The costs and benefits of grouse moor management to biodiversity and aspects of the wider environment: a review

RSPB Research Report Number 43 Murray C. Grant¹, John Mallord, Leigh Stephen & Patrick S. Thompson

ISBN: 978-1-905601-36-3 RSPB, The Lodge, Sandy, Bedfordshire, SG19 2DL ¹Dr Murray Grant, RPS, 94 Ocean Drive, Edinburgh, Midlothian,EH6 6JH © RSPB 2012

Contents Executive summary

1. Introduction 1.1 Moorland habitats in the UK 1.2 Grouse moor management 1.3 Objectives of the review

2. Literature sources 3. Grouse moor management and biodiversity

3.1 Assessing effects on biodiversity 3.2 Moorland vegetation

3.2.1 Managements of relevance 3.2.2 Grazing-related effects 3.2.3 Factors influencing the response of moorland vegetation to fire 3.2.4 Post-fire vegetation succession 3.2.5 Effects of rotational muirburn on species richness and diversity 3.2.6 Effects of rotational muirburn on habitat condition 3.2.7 Effects of drainage and drain blocking on moorland vegetation 3.2.8 Summary and synthesis of effects on vegetation

3.3 Moorland invertebrates 3.3.1 Managements of relevance 3.3.2 Moorland invertebrate communities and the main factors affecting

diversity and abundance 3.3.3 Effects of rotational muirburn on invertebrate abundance and diversity 3.3.4 Effects of rotational muirburn on species and taxa of conservation

importance 3.3.5 Effects of drainage and drain blocking 3.3.6 Summary and synthesis of effects on invertebrates

3.4 Moorland birds 3.4.1 Assessing the effects of grouse moor management on the moorland bird

community 3.4.2 Effects of rotational muirburn 3.4.3 Effects of legal predator control 3.4.4 Effects of illegal persecution of predatory birds 3.4.5 Effects of other managements 3.4.6 Summary and synthesis of effects on birds

4. Wider Environmental Effects: Grouse Moor Management, Carbon and Water 4.1 Grouse moors and wider ecosystem services 4.2 Peatland soil structure and carbon flows 4.3 Effects of grouse moor management on wider environmental issues

4.3.1 Geographical overlap of grouse moors with soil carbon stocks 4.3.2 Effects of rotational muirburn on carbon store and water quality 4.3.3 Effects of drainage and drain blocking on carbon stocks and water

quality 4.3.4 Effects of grouse moor management on water flows 4.3.5 Summary and synthesis of wider environmental effects

5. Acknowledgements 6. References

1

Executive summary 1. The sport shooting of red grouse Lagopus lagopus scoticus is a major land-use on

UK moorlands. Debate continues over the extent to which the associated managements (i.e. grouse moor management) represent beneficial or detrimental impacts to the upland environment.

2. UK moorlands are considered to be of high conservation value for their habitats and associated biodiversity, whilst they are also important stores of soil carbon (C), major sources of drinking water and determinants of water flows, and hence flood risk.

3. Grouse moor management is associated with moorland habitats in which ling heather Calluna vulgaris is a major vegetation component, notably dwarf shrub heaths overlying mineral or thin peat soils, blanket bogs overlying deep peat soils, and their intermediates. Traditionally, much of the debate on the impacts of grouse moor management has focussed on biodiversity but, more recently, the other ecosystem services mentioned above have received increasing attention.

4. At a UK level, grouse moor management has been in decline since the 1940s. The extent of declines has varied geographically and, in Scotland, they have been greatest in the west and in the far north. In some regions, there have been increases in grouse moor management since the 1970s, but there is poor understanding of both the current area of managed grouse moor (between 0.66 and 1.7 million ha) and current trends in the intensity of management. For example, published information on the extent to which ‘driven’ and ‘walked-up’ shooting is practised, and how management regimes vary in relation to these different practices, is lacking.

5. Grouse moor management has helped to limit losses of dwarf shrub dominated habitats to afforestation and conversion to grassland. The more detailed effects of grouse moor management on vegetation are more difficult to determine, with rotational muirburn and grazing regimes of most importance in affecting vegetation. The maintenance of Calluna cover on grouse moors is sensitive to grazing, although detailed information on differences in grazing regimes between grouse moors and other moors is lacking, whilst effects of grazing and burning on vegetation are often closely interlinked.

