Australia suffered substantial damage, together with barley, and loss of pasture grasses from webworms was sometimes sufficient to reduce carrying capacity by up to 2.5 sheep/hectare, with grass removed from large areas of pasture. Early accounts (Wallace and Mahon 1963, for webworms) indicated that concerns were likely to become greater as such incidences proliferated. A number of other native Lepidoptera are notable as feeding on pasture grasses or weedy species within those associations (Table 8.1). Amongst the most important in the eastern states, species of Oncopera (Hepialidae) have subterranean larvae that feed on grasses and can become abundant in sown pastures, their abundance increasing in response to pasture management and artificial drainage. Several species of these taxonomically complex swift moths, however, are rare and have very restricted distributions (Simonsen 2018). Some introduced grasses may be especially susceptible to these moths – McQuillan et al. (2007) noted that larvae of the native hepialid Oncopera intricata regularly affected introduced grasslands of Lolium and Dactylis in Tasmania, whereas only rarely did it damage native pastures of Poa or Themeda. However, whilst O. intricata is endemic to Tasmania (and the only other congeneric species in the state, O. rufobrunnea, shared with the south east mainland), several of the other mainland species are very restricted in distribution and some are rare (Simonsen 2018) and, perhaps, need conservation in grasslands. They are paralleled in New Zealand by Porina moths (Wiseana spp., Hepialidae), whose larvae live in burrows and feed on foliage of many pasture plants, changing the composition of pasture by selective grazing and reducing its carrying capacity for stock by direct consumption of forage. Wiseana larvae do not construct tunnels for their first

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Table 8.1  Representative pest Lepidoptera in Australian pasture systems Crambidae Hednota spp., significant pests on introduced grasses in pastures, and also attack wheat and barley, in particular. As ‘webworms’, larvae can remove grass from substantial areas when abundant under outbreak conditions Herpetogramma licarsisalis (Sod webworm) Hepialidae Oncopera spp., notably O. intricata (Common corbie) and O. rufobrunnea (Winter corbie) in Tasmania O. alpina (Alpine grass moth) O. brachyphylla (Round-headed pasture webworm) O. mitocera (Flat-headed pasture webworm) Oxycanus antipoda (O. fuscomaculatus is a synonym) in Tasmania and Victoria Geometridae Ciampa arietaria (Brown pasture looper), sometimes found in large numbers, but feeding mostly on broadleaved weeds Anthelidae Pterolocera amplicornis (Grass anthelid) Noctuidae Agrotis infusa (Bogong moth, Common cutworm) A. ipsilon aneituma (Black cutworm) A. munda (Brown cutworm, Pink cutworm) Diarsia intermixta (Chevron cutworm) Leucania (or Mythimna) convecta (Common armyworm) Persectania ewingii (Southern armyworm) Spodoptera mauritia acronyctoides. (Lawn armyworm) S. exempta (Day-feeding armyworm) Oecophoridae Philobota productella (Pasture tunnel moth) Pyralidae Herpetogramma chrysotricha (Sod webworm) Sclerobia tritalis (Couchgrass webworm) Compiled from various sources, including Wallace (1970), Bailey (2007), and Beehag et al. (2016) the last two are major sources of background information on incidence, impacts and management of the taxa

6–10 weeks after hatching, but live on the soil surface. At that time, ‘mob grazing’ by sheep can cause high mortality, seemingly by a combination of larval desiccation and direct trampling by the sheep (Stewart and Archibald 1987).

8.2.2  Pasture Scarabs Other groups with major native pasture pests include Coleoptera: Scarabaeidae, with several species of Acrossidius (previously Aphodius) known as ‘pasture scarabs’. Larvae of abundant taxa such as A. tasmaniae (often cited as A. howitti, now a

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Table 8.2  The most important pasture scarab pests (Scarabaeoidea) in Australian pastures Acrossidius (formerly Aphodius) tasmaniae. Blackheaded pasture cockchafer, Tasmanian grass grub, widespread in south-eastern Australia A. pseudotasmaniae. Closely related to the above, endemic to Tasmania Adoryphorus couloni. Redheaded pasture cockchafer Anoplognathus spp. Christmas beetles Ataenius imparilis. Brown-headed cockchafer Costelyra zealandica. Grass grub chafer, New Zealand grass grub Cyclocephala signaticollis. Argentinian scarab Heteronychus arator. African black beetle Lepidiota squamulata. Squamulata canegrub Othnonius batesii. Black soil scarab Plectris aliena. Imported white grub Rhopaea magnicornis. Rhopaea canegrub R. morbillosa R. verreauxi Saulostomus villosus. Hairy pasture scarab Scitala sericans. Shining pasture scarab Sericesthis consanguinea. Wheat root scarab S. geminata. Pruinose scarab S. harti. Yellow-headed cockchafer S. micans Pasture scarab S. nigra. Small pasture scarab S. nigrolineata. Dusky pasture scarab S. ocularis Compiled from various sources, as cited for previous table

synonym) generally feed on sown legumes such as Subterranean clover (Trifolium subterraneum), where their effects lead to loss of pasture carrying capacity. Larvae of most pasture scarabs feed on grass roots, and several key species are listed in Table 8.2. Native ‘pasture scarabs’, whose larvae feed underground on the roots of grasses and other pasture plants, are among the most serious pasture pests. A few species including A. tasmaniae (above) have been studied in considerable detail, and the early account of that species by Carne (1956) was the forerunner of several later parallel surveys seeking to evaluate their roles and the factors determining their abundance. Carne noted that the combination of many Acrossidius and overgrazing can lead to severe pasture damage, whilst closed paddocks or those grazed only lightly may be able to support high scarab populations. Discussing the nature of Acrossidius damage to pastures, Carne noted four main aspects, as (1) vegetation that is otherwise available to stock is eaten or buried by the beetle larvae and if infestations are severe, the carrying capacity is reduced, possibly necessitating agistment of the stock elsewhere; (2) if pastures are denuded over periods of slow growth, soil excavated by the larvae may completely cover short grass, reducing its regeneration and inducing replacement of original pasture species by others; (3) infestations characteristically occur on rising ground where soil erosion (including

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extensive sheet erosion with repeated infestations) can be increased; and (4) stock feeding on infested pastures can ingest considerable quantities of soil, with effects on health and tainting of milk. This appraisal is sufficient to indicate the variety of concerns that may arise and which are common to other scarabs, and necessitate their suppression in pastures. Both chemical and ecologically-founded control methods have been used, with varying success, and entomopathogenic fungi (notably Metarhizium anisopliae) and nematodes have potential as biological control agents. Considerable recent attention has been paid to management of the Redheaded pasture cockchafer (Adoryphorus couloni), which has become an important native pest of pasture in the south-east. It is also a pest of turfgrass and lawns, golf courses and similar environments, but is most significant as causing substantial economic damage to perennial dairy pastures (Berg et al. 2014). Because damage symptoms are not initially visible above the ground, subterranean damage is caused before the beetle is detected during its two-year life cycle – so that management must then be drastic, largely involving removing and re-sowing the affected pasture. Costs are consequently high: Berg et al. quoted average control costs attributed to A. couloni infestations in dairy pasture in South Gippsland (Victoria) approaching Au$ 803/ hectare, and a 1993 estimate of Au$ 9000/hectare for turfgrass production. The cockchafer can damage a wide range of grasses and subterranean clover, with Perennial ryegrass (Lolium perenne) perhaps especially susceptible. Control has recently emphasised uses of parasitic nematodes (especially Steinernema spp.) and fungi (Metarhizium), with various pasture management measures also considered, with varying success. As Berg et al. (2014) concluded, control of the cockchafer is a very complex task, and much further research into its biology is needed in order to refine this – with comment that future studies in native grassland might help to explain why it has become such a serious pest in introduced and improved pastures. Economic needs for control of native scarabs attacking introduced crops are exemplified by the Greyback cane beetle (Dermolepida albohirtum) and sugarcane (Saccharum spp.), which Frew et al. (2016) reported as causing annual losses of up to Au$ 40 million to the sugarcane industry. Not all pasture scarabs cause visible damage. In northern New South Wales, about half of the dozen or so species encountered commonly do so (Davidson and Roberts 1968), in some cases reflecting a range of larval feeding habits. Various agricultural practices and other management can substantially affect the abundance of pasture scarab larvae, which are influenced greatly by quality of soil, the plants present, and nutrient levels. One striking example (from Frew et al. 2016) is that larvae of Ataeniuis spretulus on a golf course were far more abundant in turf mowed to ‘fairway height’ (1.6 cm) than in adjoining areas mown to ‘rough height’ (5.1 cm). This difference was correlated with proportions infected with a bacterial pathogen: 68% of larvae were infected in the fairway regime, compared with only 34% in the ‘rough’. The various influential factors are summarised in Fig.  8.1. Approaches such as use of pathogens and host plant endophytes to suppress pasture scarabs continue to receive attention, and it is clear that ‘changes in management

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Fig. 8.1 Agricultural practices and management factors that influence plant and soil factors that affect oviposition by adult pasture scarabs, together with larval survival and feeding behaviour. VOCs volatile cues. (Adapted from Frew et al. 2016)

can potentially have a large impact in limiting damage to crops and grassland’ (Frew et al. 2016) but, as for many other grassland pests, greater ecological understanding would vastly enhance this sentiment. As in other parts of the world, broad-scale changes such as transformation of grasslands for agricultural production and clearing of forest and woodlands for pasture have accelerated the problems of native scarabs becoming pests (Frew et al. 2016), and wider ecological impacts than those to the pasture areas alone can eventuate. Thus, clearing of woodland for fertilised and highly nutritious pasture in inland New South Wales led to greatly increased larval resources for Christmas beetles (Anoplognathus spp.), and the correspondingly larger adult populations defoliating the lowered numbers of eucalypt trees becoming a contributor to eucalypt dieback in the region (Heatwole and Lowman 1986), as well as in plantations (Carne et  al. 1974). Parallels occur elsewhere  – some of them with more direct impacts on native insects. Large-scale clearing of mallee eucalypts in Western Australia was viewed as direct destruction of habitat for local jewel beetles (Buprestidae), for example.

8.2.3  Orthoptera Several native species of Orthoptera, as attested by their common names, can cause extensive damage to pastures and other grasslands. Indeed, for North America ‘it must be accepted that … grasshoppers surpass all other arthropods in their

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destructiveness to rangelands’ (Watts et al. 1982). Those impacts are paralleled elsewhere. The Australian plague locust (Chortoicetes terminifera) and the Small plague grasshopper (Austroicetes cruciata) both undergo sporadic and highly damaging outbreaks with serious damage, notably to wheat crops. A third key pest species, the Wingless grasshopper (Phaulacridium vittatum), feeds mainly on low-growing rosette-forming plants rather than grasses, and domination of pastures by grass hampers food discovery by the nymphs (Clark 1967). Particularly in tropical northern areas of Australia, the Yellow-winged locust (Gastrimargus musicus) attacks most species of pasture grasses and legumes. Concerns over the impacts of the first three of these species over vast areas of Australia, much of it remote inland country, led to the establishment of a specialist organisation, the Australian Plague Locust Commission (APLC) dedicated to monitoring, predicting migratory paths in relation to weather systems, swarm movements and behaviour and reducing locust impacts as effectively as possible over their largely inland ranges. APLC, financed largely by contributions from the affected States and Territory, was established in 1974 and became active in 1976, since when it has led research into Australian locust ecology and management. Major continuing activities have emphasised the development and movements of locust swarms, monitoring and predicting movements and likelihood of economic damage, and developing optimal ways to control both bands of nymphs on the ground and swarming adults (p. 157), in general seeking to lower insecticide uses by replacing them with more effective and less environmentally concerning approaches. APLC operates its programmes for locust management for minimising adverse environmental impacts, as discussed by Story et  al. (2005), with a primary concern in the past being the then necessary applications of insecticides over enormous areas of arid and semi-arid ecosystems. Continuing refinement of programmes and increased levels of ‘duty of care’ and ‘due diligence’ accompany incorporation of environmental research into the organisation’s core business strategy. In addition to these highly mobile species, native crickets (Teleogryllus spp.) and Mole crickets (Gryllotalpa spp.) also cause concerns from time-to-time.

8.3  Nutrition and Grass Quality The nutritional quality and suitability of native or other pasture grasses to native insect herbivores has only rarely been investigated, in contrast to far greater information on host plant species ‘preference’ or selection. Comparison of responses of Chortoicetes terminifera fed on two host plant grasses is a rare exception. On the Mitchell grasslands of interior Australia, the locust develops after rainfall (Clissold et al. 2006), in an environment where two species of C4 grasses are predominant (Hunter 1989). The perennial Astrebla lappacea (curly Mitchell grass) and the annual Dactyloctenium radulans (button grass) rapidly produce green shoots after rain, as suitable food for the locust hatchlings. Observations that if D. radulans dies

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off before locust development is completed – so that nymphs must then feed exclusively on A. lappacea  – outbreaks are prevented (Hunter et  al. 1981) have been interpreted historically by considering the annual grass to be the higher quality nutritional resource. This interpretation was queried by Clissold et  al. to demonstrate that wider aspects of nutrient balance (of proteins and carbohydrates) and actual rates of supply of these and their relationships to the physical structure of the grasses influenced the dynamics of locust plague formation. Although the period over which D. radulans remained available after rainfall influenced the population dynamics, the mechanisms of this appeared to be more complex than previously assumed. Applications of fertilisers to improve pasture productivity and quality are ­widespread, and also have implications for native insect nutritional wellbeing: indeed, the consequences of fertiliser applications may be far greater than usually presumed. For example, whilst decline of grassland Lepidoptera species in much of Europe is correlated strongly with increasing intensification of agriculture, larval food plants may suffer from applications of nitrogen-based fertilisers (Kurze et al. 2018). Trial applications at commercial agricultural rates to food plants of six representative common Lepidoptera in Germany decreased survival of larvae fed on treated plants by at least one third, and species feeding on Poa pratensis and Rumex acetosella were affected to similar extents. This outcome was contrary to much ‘conventional wisdom’ that anticipates a higher dietary nitrogen content having positive effects on consumer performance. The rates applied by Kurze et al. apparently exceeded the physiological tolerance of the Lepidoptera investigated, for which dose-dependent responses were clear (Fig.  8.2). Differences were found between species, but overall trends suggested that range-wide declines of common Lepidoptera species might be far more strongly linked with agricultural fertilisers than supposed previously. Traits among butterflies differ strongly in relation to nitrogen availability in their preferred environments, and a large proportion of species depend on nitrogen-poor

Fig. 8.2  Survival rate of the sorrel-feeding butterfly Lycaena tityrus in relation to concentration of nitrogen fertiliser applications: treatments (1–5) are control and fertiliser applications of 30, 90, 150 and 300 Kg N ha−1/year−1. (Kurze et al. 2018)

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Fig. 8.3  Schematic representation of main features that determine the community nitrogen indicator (CNI, see text) for butterflies. (WallisDeVries and van Swaay 2017)

conditions. Appraising a derived ‘community nitrogen index’ for butterflies in the Netherlands, WallisDeVries and van Swaay (2017) found that half (28 of 56) the species depend on low-nitrogen habitats, and only eight on high-nitrogen habitats. The others occurred at moderate nitrogen levels (13 species) or were indifferent to nitrogen (seven species). These traits are related also to development rate, with low-­ nitrogen associated species expected to develop slowly. The underlying principle of this index is reflecting nitrogen availability through its differential impacts on habitat quality for ‘low-nitrogen’ and ‘high-nitrogen’ species. It follows the scheme shown in Fig. 8.3, and WallisDeVries and van Swaay noted that these categories are essentially ‘habitat generalists’ and ‘habitat specialists’ in the sense used by many butterfly ecologists. The index might be a useful tool to track community responses to deposition of nitrogen from agriculture, industry and traffic. The immediate impacts of nitrogenous fertilisers on plant productivity and the long-term effects on plant and vegetation composition pose rather different outcomes, with the latter sometimes a consequence of long-term historical fertilisation regimes (Siemann 1998) through which plant species richness has been decreased without notable loss of productivity. In Minnesota, Siemann found that short-term ‘modern’ nitrogen addition increased productivity without substantial changes to plant diversity and composition, whilst long-term ‘historical’ fertilisations lessened plant diversity. Total arthropod richness and abundance, and the abundance of each major trophic group assessed (namely, detritivores, herbivores, parasites, predators) were higher in the ‘modern’ plots with greater plant productivity. Plant diversity and composition influenced diversity of consumers by both direct resource supply and influencing the numerous interactions between herbivores and their natural enemies.

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8.4  Pest Management As in other contexts, harmonising the needs for effective pest suppression and conservation of non-target native insects in Australian pastures has increased in importance and focus, and has largely involved decreasing use of chemical insecticides by substitution of more specific and carefully applied approaches. Both chemical and ecologically-based control methods have been attempted against pasture scarabs, with varying levels of success. Entomopathogenic fungi (notably Metarhizium anisopliae) and nematodes have potential as biological control agents. To a large extent, such ‘biological pesticides’ are sought, and desired, to replace the formerly far more widespread use of chemical insecticides. Those concerns are global, and well publicised. Thus, insecticide use to control North American rangeland grasshoppers was listed as a threat to the Dakota skipper butterfly on those grasslands, and herbicide use against weeds noted as possibly eliminating native forbs and nectar sources needed by this butterfly (Environment Canada 2007). In the past, the predominant mode of locust control in Australia relied largely on chemical insecticides, applied aerially (as Ultra Low Volume sprays) or by ground sprays, with their use planned and coordinated through APLC (p. 157). Both were reasonably effective, the latter approach largely as ‘barrier strips’, in which the walking massed nymphs contacted the pesticide as they moved. Treatment of nymphal bands can help prevent the later formation of highly mobile swarms of adults (Chortoicetes terminifera can migrate hundreds of Km overnight, for example), if they are detected in time – hence the need for detailed regular monitoring across large areas of breeding habitat. The sheer size of outbreaks renders control difficult, but APLC efforts have contributed much to understanding the dynamics and movements of these grasshoppers. The major alternative developed to replace broad spectrum chemicals is a Metarhizium fungus formulation, known as ‘Green guard’, an isolate of M. anisopliae var. acrida, which has been used widely for locust control in areas where chemical use is undesirable. Such areas include environmentally sensitive areas, such as remnant vegetation or organic beef farms, as well as watercourses and dams where side-effects (such as water contamination and potential non-target impacts) are clearly significant considerations. Widely considered to lack any adverse effects on aquatic organisms, tests on Metarhizium by Milner et al. (2002) confirmed that this biopesticide was very unlikely to pose risks to them at the doses it was normally applied. Before Green guard became available, the organophosphorus compound ‘fenitrothion’ was the predominant insecticide used in APLC treatments. It is used widely against a variety of insect pests (Bunn et al. 1993), and its rapid action was especially suitable for attacking locusts on and near cropping areas, where rapid pest suppression was needed. However, its lack of specificity induced widespread environmental concerns, both for direct non-target impacts and the possible fate of insectivorous vertebrates feeding on contaminated locust corpses. Investigation of impacts on non-target invertebrates on grasslands in eastern Australia implied that these would be short-term, and that the low dosage levels would further minimise impacts on other grassland invertebrates.

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In the interior of Australia, far from crops, chemicals such as ‘fibronil’ (used from 1997) with longer persistence than fenitrothion and slower kill were preferred, and were used largely against nymphal bands. Green guard can be applied aerially (Hunter et al. 2001), and its efficacy can last on vegetation for about a week in mild weather. It is effective against both nymphal bands and adult swarms, and is most economically used when these are in the early pre-outbreak phases, perhaps then over relatively circumscribed areas. Predicting and forecasting locust outbreaks is a key aspect of APLC activity, using decision Support Systems that integrate weather and habitat conditions with migration and development of the locusts (Hunter and Deveson 2002). Wider aspects of the development of biological pesticides for Orthoptera recognise that their impacts and needs for effective control can at times override ecological considerations. However, Lomer et al. (1999) envisaged the development of integrated pest management systems that focus on surveillance and monitoring, with use of Metarhizium as the main component. In conjunction, fast-acting non-persistent pesticides could be used to protect crops from swarms, and long-persistence biological control agents contribute to lowering populations over larger and more remote areas. Management of any pasture insect pests, whether on small and well-defined areas or more broadly, has some potential to be accompanied by unanticipated and unwelcome non-target impacts on non-pest taxa. The presence of insect species that are known to be rare, restricted or of conservation concern, may lead to local ­modifications of management in order to reduce or avoid harm to them. All too often, however, such species are not signalled, or are detected (often by chance rather than from planned surveys) only after such further threat has eventuated. Seeking to minimise non-target impacts is a continuing driver of pasture pest management. Realisation that pasture areas and their surrounds can harbour insects – predators and parasitoids – that can contribute to pest suppression as beneficial in pest management programmes can aid practical appreciation of that need.