6. Effects of rotational muirburn on vegetation may vary considerably according to vegetation composition and age, fire severity and post-fire grazing regimes. Overall, the available evidence suggests rotational muirburn increases species richness and diversity on dwarf shrub heath, but existing studies may not be fully representative of the conditions that occur on managed grouse moors, some of which may be of fundamental importance (e.g. rotation lengths and the extent to which degenerate Calluna stands are retained). Data for blanket bogs are even more limited, providing indications of detrimental impacts on habitat condition, but mainly for areas burnt within the last seven years. Sphagnum mosses (a key component of blanket bogs) are frequently considered to be susceptible to burning, but the full effects of rotational muirburn on them, and how these effects vary according to differences in muirburn regimes, are poorly understood.

7. Past drainage and more recent, drain blocking affect vegetation on grouse moors. Drainage was previously widespread on grouse moors (at least in some regions), and although some moors are now actively involved in drain blocking programmes, the overall extent to which this management is undertaken on

2

grouse moors is unknown. Studies suggest that drainage generally succeeds in reducing cover of plant species typical of wet conditions (e.g. cottongrass and Sphagnum spp), whilst findings from recent studies on drain blocking suggest some reversal of these effects, although responses may be inconsistent.

8. An understanding of the effects of grouse moor management on moorland invertebrates is limited. Rotational muirburn on dry dwarf shrub heath increases the diversity and abundance of some invertebrate groups (e.g. ground beetles), via increase in the structural diversity of vegetation. However, unmanaged and degenerate Calluna stands are important for some groups (e.g. lepidopteran larvae) and the effects of rotational muirburn on overall diversity of terrestrial macro-invertebrates will depend upon the nature of the muirburn regime, and the extent to which such stands are retained. It is unclear whether these benefits of increased structural diversity of vegetation also apply on wet heaths and blanket bogs, although ‘sensitive’ rotational muirburn regimes on such habitat probably benefit the large heath butterfly Coenonympha tullia (a species of conservation importance). Limited evidence indicates that muirburn may reduce soil invertebrate abundance, but does not affect overall abundance or taxonomic richness of aquatic invertebrates in moorland streams.

9. Grouse moor management has important impacts on moorland birds, through its effects on both vegetation and predator populations. Most strikingly, there is a marked contrast between detrimental effects on populations of some predatory bird species resulting from illegal persecution, and beneficial effects on wader and grouse species resulting from both legal predator control and vegetation management, notably rotational muirburn. The predator control regimes associated with grouse moors may be important in maintaining (or slowing declines of) populations of several species that are declining nationally (notably black grouse Tetrao tetrix, lapwing Vanellus vanellus and curlew Numenius arquata). There are fewer documented effects of grouse moor management on passerine species, but, rotational muirburn is associated with lower meadow pipit Anthus pratensis densities. It is likely that grouse moor management limits bird species diversity by restricting scrub and woodland cover.

10. The importance of grouse moor management (and particularly rotational muirburn) as a factor affecting the key ecosystem service of carbon (C) storage and sequestration will be greatest where it coincides with upland blanket peats. Grouse moors are rare or absent in many (but not all) of the UK’s most extensive areas of blanket peat. Some of the main areas of overlap also coincide with upland catchments that are major sources of potable water.

11. The effects of rotational muirburn on both C storage and water quality are likely to be complex. One detailed study found small annual losses of total soil C under a 10 year burning rotation, but these findings were obtained from an experimental site where burning practices may be different to those on managed grouse moors. Studies of effects on dissolved organic carbon (DOC) (a major determinant of water quality and an important component in the C budget) have obtained contrasting findings, but extensive correlative studies (examining both spatial and temporal variation in the extent of rotational muirburn) provide strong evidence that rotational muirburn on deep peat soils results in marked increases in DOC at the catchment scale. In some situations, rotational muirburn could confer indirect benefits to C stores and DOC levels by reducing the risks of

3

severe wildfires, which have the potential to cause serious detrimental effects. The risk of wildfire (and hence potential for benefit from rotational muirburn) varies regionally. The balance of evidence suggests that moorland burning impacts raw water quality and that burning results in increased water colour in raw water.