References Bailey PT (ed) (2007) Pests of field crops and pastures. Identification and control. CSIRO Publishing, Collingwood Beehag G, Kaapro J, Manners A (2016) Pest management of turfgrass for sport and recreation. CSIRO Publishing, Clayton South Berg G, Faithfull IG, Powell KS, Bruce RJ, Williams DG, Yen AL (2014) Biology and management of the redheaded pasture cockchafer Adoryphorus couloni (Burmeister) (Scarabaeidae: Dynastinae) in Australia: a review of current knowledge. Aust Entomol 53:144–158 Bunn SE, Best L, Chapman JC, Melville J, New TR (1993) Review of environmental issues arising from the Australian Plague Locust Commission’s locust control operations. Australian Plague Locust Commission, Canberra Carne PB (1956) An ecological study of the pasture scarab Aphodius howitti Hope. Aust J Zool 4:259–316 Carne PB, Greaves RTG, McInnes RS (1974) Insect damage to plantation – grown eucalypts in northern coastal New South Wales, with particular reference to Christmas beetles (Coleoptera: Scarabaeidae). Aust J Entomol 13:189–206

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Clark DP (1967) A population study of Phaulacridium vittatum Sjost. (Acrididae). Aust J Zool 15:799–872 Clissold FJ, Sanson GD, Read J (2006) The paradoxical effects of nutrient ratios and supply rates on an outbreaking insect herbivore, the Australian plague locust. J Anim Ecol 75:1000–1013 Davidson RL, Roberts RJ (1968) Species differences in scarab-pasture relationships. Bull Entomol Res 58:315–324 Environment Canada (2007) Recovery strategy for the Dakota skipper (Hesperia dacotae) in Canada, Species at Risk Act, Recovery Strategy series. Environment Canada, Ottawa Frew A, Barnett K, Nielsen UN, Riegler M, Johnson SN (2016) Belowground ecology of scarabs feeding on grass roots: current knowledge and future directions for management in Australasia. Front Plant Sci 7:321. https://doi.org/10.3389/fpls.2016.00321 Heatwole H, Lowman M (1986) Dieback. Death of an Australian landscape. Reed Books, Sydney Hunter D, Deveson T (2002) Forecasting and management of migratory pests in Australia. Entomol Sin 9:13–25 Hunter DM (1989) The response of Mitchell grass (Astrebla spp.) and button grass (Dactyloctenium radulans (R.  Br.)) to rainfall and their importance to the survival of the Australian plague locust, Chortoicetes terminifera (Walker) in the arid zone. Aust J Ecol 14:467–471 Hunter DM, McCullock L, Wright DE (1981) Lipid accumulation and migratory flignt in the Australian plague locust, Chortoicetes terminifera (Walker) (Orthoptera: Acrididae). Bull Entomol Res 71:543–546 Hunter DM, Milner RJ, Spurgin PA (2001) Aerial treatment of the Australian plague locust, Chortoicetes terminifera (Orthoptera: Acrididae) with Metarhizium anisopliae (Deuteromycotina: Hyphomycetes). Bull Entomol Res 91:93–99 Kurze S, Heinken T, Fartmann T (2018) Nitrogen enrichment in host plants increases the mortality of common Lepidoptera species. Oecologia 188:1227–1237 Lomer CJ, Bateman RP, Dent D, De Groote H, Douro-Kpindou O-K et al (1999) Development of strategies for the incorporation of biological pesticides into the integrated management of locusts and grasshoppers. Agric For Entomol 1:71–88 McIvor JG (2005) Australian grasslands. In: Suttie JM, Reynolds SG, Batella C (eds) Grasslands of the world. Food and Agriculture Organization of the United Nations, Rome, pp 343–380 McQuillan P, Ireson J, Hill L, Young C (2007) Tasmanian pasture and forage pests. Identification, biology and control. Department of Primary Industries and Water, Tasmania Milner RJ, Lim RP, Hunter DM (2002) Risks to the aquatic ecosystem from the application of Metarhizium anisopliae for locust control in Australia. Pest Manag Sci 58:718–723 Moore RM (ed) (1970) Australian grasslands. Australian National University Press, Canberra Siemann E (1998) Experimental tests of effects on plant productivity and diversity on grassland arthropod diversity. Ecology 79:2057–2070 Simonsen TJ (2018) Splendid ghost moths and their allies. A revision of Australian Abantiades, Oncopera, Aenetus, Archaeoaenetus and Zelotypia (Hepialidae). CSIRO Publishing, Clayton South Stewart KM, Archibald RD (1987) The effects of pasture management on population density and diseases of porina (Lepidoptera: Hepialidae). N Z J Exp Agric 15:375–379 Story PG, Walker PW, McRae H, Hamilton JG (2005) A case study of the Australian Plague Locust Commission and environmental due diligence: why mere legislative compliance is no longer sufficient for environmentally responsible locust control in Australia. Integr Environ Assess Manag 1:245–251 Wallace MMH (1970) Insects of grasslands. In: Moore RM (ed) Australian grasslands. Australian National University Press, Canberra, pp 361–371 Wallace MMH, Mahon JA (1963) The effect of insecticide treatment on the yield and botanical composition of sown pastures in Western Australia. Aust J Exp Agr Anim Husb 3:239–250 WallisDeVries MF, van Swaay CAM (2017) A nitrogen index to track changes in butterfly species assemblages under nitrogen deposition. Biol Conserv 212:448–453 Watts JG, Huddleston EW, Owens JC (1982) Rangeland entomology. Annu Rev Entomol 27:283–311 Zydenbos SM, Barratt BIP, Bell NL, Ferguson CM, Gerard PJ et al (2011) The impact of invertebrate pests on pasture persistence and their interrelationships with biotic and abiotic factors. Pasture Persistence Grassl Res Pract Ser 15:109–118

Chapter 9

Maintaining Ecological Integrity and Processes

9.1  Introduction Considerable attention has been given to the roles of semi-natural grasslands in sustaining and increasing the diversity of ‘beneficial insects’, notably pollinators and the natural enemies of local or regional crop pests, but with basic ecological processes often treated as secondary to more tangible economic interests. Features of both the grasslands and the local landscape are influential. Again, relevant overseas examples by far outnumber those within Australia, and illustrate many of the important parameters for more local consideration. In Ohio, United States, intensive agriculture in surrounding landscapes was a filter to composition of bee and beetle diversity in assemblages on conservation grasslands (Crist and Peters 2014). Bee species richness (within the local pool of 48 species) was influenced positively by forb cover, and predatory beetle richness (pool of 143 species) was related positively to grass cover. In that study, neither grassland age (plots sampled were aged 1–13  years) nor size (1.2–17.8 hectares) had any strong effect on these insects, other than for large patch area linking with higher abundance of small beetles. Several parallel examples, noted below, emphasise the roles of grassland structure and landscape context in promoting the wellbeing of insects with well-defined and widely-appreciated ecological functions in those systems.

9.2  Pollination Widespread declines of pollinating insects in many biomes have raised considerable concerns over consequences both for human food security and the survival of the varied ecosystems involved, as specific long-evolved and intricate interactions are lost. Grasslands are no exception, and numerous studies on the extent and causes of © Springer Nature Switzerland AG 2019 T. R. New, Insect Conservation and Australia’s Grasslands, https://doi.org/10.1007/978-3-030-22780-7_9

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pollinator declines, most notably of bees, have helped to clarify some of the wider ecology of pollinator losses. High abundance of flowers is associated with abundance of pollinators, and in part implies that grazing management of grasslands that is intended to maintain plant diversity should be beneficial. In essence, changes in grazing procedures can lead to changes in pollinator behaviour, and influence (1) their choice of habitat; (2) choice of flowers within a habitat; and (3) subsequent activities on those flowers (Sjodin 2007). All these levels can be assessed by direct observations, in Sjodin’s studies by standard timed periods on each plot and across different grazing treatments. Many grasslands naturally contain numerous bee-pollinated plants, and these can be severely reduced by intensive management, together with direct impacts on the pollinators themselves, as part of the broader syndrome of decreased site quality. Not all cases are clearcut. In a study in Hungary, management effects were not detected, and species richness and abundance of both bees and bee-pollinated plants were similar in intensively and extensively managed grasslands (Batary et  al. 2010) – possibly because management intensity was relatively low on both treatments. However, grazing intensity led to changed species composition among the plant assemblages, and richness of the plants was a good indicator of richness of the bees. Relatively simple changes to a widespread mowing regime in agricultural grasslands can enhance pollinator populations and other desirable insects such as the natural enemies of crop pests. In Switzerland, delay in mowing, or simply leaving unmown refuges had positive impacts on bee and hoverfly assemblages and populations (Meyer et al. 2017). Spillover of both groups could provide valuable services to adjacent crops. The complementarity between delayed mowing and retaining refuges implied that both practices should be implemented together for their effects to have greatest value. Land use in areas surrounding grasslands can strongly influence pollinators, sometimes in rather subtle ways such as by causing pollen limitation in the grassland plants (Clough et al. 2014). Increasing arable land use in the surrounding landscape reduced the density of plants that depended on insect pollination, so that insect-dependent grassland plants are then increasingly susceptible to increased intensity of use on adjacent land. Management to enhance pollinating insects and the wild plants they pollinate may then be especially important in arable landscapes surrounding significant semi-natural grasslands. Measures for increased representation of wildflowers on grassy buffer strips in order to benefit butterflies have been investigated in the United Kingdom, where Blake et al. (2011) compared outcomes from (1) using a selective graminicide to reduce competing grasses and (2) scarification of soil to increase germination of sown wildflower seed mixes. Combinations of these produced the greatest richness of sown flowers, with positive correlations with butterfly richness and abundance, and incorporation of such measures into farmland management may have much wider benefits for native insects. Even modest increases, using forbs and legumes, to grassland diversity can enhance richness and functional diversity of pollinators (Orford et  al. 2016). Measures that can be incorporated easily by grassland farmers include suitable seed mixes, reducing applications of phosphorus and potassium, disturbing swards, and

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spreading green hay sourced from species-rich sites. Abundance of particular host plants may induce greater variety of bees, such as diversity within both long-tongued and short-tongued taxa (Rotches-Ribalta et al. 2018). Plant community determines both hoverfly and bee communities, irrespective of grassland restoration or abandonment, and those pollinator associations are also affected by landscape connectivity. In short, diversity and accessibility of floral resources can affect bee (and other pollinator) wellbeing, with the corollary that if the insects have to forage in landscapes with very limited nutritional resources they may incur ‘nutritional deficits’ (Smith et al. 2016) with these leading to increased susceptibility to diseases. Bee ‘nutritional variables’ measured by Smith et al. to determine whether nutrition was related to grassland management, plant community characteristics, or both of these, included bee head width and wing length (sizes used as indicators of larval nutrition), lipid content (indicating adult nutrition), and abdomen mass (indicating nutrition over both stages). In conjunction, they assessed whether species abundance varied among patches because of treatment or plant community features. They concluded that there was no single ‘best’ land management strategy to increase both bee abundance and nutritional status  – but noted that high levels of grazing are likely to be detrimental to both. Direct estimates of nutritional variables, as in this study, are rare. Grazing (Chap. 10) affects bee and hoverfly abundance and richness directly (by overall supply of flowers) and also more indirectly through changes in the floral community. Lazaro et al. (2016) suggested that conservation of particular plant species pollinated by early season bees, and the bees themselves, may benefit from reduced grazing in early spring. Otherwise, modest grazing may be a useful step, in paralleling an ‘intermediate disturbance’ condition. Wider declines of pollinators and, more broadly, flower-visiting insects in biomes such as semi-natural grasslands have emphasised needs to protect, enlarge, and reduce fragmentation of such systems, with accompanying management of adjacent areas to promote plant diversity. Pollinator declines in the species-rich semi-natural grasslands of Europe are influenced by the cover of arable land in the surrounding landscape (Clough et al. 2014), this leading to reduced densities of many grassland plants that depend on bee (and other insect) pollination. High proportions of permanent grassland, conversely, led to increased relative abundance of plants dependent on bee pollination, accompanied by increased bee diversity and abundance. That study, across 239 semi-natural grassland sites, separated the influences of the enveloping landscape from those of the sites themselves, to confirm that grassland plants that depend on insects for pollination are vulnerable to increasing intensity of land use in that landscape (Clough et al. 2014). Conversely, remnant semi-natural grasslands can become valuable reservoirs for insect pollinators. In Sweden, Ockinger and Smith (2006) found that those small fragments acted as population sources for butterflies and bumblebees, from which they can spread into surrounding cultivated areas. Their trials involved comparisons of richness and diversity of these insects in (1) semi-natural grassland; (2) an uncultivated linear field margin adjacent to this; and (3) a similar field margin at least 1 Km away, and were based on the premise that the linear margins alone might not

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harbour viable populations of all pollinating insect species. Butterfly richness was greatest in the grasslands, and richness and diversity of both groups were significantly higher in the adjacent strips than the more distant ones. Semi-natural grasslands in agricultural landscapes appear to benefit these pollinator groups, and are widely considered to be important in sustaining farmland diversity, in addition to their roles in facilitating pollination services. Future restoration and conservation – even, re-creation  – of flower-rich grasslands in cultivated landscapes is likely to enhance these roles. Key grassland patches for entomophilous plant species are an important conservation focus (Herrera et al. 2017, for Argentine pampas). Abundance of nectar plants in grasslands, together with larval food plants, are the most critical resources for butterflies, with microclimate also an important influence. Landscape structure can play a relatively subordinate role to these (Kramer et  al. 2012). However, while the latter had little (if any) effect on generalist species, surrounding grassland with forest can constitute a barrier to dispersal of specialist species, with the grassland then becoming a ‘population sink’ for those species if they are unable to reproduce in the surrounding matrix or move through it to reach other suitable breeding areas. Other studies on butterflies have also found that habitat patch quality has greater importance than the composition of the surrounding matrix, but those species inhabiting more than one kind of grassland may depend on those habitats being managed differently. The Dryad butterfly (Minois dryas, Nymphalidae) occurs on both xerothermic grassland and wet meadows in Poland (Kalarus and Nowicki 2015). In assessing the butterfly’s potential for colonising habitat patches, the biggest threat to wet meadow populations was drainage, and the butterfly metapopulation may depend on preservation of the largest patches of both habitat forms. However, smaller and more fragmented patches may also have critical importance, and in some cases supported higher butterfly density. In xerothermic grassland, rotational mowing in early spring was a suggested management guideline whilst in wet meadows part or an entire patch should be mown every 2–3 years, no earlier than mid-September, although fragments with atypical floristic composition can be mown more frequently. Restoration of grasslands (Chap. 11) involves both structure (almost universally assessed by plant species variety grown or planted in relation to some reference ‘target’) and function, far more difficult to assess but vital in assuring continuity of pollination and other key processes. In this context, but relevant much more widely, Forup and Memmott (2005) proposed that the interactions between plants and pollinators on restored sites in relation to reference sites could also be a useful monitoring approach. They compared flower-visiting insects and their pollen loads across restored and ‘old’ meadows in southern England to show the intricacy and variety of interactions present, leading to interaction webs that emphasised the importance of Hymenoptera and Diptera. Restored and old meadows were overall rather similar in the processes evaluated, and in this example the restorations were thereby deemed ‘successful’. No significant differences were found between restored and reference meadows in insect or plant species richness, the proportions of flowers that insects visited, the numbers of pollen grains transported by insects, or the numbers of interaction links for each species. This study also inferred that Hymenoptera were much

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more efficient than other insect groups at transporting pollen – so that Forup and Memmott suggested a scenario that rare bee species could emerge as far more important pollinators than common fly species. Such considerations are clearly relevant to management and promoting functional integrity, but are largely themes for future development rather than embedded in current practice. The above study exemplifies the importance of ‘ecological networks’ of interactions, whose variety may be especially significant for grasslands, where a high proportion of forb species depend on insect pollination. The resilience of interactions to environmental change is likely to be reflected strongly in changes within such networks. The key insects visiting flowers and likely to be involved comprise largely members of the four largest endopterygote orders (Coleoptera, Diptera, Hymenoptera, Lepidoptera). In tallgrass prairie in Kansas (at Konza), collections from 44 flowering species included 369 mophospecies from these orders (Welti and Joern 2018) and models derived from that data showed that both bison grazing and burning changed the flower and floral visitor communities and led to greater understanding of how disturbances affected interactions, and refined management protocols. These were to ensure sufficient heterogeneity to maximise plant diversity, floral vistor diversity, and the interaction structures at all scales. Effects on pollinator networks can be complex. In KwaZulu Natal, South Africa, remnant grassland patches are continually invaded by Bramble (Rubus cuneifolius), leading to reduced areas of native flowers (Hansen et al. 2018). This trend extends to loss of specialist flower-visiting insects when bramble is present, and influences the natural networks of flower-visitors, including pollinators. In particular, Hansen et  al. noted lower abundance of butterflies in bramble-invaded areas, with clear implication that this alien invader reduced habitat quality to the level that viable butterfly populations could no longer be supported. However, in some such areas the abundance of Diptera increased, and alien honeybees strongly preferred bramble – in itself a trend likely to spread bramble and lead to reduced representation of other bee-pollinated plants. Removal of brambles was designated as the key management need to ensure robust and resilient ecosystems in these remnant grasslands. The theme of grassland restoration for pollinator conservation was explored in England for landfill sites, regarded as a significant potential reserve of semi-natural habitats, by comparisons of plant-pollinator assemblages there with those on ‘reference sites’ (Tarrant et  al. 2013). No differences were found in plant richness or abundance, and pollinator assemblages were also similar in the two treatments – but some seasonal differences were found in assemblage compositions. Reassuringly, it seemed that landfill sites were indeed being restored to levels comparable to reference sites in the East Midlands region. The most abundant pollinators on the restored sites were bumblebees (Bombus terrestris/lucorum, B. lapidarius) and various Diptera, and a number of these taxa were also abundant on reference sites. Management of restoration sites can expand the periods over which floral resources for key pollinators are available. Grassland restoration for pollinating insects has thus focused largely on providing floral resources to enhance those present or replace those lost by earlier disturbances. In arable systems, establishment of wild-flower habitats by sowing has been

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a predominant approach. Three major plant functional groups in seed mixes (grasses, legumes, non-leguminous forbs) used over a 4-year period in southern England were compared by Woodcock et al. (2014). Abundance and richness of pollinators increased with flower availability over that period, and inclusion of forbs increased the persistence and variety available: low cost seed mixes can indeed benefit pollinators by increased resources. The correlation of bee richness and density with diversity of insect-pollinated plants is widespread (Batary et al. 2010), and agrienvironment schemes might become even more effective for pollinators if that increased density of selected plants became a more general core focus, as has occurred in Switzerland (Albrecht et al. 2007). Creating uncut refuge areas on small parts of hay meadows can strongly promote the roles of bees and other pollinators in the following year (Buri et al. 2014), and has been recommended as a general means to promote these insects on farmland. In comparison with treatments such as delaying mowing to prolong flower availability (Buri et al. 2016), the approach affects hay productivity only on the small areas on which mowing is prevented, and the increased flexibility of mowing can allow for timing of haying to obtain the best quality possible. Whilst pollinators have most commonly not been a pre-defined target in grassland restoration exercises, they – and other important functional taxa – are a significant consideration. In some cases, grassland restoration has indeed led to successful restoration of pollinators. This was confirmed in native tallgrass prairie in Illinois, through a 26-year chronosequence survey in which bee richness and abundance, community composition and changes were evaluated (Griffin et  al. 2017). Restoration was by seeding with mixes of native plants. Richness and abundance of bees increased almost to the levels found in prairie remnants, within only 2–3 years after restoration and thereafter the community developed by gradually accumulating further taxa. Those surveys included 85 bee species in 23 genera, and demonstrated that pollinators may establish diverse communities rapidly and maintain these over time following restoration – but Griffin et al. noted also that the rapid response they found may have been influenced by the large and highly interconnected prairie landscape amidst which the restoration sites were embedded. Another study on restored prairies in Illinois yielded a diverse assemblage of more than 100 bee species taken by an array of trapping methods (Geroff et al. 2014); that study confirmed that multiple trapping methods should be deployed where possible to enumerate bee assemblages on prairies. And, again in Illinois, a chronosequence survey of bees (115 species, in 32 genera) confirmed their diversity in restored prairies (Tonietto et al. 2017), and that continued restoration helps to support assemblages most similar to those on remnant prairie sites. These assemblages can be supported by targeting a high diversity of flowering forbs in restoration, and adjusting seed mixtures to ensure that flowers are present throughout the bees’ actively foraging periods – in this case, from late April to early October. Ability to re-colonise easily from existing sources reflects connectivity of sites (such as by corridors; Ockinger and Smith 2008), availability of critical resources (such as nesting sites for bees), the behaviour of the species involved, and the time for movement to occur. For example, many solitary bees have low dispersal ability,

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as a trait that can limit their incidence on restored sites. Ockinger et al. (2017) found them less likely to occur on restored patches in their Swedish landscape study than were either butterflies or bumblebees. Essentially, only some of the local species pool of potential pollinators are able to reach or exploit a restored habitat  – and Ockinger et  al. demonstrated that the pollinators that did so were a non-random subset of the wider grassland species pool. Guild bias was also evident; the absence of parasitic bees from restored patches was attributed to likelihood that only small populations of potential hosts occurred there, and the bees present were largely those nesting in cavities and similar widely distributed features. Because flowers have most commonly been the primary focus for management of areas such as roadside verges (Noordijk et al. 2009), effects on insects have been overshadowed. Parts of a highway verge in the Netherlands, originally a species-­ rich hay meadow, were subjected to five levels of management, and the diversity and abundance of flowering plants and insects assessed. The management regimes provided a variety of opportunities and conditions for flower-visiting insects (Fig. 9.1) with only one treatment attracting insects of all the eight groups surveyed. Each treatment, nevertheless, attracted at least six of those groups. The treatments were (1) no management; mowing once each year in September (2) without hay removal or (3) with hay removal; and mowing twice each year, in June and

Fig. 9.1  Distribution of eight insect groups over five grassland management regimes (Noordijk et al. 2009). Management types are (O) no management, (1 M) mowing once/year in September, without hay removal, (2 M) mowing twice a year without hay removal, (1 M+) mowing once/year in September, with hay removal, (2 M+) mowing twice a year with hay removal; insect groups are Tenthredinidae (vertical hatching), Ichneumonidae (crossed diagonal hatching), Bombus spp. (coarse dots), Syrphidae (black), Diptera (open), Apidae (sparse dots), Lepidoptera (diagonal hatching), Coleoptera (horizontal hatching)

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September (4) without hay removal or (5) with hay removal. The outcomes led to suggestions for rotational management schemes that enable nectar- and pollen-­ dependent insects to persist throughout the season on verges and also enhance survival of their early stages. Rotational management on roadsides, however, can be relatively complex where (1) semi-natural grasslands need hay management twice a year in order to sustain high plant diversity, and (2) management by commercial contractors rather than by conservation organisations may lead to lack of sensitivity (Noordijk et al. 2009) through use of large vehicles and need to treat large areas each day. Grassland restoration can deliberately or fortuitously help to conserve ecological function, with restoration of native bees as key pollinators one of the more publicised outcomes of this. Studies on tallgrass prairie in Illinois examined how abundance and species richness of bees changed over a period of restoration. Over their 26-year chronosequence study, Griffin et al. (2017, above) found that bees increased rapidly from their very low levels on agricultural (pre-restoration) sites to closely approach levels on reference sites. Thereafter, species accumulated gradually, so that after 5–7 years, community composition on restored sites had converged with that on natural remnants. That comparative trial was undertaken in a mosaic landscape of remnant and restored sites, where foundation bee populations were available as sources of colonists. Similar outcomes have been reported from other studies, but most of those, as with many of the studies on grassland ant assemblages (p. 131) had not extended over more than about 5 years. Relationships between landscape elements strongly influence the movements of species among them. Thus, pollinators are threatened widely by fragmentation of grassland areas and intensified agricultural practices. In Sweden, larger semi-­natural grassland fragments are the putative refuges and sources of colonists for many insect species that are otherwise confined to small uncultivated areas such as field margins. Richness and density (individuals/unit area) of butterflies and bumblebees were significantly greater in margin strips closer to semi-natural grassland than in more distant ones (Ockinger and Smith 2007). Impacts on the two insect groups, however, were rather different: (1) for butterflies, even small fragments of semi-­ natural grassland function as population sources from which individuals disperse to other areas and augment richness and abundance present there; and (2) for bumblebees it is likely that grasslands contain higher nest density than the cultivated landscape, and density of foraging bees declines with increasing distances to the nests. Maintenance of high population density and sustaining viable populations necessitates preservation of the remaining fragments of semi-natural grassland, whilst restoration and re-creation of flower-rich grassland vegetation in intensively farmed landscapes would lead to increased pollinator abundance in nearby agricultural areas. Increased pollination efficiency is a likely outcome from agri-environment schemes that embrace this principle (Ockinger and Smith 2007). Two key pollinator groups, hoverflies (Syrphidae) and bees, were surveyed on intact and restored semi-natural grasslands in Sweden, to examine changes with time since restoration, and possible increased connectivity as restoration proceeded (Ockinger et al. 2017). Particular traits of the insects, especially those associated

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with mobility and resource use, were found to be ‘filters’ for colonisation or survival, so that the pollinator assemblages on restored grasslands were a non-random subset from the embedding species pool. Thus, only 7% of the 55 solitary bee species (as putatively poor dispersers) in that wider pool were found on restored patches, whereas more than a third of the hoverflies (total of 84 species) and bumblebees (total of 19 species) were observed there. In short, larger and more mobile species were the more likely colonists, with mobility also enabling discovery of complementary resources elsewhere. Average body size of the Syrphidae was larger in restored pasture than that of all species in the local pool. Restored pollinator communities were dominated by dispersive large hoverflies, and larger bees that can exploit existing cavities or plants as nest sites. Such trait differences between pollinators in restored grasslands and wider landscapes may have implications for conservation through interactive networks. Other studies have demonstrated that augmenting the provision of floral resources, especially at times of the year when they may be naturally rather scarce, may support enhanced pollinator populations. On old landfill and reference sites in southern England, the two categories of sites had similar pollinator complements (Tarrant et al. 2013), but the total insect assemblage was greater on restored sites in autumn and lower in spring. As a more general principle, considerations of pollinating insects (or, more broadly, ‘flower-visiting insects’) during the early stages of restoration may lead to far more successful and functional progress (Neal 1998). Permanent grasslands are vital for bees in agricultural landscapes, providing sources of wild bees for crop pollination through their provision of floral resources and nesting sites. The character of the ground, with slopes and areas of sparse vegetation, can significantly affect availability of wild bees for annual crops (Carrie et al. 2018), and crop fields surrounded by sparse nesting resources supported only low bee numbers and diversity. Management such as grazing intensity interacts with pollinators on grasslands through both direct effects and changes to the community of flowering plants. Surveys of two key families of Diptera in Greece showed that intensity of grazing by sheep and goats affected abundance (Syrphidae) or richness (Bombyliidae), both parameters peaking at intermediate grazing levels. However, impacts on bees were also seasonal, with these enhanced by grazing intensity later in the season, whilst early season grazing decreased richness but increased abundance. Moderate grazing was recommended for management, because particular early bee species – and the plant taxa they frequent  – may benefit from reduced grazing intensity in spring (Lazaro et al. 2016). The examples of pollinator incidences given above also illustrate the wealth of methodological problems that occur in attempts to evaluate floral resources in grasslands for these insects. As Szigeti et al. (2016) emphasised, no general methodology for evaluating the activities and impacts of pollinators and nectar feeders exists, and ambiguities are almost inevitable in attempting to compare inferences from different surveys, these flowing from the methods used and the precise aims of each study. Variations in blossom density in grasslands affect activity of many pollinators (Hegland and Boeke 2006) and, for many of them, may be more influential than

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plant species richness, despite the latter affording a greater diversity of resources that (in principle) may correlate with increased visitor richness. Hegland and Boeke attributed this effect to most pollinators being generalists, so that the overall amount of resources rather than their variety was most influential.