12. Drainage on grouse moors is likely to have caused increased losses of soil C, but there is little information available from which to assess the effects of drain blocking on overall C stores. Current evidence suggests that blocking reduces DOC yield, but effects on DOC concentration may be inconsistent and response is likely to vary with the time since blocking. There are concerns that blocking will cause increased global warming potential (irrespective of the overall effect on C store) from increased CH4 emissions, but this remains unverified, and over the longer-term the response of Sphagnum (which may act to reduce CH4 emissions) to blocking may be critical.

13. The way in which grouse moors are managed may have a role to play in moderating downstream flooding. Drainage has been found to have both positive and negative impacts on flood peak. Drain blocking raises the water table and slows down the water discharged through the drainage network, perhaps by reducing connectivity between drains and allowing water to escape from the drainage network and flow overland. The movement of water across a bog surface typically depends on slope, water table and surface vegetation. Water moves more rapidly over smooth surfaces (e.g. bare peat) and more slowly over rough surfaces (e.g. moss dominated). Land management that results in the loss of, or a change to, the surface vegetation (e.g. through drainage, grazing or burning), may increase or decrease overland flow. It is difficult to disentangle the multiple and interacting effects of grazing, burning, drainage and habitat restoration on water flows without conducting further research at multiple scales.

14. Despite being one of the major upland land-uses, knowledge gaps make it difficult to fully assess the costs and benefits of grouse moor management to biodiversity and the wider environment. Fundamental information is lacking on some of the key management components and their interactions, as well as on regional variation and current trends in these (e.g. burn rotation lengths). Basic information is still required on several aspects of the biodiversity effects (e.g. how vegetation and invertebrate communities differ between grouse moors and other moors), whilst there remains a poor understanding of management effects on some key species and taxa (e.g. effects of rotational muirburn on Sphagnum spp). Much of the science on the wider environmental effects is in its early stages and considerable research effort will be required to gain a general understanding of how key managements, and variation in the way these managements are applied, affect factors such as C stores, DOC levels and water flows.

4

1. Introduction 1.1 Moorland habitats in the UK The uplands form the largest extent of semi-natural habitat remaining in the UK, encompassing over 25% of the country (Haines-Young et al. 2000). The bulk of this area is open moorland, comprising mainly blanket bog, dwarf-shrub heath, acid grassland and, less commonly, calcareous grassland habitats, and is usually found from the upper edge of the enclosed farmland up to the limits of the original climax upper tree-line and sometimes into the alpine zone. Dwarf shrub heath tends to be defined as occurring on mineral and thin (<0.5 m) peat soils, and is distinguished from grassland habitats by having at least 25% cover of dwarf shrubs. Typically, this habitat is characterised by the extensive cover, or dominance, of ling heather Calluna vulgaris (subsequently referred to as Calluna1

www.ukbap.org.uk/library/UKBAPPriorityHabitatDescriptionsfinalAllhabitats20081022.pdf#BB

), and sometimes other dwarf shrubs (e.g. blaeberry Vaccinium myrtillus and bell heather Erica cinerea), but vegetation composition can vary considerably according to such factors as soil wetness and management regimes. For example, wet heath may be dominated by mixtures of cross-leaved heath Erica tetralix, deer grass Trichophorum cespitosum, Calluna and purple moor-grass Molinia caerulea, and so may have close affinities to blanket bog vegetation. Blanket bog is generally defined by occurring on deeper peat (≥ 0.5 m), and so the distinction from wet heath may be arbitrary and ill defined in some instances. Typical species include Calluna, cross-leaved heath, deer grass, cottongrass Eriophorum spp and several of the bog moss Sphagnum species, although species composition is highly variable (

). In high rainfall areas, as well as at high altitudes and in the far north of Scotland, it is likely that some of these moorlands are of natural origin (Rackham 1986). However, most are habitats of anthropogenic origin, resulting from the clearance of woodland by the use of fire and grazing animals, a process which began in earnest in the Neolithic period and continued into more recent historical times (Stevenson & Birks 1995, Tucker 2003). Today, our moorland habitats are maintained through the combined effects of grazing by large herbivores and burning, being managed primarily for sheep Ovis aries grazing and for the sport shooting of red grouse Lagopus lagopus scoticus and red deer Cervus elaphus. Despite their anthropogenic origins, the UK’s moorland habitats are considered to be of high conservation value, with the widespread occurrence and frequent dominance of Calluna making them particularly distinctive. Several of the main vegetation communities are either virtually confined to Britain and Ireland, or are better represented here than elsewhere, and 13 of the communities are listed under the EC Directive on the Conservation of Natural Habitats and Wild Flora and Fauna (Thompson et al. 1995). Upland dwarf-shrub heath, active blanket bog and upland calcareous grassland are also UK BAP habitats (www.ukbap.org.uk/library/UKBAPPriorityHabitatDescriptionsfinalAllhabitats20081022.pdf#BB). Furthermore, UK moorlands hold important invertebrate and breeding bird assemblages, including a high proportion of bird species of moderate to high conservation priority in the UK, as well as several that are listed on Annex 1 of the 1 Calluna is used for specific references to the plant species but the generic term ‘heather moorland’ is retained

5

EC Birds Directive (Thompson et al. 1995, Brown & Bainbridge 1995, Pearce-Higgins et al. 2009). The period since the Second World War has seen substantial declines in the extent and perceived ‘quality’ of moorland habitats, as a consequence of various land-use changes. Losses of Calluna dominated moorland have been particularly severe, with an estimated reduction of 28% across Scotland between the 1940s and 1980s (Mackey et al. 1998), and of 22% across England and Wales over a similar period (Huntings Surveys 1986). These changes are the result of either afforestation with exotic conifers (Ratcliffe 2007) or conversion to grassland following either re-seeding at the moorland edge or the indirect consequences of grazing by sheep and (in the Scottish Highlands) red deer (Staines et al. 1995, Thompson et al. 1995, Mackey et al. 1998, Fuller & Gough 1999). Management for the sport shooting of red grouse (herewith termed grouse moor management) has been important in preventing further losses of Calluna, with losses to forestry and grassland between the 1940s and 1980s greater on areas where grouse shooting had ceased than on those where it had been retained (Robertson et al. 2001). In recent years, there has been concern over the condition of moorland habitats, and the extent to which they have failed to meet favourable condition status on designated sites, as defined by the Common Standards Monitoring (JNCC 2006, Crowle & McCormack 2009). Heavy grazing pressure and inappropriate burning have been identified as major factors in causing ‘unfavourable condition’ of these habitats on such sites (Williams 2006, Crowle & McCormack 2009, Scottish Natural Heritage 2010). 1.2 Grouse moor management Grouse moor management is one of the main land-uses on UK moorlands. Heather moorlands are the main habitat of red grouse, with Calluna being the species’ main food plant. The management of heather moorlands specifically for red grouse shooting began in the 1850s, with large areas devoted to this land-use by the early 1880s when driven shooting had become fashionable (Lovat 1911). Estimates of the extent of this land-use vary depending upon definitions and methods used. Complicating factors include the difficulty in determining the contribution of management for red grouse relative to that for red deer in some areas of the Scottish Highlands, and the fact that some of the associated managements extend onto adjacent non-Calluna dominated habitats. Existing published estimates suggest that the total area of moorland managed principally for this purpose is between 0.5 and 1.7 million ha (Bunce & Barr 1988, Hudson 1992), representing some 5 – 15% of the UK uplands, and about 20 – 60% of Calluna dominated moorland (Ball et al. 1983, Brown & Bainbridge 1995). However, estates on which any grouse shooting is conducted may cover as much as 3.8 million ha, not all of which will be heather moorland (Hudson 1992, 1995). Figures from the Moorland Association (for the area of ‘moor’ managed by its members - www.moorlandassociation.org/economics2.asp) suggest that in England