9.3  Nectar Supply The availability of floral nectar, a critical resource for many adult insects – not all of them pollinators – may be strongly reduced by intensification of agricultural activities, and remnants of native vegetation including herb-rich grassland are a critical need in such altered landscapes. The influences of changed nectar quality and quantity on insect consumers can become complex, but are difficult to evaluate in detail. Some of the subtleties and difficulties of extrapolation emerged from a comparative study of adult Meadow brown (Maniola jurtina) butterflies kept in flight cages in Belgium (Lebeau et al. 2018), where they were subject to conditions of intensively-­ managed (nectar-poor) and extensively-managed (nectar-rich) landscapes in the vicinity, reflecting very different sources of energy. Survival was higher in the nectar-­rich conditions, but female butterflies collected from intensive regimes were heavier and survived better than those from extensively-managed landscapes. The variety of the subtle, even unexpected, differences illustrates the difficulties of detecting and explaining these at levels at which they can be used in conservation planning. In most cases, they are not considered constructively in other than very general terms of resource supply. The importance of nectar supply in grassland, however, becomes a key consideration for maintenance of abundant parasitoid Hymenoptera, Coleoptera, Diptera and others that have values as ‘natural enemies’ of pest host or prey species found on adjacent crops. It is a major component of augmenting values of grasslands as refuges or reservoirs for such beneficial insects, and the wider benefits to less-­ heralded insect taxa constitute considerable conservation value. In general, maintenance and augmentation of floral resources in grasslands is significant in their management for insect conservation.

References Albrecht M, Duelli P, Muller C, Kleijn D, Schmid B (2007) The Swiss agri-environment scheme enhances pollinator diversity and plant reproductive success in nearby intensively managed farmland. J Appl Ecol 44:813–822 Batary P, Baldi A, Sarospataki M, Kohler F, Verhulst J et al (2010) Effect of conservation management on bees and insect-pollinated grassland plant communities in three European countries. Agric Ecosyst Environ 136:35–39 Blake RJ, Woodcock BA, Westbury DB, Sutton P, Potts SG (2011) New tools to boost butterfly habitat quality in existing grass buffer strips. J Insect Conserv 15:221–232

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Buri P, Humbert J-Y, Arlettaz R (2014) Promoting pollinating insects in intensive agricultural matrices: field-scale experimental manipulation of hay-meadow mowing regimes and its effects on bees. PLoS One 9(1):e85635. https://doi.org/10.1371/journal.pone.0085635 Buri P, Humbert J-Y, Stanska M, Hajdamowicz I, Tran E, Entling MH, Arlettaz R (2016) Delayed mowing promotes planthoppers, leafhoppers and spiders in extensively managed meadows. Insect Conserv Divers 9:536–545 Carrie R, Lopes M, Ouin A, Andrieu E (2018) Bee diversity in crop fields is influenced by remotely-­ sensed nesting resources in surrounding permanent grasslands. Ecol Indic 90:606–614 Clough Y, Ekroos J, Baldi A, Batary P, Bommarco R et  al (2014) Density of insect-pollinated grassland plants decreases with increasing surrounding land-use intensity. Ecol Lett 17:1168–1177 Crist TO, Peters VE (2014) Landscape and local controls of insect biodiversity in conservation grasslands: implications for the conservation of ecosystem service providers in agricultural environments. Land 3:693–718 Forup ML, Memmott J (2005) The restoration of plant-pollinator interactions in hay meadows. Restor Ecol 13:265–274 Geroff RK, Gibbs J, McCravy KW (2014) Assessing bee (Hymenoptera: Apoidea) diversity of an Illinois restored tallgrass prairie: methodology and conservation considerations. J  Insect Conserv 18:951–964 Griffin SR, Bruninga-Socolar B, Kerr MA, Gibbs J, Winfree R (2017) Wild bee community change over a 26-year chronosequence of restored tallgrass prairie. Restor Ecol 25:650–660 Hansen S, Roets F, Seymour CL, Thebault E, van Veen FJF, Pryke JS (2018) Alien plants have greater impact than habitat fragmentation on native insect flower visitation networks. Divers Distr 24:58–68 Hegland SJ, Boeke L (2006) Relationships between the density and diversity of floral resources and flower visitor activity in a temperate grassland community. Ecol Entomol 31:532–538 Herrera L, Sabatino M, Gaston A, Saura S (2017) Grassland connectivity explains entomophilous plant species assemblages in an agricultural landscape of the Pampa Region, Argentina. Austr Ecol 42:486–496 Kalarus K, Nowicki P (2015) How do landscape structure, management and habitat quality drive the colonization of habitat patches by the dryad butterfly (Lepidoptera: Satyrinae) in fragmented grassland? PLoS One 10(9):e0138557. https://doi.org/10.1371/journal.pone.0138557 Kramer B, Poniatowski D, Fartmann T (2012) Effects of landscape and habitat quality on butterfly communities in pre-alpine calcareous grasslands. Biol Conserv 152:253–261 Lazaro A, Tscheulin T, Devalez J, Nakas G, Petanidou T (2016) Effects of grazing intensity on pollinator abundance and diversity and on pollination services. Ecol Entomol 41:400–412 Lebeau J, Wesselingh R, Van Dyck H (2018) Impact of floral nectar limitation on life-history traits in a grassland butterfly relative to nectar supply in different agricultural landscapes. Agric Ecosyst Environ 251:99–106 Meyer S, Unternahrer D, Arlettaz R, Humbert J-Y, Menz MHM (2017) Promoting diverse communities of wild bees and hoverflies requires a landscape approach to managing meadows. Agric Ecosyst Environ 239:376–384 Neal PR (1998) Pollinator restoration. Trends Ecol Evol 13:132–133 Noordijk J, Delille K, Schaffers AP, Sykora KV (2009) Optimizing grassland management for flower-visiting insects in roadside verges. Biol Conserv 142:2097–2103 Ockinger E, Smith HG (2006) Landscape composition and habitat area affects butterfly species richness in semi-natural grasslands. Oecologia 149:526–534 Ockinger E, Smith HG (2007) Semi-natural grasslands as population sources for pollinating insects in agricultural landscapes. J Appl Ecol 44:50–59 Ockinger E, Smith HG (2008) Do corridors promote dispersal in grassland butterflies and other insects? Landsc Ecol 23:27–40 Ockinger E, Winsa M, Roberts SPM, Bommarco R (2017) Mobility and resource use influence the occurrence of pollinating insects in restored seminatural grassland fragments. Restor Ecol. https://doi.org/10.1111/rec.12646

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Orford KA, Murray PJ, Vaughan IP, Memmott J  (2016) Modest enhancements to conventional grassland diversity improve the provision of pollination services. J Appl Ecol 53:906–915 Rotches-Ribalta R, Winsa M, Roberts SPM, Ockinger E (2018) Associations between plant and pollinator communities under grassland restoration respond mainly to landscape connectivity. J Appl Ecol 55:2822–2833 Sjodin NE (2007) Pollinator behavioural responses to grazing intensity. Biodivers Conserv 16:2103–2121 Smith GW, Debinski DM, Scavo NA, Lange CJ, Delaney JT et  al (2016) Bee abundance and nutritional status in relation to grassland management practices in an agricultural landscape. Environ Entomol 45:338–347 Szigeti V, Korosi A, Harnos A, Nagy J, Kis J  (2016) Measuring floral resource availability for insect pollinators in temperate grasslands – a review. Ecol Entomol 41:231–240 Tarrant S, Ollerton J, Rahman ML, Tarrant J, McCollin D (2013) Grassland restoration on landfill sites in the east midlands, United Kingdom: an evaluation of floral resources and pollinating insects. Restor Ecol 21:560–568 Tonietto RK, Ascher JS, Larkin DJ (2017) Bee communities along a prairie restoration chronosequence: similar abundance and diversity, distinct composition. Ecol Appl 27:705–717 Welti EAR, Joern A (2018) Fire and grazing modulate the structure and resistance of plant-floral visitor networks in a tallgrass prairie. Oecologia 186:517–528 Woodcock BA, Savage J, Bullock JM, Nowakowski M, Orr R, Tallowin JRB, Pywell RF (2014) Enhancing floral resources for pollinators in productive agricultural grasslands. Biol Conserv 171:44–51

Chapter 10

Grassland Management for Insect Conservation: Grazing, Mowing, and Fire

10.1  Introduction ‘It is a great travesty how a world grassland type that has nurtured the needs of humans for millenia has received so little appreciation, attention and protection’ (Carbutt et al. 2017, writing on temperate grasslands). The numerous calls for grassland conservation, in many parts of the world, and the now wide acceptance that grasslands have wide ecological and practical importance to humanity, as well as for sustaining native biota, constitute one of the best prospects for grassland insect conservation. At one level, those insects are then ‘passengers’ within the wider targets of conserving grassland biomes and resources – but are also recognised progressively as predominant functional components of sustaining those ecosystems, and that wider management of grasslands may be critical in their conservation. Strategies for the conservation of grasslands, predominantly those of the temperate regions, have proliferated, and range from global perspectives (Peart 2008; Henwood 2010; Carbutt et  al. 2017) to those dealing primarily with particular regions (South Africa: Carbutt et al. 2011; South America: Alba Mejia n.d. [2015]; North America; Gauthier et al. 2003) so that a range of scales is targeted. That wider perspective of grassland conservation may benefit many little-known insect taxa, as well as the relatively few specifically attended ‘flagships’. Thus, the detailed ‘Native Grassland Conservation Strategy and Action Plans’ for the Australian Capital Territory (ACT Government 2017) includes separate Action Plans for this ‘Endangered Ecological Community’ as well as for selected key species, including the Golden sun-moth and the Perunga grasshopper. The vision statement for that ambitious plan (‘Healthy native grasslands supporting a diverse flora and fauna for now and the future’) is accompanied by considerations for all primary and derived grasslands in the Territory, with the objective to both ‘Conserve all remaining areas of Natural Temperate Grassland in the ACT that are in moderate to high ecological condition’ and ‘Retain areas of native grassland in lower ecological condition that

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serve as ecological buffers or landscape connections, or contribute significantly to threatened species conservation, or are a priority for restoration’. Formation of the World Conservation Union’s ‘Grassland Protected Areas Task Force’ (later a ‘Specialist Group’) in 1996 was focused heavily on temperate grasslands (Henwood 2010), with a major aim of increasing their representation in effectively protected areas. Much of the endeavour was reflected in a declaration (the ‘Hohhot declaration’, after the venue in China where the Task Force workshop leading to this was held). That workshop, organised through the Temperate Grassland Conservation Initiative, brought together a geographically wide range of grassland specialists. The intention of the declaration was use as a key statement advocating for the improved conservation of temperate grassland (Peart 2008). Following a series of concerns and justification, the declaration affirmed that ‘temperate indigenous grasslands are critically endangered and urgent action is required to protect and maintain the many valuable ecological services they provide to sustain human life’. The tri-nation strategy for central North American grasslands emphasises the needs for such inter-nation cooperation to achieve sufficient and well-coordinated measures for conserving biodiversity and, whilst specifically highlighting the plight of mammals and birds, Gauthier et al. (2003) also noted that ‘many other species may have disappeared without recognition’. Nevertheless, declines of biodiversity involving losses of species and/or changes in grassland structure were identified as the primary issue in all three countries involved (Canada, Mexico, United States), with grassland loss and fragmentation the key concerns. For Canada and the United States, the primary short-term needs (with the overarching purpose of promoting habitat conservation) overlap substantially with those potentially of critical importance for Australia (Table 10.1). These needs are essentially universal – but some, especially in relevance to insects, are perhaps more difficult to achieve in Australia than in North America, because of the poorer faunal documentation and high proportions of undiagnosed species present. Lessons from the management of grassland pasture pests (Chap. 8) illustrate the detailed ecological background that may be needed in individual species conservation programmes – and also the difficulties of obtaining this for localised species that occur in very low abundance: almost by definition, insect species of primary conservation concern are rarely sufficiently secure for experimental manipulations that may lead, through lack of understanding, to mortality or extirpations in the search for more informed conservation management. The general purpose and design of grassland habitat management for conserving those taxa necessarily draws on more general principles and the ‘best available’ comparative studies. Surveys of the insects of relatively pristine grasslands in southern Australia can help to provide a template against which the assemblages of changed or disturbed grasslands may be compared (Gibson and New 2007). However, some cautions are necessary – near Melbourne, for example, New (2000) found the ant assemblages of grasslands dominated by native and alien grasses to be rather similar, suggesting that changes in ant assemblages might be influenced more by changes in plant architecture (such as successions to greater herb variety) than by simple native-alien

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Table 10.1  The needs identified for biodiversity conservation within a strategy for conservation of central North American grasslands, as listed by Gauthier et al. (2003), following acceptance that the primary need is to promote habitat conservation Common secondary short-term needs in the United States and Canada  1. To achieve complete identification, understanding and representation of biodiversity  2. To identify target species, high value habitats and natural corridors for wildlife and create a joint database  3. To determine the biotic and abiotic requirements of native prairie species and communities  4. To counteract excessive removal of flora and fauna Common important and mid-term needs to address biodiversity and habitat issues identified by United States and Canada respondents  1. To restore extirpated wildlife populations  2. To effectively manage endangered and threatened species and habitats to prevent extirpations  3. To reverse declines in grassland species  4. To prevent exotic plant invasions  5. To significantly improve the promotion of habitat conservation Secondary mid-term needs important in the United States and Canada  1. To achieve complete representation of biodiversity, identification and understanding  2. To identify target species, high value habitats and natural corridors for wildlife and create a joint database  3. To counteract excessive removal of flora and fauna  4. To determine the biotic and abiotic requirements of native prairie species and communities

grass transitions. Similar inferences have been made for some Orthoptera in Europe (Hochkirch and Adorf 2007), in which some assemblages may respond more strongly to structure than to species composition of grasslands. Wherever they have been studied in any detail, the high species richness, whether of plants or insects, in semi-natural grasslands is associated with spatial and temporal heterogeneity, and conservation management necessarily seeks to preserve or re-instate that diversity. Much land management has historically been imposed largely for other purposes – notably for improving agricultural productivity – and the needs to combine this priority with management for conservation commonly involve seeking compromise (Mazalova et al. 2015). Nevertheless, grasslands support much of the non-forest biodiversity over large regions (such as semi-natural grasslands in Europe), and much of the background to enhancing their survival flows from awareness of responses to grazing, mowing, or other landscape interventions associated with agriculture. Mazalova et al. noted that relatively little is known of the functional consequences of agricultural management, and the extent to which fundamental ecological processes are eroded. Their studies on farmland in the Czech Republic recognised the importance of linear features such as uncut strips, with diversity of arthropod communities highest in strips of grassland or belts of trees. Conservation recommendations listed were: (1) the time of intervention was an important influence on some invertebrates, and the seasonal timing may be a critical consideration – highest arthropod richness was found

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in July: mowing should thus take place only from the second half of July and in two stages in order to avoid the whole area being disturbed at the same time; (2) uncut strips left until the following season may help counter the negative effects of mowing; (3) most grassland arthropods are associated with those strips, especially when they constituted boundaries, an outcome attributed to the greater variety of vegetation present; so that (4) establishment of linear landscape features can markedly increase the biological representation of agroecosystems in this grassland formation. ‘Land-use intensity’ for grasslands is a measure of impacts on biodiversity and ecosystem functions and, as Bluthgen et al. (2012) emphasised, management treatments of fertiliser applications, mowing and grazing are all variable, and not independent. Various outcomes are sought – hay yields in meadows can be increased by fertiliser use, for example. However, quantifying the interactions between the various management measures and the horde of variables within each (notably amongst intensity, timing and frequency) and of individual site characteristics remains highly incomplete, and their interactive effects uncertain (Koerner and Collins 2014). Each facet of land use presents a gradient of intensity that can be linked, albeit sometimes only tenuously, with ecological parameters and impacts. Thus, comparison of invertebrate herbivory across 146 temperate grasslands in Germany in relation to landuse intensity showed herbivory to decrease as intensity increased (Gossner et  al. 2014). Frequency of mowing and level of herbivory both decreased herbivory, but vertebrate grazing showed no effect. In essence, land managers must continually seek more cost-effective and ecologically sound ways to conserve grassland biodiversity, counter threats to the grassland and rehabilitate degraded areas to restore and increase their conservation values. That management may relate to particular threats – such as natural succession leading to losses, to a variety of anthropogenic intrusions such as advent of invasive plants or animals – and to practical conservation targets ranging from single species management to sustaining or restoring entire communities. Very broadly, the three most widespread management components, used either singly or in various considered combinations, are grazing, mowing or slashing (‘cutting’), and prescribed burning, with their relative values and applications (below) determined by features of the different grassland regimes and sites, and the purpose and constraints of the management. Traditional site history and management may also influence current practices. Thus, Valko et al. (2014) compared use of prescribed burning on grasslands in North America and Europe, and concluded that North American practice is ‘fine-­ tuned’ in terms of frequency and timing of use and integration with other methods. They suggested that those lessons could help to apply prescribed burning more effectively in European grasslands. Evaluating effects of fire on the structure and species composition was deemed crucial, including its application under different weather and local conditions. Valko et al.’s survey revealed that much European study of prescribed burning of grasslands had been based on the relatively simple approach of annual burns of the entire site studied. In contrast, North American work was more selective and integrated into wider grassland management  – in

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consequence, well-designed prescribed burning can be used against selected invasive species. European grasslands are largely composed of C3 grass species in contrast to the more fire-prone C4 grasslands of North America. Degraded grasslands, commonly with a mixture of native and alien plant species, present problems for managers and others concerned with biodiversity conservation, and their changes from management are varied, and often complex. They are also not always predictable. Optimal ways to manage and restore grassland environments for conservation continue to be developed, sometimes amid considerable controversy over the relative values of different approaches. Thus, and perhaps especially in North American tallgrass prairie systems, use of fire has become highly contentious, together with concerns over burning small remnant tracts of prairie regarded as refuges for native taxa (Panzer 1988, 2002). Recognition that burning and grazing can strongly influence the insect complement on any such habitat unit, and that both processes are to a large extent manipulable, has generated studies either advocating or opposing their uses in insect conservation on grasslands – or, commonly, suggesting some compromise or combination of these known or presumed to reduce harmful impacts and/or confer benefit. That current management necessarily builds on scenarios created by prior patterns of land use. However, the general comment that ‘Grassland species are particularly vulnerable to herbivore activity through trampling and habitat destruction by livestock’ (Eldridge and Delgado-Baquerizo 2017) is a salutary caution that grazing regimes may need to be regulated carefully. Debinski et  al. (2011) considered that those ‘land-use legacies’ can influence composition of grassland insect assemblages more than much current management does  – but that components of those legacies may differentially affect different insect groups. Their surveys in Iowa and Missouri involved comparisons of ants, butterflies and chrysomelid beetles across three treatments (burn-only, graze and burn, patch burn graze) together with pre-treatment variables of grazing history, remnant history, fire history, and pre-treatment vegetation composition and structure – recognising that vegetation present at the commencement of their surveys is the legacy of earlier land use. Even after three years of treatment implementation, influences of those legacies on insect abundance and distribution were far greater than the imposed treatments. However, those effects were not equivalent across the three insect groups, so that one taxon cannot ‘automatically’ act as a valid surrogate for others. Each responded to different components of the legacies. Debinski et al. emphasised the importance of acknowledging land-use legacies in evaluating grassland conservation and restoration. Manipulated grazing by cattle can sometimes reduce predominance of alien plants, as discussed for tallgrass prairie in Iowa and Missouri (Delany et al. 2016). Comparison of butterfly and plant community composition on grasslands dominated by either alien or native plants were assessed for their responses to two treatments: (1) patch-burn grazing, managed for heterogeneity with a third of the pasture burned each year and cattle grazing over the whole areas, and (2) graze-and-burn, where the entire pasture was burned every three years, and grazed every year. Over seven years, communities on degraded pastures of both treatments became more similar to those on control sites, and the treatments

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may help to shift butterfly assemblage composition to more closely resemble those on reference tallgrass prairie. In this example, butterfly diversity increased in the presence of alien plant dominance, whilst allowing an economic return to managers through cattle production. Past land use relates also to ‘extinction debt’, reflecting the time lag in the responses of species to habitat changes. Effects of habitat history may persist for substantial periods, with some studies implying periods of up to a century for ground beetles (Sang et  al. 2010) on calcareous grasslands in northern Europe. Species richness of habitat specialists correlated positively with both past (75 years previously) and present habitat areas, with the independent effect of past habitat area presumed to indicate the presence of extinction debt  – with those species requiring large areas most likely to exhibit this. Relationships among many other species of butterflies and Zygaenidae were not associated with habitat area. Comparative study of several insect taxa (butterflies, bumblebees, ground beetles, dung beetles) and others (plants, birds) across 31 semi-natural grasslands in Sweden (Vessby et al. 2002) revealed few correlations among richness of these different higher taxa, with different groups correlated more with grassland area and vegetation structure than with each other. Lack of patterns of species composition among different taxa also hampered any definition of cross-taxon functional groups with habitat needs in common. Management to maintain a wide variety of semi-­ natural grasslands was considered crucial to conserve the greatest variety of species on these grasslands. However, some insects are susceptible to even very low levels of management disturbance, a scenario that can become problematical for their conservation. Larvae of the European Woodland ringlet butterfly (Erebia medusa) feed on various grasses and, although it is widely distributed on nutrient-poor grasslands and woodland glades lacking intensive management, E. medusa has also declined widely from a combination of habitat loss and climate change (Stuhldreher and Fartmann 2014). The butterfly depends on low intensity or zero land use, on which the litter layer essential for overwintering larvae can persist. Conservation thereby depends on very low levels of habitat management  – Stuhldreher and Fartmann noted rotational grazing or mowing only of small areas, and that reintroduction of grazing or mowing into abandoned grasslands has the potential to drive populations to extinction. The measures that they suggested, above, prevent the succession that would inevitably render the habitat unsuitable as taller and woody vegetation encroached. Even very small patches can sustain metapopulation units of the ringlet, which depend also on sufficient connectivity of suitable habitat patches. Preservation of grassland remnants raises two important management issues. Discussed for the Dakota skipper butterfly (McCabe 1981) and for the related Ottoe skipper (Hesperia ottoe), which also has a long history of conservation interest on tallgrass prairie (Swengel and Swengel 2007, 2013), management should generally be rather cautious because the vulnerability of the target insect populations is generally unknown. The issues noted are (1) the management before deliberate conservation issues were imposed might have been favourable, but is stopped; and (2) a new management regime may or may not be as favourable as that previous management. In the case of H. ottoe, very little information was available on pre-conservation

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management impacts, with the rapid adoption of fire as a primary conservation management component and subsequent need for long-term monitoring to clarify its impacts. A variety of different management regimes for grasslands over a landscape level may be a useful strategy to increase and sustain insect diversity. For diurnal Lepidoptera in Europe, different kinds of management selectively benefitted individual taxa – but a combination of management approaches maximised their diversity (Fiedler et al. 2017) and helped harmonise the different impacts of grazing and mowing. Their study area (in Austria) allowed comparison between three major management treatments in semi-natural grassland, as (1) pastures grazed annually by cattle from April to October; (2) areas mown once in early summer and rarely with a second cut in late summer, for hay; and (3) fallow land usually mown once a year in late summer to prevent succession, with hay collection. Fiedler et al. suggested that such ‘pluralism’ in management regimes essentially provides for complementarity of resources and ‘offers a kind of insurance to maintain high species richness at the landscape level’. Many commentators have implied that maintaining diverse grassland butterfly assemblages is an important conservation target and, in addition to management of larger grassland areas, maintenance of grassland buffers and conservation headlands in wider agricultural landscapes diversifies the range of management options available (Dollar et al. 2013) through the parallel principle of providing the resource variety needed by each developmental stage of the species present. In their example, in Mississippi, the variety of tolerances to disturbance and habitat needs among the 45 butterfly species present rendered it difficult to recommend either of the two main management methods (burning or disking) for exclusive treatment of conservation grasslands. Debates over relative importance and values of major grassland management approaches are outlined in this chapter, with the uses of grazing, mowing and burning all garnering much practical attention from both traditional and potential future uses. Studies for maintaining diversity and for needs of individual rare or threatened species may not always be compatible (butterflies in Germany: Dolek and Geyer 1997), and the landscape context of the grassland may be influential. As Schneider and Fry (2001) noted for butterflies in southern Sweden, isolated grasslands in an arable landscape can be unsuitable for species with low mobility, again emphasising the importance of mosaic landscapes with high heterogeneity. As a rather different example, richness of grasshoppers in Tanzania across all land-use categories surveyed was consistently higher than that on any single land-use category – so that, again, a mosaic agricultural landscape with areas of natural savanna grassland was needed to maintain highest diversity (Kuppler et al. 2015). Earlier chapters include references to many insects for which grassland management by one or more of the several key processes – grazing, mowing and fire – has played active roles in conservation or, conversely, led to increased threat and loss. Other processes occur. Invasive alien grasses can become severe threats to native grassland insects as their more natural food plants and other needs are lost or displaced, and controlling such grasses is a widespread need in habitat management and restoration. Selective herbicides, applied carefully and with optimal timing, can

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be an effective tool (Marushia and Allen 2011), notwithstanding that non-target effects on insect behaviour, survival or development might occur. Several examples were discussed by Schultz et  al. (2016) in a study of impacts of several specific graminicides used widely for prairie restoration in north western North America on larvae of three species of Euphydryas butterflies (Nymphalidae). Many species of this genus have high conservation profiles, with effects of alien grasses on larval food plants and adult nectar sources linked strongly with their declines. The trials suggested that some graminicides led to larval mortality and others harmed native forbs and grasses. Behavioural changes, such as foraging and increased exposure to natural enemies, were also implied. It remained unclear whether such detrimental impacts might outweigh improvements in habitat quality from the use of the herbicides. More widely, management of invasive weedy plants in grasslands involves practices that have some potential to harm a range of grassland insects. As LaBar and Schultz (2012) noted, herbicides have been used most commonly in agricultural contexts, with most evaluations of their non-target impacts being for pollinators, ‘natural enemies’ (mainly predators) or decomposers, with all effects both site-­ specific and variable across different taxonomic groups. They also commented that far more information is needed on herbicide impacts on non-target species – and this remains a significant gap in conservation studies. It implies need for caution over pesticide uses in grasslands on which significant insect species occur. The effects of sethoxydim (a grass-specific herbicide used widely to control invasive grasses without harming native forbs in North America) were investigated on four prairie butterfly species in Washington, for all of which conservation necessitated management of invasive grasses (LaBar and Schultz 2012). These butterflies, the most abundant species on the sites selected for treatments, which were also those sites with low densities of invasive perennial grasses, were Puget blue (Icaricia icarioides blackmorei), Silvery blue (Glaucopsyche lygdamus) (both Lycaenidae), Ochre ringlet (Coenonympha tullia) and Wood nymph (Cercyonis pegala) (both Nymphalidae), with somewhat differing phenologies. Although butterflies passed significantly less time in sprayed plots than in unsprayed control areas, the herbicide had little impact on larval performance or on the range of flower species available to the insects and, reassuringly, sethoxydim exhibited no short-term visible impacts on native forbs. However, findings such as different butterfly ‘residence times’ related to the herbicide use might have wider effects. LaBar and Schultz noted the potential to affect oviposition by reducing the time that female butterflies were present, but factors such as host plant volatile attractants might also be influenced, and were not appraised in their trials. They went on to suggest avenues of research that could help to determine wider aspects of management involving specific herbicides (Table 10.2), and also urged the importance of early detection of important invasive weeds at stages where they might be controlled with only minimal inputs – measures such as hand-pulling and ‘spot’ herbicide applications may then be adequate for small areas, and over larger areas processes such as disking may help to eliminate the seedbanks of some problem species (Marushia and Allen 2011).

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Table 10.2  Recommendations for future research themes to aid future management of grassland butterflies in relation to herbicide use to control invasive grasses (LaBar and Schultz 2012) 1. Assess survival of immature and adult butterflies exposed to herbicides used in or near at-risk butterfly habitat 2. Conduct chemical assays of butterfly host plants to help determine the cause of reduced adult butterfly residence time 3. For ant-tended Lycaenidae, evaluate the relationship between larvae and the attending ants, and document if herbicides disrupt this relationship 4. Assess oviposition at larger scales to determine if reduced female residence time in large sprayed areas influences oviposition 5. Conduct butterfly behaviour assessments to determine whether observed changes in residence time generalise to other species and systems

Nevertheless, use of herbicides to control invasive plants within butterfly grassland habitats is frequently recommended in restoration planning. The wider impacts of sethoxydim on Pieris rapae (reducing survival by 32%) and I. i. blackmorei (21% reduction in development time from treatment to eclosure) noted by Russell and Schultz (2010) raised concerns over uses of such chemicals in areas supporting known threatened species – and butterfly larvae are among the many insects susceptible to exposure by direct contact, indirect contact with residues on plant surfaces, and ingestion of treated plants. Timing of herbicide applications in relation to likely susceptibility of key insect herbivores may be critical in decreasing possible adverse effects. Induced phenological changes, as implied for I.i. blackmorei, for species with short flight and breeding seasons may lead to longer-term population trends. Many of the world’s grasslands evolved with fire and grazing, and both are key natural processes that influence grassland composition and condition. Human interventions have led to vastly changed patterns of burning and grazing, and those disturbances have widely disrupted long-associated species and their interactions and resilience. In consequence, much grassland management in part seeks to emulate or restore those more natural disturbance regimes, and overcome the impacts of unnatural fire, grazing and, latterly, mowing. All three processes were listed as threats to the Dakota skipper butterfly in Canada, for example (p. 126, Environment Canada 2007). Defining the ‘natural regimes’ from long-term historical disturbances is commonly uncertain, but some general ‘mantra’ about the desirability of imposed management emulating natural disturbances is not uncommon in conservation protocols. However, the mechanisms leading to impacts of any management treatment of grasslands can differ across insect groups. Reductions in diversity of Orthoptera and butterflies in meadows in the Italian Alps were both related to high fertilisation and cutting frequency, but with somewhat different processes involved. Orthoptera suffered from an unsuitable sward structure, and butterflies from the creation of disturbed plant communities with lowered availability of flowering forbs and larval food plants (Marini et al. 2009a, b). Conservation measures suggested for those hay meadows were those likely to promote species-rich, short and sparse plant communities in which the needs of both the above taxa (and others) are included.

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Many more specific or limited disturbances known or suspected to benefit grassland insect species occur, as inducing changes in resources or microclimates, the latter sometimes important in considering effects of climate changes. The Palaearctic Grizzled skipper butterfly, Pyrgus malvae, prefers warm open microhabitats for oviposition. On calcareous grasslands in Germany, molehills are significant localised habitats for larvae, as supporting higher cover of the principal larval food plant (Agrimonia eupatoria, Rosaceae). The molehills (formed by the European mole, Talpa europea) comprised a more open and warmer environment, and opportunity for Agrimonia to thrive away from the more diverse and competitive surrounding sward (Streitberger and Fartmann 2013). Local mechanical soil disturbances are a common management component in promoting low-competitive plants in degraded grasslands but, in this example, the localised changes made by moles provided important habitat components for P. malvae on calcareous grasslands with sufficiently deep soils to support their activity. Many forms of direct local disturbance to grasslands occur. Military training areas, used for manouevres, vehicle testing, shooting and related activities and located on large, uninhabited open spaces can create changes ranging from minor soil compressions and flattening of vegetation to severe compaction and total or near-total loss of vegetation cover. Many such areas have – or have earlier had – considerable ecological values and biodiversity. As Kim et al. (2015) commented, precise information on those values, even on some very large areas, is limited because public access is prohibited or regulated strictly – for reasons such as presence of unexploded ordinance. Studies of the butterflies of several habitats were undertaken near a military training area in South Korea, and with which the fauna of the training area itself was compared, and within which the species (n = 112) were categorised as grassland (40), forest edge (41) or forest interior (31) species and with the overlapping species allocated to the grouping in which they were most abundant during transect surveys. Higher richness, and a larger number of red-listed species occurred in the training area (Table 10.3), in contrast to the Gwangnenung Forest, itself regarded as one of the most significant conservation areas in the country – and where butterfly communities were changed considerably after clear-­cutting (Lee and Kwon 2014). There, grasslands created by localised clearcuts helped to Table 10.3  Butterfly species richness and abundance recorded in three study sites in South Korea over four years (2008–2011, overall totals of 112 species and 3151 individuals; no. of red-listed species 16)

Total richness Total abundance Grassland species Forest edge species Forest interior species Red-listed species From Kim et al. (2015)

Numbers recorded in site Military training area 82 1956 36 33 13 13

Inje forest 62 383 24 25 13 9

Gwangneung forest 57 812 9 23 23 0

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increase butterfly diversity. Grassland species were most abundant in the military training area, implying that open habitats made by military activity may foster their wellbeing. Parallels have been reported elsewhere, with a variety of disturbance-dependent insects benefiting directly from military training operations (Warren and Buttner 2008). Two such species in Europe, both categorised as ‘endangered’, are the Blue-­ winged grasshopper (Oedipoda caerulescens) and the Northern dune tiger beetle (Cicindela hybrida), both favoured strongly by habitats made available on military training areas in Germany, where those areas represent some of the largest remaining areas of dry sandy sparse grasslands (Warren and Buttner 2008). Oedipoda requires bare ground for oviposition, but also vegetated patches for food, and optimal cover is likely to be in the range of 50–70%. C. hybrida occurs in barren or sparsely vegetated areas, and is known to use shade to help regulate its body temperature. Both species also depend on connectivity to sustain their metapopulations. The significance of military training areas is also recognised by Germany’s nomination of nearly half of these as ‘special areas of conservation’, with equivalent proportions signalled as such in some other European countries. Measures such as prohibiting motorised vehicle use on military areas, avoiding destruction of sensitive vegetation and creation of bare soil patches, benefited Speyeria idalia (p. 79) in Pennsylvania (Ferster and Vulinec 2010). Mechanical soil disturbance can sometimes lead to short-term increase in larval food plants (Viola spp.), but other key resources for the butterfly – tussock grasses (providing protective shelter for larvae and pupae) and nectar plants – are far slower to recover. Local extinctions of S. idalia have been attributed to reduced plant diversity from disturbances. Ferster and Vulinec suggested that large track vehicles may be a surrogate for historical habitat grazing by megafauna, and noted earlier references to those vehicles as ‘iron bison’. Warren and Buttner (2008) also noted that more threatened and endangered species in the United States occur on lands managed by the Department of Defense than on those managed by any other major federal land management agency. The variety and patchiness of disturbances plays a large part in these values, with all gradations from heavily disturbed to essentially undisturbed grassland patches and other vegetation. The nature of changes in North America were illustrated for the Fort Riley Military Reservation in Kansas – cited by Quist et al. (2003) as encompassing ‘over 40000 hectares in the largest remnant of tallgrass prairie in North America’. High training use was associated with increased bare ground, reduced plant cover, and changes in plant community composition. The last included reduced cover of perennial matrix-forming grasses and native species, with increases in annuals and introduced plant species. Impacts extended to water bodies, with increased sediment loads leading to changes in aquatic invertebrate faunas. In Europe, many former military training areas are being abandoned and their management transferred to civic rather than military governance and, in some cases, to development (Cizek et al. 2013). Their previously high conservation values may thus become threatened. Surveys in the Czech Republic showed that military training areas harbour half of the country’s red-listed butterfly species – 42 of

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the 118 species found are recognised as endangered, and there was also high representation of endangered plants. Even small military training areas (compared with the much larger Fort Riley, above) can support high and significant richness of these groups and, undoubtedly, others. Cizek et al.’s (2013) study showed that the characteristic endangered butterfly species found depend on barren ground, mechanical disturbances that promote heterogeneity, or are typical of neglected grasslands and transition zones – so benefiting from the successional processes that occur. In Cizek et al.’s words, these military areas ‘represent a priceless biodiversity conservation opportunity’, and provision of appropriate disturbances should continue. Some of the largest maintained grassland areas in Japan are in military training areas, which Ishii (1994) noted as valuable haunts for grassland butterflies. In particular, Fabriciana nerippe now occurs in those areas, whilst it has undergone declines and regional extinction elsewhere, due to changes to the traditional Satoyama management system. In Japan, grassland butterflies are the most threatened group of these insects, and Nakamura (2011) reported a decline of more than 80% over 40 years, largely because of changes to this traditional landscape maintenance. Needs to restore grasslands are recognised widely in the country, and pursued through the Grassland Restoration Network founded in 2007. ‘Satoyama’ comprises a secondary landscape made up of grassland, rice paddies, secondary woodlands and others. About 15% of Japan’s approximately 240 butterfly species are listed as threatened, and 10 of the 12 critically endangered species are grassland taxa. Their declines are attributed largely to degradation of Satoyama systems, and contrast markedly with the largely more secure forest taxa. Whilst reduced demands for timber have led to less forest clearance and afforestation, this has itself adversely affected grassland butterflies as forest clearings give way to succession transitions. In an example of succession, Nakamura (2011) noted the wellbeing of the skipper Pyrgus malvae, for which forest clearings remained suitable for early stages for only five years before they were lost to encroaching pine plantations. Wider abandonment of traditional Satoyama management is the major cause of grassland butterfly losses. Each of the three major treatments of grazing, mowing and burning, used singly or in combination, determines the characters of the grassland, from broad categorisation as ‘semi-natural’ or ‘improved’ to features of individual sites and their roles in harbouring biodiversity. Understanding impacts of these manipulable processes on grasslands and the species they support is central to advancing conservation practice. Each has generated an enormous literature, collectively demonstrating a diversity of outcomes in different ecosystems, ecological contexts and places, and in response to the variables inherent in use of each practice. Each process has also figured actively in insect conservation programmes and prescriptions, and in wider efforts to restore grassland environments. An underlying theme of most management treatments is to create or maintain habitat mosaics, at local or landscape level, and which can collectively support the greatest possible biodiversity. Thus, management of mosaics of grassland and shrubland in the garrigue ecosystems of Cyprus may facilitate conservation of the greatest richness of local butterflies (Ozden and Hodgson 2011). Richness was rather similar in both systems; 14 of the 16 species occurred in grassland patches, and 15 species in shrub-dominated areas. However,

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the importance of Poaceae as larval food plants was reflected in consistently greater abundance of some species in grasslands, which were needed to sustain most of the species present. Mowing and grazing lead widely to increased cover and richness of flowering plants in grasslands, as key resources for butterflies and other insects. Many such plants – and, hence, herbivores restricted to them – are restricted to early successional stages, and disappear unless management prevents this. Long-term abandonment of systems such as calcareous grassland (Ernst et al. 2017) can thus lead to loss of these open habitats and their dependent taxa. However, distinguishing the impacts of any one process (as a basis for understanding and refining remedial management) when two or three options are present, is both difficult and beneficial. For grassland butterflies, as a commonly studied example, fire and/or grazing frequently co-occur – both naturally and as prescribed grassland management. Either or both affect habitat suitability through influences on plant community composition, nectar supply, vegetation height, litter cover and character, nesting conditions for mutualist ants and, indeed, most resource needs. Most of those parameters are also amenable to evaluation through comparative studies of butterflies across different treatments in which the disturbance factors vary, or can be imposed in a reasonably controlled manner covering variety of intensity, extent and seasons. Many attempts have been made to distinguish impacts of fire and grazing on grasslands and selected constituent species, as aids to focusing management and optimising regimes that conserve heterogeneity, in large part through increasingly resembling putative natural disturbances. Butterflies of prairie pasture in the central United States (Iowa, Missouri) were sampled across three management treatments (Moranz et  al. 2012), namely (1) ‘patch-burn graze’, rotational burning of three patches within a pasture, and moderate cattle grazing; (2) ‘graze-­ and-­burn’, burning the entire pasture every three years, and moderate cattle grazing; and (3) ‘burn-only’, burning the entire pasture every three years but with no cattle grazing. Treatments affected vegetation height and the amount of bare ground, but evidence that historical treatment of the pastures – notably introductions of alien plant species – had greater influence on butterflies than did the above treatments suggested that conservation might require further measures to restore native plants, as well as restoring disturbance processes. Butterfly species richness (from a pool of 36 species) was not affected by treatment. The four most abundant species found comprised two habitat generalists (Eastern tailed-blue, Cupido comyntas; Clouded sulphur, Colias philodice) and two prairie specialists (Regal fritillary, Speyeria idalia; Common wood nymph, Cercyonis pegala). Moranz et al. hypothesised that the finding that the two specialists occurred at higher densities in the burn-only areas – a condition they described as ‘surprising’ – could be for one or more of the following reasons: (1) cattle grazing reduced habitat quality by reducing vegetation height; (2) historical variation in vegetation was more important than anticipated; (3) fire was less harmful than expected; and (4) patch-burn grazing did not generate sufficient structural heterogeneity. Nevertheless, the finding was counter to some previous studies in which prairie specialist butterfly responses were examined. Swengel (1996) found them taking 3–5 years to recover from fire.