6

and Wales alone, the total area of land managed as grouse moor is at least 0.28 million ha. This represents c.50% of the combined area of dwarf shrub heath and blanket bog (the two habitats likely to have moderate to high Calluna cover) in the English and Welsh uplands (Haines-Young et al. 2000). A smaller proportion (but greater total area) of the Scottish uplands is managed as grouse moor than in England and Wales, but if it is assumed that 50% of dwarf shrub heath and blanket bog in the Scottish uplands is also grouse moor (i.e. a similar proportion as in England and Wales, and therefore a likely overestimate), then this gives a total grouse moor area of c.1.7 million ha. A questionnaire survey in 2009 to 267 estates in Scotland that have heather moorland, and were known or considered likely to have shot red grouse in the past five years, produced a 30% response rate, with the total area of moorland covered by the 92 responding estates being 0.55 million ha, of which 0.38 million ha was considered to be heather moor (Fraser of Allander Institute 2010). Therefore, the total area of grouse moor in the UK must be greater than 0.66 million ha, but seems likely to be less than 1.7 million ha. Further uncertainty exists over the extent to which the total grouse moor area is managed intensively as ‘driven’ shoots (i.e. where lines of beaters flush grouse over static guns – Tapper 1992), as opposed to ‘walked-up’ shooting or shooting over dogs (presumably requiring less intensive management). The 2009 questionnaire survey of Scottish estates found that approximately 80% of the heather moorland on the 92 responding estates was used for ‘driven’ shooting, with 41% used for ‘walked-up’ or other shooting, so that some land was used in both ways (Fraser of Allander Institute 2010). It is unknown whether these figures are biased by differing propensity of estates favouring different forms of grouse moor management to respond to the survey. Previously, much greater areas of land were managed for red grouse shooting, and the land-use has been in decline since the 1940s (Hudson 1992, Robertson et al. 2001). The extent of these declines has varied geographically and, in Scotland, they have been greatest in the west and in the far north (Robertson et al. 2001). In some other regions, there is less evidence of decline and even evidence of increases since the 1970s in parts of northern England (Yallop et al. 2006, Natural England 2009). For example, a recent analysis of aerial images of the North York Moors found that the area of new burn had increased from 58.6 km2 in 2006 to 66.9km2 in 2009 (Natural England, unpublished). Generally, there appears to be poor understanding of current trends in extent and intensity. The regions where grouse moor management remains most intensive are the South and North Pennines, the North York Moors, the Southern Uplands of Scotland, and the central and north-eastern Highlands (Fig. 1).

7

Fig. 1. The distribution of ‘strip’ burn intensity in Britain, after Anderson et al. (2009). The authors measured the extent of ‘strip’ burning at a 10 km resolution (mainly from 2005 – 2006 satellite images), and used the data to generate an index of grouse moor management intensity. Reprinted from Biological Conservation, Volume 142 (3), Anderson et al., Using distribution models to test alternative hypotheses about a species’ environmental limits and recovery prospects, pp488-499, 2008, with permission from Elsevier. Grouse moors are characterised by a number of managements that directly affect the moorland habitat and the associated fauna. Principal amongst these are the rotational burning of Calluna (rotational muirburn) and the control of generalist predators, primarily foxes Vulpes vulpes, mustelids, and various corvids (mainly carrion Corvus corone and hooded crows C. cornix). Muirburn improves the food supply of red grouse by providing a source of young Calluna, which is more nutritious and easier to access than old plants (Savory 1978). It is undertaken in small patches, or strips, in rotation, with the aim of creating a patchwork of different aged stands, so that feeding areas are close to areas of older, taller, Calluna that provide cover. Muirburn is permitted between 1st October and 15th April in the uplands of England and Wales (although farmers in Wales cannot burn after 1st April), and between 1st October and 30th April in Scotland. Patches of 0.5 – 1 ha are generally recommended (Defra 2007, www.moorlandforum.org.uk/assets/documents/3102008144745_muirburn_code_mar08.pdf), whilst the length of rotation is dependent on Calluna growth rates and the time taken to return to the pre-burn height. Classically, rotational muirburn on dwarf shrub heath produces a mosaic of different aged Calluna stands, and arrests the succession of Calluna as it progresses from pioneer (6 – 10 years after burning), through to the building and mature phases (13 – 15 and 20 – 25 years after burning, respectively), before becoming degenerate, approximately 30 years after burning (Gimingham 1975). Recommended rotations are often of 10 – 15 years (when Calluna is in the building phase and usually 20 – 30 cm tall), but may be over 20 years where