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Whilst re-colonisation of burned sites from surrounding unburned prairie might be responsible for the raised densities, Moranz et  al. also noted discovery of ­freshly-­eclosed S. idalia on unburned patches on the burn-only sites, implying that such species may avoid extirpation through such mosaic heterogeneity. Comparison of butterflies on Iowa prairie remnants that were (1) grazed only; (2) burned only; or (3) grazed and burned, revealed butterflies to be most abundant on the last of these treatments, and lowest on sites that were only burned (Vogel et al. 2007). Again, butterfly species’ responses to any practice were highly individualistic. The total 51 species observed included 16 habitat specialists, 28 generalists, and 7 woodland species. As in Moranz et  al.’s (2012) study, two habitat specialists (S. idalia, C. pegala) were among the most abundant species. The three treatments each led to the site categories each supporting a different suite of butterfly species and, from this example, Vogel et al. suggested that multiple types of management should be used where possible to provide a wide array of microhabitats for different assemblages. The equivalent of ‘rotational management’ can be provided in nature reserves by maintaining a variety of conditions through carefully regulated grazing and/or mowing or burning. Complexities of habitat management for prairie specialist butterflies were also discussed for S. idalia by Henderson et al. (2018), who confirmed that fire may have short-term harmful impacts on butterfly abundance, but that the most imortant factor (habitat quality) was associated positively with prescribed fire. Further, although abundance was greatest with burning every 3–5 years, even annual burning was better than no burning at all. Henderson et al. concluded that ‘the key to (Regal fritillary) recovery seems to be more and better quality habitat, not necessarily less fire’. Much understanding of interactions between fire and grazing is based on small-­ scale studies in which the treatments are applied separately to areas that are considered homogeneous (Fuhlendorf et al. 2009). Grazing (and wider herbivory) and fire interact in many ways, leading Fuhlendorf et al. to suggest that they might best be viewed as a single disturbance process –‘pyric herbivory’ – that creates a shifting mosaic of disturbance conditions (equating to a range of insect habitats) across a complex landscape. Their models (Fig. 10.1) emphasised the different approaches needed to do this, and to investigate influences on biodiversity and ecosystem functions and their patterns in time and space. Fuhlendorf et  al. accepted readily the validity of the basic mechanisms and relationships from fire and grazing trials undertaken alone, but also that extending these initial steps to accomplish greater understanding needed more complex approaches, including innovative statistical analyses. Pyric herbivory, they suggested, might be an important component of more informed conservation, not least in restoring evolutionary disturbance patterns for grasslands. Grazing and fire both span anthropogenic and natural processes. Grazing by a wide range of mammal size classes, notably domestic stock, feral species such as rabbits, and by native mammal herbivores such as kangaroos, all contribute to changes in grassland conditions in Australia, and understanding impacts of manipulable grazing superimposed on natural levels of herbivory and disturbance is a key need for management of both native and sown pastures. Invertebrate grazing also

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Fig. 10.1  Conceptual models of approaches to studying fire, grazing and their interactions on grassland: (a) the conventional ‘factorial’ approach used to study separate effects of fire and grazing, and their statistical interaction; (b) a more dynamic landscape approach use to study effects of ecological interaction of grazing and fire that forms a shifting mosaic of disturbance patches across the landscape that, in turn, influences ecosystem function and biodiversity (A–F represent components that could be studied separately). (After Fuhlendorf et al. 2009)

contributes to grassland condition and floristic composition but, other than for some pest species, most native insects have evolved with the grassland ecosystems they inhabit, and are largely disregarded in management. Likewise, other environmental variables may be overlooked, or their significance minimised. Precipitation, for example, can be influential. Responses of tenebrionid beetles to cattle grazing in Colorado shortgrass steppe were more pronounced in a year with low spring-­ summer rainfall, perhaps implying that the greater heterogeneity resulting from combined grazing and precipitation may influence herbivore responses (Newbold et al. 2014). In that study, grazing significantly affected vegetation structure, plant species composition, and extent of bare ground cover. Tenebrionidae are ground-­ active detritivores, with 12 species represented in this survey. Abundance of some species increased with grazing, but richness was highest in long-term grazing exclosures. Likewise, wildfires are a natural occurrence in much of Australia, and accompany deliberate ‘control burning’, the latter sometimes undertaken to reduce fuel

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loads in expectation of reducing severity of any later wildfire. In short, interactions between grazing, mowing and burning on grasslands are complex, their impacts ­difficult to tease apart and daunting to evaluate in predicting management outcomes for conservation purposes. All, however, are used widely, even if uncritically, in conservation management. Importance of managing small grassland areas, such as grassy conservation headlands around intensively managed areas such as silage fields, as habitats used by numerous invertebrates, is easily overlooked in relation to the attention given to larger and putatively more natural sites. Grassy headlands can be important permanent refuges or seasonal retreats (such as aestivating sites) for insects, and have particular values in harbouring predatory carabid beetles and other natural enemies desirable for pest suppression in the adjacent agricultural environments (Cole et al. 2007). Essentially, ‘conservation headlands’ involve the edge of the crop, in which management is modified – for example by not using pesticides there – so that resident biota are not harmed. Modification of headland management to enhance biodiversity has received considerable attention. The responses of carabids to three cutting regimes (uncut, cut once a year, cut three times a year) on grassy headlands in Scotland reflected changes in activity/density of individual species rather than presence/absence changes (Haysom et al. 2004). After three years, the diversity of carabid assemblages increased on plots given the first two treatments, so benefitting from the lower intensity management regimes. Headlands exemplify the wider variety of ‘grassy strips’ promoted in agricultural landscapes as refuges and reservoirs for natural enemies and pollinators, and studies of ‘beetle banks’ and similar constructions based on grassland can contribute to the wider wellbeing of grassland insects. Influences vary considerably. In Germany, the bees and wasps occurring in narrow (3 m wide) grassy strips connected to larger areas of calcareous grassland or separated from them by varying distances (100, 300, 750 m) were surveyed by pan traps and trap nests (Krewenka et al. 2011). Isolation of these strips from larger grassland patches negatively affected both these groups and their parasitoids  – so that habitat isolation and increasing distance from likely source areas was associated with lowered abundance and richness. The larger areas were in particular need of protection for these insect groups to be adequately conserved. Planting native prairie grasses alongside potato crops in Wisconsin led to increased predation of Colorado potato beetle (Leptinotarsa decemlineata, Chrysomelidae) from the increased populations of spiders and harvestmen, but the limited mobility of these predators largely restricted those increased impacts to near the grassland (Werling et al. 2012). Any form of non-crop vegetation may increase local heterogeneity and help to sustain populations of insects of parallel applied values to agriculture. Grass margins thus have potentially wide roles in promoting conservation biological control, perhaps especially by enhancing resources needed by generalist arthropod predators capable of moving into nearby crops. Those predators may depend on presence of areas of semi-natural or other grassland in order to persist in the wider agricultural environment. In surveys of maize fields and their margins in South Africa, for example, Botha et al. (2018) found that predator species richness was far lower in the fields and that many species occurred almost exclusively in

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marginal areas. Those margins thus had clear values as sources for the natural enemies of maize pests. There is considerable scope for establishing grassy field margins and ­manipulating them to ‘adjust’ sward condition and composition to benefit many grassland insects whilst allowing adjacent grassland areas to be managed conventionally. The commonly cited state of ‘sward architectural complexity’ reflects the structures and seasonal differences between plant species, and is influenced heavily by management, with changes in richness and abundance of insects associated with that flora. Some species of Auchenorrhyncha, for example, benefit from simple short swards with little structural complexity, whilst many others are disadvantaged by cutting and/or grazing (Morris 1981a, b; Morris and Lakhani 1979). Blake et al. (2011) compared the Auchenorrhyncha along seven stages of a management intensification gradient in southern England, showing that abundance and richness increased greatly as management intensity decreased, by progressively reducing cutting frequency and/ or intensity, stopping fertiliser applications, and stopping late summer grazing. In this study, cutting and grazing both had harmful effects on Auchenorrhyncha, with increased cutting frequency reducing their abundance. Parallel trends have been reported for beetles (Woodcock et al. 2007). Rather than the more usual approach of focussing on Carabidae alone, Woodcock et  al. surveyed a far wider taxonomic range of beetle families, comprising a total of 225 species. One implication of the reduced management of field margins was change in the successional pattern of the beetle assemblages to include a greater proportion of seed- or flower-feeding taxa, following from the increases in abundance of plant structure and of sward complexity. Sown wildflowers on arable field margins can markedly increase abundance of many nectar- and pollen–feeding insects (Meek et al. 2002), and the type of margin produced strongly influences the nature of that biodiversity. Initial dominance by relatively vagile generalist insect species may eventually give way to more stable assemblages that include longer-lived habitat specialists. A wider context that is analogous to such field margins was explored for butterflies in a prairie grassland context in Nebraska and Iowa by Farhat et al. (2014), in which grassland specialists (13 species) and generalists (34 species) were compared in marginal grasslands within an agricultural matrix and grasslands managed for conservation of prairie species. Specialist species, including some of conservation concern, occurred in those margins associated with agricultural fields, and moves to increase the conservation values of those margins in anthropogenic landscapes could be beneficial for specialist butterflies. Grassland specialist species were more common on conservation grasslands, and generalist species on the marginal grasslands, and restoration of the latter could increase resources for the specialist species. In a parallel context, restoration of prairie vegetation along roadsides also gave positive responses by grassland butterflies (Ries et al. 2001). Indeed, management of roadside vegetation has a ‘profound effect on butterfly communities’, but weedy and prairie roadsides harboured more species than grassy roadsides, so that increasing diversity of forbs was a key issue (Ries et al. 2001).

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10.2  Grazing Grazing can be thought of as the processes and conditions in which plants are consumed by herbivores, these ranging from small insects to large mammals, and which vary greatly in their feeding specificity and selection, and their patterns of activity. The conversion of so much previously high quality grassland in Australia for agricultural or pastoral uses (Moore 1970) exemplifies the principle that extensive grazing is ‘a way of making economic use of grassland that is not suited to more intensive agricultural enterprises’ (Suttie et al. 2005). Many characteristic patterns and interactions between native grazers and grasslands have co-evolved over long periods. The more specialised of these interactions, such as insects feeding on single host plant species on grasslands, can break down from impacts of changed management or other disturbances. In general, the activities of many grazing species help to promote grassland heterogeneity, and Blair et al. (2014) summarised their impacts as affecting (1) plant populations and community composition; (2) energy flow and nutrient supply; and (3) landscape heterogeneity and movement of materials. However, it seems that those grasslands (predominantly of Africa and North America) that have very long evolutionary associations with nomadic grazing mammals are very resilient to grazing – in part because, despite causing an immense local impact in any location, this pressure occurs over only a short time before the grazers move away, and allow a long recovery time for the grazed sites. Impacts of domestic stock and more intensive grazing regimes manipulated for more intensive, economically-driven, production have removed or reduced those opportunities to sustain natural balances. Overstocking and overgrazing are prime causes of grassland degradation. Changes to insect communities from grazing are mediated through changes to the plant community composition and structure. Heavy grazing can lead to an impoverished insect community (Argentine montane Pampa: Cagnolo et al. 2002), with changes in richness, biomass and guild structure (Fig. 10.2), and such effects may be far less evident under less intensive grazing. The most heavily grazed regimes had far fewer secondary consumers, and ‘chewers’ replaced ‘suckers’ as the dominant insect herbivores. Impacts of grazing can differ considerably between earlier and later successional phases of grasslands or heathlands treated, so that different assemblages depended on high or low grazing intensity. In the Netherlands, all ten butterfly species assessed by WallisDeVries et al. (2016) were influenced by grazing management. High intensity grazing helped to define a community of thermophilous species – with varied example species noted by WallisDeVries et al. being a butterfly (Hesperia comma), a grasshopper (Oedipoda caerulescens) and an ant (Solenopsis fugax). Only very light grazing favoured the needs of the contrasting species assessed in these orders; Heteropterus morpheus, Stethophyma grossum, and Formica picea. In practice, differences such as these imply needs for different management regimes along grassland successional gradients. In this example, WallisDeVries et  al. suggested that insect diversity would be promoted by (1) low intensity year-round grazing by cattle or herded sheep grazing over large heterogeneous areas, and (2) targeted or rotational grazing in smaller areas where risks from overgrazing are greater.

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Fig. 10.2  Insect density (a) and biomass (b) in different trophic guilds, from samples collected at four sites in montane grassland in Argentina under different grazing regimes and periods of cattle exclusion. Regimes are (1) heavy grazing; (2) light grazing; (3) young (7 years) exclusion; and (4) old (19 years) exclusion. Feeding categories are parasites (top dots), (predators (vertical hatching), scavengers (open), suckers (black), chewers (bottom dots), and scrapers (diagonal hatching). (After Cagnolo et al. 2002)

There is, of course, universal acknowledgement that grazing affects grasslands of all kinds, and that grazing regimes can be imposed for both management (general or specified) and agricultural production. Examples proliferate of both beneficial and harmful impacts of grazing by both native and domestic herbivores, but most studies have not focused on invertebrates. In reviewing impacts of large herbivore grazing, from studies mostly undertaken in Europe or North America, studies of overall impacts on arthropod diversity were scarce amongst the 141 studies published over 1940-May 2013 (van Klink et al. 2015). One essential ecological role of large herbivores is to prevent grassland succession to woody vegetation, and so to sustain grassland habitats. Highly variable responses of arthropod diversity to grazing relate to defoliation reducing the plant resources needed by insects and other consumers, but in some cases structural heterogeneity may be increased, leading to increased plant diversity. The other key impact was generally increased disturbance, including unintentional predation, and an important conclusion from van Klink et al.’s review was that large herbivores only increased arthropod abundance if they also cause increased biotic heterogeneity  – if that is sufficient to compensate for resource losses and the increased mortality rate(s). That scenario was considered likely only under conditions of low herbivore density or with patchiness in herbivore densities, as a component of the grassland management. The large and complex literature on impacts of grazing on Australian vegetation, whether by domestic stock, alien mammals (notably rabbits) or native marsupials contains very little information based on experimental studies of insects. Changes

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in vegetation composition and structure are the most common focus, and many authors discuss the interactions between grazing and fire in attempts to attribute changes to particular causes. Thus, Kirkpatrick and Bridle (2013) summarised the continuing history of fire in the alpine areas of Tasmania and the Australian mainland, with far more extensive burning in the latter region, and any such differences also influencing marsupial grazing patterns. It is self-evident that herbivorous insects may be influenced by such processes, and the impacts vary geographically in relation to climate, grazing and other influences. Earlier, Bridle and Kirkpatrick (1999) demonstrated differences in outcomes from three grazing-related treatments of treeless subalpine vegetation in Tasmania. Details of different impacts from grazing by different species, stocking intensity, seasonality of grazing and other variables raise many issues in honing grassland management for optimal outcomes for all interested stakeholders. Many outcomes are context-specific. The use of domestic livestock for conservation grazing, rather than relying on native mammals to achieve this, can occasionally pose unexpected problems – such as introducing diseases that could be transmitted to wildlife (Aguirre and Starkey 1994), but this is relatively unlikely to occur in Australia, where native grasslands lack herds of large native herbivores, and is of greatest significance in North America or Africa. In contrast to, for example, North American prairies where large grazing mammals are an integral feature of the systems’ evolution, the implementation of livestock grazing in Australia’s grasslands can be considered ‘exogenous’, with impacts likely to be more severe and differ from those found in long-term coevolutionary ‘endogenous’ relationships (Newbold et al. 2014). Nevertheless, in the latter systems imposition of further livestock pressures and markedly altered grazing intensity can have substantial adverse effects on native biota, and much conservation of praire ecosystems incorporates carefully regulated grazing regimes, together with fire regimes (p. 217), that can create and sustain suitable habitat mosaics amongst which native plants and their associates can thrive. Understanding the roles of grazing is a key component of prairie restoration (Samson et al. 2004). In South Africa, Pryke et al. (2016) demonstrated that presence of varied wild megaherbivores improved diversity of three focal insect groups (dung beetles, butterflies, grasshoppers) when compared with domestic grazers on the grassland/Eucalyptus plantation ecological networks they surveyed. They recommended that a variety of large native mammals should be maintained (or introduced) into these networks in production landscapes in order to sustain the natural heterogeneity paralleling that found in local protected areas. The type of grazing in this trial was a significant driver of the assemblages of all three insect groups whilst presence of wild grazers enhanced their diversity, reflecting the increased variety of grazing impacts. Thus, varied grazing may have linked with a variety of nectar sources and larval food plants for butterflies, with the grassland species an important subset of these. Pryke et al. (2016) emphasised the likely importance of varied grazing activities for grassland insect conservation and that, if this was not possible, a range of species of domestic stock may in part provide equivalent variety of outcomes.

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Large mammal grazers clearly have very varied impacts on arthropods, ranging from strongly positive or negative, to non-significant (or, at least, non-detected), and some studies have implied different impacts on different ecological guilds of arthropods. Thus, for grazing by American bison (Bos bison) in tallgrass prairie in Oklahoma, Moran (2014) found that herbivore and carnivore assemblages increased in abundance, with herbivore increases largely due to sap-sucking taxa, and the chewing taxa little affected. Increase in Hemiptera on grazed over ungrazed plots was especially significant (Fig.  10.3). However, no increases were found among detritivores. Underlying effects of bison grazing included (1) large reduction of total plant biomass; (2) changes in plant community composition from dominance by grasses to greater balance between grasses and forbs; and (3) increased nutrient cycling of plant nitrogen. Changes in physical structure occurred from bison activities such as ‘wallowing’. Moran suggested that increased nitrogen levels are likely to be the predominant cause of increased arthropod abundance, with community transformation to include greater plant variety secondary to this. Indeed, declines of this dominant herbivore and of tallgrass prairie have thwarted understanding of the roles of bison, as a supposedly ‘keystone grazer’. Knapp et al. (1998) noted the recovery of bison by conservation efforts and that these recently-­ formed herds are the basis for increasing understanding of that supposed historical role of bison, and of interactions between burning and grazing. Reflecting a perhaps more general condition (from surveys on Konza Prairie, Kansas) bison exhibit two grazing patterns to create (1) distinct grazing patches, often 20–50 m2 in area and (2) more extensive ‘grazing lawns’ of >400 m2, but are selectively grass-feeders and avoid forbs and woody vegetation. That selectivity can lead to changes in floristic composition – but in favour of the rich forb component that is so important for fostering insect diversity. On Konza, re-introduced bison fed largely on the dominant C4 grasses (mainly Andropogon gerardii and Sorghastrum nutans, with dominants comprising >90% of herbaceous material: see Koerner and Collins 2014). Before the bison re-occurred, fire regime was a critical structuring process – Knapp et al. (1998) commented that without fire, tallgrass prairie disappears – with frequently burned sites highly productive but with relatively low plant diversity. A conceptual Fig. 10.3  Abundance of different orders of arthropods on grassland plots grazed (black) and ungrazed (open) by bison (Moran 2014). Orders are: (1) Hemiptera, (2) Diptera, (3) Hymenoptera, (4) Coleoptera, (5) Araneae, (6) Orthoptera

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model (Fig. 10.4) devised by Knapp et al. illustrates some of the dynamics of the interactions within this grazing system. Selection of grasses by bison was driven in large part by frequent low-intensity fires (Scasta et al. 2016), and recent attempts to equate bison and cattle for their equivalence as ‘ecosystem engineers’ are a complex task that necessitates comparisons across variations in stocking rates and burning regimes. Because stocking rates are controlled by managers, Scasta et  al. asserted that managers are the primary ecosystem engineers, in altering grazing intensity, and recognising that effective stocking rate is essential for application of patch-burning in management. The vegetation structure established at three levels of grazing and fires at different intensities are shown in Fig. 10.5, and indicates how the interaction (as patch-burn grazing) may be beneficial. In tallgrass prairie at Konza (Kansas), the density of grasshoppers differed significantly in adjacent watersheds that differed in bison grazing and fire frequency (Joern 2004). Grazing was associated with increased density of seven of the nine most abundant species; fire frequency influenced two others, and only one failed to respond to either treatment. No interactions between grazing and fire were detected. However, exclusion of grazing, such as by avoiding the selective impacts of bison grazing on tallgrass prairie restorations, is a feature of many such exercises (Camill et al. 2004) and may influence the convergences between restored and natural vegetation structures. Grazing by red deer (Cervus elaphus) and Chamois (Rupricapra rupricapra) on subalpine grassland in the Swiss Alps, in one of few studies that have appraised impacts of large non-domestic mammals on Orthoptera (Spalinger et al. 2012), created mosaics of short and tall grass patches However, orthopteran abundance and diversity did not differ between these, and Spalinger et al. believed that no direct effects were evident from wild ungulate grazing – but that some indirect impacts occurred through changes in plant nitrogen and small-scale habitat structure. Other studies, in contrast, imply that herbivores of different body sizes might have different effects on grasshopper abundance. Huntzinger et al. (2008), for example, found different trends from grazing by Black-tailed deer (Odocoileus hemionis columbianus, from which grasshopper abundance increased) and Jackrabbit (Lepus californicus, with grasshopper abundance decreased). Grazing by either size class of grazer can thus become a transformative agent and, in general, vertebrate herbivory can substantially change or modify vegetation, with concomitant changes to associated invertebrates (Rambo and Faeth 1999). Relevant comparisons from Huntzinger et al.’s study in California coastal dunes (using replicated ‘exclusion blocks’ with one or other of the two grazers present, both species present, or both species excluded) are (1) that deer had no measurable impact on grasshopper abundance, but jackrabbits reduced this by 62%, and (2) abundance of lepidopteran larvae increased by 31% with deer, but were reduced by 36% from jackrabbits (Fig. 10.6). Those trends emerged from surveys of single species of each taxon: Platyprepia virginalis, an arctiid moth, and Microtes occidentalis, a localised and specialised grasshopper. Both species are polyphagous, so that food may not become limiting through grazing, but either herbivore could influence the kind and quality of food available.

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Fig. 10.4  Conceptual model of dynamics of bison grazing and responses of tallgrass prairie under different management regimes to the reintroduction of bison. (Simplified after Knapp et al. 1998, who incorporated information from many published studies into their original diagram)

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Fig. 10.5  Generalised model for grassland ecosystems based on studies of responses of vegetation structure to a gradient of stocking rate in patch-burn grazing management in Iowa, 2008–2013. (Burn patch structure, dashed line; structural heterogeneity, solid line) (Scasta et al. 2016)

Fig. 10.6 Abundance (mean and SE) of grasshoppers on grassland plots in response to presence of (a) jackrabbits and (b) deer (grazer present, black; grazer absent, open). (Huntzinger et al. 2008)

Debates continue, also, on the relative importance of native and introduced mammals as grazers in achieving conservation (Kohl et  al. 2013), as associated with ‘more natural’ versus ‘more imposed’ disturbances – with some implications that effects depend more on management than the species involved. However, different grazers have different impacts – grazing by cattle and sheep, for example, leave ­different sward heights. Sheep may also be more selective, but also tend to avoid taller fibrous plants. Reports that different breeds also produce differences in

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grazing (Rook et  al. 2004) probably in part reflect body sizes, with some large breeds less selective. Rook et  al. noted differences in grazing between mammal species, and pointed out that whilst some species-related effects on selectivity of food are reasonably clear, there is far less experimental evidence of impacts of grazing animals on biodiversity. The major research need identified was to provide a more informed basis for choice of stock used to manage diverse degraded grasslands, recognising that individual empirical investigations may have only limited values. They identified five such areas considered critical to future progress (Table  10.4). The current most practical perspective simply acknowledges that grazing has profound effects on grasslands by (1) selective defoliation; (2) treading, creating opportunities for gap-colonising plants; and (3) nutrient cycling, providing local nutrient supplies through dung and urine patches, affecting plant competition and distribution of grazing activity. The wider beneficial aspects of mammal grazing include restoration of plant diversity, but increased awareness of the variety of possible outcomes (Metera et al. 2010; Codron et al. 2011) may lead to conservation-targeted management that incorporates carefully regulated grazing regimes. An intriguing hypothesis discussed by Debano (2006) has suggested that grasslands and their associated fauna that evolved in the presence of large grazing herbivores should be expected to be relatively insensitive to  – and even, dependent on – grazing by domestic stock in comparison with grasslands without recent evolutionary history of large herbivores. Thus, Debano pointed out that most European grasslands had been grazed (both by native species and domestic stock) for millennia, and that most studies on North American prairies have been within the historical range of bison. In contrast, much of the south western United States has no such associational history, as herds of native grazers have been absent since the Pleistocene. She anticipated examining how insects on these potentially more ­reactive grasslands were affected by cattle grazing, using a large research ranch ungrazed since 1967, but surrounded by active cattle ranches, in Arizona. This Table 10.4  Five key future research areas suggested to enhance progress in choosing the best types of animal to manage biodiverse grazed pastures (Rook et al. 2004) 1. Determining the extent to which elements of foraging behaviour and selection are genetically determined or learned, allowing for better use of animals intended for use in conservation schemes 2. For elements that are determined genetically, utilising advances in genetics to identify the genetic basis of observed differences, to facilitate breeding of animals more suited to management objectives but also enhance understanding of the co-evolution of grazing animals and their food plants 3. Identifying the factors that determine choices made by different animals in different situations, to allow more accurate prediction of choice animals will make without needs for multiple individual preference testing 4. Improving and generalising understanding of the effects of animal type on the spatial distribution of foraging, leading to greater understanding of effects of grazing on structural heterogeneity 5. Integrated modelling of biological and socio-economical outcomes of using different animal types, for better application of decision support systems for grazing management

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comparison provided ‘one of the greatest contrasts’ between grazed and ungrazed grasslands in the region. Sweep net samples, analysed for ‘all insects’ and for each of the six most abundant orders, showed insects more abundant on ungrazed sites but with overall richness not differing significantly between treatments. However, Coleoptera were richer on ungrazed sites, Diptera more diverse (Shannon-Wiener index), and Hymenoptera both richer and more diverse. In contrast, Hemiptera were less diverse on ungrazed sites. Overall, substantial differences were found between the insect communities, correlated significantly with proportion of vegetation cover and abundance of shrubs. There was some suggestion that insects of grasslands without large herds of grazing mammals in recent history might indeed be more sensitive to grazing pressure – and the multiple taxa affected implied that diversity of insects might be maximised by protecting focal grasslands from livestock grazing. The most sensitive of grassland butterflies to unsuitable grazing or mowing regimes include those that are ecological specialists with short summer flight periods, and for which seasonality of resources is especially critical. The dilemma of conserving those species whilst maintaining grassland productivity has been addressed most effectively in Europe, where Farruggia et al. (2011) suggested a strategy they termed ‘alternative rotational grazing’ (‘arg’) and compared outcomes for butterflies with a ‘continuous stocking’ regime in France. The principle of arg is removal of stock from a rotational paddock, so that local grazing pressure is increased by that stock being located elsewhere, during the main flowering period. In this context, that period spanned two months from early June to early August, so (1) preserving diversity and abundance of flowering plants over the period of their greatest values to butterflies, and (2) maintaining greater structural variety of the sward, with shorter or longer grass in the different plots. ‘High’ and ‘lenient’ stocking rates by cattle were also compared, and the arg can produce biodiversity benefits, these clearer at high stocking rates, whilst also meeting animal production targets. Species-rich semi-natural upland grass pastures in France were also the arena for comparison of three cattle stocking regimes on functional diversity of vascular plants, butterflies and grasshoppers (Dumont et al. 2009), this study implying that low or intermediate stocking rates were the most suitable for conserving high diversity of all three focal groups. Together with other studies – showing, for example, that abundance of dung–associated beetles (Staphylinidae, Scarabaeidae) was considerably greater at high stocking rates – Dumont et al. suspected that responses to different stocking rates varied, perhaps consistently, across taxa. Grazing regimes can assume critical importance in relation to conservation of monophagous insects that, by definition, depend on single plant species within affected swards. The Duke of Burgundy butterfly, Hamearis lucina (Lycaenidae) has a number of other specialisations  – such as preferring west-facing slopes for oviposition in calcareous grasslands in Europe  – but relies on Primula spp. as its sole larval food plants (Goodenough and Sharp 2016). Individual plants become more suitable if they have large near-ground ‘rosette’ form and increased leaf length, traits that are in part related to autumn and winter grazing. However, counter effects from winter grazing

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include decreased height of the sward around Primula plants. From studies in England, Goodenough and Sharp recommended moderately high autumn grazing to produce numerous food plants, with later cessation of grazing so that suitable individual plants are generated. In Switzerland, Orthoptera abundance was correlated negatively with habitat diversity, and positively with the proportion of grass on the study site. More grasshoppers, especially of grass-feeding species, occurred in homogeneous grass-­ dominated areas. As an example of likely grazing impacts on British Orthoptera, Haes (1987) noted the loss of short turf, normally maintained by rabbits, as the major threat to the Field cricket, Gryllus campestris. This always localised species depended on hot sheltered sites where its grass food plants could thrive, and its distribution decreased considerably to, at that time, leave only a single colony in England (p. 248). For some other insects, floristic diversity, especially of forbs, within grassland is vital. Different species of Bombus bumblebees on chalk grassland in southern England responded individualistically to features such as diversity and abundance of flowering plants, and height and structure of vegetation, all influenced by grazing. Regulated grazing by cattle was preferable over either sheep grazing or no management in producing locally abundant forage for bumblebees (Carvell 2002), but intensity and timing of grazing were important. Six management regimes were compared (Table 10.5), and 11 species of Bombus were recorded over a five-week period, together with 55 flowering plant species. Recently grazed grasslands (by cattle over the previous year) were clearly important for the bees, and abundance of

Table 10.5  Management regimes and habitat types for bumblebees in Britain. The six categories of grassland management represent the main features of the grassland mosaic habitat available to the bees (Carvell 2002) Reverting arable land Previously improved: last ploughed eight years before study, then lightly grazed by cattle approximately each year Unmanaged edges Habitats bordering major tracks or small plantations, generally consisting of taller late-successional grassland stages Disturbed edges Track edges heavily disturbed by vehicle activity, with plant species known to be indicators of disturbance Sheep-grazed grasslands Grazed over winter on yearly basis, at defined rates (temporary pennings of maximum size 8 ha plots with maximum of 500 sheep for two weeks, or pro rata); occasionally mown in rotation depending on herbage growth in any given year Recently cattle-grazed grasslands Cattle-grazed within the last year (temporary pennings of maximum size 8 ha plots with maximum of 120 cattle for two weeks, or pro rata) Previously cattle-grazed grasslands Ungrazed by cattle for more than a year, previous grazing as above stocking rate

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flowering plants and bees was correlated positively. Several of the most common bumblebees occurred more frequently in shorter (30–50 cm high) flower-rich grasslands, and Carvell’s study endorsed earlier work to confirm that perennial plants more typical of later succession are more valuable than annual plants to these bees. Carvell (2002) also showed that military vehicles (the main study site was on Salisbury Plain Training Area, one of the largest local grassland areas and used primarily for military training) created locally abundant forage patches on small disturbed sites resulting from vehicle activity, as further evidence of benefits from disturbance creating heterogeneity and open structure within grassland. Grazing is sometimes justified over mowing because the varied activities (trampling, defaecation, grazing) by stock are expected to foster greater spatial heterogeneity, in contrast to the more uniform outcomes from mowing. However, several studies on Palaearctic grasslands counter this assumption – for example, Turtureanu et al. (2014) found plant richness far higher in mown dry grasslands than in grazed but otherwise similar areas in Transylvania. Grazing is regarded widely as a key management process in creating sward heterogeneity, through several processes, as outlined by Rook and Tallowin (2003). These are (1) selective defoliation between plant species and plant parts within a species, affecting plant competition directly and by influencing local microclimates; (2) ‘treading’ by stock, opening up spaces for regeneration of gap-colonising species; (3) nutrient cycling, with dung and urine as localised nutrient enrichment (enhanced production of grasses on urine patches stimulate initiation and re-selection of grazing patches by bison: p. 79); and (4) dispersal of plant propagules. Each of these processes affects food and breeding site abundance and accessibility. One outcome of selective grazing can be formation of grassy tussocks, resulting from heavy grazing interspersed with clumps of ungrazed vegetation, but the major outcome is simply to reduce the complexity of living space and resources for the insects present. Because cattle generally avoid grazing directly around their dung, and that avoidance can persist even after the dung has decomposed, those areas can produce taller grass patches (‘islets’, reflecting their structural contrast with surrounding grazed sward). These localised resources may support high proportions of the arthropods in the area. Pasture ‘latrines’ avoided by grazing stock may become preferred feeding patches for some Orthoptera (Gardiner 2018). Dittrich and Helden (2012) believed that the presence of islets in grazed grassland is an important aspect of structural heterogeneity for biodiversity conservation. Trials in England involved comparisons of Hemiptera and spiders on artificially-generated islets treated with dung, fertiliser, a cutting regime to mimic grazing, and a control (fallow) area. Densities of both groups were affected by the increased nutrients available through dung or fertiliser applications, and those effects were independent of sward height. Most attention to islets has historically emphasised that the areas they cover (Helden et al. 2010, cited examples of between 10% and 47% of pasture areas) and persistence over months to, in some cases, more than a year lead to loss of productivity and financial returns. Comparison of arthropod densities in islets and ‘non-islet’ areas, the former treated as areas with longer sward, showed enhanced densities in islets, flowing from effects such as nutrition, physical complexity and microclimate suitability. In general, islets

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tend to be less disturbed than other parts of pastures, and Helden et al. suggested that they might furnish refuges for species that cannot tolerate the close grazing nearby. They may thus be functional analogues with the more widely accepted measure of leaving some areas unmown in managed grasslands, advocated by Humbert et al. (2009) in providing structural heterogeneity. Accumulation of cattle dung on Australian pastures as grazing intensity increased produced the problems of its persistence leading to pasture staling, and it being a breeding resource for native nuisance flies. Native dung beetles, adapted to the smaller faecal pellets of native mammals were unable to degrade this new resource effectively and need to alleviate the problems led to a large-scale and long-term program of classical biological control involving importation and release of an array of dung beetles native to Africa or Europe: 43 species had been released by 1984, of which more than half became established (Edwards 2007) and persistent residents in Australian pastures. The twin objectives of improved nutrient recycling and control of dung-breeding flies have been widely achieved (Ridsdill-Smith and Edwards 2011). Impacts on most native dung beetle species are poorly known, but many are unlikely to encounter the introduced species in pastures – in general, very few of the diverse scarabaeine beetles of Australian woodland or open biomes have become abundant in improved pasture in which cattle dung has accumulated. However, in Western Australia, Onthophagus ferox is common in pastures. Its numbers in trap catches declined substantially following introductions of the alien Onitis alexis and Euonoticellus intermedius, and decline was attributed to competitive exclusion by these aliens, even though they showed different seasonal activity patterns. O. ferox is predominantly nocturnal, in contrast to Onitis being mainly active by day. As cattle largely produce dung by day, dung pads may be already occupied by the alien species before thay are exposed to O. ferox. Sheep grazing is not generally considered the best method for pursuing conservation in temperate grasslands (Scohier and Dumont 2012), and their review of relevant literature on sheep impacts on plants and arthropods revealed very mixed inferences and study outcomes. Scohier and Dumont noted that sheep grazing differs from that of large ruminants, not least in sheep tending to feed more selectively, and selecting legumes and forbs over grasses, this in turn leading to increased grass dominance. Taller swards in sheep-grazed pastures also affect floristic composition by imposing competition for light on smaller plants  – and several studies have implied that abundance of nectar-feeders and pollinators may be reduced by sheep preferring flowering plants. Thus, butterfly richness can be lower in sheep-grazed pastures than those grazed by cattle (Ockinger et  al. 2006). Decreased grazing intensity may benefit Orthoptera and Auchenorhyncha, as well as Lepidoptera that then gain from reduced losses of flowering plants. Season of sheep grazing may also be significant – the limited information discussed by Scohier and Dumont (2012) suggested that plant species richness was significantly greater in autumn-grazed areas than in ungrazed pastures. Such information has wide relevance in development of grazing systems. One approach, advocated widely in some sectors, is to pursue systems for increased grazing intensity in parts of the landscape in order to enable land-saving for biodiversity

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conservation (Dorrough et al. 2007). Whilst traditional practice in southern Australia over the last two centuries and associated removal of perennial vegetation and ‘pasture improvement’ has been a major contributor to biodiversity loss, trends toward intensification through application of phosphate fertilisers have been associated with severe declines in native plant diversity. Dorrough et al. suggested that continuing those practices could become even more detrimental – in contrast, they considered that improved grazing management over broader scales may be far more beneficial, so that ‘extensive’ rather than ‘intensive’ pasture management may be clearly better for Australian biota. Without that, set-aside areas in which native plants and insects are protected from unknown grazing impacts are essential, but with the practical difficulties of buffering these against wider environmental threats such as weed invasion, increasing salinity and nutrient enrichment. Formal controls and prohibitions of grazing by domestic stock, for example the banning of cattle grazing, through non-renewal of long-held grazing leases, in Victoria’s ‘high country’ (now largely within a national park) led to strong and continuing debates in which polarised views continue for and against the practice in ecologically specialised alpine/subalpine grasslands. Cattle grazing commenced there in the 1850s, and from then on graziers also used regular burning of alpine vegetation to promote forage production. Both practices remain contentious. Wahren et al. (1994) showed that cattle grazing had severe impacts on composition of subalpine vegetation, as well as leading to increased bare ground in grassland, this increasing risks of soil erosion. The rate of recovery from disturbance is ‘unquestionably slower in areas grazed by cattle’. Those inferences draw on one of the longest-running monitoring investigations of Australian vegetation changes, based on plots established on the Bogong High Plains, Victoria, where in 1945/1946 Carr and Turner (1959a, b) constructed fenced plots to prevent grazer activity and marked out nearby unfenced plots for comparison. Those plots have been monitored intermittently since then, enabling Wahren et al. to discuss comparative changes over almost half a century. Cover increased considerably on ungrazed plots. A variety of insect responses to grazing is expected to occur amongst larger insect groups, each containing species with very different biological features and with a corresponding variety of resource needs and ‘preferences’ (Littlewood 2008). Thus, the ground-active Carabidae and Staphylinidae of Nardus stricta grassland in northern Britain were compared across ungrazed, sheep-grazed, and sheep plus cattle-grazed regimes. The assemblages of 32 species constituting 99% of the catches (17 Carabidae, 15 Staphylinidae) were dominated by five species that together comprised 84% of the total catch, and 16 species summed to 95% of the total 36,176 individuals trapped (Dennis et al. 1997, 1998). Grazing regimes significantly affected five species, but 24 of the 32 species were correlated with effects of different grazing treatments. Greater abundance of four was indicative of upland grassland/heathland assemblages, but only one of those (Carabus violaceus) was in the most abundant cohort of five species. Dennis et al. (1997) suggested that these species could be monitored to assess balance between impacts of grazing management with needs to restore the dominance of N. stricta in upland grasslands – in this context, achieved through summer grazing of sheep and cattle to produce an average

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sward height of 6–7  cm. The variety of responses to grazing is illustrated well through Hoffmann’s (2010) review of ants, discussed on p. 133. Direct trampling of larvae of the Spring webworm moth (Ocnogyna loewii, Erebidae) in Israel by cattle was negligible, with 90% of caterpillars surviving (Berman et al. 2018). These urticaceous caterpillars may cause severe irritation or allergenic reactions, and Berman et al. speculated that cattle actively detect larvae and avoid ingesting them. However, by reducing vegetation height by grazing, cattle expose the sward to insolation and higher temperatures. Larvae actively seek such warmer conditions  – so the cattle grazing produces regimes favourable to webworms during their winter development period. Larvae can become very abundant at intervals – Berman et al. noted larval densities as high as two communal nests and 400 solitary larvae a square metre. Both nested and solitary larvae can move actively, and numbers of both were more than twice as high on grazed plots than on plots fenced to exclude grazing. Such high densities, with their impacts on forage persisting after feeding, may influence other insect herbivores and food webs. In contrast to the more usual inferences on grazing impacts over longer periods, cattle grazing had a facilitating effect on Spring webworm larvae within only a few days. In a rather different context, also involving Lepidoptera, intensity of livestock grazing in Mongolian rangelands was associated with moth diversity (Enkhtur et al. 2017), reflecting pasture degradation and loss of plant species. Three categories of pasture compared were (1) lightly grazed, with typical steppe vegetation largely intact and with trends for plants that tolerate grazing to be increased and overall richness to be decreased; (2) medium-grazed pasture, with plants in the main steppe community reduced in abundance and grazing-resistant plants becoming predominant; and (3) heavily grazed pastures, in which typical steppe plants are scarce and grazing-tolerant plants are dominant. Plant and moth data were assessed on the same replicated plots and in both localities surveyed, moth species richness was 1.3–2 times higher on lightly grazed than on more intensively grazed areas, leading Enkhtur et al. to assert that moth richness is a suitable indicator of grazing pressure in Mongolia, with clear relationship to plant species richness across the three treatments. Some moth species there could possibly become endangered through heavy grazing. Although sward height is often stated to be the key element of faunal differences, and some grassland insect groups (such as Auchenorrhyncha, p. 113) show strong vertical stratifications, the main driver of such apparent restrictions in species distributions is often less clear. Thus, sward height was the most important correlative factor for European Hemiptera (Korosi et  al. 2012), reflected in the trends for grassland leafhoppers (84 species) and true bugs (140 species) in Hungary. The regional differences between assemblages were considerable and endorsed earlier studies (such as those summarised by Morris 2000) in showing that local environmental factors may lead to substantial site-specificity in the assemblage composition  – so that local management effects may easily be confounded by factors that are only poorly understood. ‘Preferences’ and ‘needs’ for particular sward heights are sometimes quite precise – different species of European chalk grassland butterflies, for example, variously need heights from 30 cm (BUTT 1986), essentially dictating need for small-scale mosaics that incorporate ages, heights, successional stages and management suitable to regenerate and sustain that variety. Sward height and patchiness, reflecting grazing intensity and species, can influence insect diversity considerably and many workers have urged the wisdom of creating heterogeneity on grazed areas. Retaining small patches of ungrazed sward in cattle pastures can have considerable values for insects (Helden et al. 2010). In Germany, grasshoppers and butterflies on plots of three different grazing intensities differed in their responses (Jerrentrup et al. 2014). The less mobile grasshoppers (nine species) benefited from the structural heterogeneity resulting from ‘lenient’ cattle grazing, and such intermediate grazing produced the most distinct sward patchiness – so that no further reduction in intensity from this 1.14 standard livestock units (one unit is 500 Kg) per hectare was needed for their effective conservation. Butterfly responses (20 species) were more variable over this relatively long-term (2002–2011) trial. Grazing was adjudged more suitable management than mowing for maintaining orthopteran diversity on calcareous grassland in Germany (Weiss et al. 2013), but site aspect also strongly influenced distribution of many species, with abundance and richness higher on south-facing, warmer sites. All four of the red-listed species present occurred on south-facing pastures, emphasising the importance of preserving those sites  – not only for richness but also as the needs for selected threatened species. Different grazing approaches affect insects in many different ways, and Poyry et al. (2004) noted that the relative effects of practices such as rotational grazing (with a cycle of several years), extensive grazing to create variable sward heights at the pasture scale, and varying grazing intensities across a landscape all need further investigation to determine impacts on a given insect fauna, and species within it. In Finland, the diurnal Lepidoptera of three sets of mesic semi-natural grassland sites were surveyed by transect counts across treatments of (1) old pastures grazed continually for at least several decades; (2) abandoned pastures, on which grazing had been abandoned for >10 years; and (3) restored pasture, where grazing had been re-introduced around five years previously after >10 years of abandonment. Across two seasons, 96 species were obtained. Species richness and abundance were highest on abandoned pastures, both for ‘all species’ and ‘grassland species’, and only minor differences occurred across the two grazing treatments. Of the 32 species for which a minimum of 15 individuals were recorded (Poyry et al. 2005), only three were most abundant on old pastures and 12 were most abundant on abandoned pastures. None of the ‘old pasture species’ became more abundant on restored pastures, and Poyry et al. (2005) considered that the five years of restored grazing was still insufficient for old pasture species to colonise, and that varied management techniques are needed to sustain full regional diversity. Short, intensive bursts of rotational grazing in alternate years might be useful in management of calcareous grassland in Britain. Species-level comparisons of beetles confirmed differences in communities between ungrazed, long-term cattle-grazed and long-term sheep-­ grazed treatments (Woodcock et al. 2005). Whilst no effects were evident on beetle

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richness or abundance, their community structure was clearly influenced by the changes in plant community structure from these regimes. Grazing promotes species needing short turf, and may progressively disadvantage or eliminate those that need tall vegetation (WallisDeVries and Raemakers 2001), a trend clarified by comparing butterflies across grazed and ungrazed areas, and areas managed by annual cutting. Butterfly abundance increased in both grazed and ungrazed areas and, whilst species richness was not affected greatly, individual species differed in their responses. The mosaic structure generated by traditional grazing provides habitat in which about 78% of the north-west European butterflies can live (Bink 1992). Species that depend on open grassland benefit most from grazing: WallisDeVries and Raemakers noted the Queen of Spain fritillary (Issoria lathonia) and the Small copper (Lycaena phlaeas) as especially benefited in the Netherlands (Fig. 10.7). Both were associated closely with abundance of their larval food plants, which are restricted to open pioneer grassland associations, and gain from even low stocking rates. Extensive grazing was regarded as a valuable management process on those coastal dune grasslands – but refinements such as determining optimal stocking rates and timing of grazing are not yet fully understood. For Lepidoptera, Poyry et al. (2004) suggested that one reason for declines in richness and abundance in grazed pastures is that many of the abundant species were confined to abandoned pastures, and declined rapidly there once grazing commenced. Interactions of grazing stock and plant diversity in grasslands are complex. Severe declines of the Bay checkerspot butterfly (p. 142) followed removal of cattle grazing, but did not occur on continuously grazed sites. Weiss (1999) attributed this to dense swards of alien grasses forming once grazing ceased, crowding out the larval food plant, Plantago erecta. Those grasses were fostered by high nitrogen levels from cattle, with those levels over several years of grazing reaching those from usual fertiliser applications – nitrogen levels (including contributions from air pollution) were an important determinant of grassland suitability for the butterfly. With acknowledgement that livestock grazing is a key management tool for grasslands, steps to clarify this complexity also for native insects are important. In China, Zhu et al. (2015) compared outcomes for plots of meadow steppe grassland s­ ubjected Fig. 10.7  Numbers of butterfly species on grassland in coastal dunes in the Netherlands subjected to three different management regimes: grazed (black), unmanaged (open), cut (dotted). (WallisDeVries and Raemakers 2001)

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to grazing by cattle, goats, and sheep as well as control sites, and low, intermediate or high plant diversity. Grazing led to increased abundance across the whole insect community present – but with the grazing treatments leading to changes in taxonomic representation and composition of the samples, taken by sweep-netting. Orthoptera and Homoptera increased in abundance with sheep grazing, whilst Hemiptera abundance decreased, and Lepidoptera abundance increased under goat grazing. The six most abundant insect species all responded, but in different ways, to large herbivore grazing. The insect community was clearly influenced by these two major factors, and better understanding of the interactive effects of large herbivores and plants on grassland insect abundance – at both species and ordinal levels – is needed to appraise fully the impacts and importance of grazing. Developing grazing strategies that can sustain and enhance local biodiversity whilst also maintaining long-term productivity depends on sound understanding and, as Dorrough et  al. (2004) noted, for Australia there is only relatively poor understanding of the real impacts of many grazing regimes. They made the general presumption, as iterated elsewhere in this book, that compatibility is most likely in low-input (‘fragile’) systems whilst less likely in highly productive (‘robust’) landscapes where management intensity and extent of divergence from natural systems is greatest. In general, biodiversity conservation was largely ignored as a primary focus of much early Australian research on pasture and livestock production. Emphasis was, rather, largely on (1) increasing the persistence and drought tolerance of pastures; (2) reducing losses of soil and nutrients, acidification, and extent and impacts of salinity; (3) minimising and preventing invasion by undesirable (toxic, low productivity) plants; and (4) controlling immediate pest outbreaks. The gradient from ‘fragile’ to ‘robust’ landscapes reflects structure of a suite of abiotic variables (Table 10.6) that can be related to ecosystem function and level of stress (here, from grazing), with major causes of decline including loss of perennial grass cover, soil nutrients and organic matter, and deterioration of soil structure – all most likely in fragile landscapes (Dorrough et al. 2004). Implied relationships between ecosystem function, stress and native biodiversity (Fig. 10.8) display clear differences between the two contrasted landscape categories. Timing and intensity of grazing across the landscape can be varied to ensure a wide variety of regimes under which both grazing-tolerant and grazing-susceptible species can thrive. Subdividing the landscape into manageable units enables some units to be managed primarily for productivity, and others for particular products, weed control, or ­conservation of rare species and communities. However, as Dorrough et al. emphaTable 10.6 Contrasting abiotic characteristics of robust and fragile landscapes (Dorrough et al. 2004)

Variable Soil nutrient availability Soil structure Erodability Temporal variability in soil moisture Mean monthly soil moisture availability Length of plant growing season

Robust High Heavy Low Low High Long

Fragile Low Light High High Low Short

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Fig. 10.8 Relationships between ecosystem function, (a) stress and (b) biodiversity in ‘fragile’ and ‘robust’ pasture landscapes: (a) as stress increases, ecosystem function declines as a result of loss of perennial grass cover, soil nutrients and organic matter, and deterioration of soil structure – likely to be most rapid in fragile landscapes; (b) if relationship between stress and native biodiversity is similar in most landscapes, dependence of ecosystem function on native biodiversity is likely to be greater in fragile landscapes. (Dorrough et al. 2004)

sised, there is little empirical evidence of the effects of different grazing regimes on biodiversity of remnant temperate grassy vegetation in Australia. That limited information implies that intermittent grazing strategies are more likely to provide long-term benefits than year-round non-varying regimes that affect persistence of native species. In Germany, intensive grazing interrupted a range of plant-insect interactions, based on analyses of floral diversity and that of several insect groups (grasshoppers, butterflies, lepidopteran larvae, trap-nesting solitary bees and wasps). Outcomes (Kruess and Tscharntke 2002) showed reduced grazing pressure to be beneficial, and absence of grazing led to increased diversity of all taxa studied. Higher insect diversity in plots with no or reduced grazing was attributed to increased vegetation height (butterflies, bees, wasps) but to vegetation heterogeneity for grasshoppers, with needs for mosaics including bare ground for oviposition. Needs for only restricted impacts from grazing were evident for grasshopers in KwaZulu Natal, where Joubert et  al. (2016) recommended restricting grazing to moderate levels with fire intervals of 2–4 years, and on a landscape scale in order to cater for needs of both disturbance-tolerant and disturbance-sensitive species. Conversely, specific grazing regimes may be employed deliberately to reduce populations of rangeland grasshoppers (O’Neill et al. 2003) to below an economic damage threshold of abundance. In their study, on assemblages that included five of the worst pest grasshopper species in North America, most species were more abundant in ungrazed pastures. Over a five-year study, however, results were somewhat

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inconsistent – and each local assemblage may respond individually to grazing. As Bazelet and Gardiner (2018) noted, ‘the effects of grazing on Orthoptera are multi-­ faceted’, and can be positive, negative or neutral. Proportions of disturbance-­tolerant and disturbance-sensitive taxa can differ substantially, and most species may be well-adapted to withstanding disturbances. As such, as in KwaZulu Natal (Joubert-­ van de Merwe and Pryke 2018), they do not usually need specific conservation measures as long as ‘normal’ grazing (and/or burning) occurs. The considerable interest in interpreting impacts of grazing on Orthoptera (above and Chap. 6) has emphasised the complexities of incorporating effects of intensity and seasonality as management tools in relation to the outcomes of heterogeneity and biodiversity conservation. Thus, rotational grazing can benefit grasshoppers (Fonderflick et  al. 2014) but responses of the species studied in Mediterranean grasslands were highly specific and varied greatly over the grazing season. Some species were very sensitive to reductions in plant volume by high grazing pressures and, so, higher in ungrazed plots. Representative results from that study (Fig. 10.9) illustrate differing impacts among grasshopper species of grazing pressure and ‘phytovolume’.

10.3  Mowing Although avoidance of mowing has been advocated commonly to improve conditions for insects in hay meadows, with the claim that unmown areas are then refuges or reservoirs for insects, this premise has only rarely been tested (Lebeau et  al. 2015). Clearly, the widespread short-term outcomes of providing more flower variety and higher floral density as pollen/nectar resources for insects are likely to be very positive – but consequences for local insect populations may be less obvious. For the Meadow brown butterfly (Maniola jurtina) in Belgium, density increased in unmown refuge areas after mowing was undertaken nearby, but overall this increase was insufficient to compensate fully for the likely wider losses, and local decline in butterfly abundance was implied (Lebeau et  al. 2015). For linear grassland strips such as roadsides, management by mowing is attractive because of its apparent simplicity, and the regime used is amongst the easiest management methods to manipulate, with the direct effects of any regime also clear. Timing (season) and frequency are the most commonly manipulated variables, and partial mowing augments the mosaic of habitat conditions produced (Valtonen et al. 2006). Wide acceptance that over-intensive mowing regimes, which are both expensive and harmful to many grassland insects, are undesirable has led to many attempts to modify mowing programmes for greater efficiency and conservation benefit, based on comparative surveys of key insect taxa on sites subjected to different treatments. Comparative incidence of butterflies and diurnal moths (pool of 107 species) in relation to three different roadside mowing regimes in Finland (mid-summer ­mowing, late-summer mowing, and partial mowing of only a narrow strip) suggested that lowering mowing intensity (partial mowing) and delaying treatment

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Fig. 10.9  Abundance of Orthoptera in pasture areas in southern France in relation to grazing pressure and amount of plant material (phytovolume), indicating varying responses among species: (a) Stenobothrus lineatus; (b) Omocestus petraeus; (c) Euchorthippus sp.; (d) Chorthippus bigutulus. (After Fonderflick et al. 2014)

until late summer may benefit Lepidoptera without any increased costs or safety issues arising (Valtonen et al. 2006). Kelemen et al. (2014) suggested that regular mowing of sites after restoration was necessary to sustain the quality of restored grasslands which, without this, can decline rapidly as litter accumulates and undesirable species increase. Mowing can directly destroy insects, especially eggs and relatively immobile larvae on low-­ growing vegetation, and also lead to indirect mortality though changes in plant (food) availability and microclimates. It is viewed widely as non-selective, with impacts restricted only by leaving unmown areas, changes in mowing height and frequency, and season of treatments. The dilemma over relative suitability for grasshoppers of mowing or burning continues to be debated. Discussion for South African taxa by Chambers and Samways (1998) implied that mowing might be preferable because it leaves insects exposed to less abrupt microclimate changes

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and food shortage, with relatively short-term impacts, and likely to recover more rapidly than on burned plots. Whatever management is employed, rotational use is likely to cater for the greatest species richness and density – essentially, as emphasised repeatedly above, creating mosaic conditions with representative heterogeneity of the grassland remaining. Chambers and Samways (1998) endorsed Morris’ (1989a, b, c) sentiment that a broader community of grassland specialists can be conserved through rotational than by continuous management. Comparisons of Orthoptera on mown and unmown areas have sometimes implied that mown areas are less suitable and support lower richness and density of these insects. The effect has been attributed variously to mortality from mowing and by migration of grasshoppers to adjacent undisturbed habitats (Braschler et al. 2009). Mowing prevents succession of grasslands to more mature, woody, biotopes, in which many ‘typical’ grassland insects may not persist. Thus, much biodiversity on remaining semi-natural grasslands in Europe is regarded widely as needing some form of mowing management, with differences in frequency and intensity of the treatments continuing to be debated (Talle et al. 2018). Largely in seeking to determine the more general applications of any site-focussed investigation or study of particular taxa, Talle et al. concluded that ‘there are no clear guidelines for choosing the most beneficial mowing frequency’, although their survey suggested that in areas such as national parks decreased management intensity (such as mowing every second year, rather than annually or even more frequently) might have both conservation and economic benefits. However, they also suggested that, if funds allow, the weak beneficial effect they detected from mowing more than once a year might encourage that treatment – but each case depends on site conditions and individual management priorities and needs. The intricacies of the impacts of mowing on individual insect species were exemplified by investigation of a threatened Large blue butterfly (Maculinea [or Phengaris] teleius) in Hungary, where its grassland habitats must support a complex suite of requirements for this specialised lycaenid (Korosi et al. 2014). As for other European Maculinea species, obligatory associations occur with Myrmica ants. Female M. teleius oviposit in flowerheads of Great burnet (Sanguisorba officinalis), and larvae initially feed on the developing seeds. They then leave the plant and descend to the ground, where they are ‘adopted’ by worker ants and carried into the ant nests. They complete development by feeding on ant brood. Suitable microclimates for specific ground-nesting ant hosts are influenced by sward height (BUTT 1986). In this Hungarian study, in meadows of the Orseg National Park, the primary host ant was Myrmica scabrinodis, although four other congeneric species have also been identified as hosts (Korosi et  al. 2014). Conservation management involved devising mowing regimes of timing and frequency suitable for sustaining the components of the conservation module comprising the butterfly, its specific larval food plant, and primary host ant species, within hay meadows for which traditional management has for long included mowing. Korosi et al. attempted to find how t­ raditional management could be revived for preservation of the butterfly fauna present, primarily to determine optimal timing of treatment for the wet meadows frequented by

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M. teleius. Those mowing regimes were constrained by need to be economically viable. The four replicated treatments were (1) control, left as abandoned; (2) mown once a year, in May; (3) mown once a year, in September; and (4) mown twice a year, in both May and September. These regimes were initiated in 2007, with hay baled and collected within a month of mowing. Adult butterflies, host plants and ants were assessed in 2007 and 2010. The trends detected (Fig. 10.10) included: (1) total and mean daily counts of butterflies declined from 2007 to 2010, and in 2010 numbers were significantly higher in all managed treatments than on abandoned sites, but had also declined since 2007 on all except the plots mown only in September; plots mown only in May were also preferred over abandoned plots, but significantly less so than the September plots; (2) numbers of flowerheads of S. officinalis increased over this interval and again were greater in all managed plots in 2010 than in abandoned plots; and (3) changes in frequency of M. scabrinodis showed little variation, and were not influenced obviously by management. The butterflies avoided abandoned plots and ‘preferred’ plots cut in September only, a trend consistent with supply of S. officinalis for oviposition. The lack of responses by ants was contrary to an earlier study in Germany in which Grill et al. (2008) found that September mowing benefited the Myrmica hosts of M. teleius there – and Korosi et al. noted the possibility that a three-year survey interval might be insufficient to reveal any response. Cessation of mowing could lead to rapid decline in habitat quality, and to increased weed invasions. Importantly, the regimes they tested all complied with the current legislative constraints in Hungary, and other less intensive mowing regimes not currently permitted (such as mowing only every second or third year) might be suitable for butterfly conservation. The context of mowing urban green areas such as residential gardens reflects the importance of those areas as habitats for bees and other essential pollinators, and as a substantial component of the wider urban grassland estate (p. 89) that also Fig. 10.10  Changes of butterfly index (2007– 2010) in abandoned (control) grassland and grasslands managed by three different mowing regimes. (Korosi et al. 2014)

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encompasses parks, roadside and median strips, community gardens and others (Lerman et al. 2018). Mowing of lawns retards growth of weedy species such as clover, daisies and dandelions that are commonly sources of nectar or pollen for insects throughout their growing season, together with a wide variety of other dicotyledonous plants regarded as undesirable in lawns. Presence of those resources was compared across urban lawns in Massachusetts mown at intervals of one, two or three weeks, (using a standardised regime of 6.35 cm sward height and all clippings left on site), in each of which bees were sampled over two years by pan traps and direct netting over five periods/site/year (Lerman et al. 2018). Major outcomes were: (1) sites with the 3-week mowing regime had significantly greater floral abundance than others, and (2) the 2-week regime had greater bee abundance, but lowest richness and evenness. Altogether, 93 bee species were collected, with the 10 most abundant species comprising 78% of captures. The high abundance of small Lasioglossum spp. (‘sweat bees’) implied presence of suitable nesting sites, as well as food. The taller grass (reaching 15.1 cm in the third week) was suggested as a factor associated with lower bee abundance, as access to flowers may have been hampered by them becoming overgrown, but intensity of disturbance from the 3-week mowing might also have contributed to this. Because these regimes were undertaken over a short period only, carryover effects across years could not be evaluated, but Lerman et al. also targeted sites with few cultivated floral resources, so as to better reflect the resources provided directly from the lawn areas. Their comment that unintentional and spontaneous flowers growing in lawns serve as ‘unsung heroes for bee conservation’ deserves wide attention, together with their suggestion of a ‘lazy lawnmower’ approach in which such weeds are accepted and encouraged through less frequent mowing. A major purpose of broader-scale mowing is to provide forage for winter feeding of stock. As such, it is a central component of animal husbandry, especially in regions where year-round grazing is not possible. Both mown meadows and grazed pastures can support numerous biota, and their conservation values reflect the traditional practices applied in the past, and at present. As Kirby (1992) noted, the non-­ selective practices of mowing result in different habitat mosaics from grazing, and a number of studies in the northern hemisphere have compared their impacts on grassland insects. In parts of Europe, concerns have arisen as mowing increases in extent, reflecting strong declines in livestock production linked directly with difficulties of assessing optimal grazing management (Saarinen and Jantunen 2005). Comparison of mown and grazed sites in north-west Russia and adjacent Finland revealed the butterfly fauna of the two treatments to be very similar, but of the 37 most abundant species (of 48 species in total), three showed significant preferences as indicators for mowing management. All three (Polyommatus amandus, Ochlodes sylvanus, Aphantopus hyperantus) depend on Poaceae for larval food and may be thought of as ‘open grassland species’. The precise resources needed by any focal species are, of course, critical considerations in predicting and refining management. Thus, in comparing the influences of mowing on two endangered species of Large blue butterflies (Maculinea nausithous, M. teleius) by simulation modelling, Johst et al. (2006) noted that popu-

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lation dynamics of both are affected mostly by availability of the larval food plant (Sanguisorba officinalis) and nests of the appropriate host Myrmica ant, in which later larval life is passed. Mowing affects both these resources directly, and mowing at particular seasons may cause substantial mortality. The models showed that mowing to the traditional regimes (twice a year, and the second mow within the butterfly flight season) was always detrimental. The alternative of mowing only once a year, or even only every second or third year and not during the flight season, was appropriate for conservation of both species. It was also essential for mowing to be undertaken across meadows that were interconnected and within reach of dispersing butterflies, to enhance wider regional conservation. One practical impediment was the need to depart from traditional regimes, a step that might incur financial compensation to farmers for losses of the hay yield experienced. Reduced mowing intensity, from annual mowing to mowing only every three years, led to loss of biodiversity from grasslands in southern Sweden (Milberg et al. 2017). The changes in vegetation after mowing every third year included lower plant species richness, increase in ‘woody’ species, and also in tall-growing species.

10.4  Fire Deliberate prescribed or planned burning of grassland is very common, but in many contexts is still contentious, and the outcomes and impacts unpredictable or unknown in detail. ‘Fire regimes’ encompass a range of highly variable parameters that can variously benefit or harm insects – fire frequency, extent, season, and intensity are the most important ecological considerations in seeking to tailor a fire for any particular purpose, and its possible juxtaposition with other measures for insect conservation management (New 2014). However, stated aims of fuel reduction, stimulation of nutrient recycling and control of invasive plants are far more widespread than primary conservation concern or motive. Much fuel reduction burning in southern Australia has been undertaken directly for future protection of human life and property from wildfires. Knowledge of the impacts of management fires on many groups of insects is fragmentary or non-existent, so they are almost invariably ignored in predictions or assessments of effects. In a survey of Thysanoptera (thrips) responses to patch-­ burning in Brazilian campos, Podgaiski et al. (2018) examined responses of several feeding guilds within their sample pool of 44 species. Resprouted nutritious vegetation in recently burned patches benefited some flower-feeding or leaf-feeding species, and richness of fungivorous species increased a year after the fire – probably reflecting increased soil fungal density. Podgaiski et al. commented that the burned patches ‘seemed to represent high quality resource-rich habitat spots’ favourable for increased thysanopteran diversity. Body sizes were also larger in burned patches than in collections from adjacent unburned areas, perhaps related to greater dispersal capacity of larger individuals. Positive influences of increased vegetation quality

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after prescribed burning had been noted previously for thrips in New Zealand (Barratt et al. 2009). ‘Natural’ fires are an important component of maintaining the integrity and diversity of grasslands in many parts of the world, but later well-intentioned imposed burning to reduce alien plant invasions, reinstate or conserve early successional stages and reduce fuel loads in anticipation of lessening the severity of future wild fires, have led to many impacts on insects and their relatives. The potential flexibility within ‘fire regimes’ can provide opportunities to reduce harm to invertebrates and foster conservation and recovery of notable native species. The legacy of past land-use may be influential, as Moranz et al. (2012) found the highest density of Speyeria idalia (p. 79), and of another prairie specialist butterfly (Cercyonis pegala) in their ‘burn only’ treatment in Iowa and Missouri, in which the entire pasture was burned every three years. For these two species, Moranz et al. suggested that the development of denser swards was important, that high densities persisted from re-colonisation from nearby unburned patches, and that unburned refuges remained in the burned patches. Spatial heterogeneity of burns was a critical need, with recognition that prescribed burns can be tailored to help conserve species such as S. idalia. Moranz et al. (2014) used the term ‘prairie butterfly paradox’ to reflect the sensitivity of such butterflies to disturbance that is necessary to maintain their habitat – so that, more generally, of disturbance-­sensitive species being able to thrive in disturbance-dependent systems and likely to be paralleled in many other grassland insects. Although somewhat cautious about extrapolating from studies on S. idalia to other taxa, Moranz et  al. hypothesised that disturbance regimes that affected nectar availability would often have indirect effects on specialist butterflies. Scale of burning, and its interactions with grazing, as pyric herbivory, in their study arena of mixed patch-burn grazing also implied that persistent grazing over many years can lead to losses of beneficial nectar plants, and urged need to foster habitat heterogeneity through mosaic treatments. Indeed, advocacy for mosaic or ‘micromosaic’ burning for insect conservation is widespread (New et al. 2010). Until recently, many ‘control burns’ in Australia were intended to burn large areas – largely for fuel reduction and most commonly on the floors of forests – and attempts to refine this practice in favour of smaller patch-­ burning were thwarted by ‘tradition’, regulatory demands and logistic problems, and also by lack of sound information over when and where to burn (Parr and Andersen 2006). More broadly, the ecology of fire in grasslands continues to receive considerable attention in relation to use of fire in management, following Daubenmire (1968). Rather than direct incineration, indirect effects, such as changes in microclimate, soil exposure, soil moisture regimes, and losses of shelter and food, may be the major influences on many insects. For grasshoppers in South Africa, the general structure of grassland (tall or short grass) and presence/absence of particular food plants influenced local plot suitability (Chambers and Samways 1998). Because of the long history of natural fires moulding prairie ecosystems, many of their endemic insects are likely to have evolved to survive burning (Swengel 1998). Effects are also likely to be species-specific, and interpreting impacts on individual life stages may become necessary to assess the overall fate of the species.

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The diurnal moth Hemileuca eglanterina (Saturniidae) lays more than twice as many egg masses in burned than in unburned prairie during the burn year, but no such bias was observed in numbers of active adult moths (Severns 2003), suggesting that the usual outcomes in such studies  – based on assessed differences in adult abundance alone – may be an insufficient guide to impacts. Loss of eggs and larvae of H. eglanterina from prescribed burning is highly likely and, as for parallel cases involving other species, can be a major factor in population loss. Thus, complete mortality of eggs and larvae of Fender’s blue (Icaricia icarioides fenderi, Lycaenidae) can occur from prairie fires (Schultz and Crone 1998). Less obvious impacts can occur through changes to normal development patterns. Thus, burning can influence times of egg hatch of grasshoppers, probably related to soil temperature differences between bare and vegetated surfaces (Meyer et al. 2002). The immense variety of responses by insects to the burning of their habitats is displayed in several overviews (see Warren et al. 1987; Swengel 2001; New 2014). Collectively, the insects have three main options: perish, move away, or find shelter on the site. Immediate direct mortality by incineration as the fire occurs reflects extent of exposure to the flames and the insects’ capability to disperse or survive in any available refuge as the fire passes. In a parallel context, much has been learned from the common use of fire to control populations of some insect pests of pastures or rangelands. Warren et al. (1987) noted a number of such species across a range of orders but for which control burning had been suggested or used as an effective and economical method of suppression. Deliberate burning can then be timed to affect key life history stages, and may need to be applied at limited periods over the growth of a crop or when especially susceptible pest stages occur – such as early or late summer, autumn or winter burns for particular contexts. Amongst other examples, Warren et al. noted (1) a mealybug, Heterococcus granicola, which attacks at least 67 grass species in North America controlled by late summer/autumn burns before the insects move to overwintering refuges around the base of their host plants; and (2) the vulnerability of pest grasshoppers to burning during their wingless nymphal stages. As in much insect conservation, valuable lessons in approaches and strategy flow from experiences of pest management but in which the aim is to conserve rather than destroy the target taxa, and protect them from threats. In both contexts, the success of the treatment depends on sound knowledge of the phenology and biology of the target species. Although now somewhat outdated in detail, a list of advantages of using fire to manipulate arthropod populations in pest management (Table 10.7) indicates the principles that have guided that interest, with management extending to using fire to suppress woody vegetation and other weeds. In short, the contention noted by Warren et al. (1987) ‘that prescribed burning can be used to manipulate systematically certain arthropod populations in rangelands’ has the dual potential for pest suppression and conservation salvation. Tracing fire impacts on Illinois prairie Auchenorrhyncha showed relationships between species richness and both time-since-fire and frequency of burning (Wallner et  al. 2012, Fig.  10.11) across the 19 hill prairie sites surveyed, all dominated by prairie grasses. Almost all the bug species (64 of 65) were native, and 28 species were believed to depend on grassland remnants and be sensitive to disturbance; 15 species

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Table 10.7  Advantages of using fire to manipulate arthropod populations in pest management, once their biology and seasonal development pattern is known 1. Burning is relatively inexpensive 2. Unlike in pesticide use, insects are unlikely to develop any form of genetic resistance to fire 3. There are no pesticide residue problems 4. Fuel for fire is provided by naturally renewable plant material, so no use of fossil fuels 5. Burning may simultaneously suppress woody plants and enhance vigour and seed production of more desirable perennial grasses 6. Burning aids in recycling nutrients from dead and senescent plant material As listed by Warren et al. (1987)

Fig. 10.11  Responses of Auchenorrhyncha species richness to prescribed burning, on 19 hill prairies in Illinois. Values are based on their ‘coefficient of conservation’ scores and are shown as response to (a) time since burning and (b) frequency of burning. Categories within each column are for highly conservative (coefficient values 11–18, in black), moderately conservative (6–10, open) and adventive (0–5, dotted) for each treatment. (Wallner et al. 2012)

were moderately tolerant of disturbance and associated with prairies and other habitats; and 22 species, not depending on prairies, are adapted to frequent disturbance. Unburned sites supported the highest richness. The ‘coefficient of conservation’ (see also Wallner et al. 2013) reflected the above categories as ‘highly conservative’ (remnant-dependent, fire-sensitive, restricted to native grassland), ‘­moderately conservative’ (found in edge habitats and native grasslands, moderately tolerant of disturbance) and ‘adventive’ (adapted to frequent disturbance, feed and/or overwinter on a variety of host plants) species. That coefficient was based on several life-history

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Table 10.8  The ecological traits and scoring system (scores in parentheses, for each condition) used to formulate the ‘Auchenorrhyncha Quality Index’ as a grassland monitoring tool Trait Voltinism Host plant specificity Overwintering preference Wing length Origin Habitat fidelity

Condition (score) >2 generations/year (0); 2 generations/year (1.5); 1 generation/year (3) Polyphagous or feeding on non-prairie plants (0); 2 families of prairie plants (0.5); 1 family of prairie plants (1); 1 genus of prairie plants (2); species within 1 genus of prairie plants (2.5); 1 species of prairie plants (3) Migrating and not overwintering locally (0); non-prairie plants (1); soil/duff (dead vegetation) (2); prairie forbs, grasses, sedges (3) Macropterous (0); macropterous/brachypterous (1); submacroptrerous (2); brachypterous (3) Exotic (0); native, occurring over broad geographical range (1.5); native, restricted to small geographical area (3) Remnant independent or adventive (0); other than prairie associates or ecotonal (0.75); prairie associates or found in prairie remnants (1.5); remnant-dependent species (more than one prairie ecotype) (2.25); prairie-­ dependent (one prairie ecotype (3)

From Wallner et al. (2013)

traits (Table 10.8) with each species given scores that can be summed. Each trait was given a score from ‘0’ (non-conservative) to ‘3’ (conservative), summing to a value for each species within the range of 0–18. The approach is noted in detail here because of its probable relevance for parallels with other taxa, as exemplifying the kinds of biological features that might correlate with vulnerability and conservation interest. Wallner et al. (2013) noted that the index might be a model for developing habitat quality indices based on other diverse grassland insect groups and, more broadly, contribute to more complete assessment of habitat quality. They also (Wallner et al. 2012) recommended infrequent (minimum of 3–5 years) rotational burning to conserve most native prairie Auchenorrhyncha and reduce adventive species likely to become abundant after a fire, and conserve prairie vegetation. Mortality is not necessarily immediate or short-term, and intermediate-term (over 1–2 to 12 months) impacts of fire are also very varied (Swengel 2001), reflecting destruction of key resources, or seasonal burns over periods in which insects are not vulnerable (for example whilst life stages are underground). Minimal impacts of fire on wingless invertebrates have been reported, with strong implications that such suitable refuges are present, and used (Uys et al. 2006). Soil may also protect ant nests from direct fire effects, and other variables affecting fate of insects include distance from the fire edge and the size of the area burned. The most common logistic context in assessing fire impacts on insects, however, is one in which diversity can be presumed to be reasonably high, and knowledge of their taxonomy and ecology low – so that whilst it may be possible to change an imposed fire intensity, area, timing and so on, a precautionary approach to avoid or minimise damage is largely advocated. Such an approach involves diversity of fire management both within and across successive seasons as ‘patch mosaic burning’, anticipated to promote heterogeneity and ecological resilience, and facilitating chances of refuge presence and source populations as the sources for colonists that

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can later occupy burned areas. In many contexts, such as managing protected areas for the greatest range of conservation values, managers need to determine the most suitable fire regime to achieve their aims – in Australia, commonly involving protection of human life and property. As recommended for prairie arthropods (Harper et  al. 2000), the regime should provide for unburned refuges to persist and that intervals between successive burns are long enough to allow re-colonisation of areas affected. Overly frequent burning can lead to losses and population reductions of many insects and, at least for prairies, burn rotations of at least two years, or alternatives to prescribed burning, may need to be considered seriously. No single management technique was clearly beneficial to all specialist butterflies in tallgrass prairies surveyed by Swengel (1998). For any taxa, recovery will depend on regeneration or survival of critical resources, and the presence of populations from which re-colonisation of affected sites may occur. Reports on prairie insects show patterns of changed densities and richness of butterflies, grasshoppers and others that reflect changes to the habitat, with the widely predicted outcomes that specialised species may be far more vulnerable to fire than are relative generalists. Evaluating the impacts of fire necessitates evaluation of recovery over realistic time periods for this to occur, rather than the more usual short periods limited by funding and that may constrain realistic interpretations. Butterfly recovery times after prescribed fires in tallgrass prairie were examined across a series of species in Iowa (Vogel et al. 2010), demonstrating that total richness (pool of 49 species) and richness of habitat specialists (18 species) were correlated positively with time since burning. Recovery time for habitat specialists was >70 months, potentially longer than previously reported intervals. That study was innovative in also examining the relative contributions of direct effects of the fire and the indirect effects of time since fire through changes in vegetation composition on butterfly abundance. The path analysis used the scheme summarised in Fig. 10.12, employing separate multiple regressions for each abundance variable. In contrast to Swengel’s (1996) finding that effects of fire on specialist grassland butterflies lasted for 3–5 years, outcomes from Vogel et al.’s analysis suggested that recovery for some species may take substantially longer. On managed prairie remnants, needs to control encroaching woody vegetation often necessitate burning at intervals of less than five years, so that information on recovery times of habitat specialist insects may be critical in safeguarding them. Preventing the transition from grassland arthropod communities to woodland communities on the Great Plains necessitates use of fire, creating what Kral et al. (2017) referred to as a ‘disturbance paradox’ because responses to fire depend on many different variables. That review found that most published studies focused on arthropod responses within the first six months after fire, and Kral et al. also emphasised that three main traits of species (mobility, life stage and feeding guild) may help to predict responses in the absence of more specific information. The important gaps in current knowledge include investigations of longterm responses by arthropods, within the reality that in the Great Plains fire management can not be replaced entirely by other measures – but with improved variations in patch-burn frequency, dispersion and season of fire all leading to

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Fig. 10.12  Paths of proposed direct and indirect relationships between ‘time since burn’ and vegetation characters and butterfly abundance: independent variables chosen for inclusion in these models were floral resources (number of flowering ramets), warm season grasses (percent cover of warm season grasses) and bare ground (percent cover). (Vogel et al. 2010)

mosaic treatments and fostering beneficial heterogeneity. Discussing prairie restoration methods, Rowe (2010) commented that ‘Tools such as fire, mowing, and grazing are complementary disturbance management techniques’ that can lead to increased plant species richness. Nevertheless, fires (whether deliberate or from natural causes) can be severe threats to specialist butterflies and others that are associated with grasslands or depend on other species there, such as mutualistic ants. Swengel (2001) noted the high proportion of open habitat Lycaenidae of major conservation concern, and for which ‘fire’ had been cited as a threat. The converse scenario, of fires benefiting the butterflies, occurred largely in contexts of infrequent fires creating new habitat patches that the insects could potentially occupy for long fire-free periods, rather than more frequent burns of already-occupied patches. In tallgrass prairie systems, Panzer and Schwartz (2000) tested a hypothesis (the ‘fire-attrition hypothesis’) that frequent fires (the prairies are generally burned on a 2–5 year rotation system, with each fire burning up to half the site) will reduce species richness and lead to low population densities. In contrast, sites from which fire is excluded should continue to support a suite of fire-sensitive species absent from fire-managed sites. Their study involved comparisons of butterflies and leafhoppers, among the species surviving on prairie remnants and that must experience the fires. This hypothesis was not supported. Across the 27 remnant-dependent butterflies and 67 remnant-dependent leafhopper species, local richness and mean population densities were similar across the treatments and in some cases were higher on the frequently-­burned sites. These outcomes countered the common claim that frequent (or rotational) burning of prairies would increase levels of extirpation among remnant-­dependent insects. Panzer and Schwarz suggested the converse – that rotational burning had contributed to the protection of several such species.

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Fire is used widely in North America to maintain grassland ecosystems but, unlike most historical wildfires, many deliberate burns do not produce a natural mosaic whereby areas are burned unevenly and patchily to leave both refuges and areas burned to different extents and intensity. The imposed regimes are often purported to mimic natural disturbance patterns, so that burn intervals of 2–3 years for tallgrass prairie is a common recommendation (Fuhlendorf et al. 2009). Both dormant season and grazing season burns may be adapted for particular outcomes or purposes, and season may be integrated with grazing needs for the sites, on which recent burning and regeneration of palatable grasses may be needed, either broadly or as parts of a mosaic pattern. The need for such mosaic habitats is acknowledged widely as a major component of maintaining grassland heterogeneity. Variations in any parameter of a fire regime produce equally varied and species-characteristic responses by insects (Warren et al. 1987). Thus, carabid beetles on Konza Prairie, Kansas responded in complex ways to fire frequency, site topography (whereby variations in fire intensity lead to mosaic impacts, including unburned refuges), season, and wider landscape heterogeneity (Cook and Holt 2006). Prescribed fire can maintain prairies, but conflicts can arise between advocacy for (1) widespread annual burning of large areas to generate fodder for livestock, and (2) small scale mosaic burns at different frequencies to ensure that heterogeneity in age and structure is maintained. For insect and other conservation, deliberate creation of spatial and temporal mosaics is preferable. However, as Keith et  al. (2002) noted for Australian grasslands, management decisions over burning are more likely to reflect costs and other logistic concerns than any specified ecological criteria. Swengel et al. (2011) supported earlier recommendations on use of fire for insect conservation in Australia from their experiences with prairie butterflies in North America. In both regions with the parallel of highly incomplete information of fire impacts on insects, they recognised needs for a cautious scientific approach to optimising outcomes for biodiversity whilst still accommodating other management needs. The major recommendations made for Australia (New et al. 2010) were: (1) small or isolated sites or those that support listed invertebrates should never be burned without careful consideration of advice from specialists; (2) ‘micromosaic burns’ of no more than a few hectares and staggered across years on larger sites should be the usual method, rather than large-scale burns; (3) refuges should always be considered, with recommendation that at least 20% of a site should be protected permanently from deliberate burning; and (4) small sites (80% of all ants found), ant species were not considered to be useful indicators in that study. Whilst much attention has been paid to improving conservation practice through restoring native grasslands, and improving the quality of degraded remnants, much background knowledge on process has derived from the needs from (generally larger-scale) interests in restoring cropping areas to pasture. This measure has become widespread as declines in cropping progressively reduce economic returns and environmental performance. Especially on poorer soils, cropping can become unsustainable due to erosion, low water use efficiency, excessive deep drainage, restricted productivity and declines in nutrients. The last of these can be redressed, in part, by significant nitrogen from pasture legumes – so that marginal cropping soils can be rehabilitated by using pastures, and favoured by pasture-crop rotation systems. However, as Silburn et al. (2007) noted, many of the poorer soils are better suited for permanent pastures. Continuing pressures for food security have driven the incorporation of many semi-natural and abandoned agricultural grasslands to a highly uncertain future.

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Many can readily be incorporated into production systems, and the modes by which this is achieved are often largely not understood. Stoate et al. (2009) urged the needs for policy for High Nature Value systems in order to prevent ‘large-scale abandonment of the most biologically valuable systems’, and understand the environmental effects of restoring open spaces. Many degraded grasslands are deemed so because of mismanagement or abandonment, when succession, loss of species and (sometimes substantial) structural and floristic changes become almost inevitable. Recovery of more natural species-­ rich grasslands is a frequent conservation objective, and Biedermann et al. (2005) noted the three main trajectories advanced for this to occur: (1) re-creation of grasslands from cultivated land; (2) reclamation from scrub, essentially restoring earlier successional stages; and (3) restoring from species-poor intensively managed grasslands. Each may produce very varied results, but insects can be valuable monitors of any success obtained. For leafhoppers and others it is more likely that such restored areas will be colonised by common and widespread species than by rare specialist taxa – unless the latter are a clear target for which the particular restoration exercise has been designed, perhaps planned also to receive translocated individuals. Any exercise in grassland restoration can provide opportunity to clarify the dynamics of communities in terms of species diversity, incidence of individual taxa, and how functional roles change and in many cases converge toward their states on more natural remnant sites. Some insect groups, such as Carabidae, have well-­ recognised ‘functional traits’ which can be evaluated and compared across restoration treatments and stages from parameters such as body size, presence and function of wings, diet, activity periods and seasonality. Thus, a chronosequence survey of carabids in restored North American tallgrass prairie (Barber et al. 2017) compared their representation in regimes of restoration by seeding with local plants, 1–3 year rotation prescribed burns, and herbicide control of alien invasive weeds. Restoration, from agricultural fields, commenced in 1987, and carabids were sampled by pitfall trapping over 2013–2015, collectively yielding 56 species/morphospecies. Initially high species richness and abundance declined later to levels similar to those on adjacent grassland remnants, and species composition gradually approached the remnant level. Younger restoration sites contained mostly small, macropterous plant-feeding beetles, with a transition to older sites containing larger species, many of them flightless and carnivorous. Burning decreased the predominance of larger species  – taken as evidence that prescribed fire intended for a particular purpose (here, weed control) may also change functional composition of insect assemblages, as an undesirable and unpredicted outcome. Such transitions in insect functional compositions amongst assemblages with varied ecological traits are likely to be widespread. As in many aspects of practical conservation involving restoration of ‘habitat’, a key decision is ‘what to manage for’ in grasslands. On remnant calcareous grasslands protected as nature reserves in Luxembourg, management to reserve diverse plant assemblages indeed led to increased richness and abundance of plants – but also some decreases in abundance amongst Carabidae, especially red-listed and

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forest species (Habel et al. 2016). Comparisons between ‘pre-restoration’ (1987– 1992) and ‘post-restoration’ (2008–2011) gave very similar carabid richness (56, 55 species, respectively) over the six sites examined, and representation of thermophilous species (10, 12) and forest species (12, 10) was also similar. Comparisons were made also between carabids and butterflies over the above period (beetles) and the longer 1972–2001 (butterflies) (Augenstein et al. 2012). Differences between the taxa were substantial, apparently reflecting that carabid assemblages mirrored local changes in habitat structure, while butterflies showed variable and individualistic effects related to stochasticity, local extinctions and level of subsequent re-­ colonisation from nearby populations. No losses of rare carabids were detected, but butterflies showed extinctions of species that had specific habitat needs, high levels of endangerment and/or low dispersal ability. In this study, it was clear that the common landscape could have very different functions, as a corridor or barrier respectively for members of these insect groups. The siting of areas selected for restoration can clearly influence chances for species colonisation and recovery through countering effects of habitat fragmentation. Discussed initially by Huxel and Hastings (1999), restoring habitat patches adjacent to more natural patches, or actively returning key focal species into restored patches instead of relying on possibly low chances of natural colonisation may markedly increase success levels of recovery efforts. In practice, immigration is a vital process in restoration, of grassland or any other ecosystem, and site situation can determine the levels at which this process occurs. A widespread (and often unspecified) assumption of restoring prairies or other degraded grasslands for their plant complements as often the sole immediate objective of the exercise is simply that non-target organisms (including insects, birds and others) will subsequently colonise once a suite of suitable resources and environments are made available. However, as Griffin et  al. (2017) noted, around twothirds of all the restorations they surveyed lacked information on the status of non-target taxa, and the success of many grassland restorations in establishing functional ecological communities is largely unknown. Of those that have included insects in post-­restoration monitoring, most have been restricted to short-term observations, perhaps too short to provide any definitive evaluation of status and operational success. Quantifying the effects of restoration attempts from ‘real-world’ examples that also interpret the rationale for those outcomes is an important theme in grassland insect conservation. It is also an exercise that demands careful long-term evaluation, preferably using replicated sites, and carefully defined restoration processes and taxon monitoring. As Woodcock et al. (2012b) noted, long-term studies on invertebrates in restoration are scarce and  – even though a decade between habitat re-­ creation and subsequent suitability for butterflies is by no means large – time lags even longer may need to be considered in realistic strategic conservation planning. Using data sets on British butterflies obtained over restoration exercises from 9 to 21 years, changes in assemblages were examined for (1) ‘arable reversion sites’, where grassland was established on bare ground by using seed mixtures, and (2) ‘enhancement sites’, on which degraded grassland was restored by removing scrub

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and re-introducing cutting/grazing management. The most rapid rates of increase with restoration were on arable reversion sites, over the first 10 years. No consistent increases over time occurred on the enhancement sites. Times needed for grassland restorations to provide suitable habitats for insects vary greatly, but knowledge of those delays may be a key component of assessing ‘success’ and guiding selection of the methods likely to produce the most rewarding outcomes (Woodcock et al. 2012b). Thus, butterfly colonisation was most rapid for those species with widespread host plants or with suitable host plants established readily during restoration, and low mobility butterflies took longer to arrive. From those British data, dispersal ability was the only trait that strongly limited colonisation times, and emphasised the need for landscape context of restoration sites to be considered, in order to minimise their isolation. The restoration of intensively-used agricultural fields, whether arable or pasture and often with decades of management to produce monoculture or other low diversity systems, to species-rich semi-natural grasslands with biota approximating that of undisturbed grasslands in the same region, is a complex process. It is, however, one essential to practical conservation enhancement. Such aims are stated targets of many agri-environment schemes, but the precise procedures to follow are often unclear, and management couched in general, well-intentioned terms. One basic principle is that enhanced local plant communities can increase diversity and abundance of target insect groups, with chances of re-colonisation increased by site proximity to areas with source populations. Measures such as maintaining grass margins on arable fields can help promote such connectivity, and host possible reservoir populations of local grassland insects. In southern England, abundance and species richness of moths were compared across existing semi-natural calcareous grassland and unrestored or restored arable fields – so that moths were adjudged in three habitat groups, as ‘calcareous grassland moths’, ‘grassland generalist moths’, and ‘other moths’ on (1) arable fields; (2) former arable fields restored to species-rich grassland; and (3) semi-natural calcareous grassland (Alison et al. 2017). A total 11,252 moths included 244 species, with representation across the above habitat categories, in sequence being 28, 46 and 170 species. The abundance of calcareous grassland moths on restored grasslands was almost eight times that on arable fields, and richness and abundance did not differ from that on semi-natural grasslands. Some moths associated with later successional stages, such as woodland, were more abundant on calcareous grassland than on restored areas, and emphasise the importance of sporadic woody vegetation in increasing diversity. However, using calcareous grassland indicator plants on restored grassland sites – as ‘active enhancement’ of that particular plant community – was associated with greater abundance of calcareous grassland moths, and restoration led to benefits for those moths over as little as three years, and at distances of up to 7 km from semi-natural grasslands. Restoration of a plant community in grassland, of course, does not guarantee parallel restoration of the typical/expected/original insect community. Following the preservation of some abandoned semi-natural calcareous grasslands in Luxembourg as nature reserves, assemblages of vascular plants and ground beetles

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were assessed before (1988–1992) and after (2008–2010) restoration was initiated (Habel et al. 2016). After a period of natural succession, restoration involved removing trees and shrubs to re-establish open grassland, as well as grazing by sheep and cattle. Trends differed. Richness and abundance increased for plants, but decreased for Carabidae, especially among the forest species as their relatively newly-acquired resources were lost, and many were not adapted to life in open habitats. The ‘lag time’ between grassland restoration and the recovery or establishment of representative insect assemblages is usually unknown and largely unpredictable, even when the target assemblages can be defined reasonably well by the insects present on reference grassland sites that are relatively undisturbed or deemed ‘sufficiently natural’ for emulation. Phytophagous beetle assemblages (of Curculionidae, Apionidae, Chrysomelidae) of calcareous and mesotrophic grasslands in England were recorded over periods from 3 to 14 years at different sites, with each site reasonably close (ca 0.5–1.5 Km) to possible source areas of species-rich grassland, together with availability of key host plants (Woodcock et  al. 2012a). The latter, together with proximity of source populations influenced colonisation rates. Restoration through use of seed mixtures was also important, but lag periods of more than 10  years were suggested and, extrapolating from the data available, Woodcock et al. suggested that colonisation success would plateau after 20 years, to then represent about a 60% increased assemblage similarity to that on the reference grasslands. Integration of the needs of flora and fauna on European calcareous grasslands (as dry grasslands on chalk and limestone, or on calcareous loess) shows several features in common (WallisDeVries et al. 2002) on these varied grasslands showing historical land use patterns in various forms. Those features are (1) that their characteristic biodiversity is high, with many rare species, collectively forming one of the richest plant communities with correspondingly rich attendant or dependent insects; (2) this reflects their semi-natural genesis that requires management by mowing or grazing to provide a variety of possible treatments of varying values for conservation; and (3) they are severely endangered, both from agricultural intensification and abandonment, and associated isolation and fragmentation, with increasing risks of species extirpations or wider loss. Inclusion of invertebrates in plans for management of these grasslands is relatively recent, lagging greatly behind attention given to plants. Botanically–centered management alone can disadvantage invertebrates and lead to impoverished fauna, so that management integrated to consider both plants and invertebrates is necessary. In the Netherlands, butterfly distributions have declined far more than those of plants: 8 of the 10 targeted butterfly species had been reduced to occupy