8

growth rates are low, particularly on deep peat sites (Muirburn Working Party 1977, Defra 2007). In reality, burning rotations may often be longer than this in some areas (Hester & Sydes 1992, Rhodes 1996, MacDonald 2000), although in parts of the English uplands they appear to be closer to the recommended lengths, with a trend towards shorter rotations in some parts in recent decades (Yallop et al. 2006, Natural England, unpubl. data). Importantly, burning as practised for red grouse differs from that undertaken to improve sheep or deer grazing, which involves burning larger areas, is not undertaken on a rotation and is more haphazard in its occurrence, and may not be aimed at maintaining Calluna (Tucker 2003). The control of generalist predators is undertaken to reduce predation on grouse and their nests, and aims to increase the harvestable surplus of birds available in the late summer and autumn (Hudson 1992). Predator control is generally undertaken throughout the year by game keepers using a wide variety of trapping and killing methods. Other managements that are, or have been, associated with grouse moors include: • Persecution (i.e. killing, nest destruction, nest-site removal and disturbance) of

predatory birds (Newton 1979, Etheridge et al. 1997), with all birds of prey having been fully protected by law since 1954 (except for sparrowhawk Accipiter nisus which received full protection in 1961).

• Influencing moorland grazing regimes to help maintain Calluna cover, by direct influence on sheep and deer numbers and indirectly through effects of rotational muirburn on the distribution of grazing pressures (Grant & Hunter 1968, Sotherton et al. 2009a).

• Drainage of wetter moors using open ditches (grips) to encourage heather growth (Hudson 1986) and, more recently, the blocking of such grips to restore wet habitat (Waddell 2008).

• Vaccination of sheep, ‘dipping’ of sheep to act as tick ‘mops’ and the culling of mountain hares Lepus timidus to reduce the incidence of louping ill (a tick-borne disease) in grouse (Hudson 1992, Game Conservancy Trust 2001, Laurenson et al. 2003, Newborn & Baines 2011).

• Provision of medication (via medicated grit and/or catching/dosing adult birds) to treat grouse for infection by the nematode Trichostrongylus tenuis (Hudson 1992).

• Construction of vehicular access tracks on moorland (Scottish Natural Heritage 2005, P. Leadbitter, pers. comm.).

1.3 Objectives of the review The first detailed studies of red grouse biology to inform grouse moor management were commissioned over a century ago (Lovat 1911), and long-term research into the population biology of red grouse (e.g. Jenkins et al. 1963, Hudson 1986, 1992, Watson & Moss 2008, Watson 2011) is a cornerstone of grouse moor management that has provided detailed insights into the role of both extrinsic environmental factors such as habitat condition, predation and parasitism as well as intrinsic population factors such as territoriality and social behaviour in driving the complex population dynamics of the species.

9

Our purpose is not to review again this comprehensive evidence base on the relationship between grouse moor management and the production of grouse, but to review the wider evidence base that may inform continuing debate over the impacts of grouse moor management on the upland environment beyond red grouse (e.g. Glaves & Haycock 2005, Yallop et al. 2006, 2009, Sotherton et al. 2009a, 2009b, Thompson et al. 2009). Traditionally, much of this debate has focussed on biodiversity effects (particularly those concerning the maintenance of dwarf shrub habitats and the associated bird populations - Brown & Bainbridge 1995, Robertson et al. 2001), but more recently wider environmental effects (e.g. carbon (C) store and water quality) have received increasing attention (Holden et al. 2007a, Yallop et al. 2009, Ramchunder et al. 2009). This current review assesses the scientific evidence for impacts of grouse moor management on both biodiversity and ecosystem services, and identifies key gaps in knowledge that prevent conclusions being reached. This will facilitate future policy and decision making, and determine where future research is most urgent. Specifically, we undertake this assessment for:

1. Moorland vegetation, invertebrates and birds. 2. C balance, water quality and water flows on moorlands.

In terms of biodiversity, plant communities, invertebrates and birds represent the groups for which UK moorlands are considered to be of greatest national and international conservation importance. Moorland assemblages of mammals, reptiles and amphibians are particularly impoverished and, with few exceptions (e.g. mountain hare), lacking species of conservation importance (Thompson et al. 1995, Tucker 2003). Nonetheless populations of some mammals, notably mountain hares and field voles Microtus agrestis are critical elements of the food chain in supporting top predators such as golden eagles Aquila chrysaetos, hen harriers Circus cyaneus and short-eared owls Asio flammeus (Watson et al 1993, Redpath et al. 2002b, Wheeler 2008) ,and are directly affected by aspects of grouse moor management such as grazing, burning and disease control (Hope et al. 1996, Evans et al. 2006, Laurenson et al. 2003, Amar et al. 2011). For wider environmental effects, the focus on carbon and water reflects services that have been highlighted as important in an upland context and for which some of the managements associated with grouse moors may have considerable influence (Tucker 2003, Yallop et al. 2006, 2009, Ramchunder et al. 2009). Throughout the review, the major focus is on those managements that are likely to have greatest and most widespread effects and which have been subject to most study. Grazing regimes are the exception in this respect, despite their potentially widespread importance. This is due in part to the lack of detailed information available on sheep and deer densities and grazing practices on grouse moors as compared with those on non-grouse moors, but also to the interactions that occur between grazing and burning, and which may make it difficult to distinguish the additive effects of grazing regimes on grouse moors over and above those of burning (see 3.2.2).

10

The assessment of whether or not particular managements are beneficial or detrimental is generally considered in the context of moorland habitats, as opposed to comparing the benefits of moorland relative to alternative upland habitats (e.g. scrub and woodland, or plantation forests). Thus, with a few exceptions (e.g. effects on black grouse Tetrao tetrix populations), we do not attempt to consider in any detail whether an absence of grouse moor management is likely to lead to a reversion of moorland to scrub, or woodland, and whether or not the consequences of such change would be beneficial or detrimental. Finally, the review does not consider in detail the wider impacts of the provision of medication to treat parasites and disease as we could find no evidence of relevant studies on grouse moors, other than on grouse themselves (e.g. Newborn & Foster 2002). 2. Literature sources Both published and unpublished (e.g. university theses and unpublished reports) information sources were used. Relevant literature was obtained; (i) by conducting searches of the ISI Web of Knowledge and Google Scholar search engines; (ii) by contacting other ecologists involved in moorland research; and (iii) from the authors’ existing knowledge of relevant studies. The reference lists of all papers, theses and reports found in this way were searched for any further relevant material. Additionally, several reviews of the effects of grouse moor management on specific aspects of moorland habitats, biodiversity and the wider environment have been produced over the past 10 to 15 years - e.g. concerning effects of predator control on bird populations, or of burning on moorland habitats (MacDonald 2000, Tucker 2003, Stewart et al. 2004a, 2004b, 2005, Glaves & Haycock 2005, Gibbons et al. 2007, Holden et al. 2007a, Smith et al. 2007, Holt et al. 2008, Ramchunder et al. 2009). In producing the current review, we have drawn heavily on the material contained within several of these existing reviews. 3. Grouse moor management and biodiversity 3.1 Assessing effects on biodiversity Costs and benefits to biodiversity can be measured in different ways, creating potential problems in determining the overall effects on biodiversity. For example, benefits could include increased species richness, increased abundance of rare or declining species, or of species critical to ecosystem function, or improvements to the perceived condition of the habitat (e.g. in terms of the abundance or diversity of diagnostic species, or of reductions in invasive species). Particular managements may have contrasting effects on these different biodiversity measures, and so produce net benefits by some measures and net costs by others. Moreover, the nature of the evidence base varies between the different main taxa considered in this review. Thus, much of the evidence base for moorland vegetation concerns effects on overall species richness or diversity, and on certain plant species or taxa of functional importance in moorland habitats. By contrast, the evidence base for moorland birds or mammals focuses on the impacts on individual species, particularly those of

11

conservation importance, with few studies considering management impacts on species richness or diversity. For invertebrates, there are few detailed studies of individual species of conservation importance. These differences in the nature of the evidence base mean that the assessment of biodiversity effects differs across the main taxa, with different biodiversity measures being the focus of attention. Thus, direct comparisons of effects between the different main taxa, and a simple, overall assessment of the overall effects on biodiversity is not possible. 3.2 Moorland Vegetation 3.2.1 Managements of relevance Of the managements associated with grouse moors, rotational muirburn is of most importance in influencing moorland vegetation, and therefore this section focuses on considering its effects. Moorland grazing, drainage and drain blocking will also have direct effects on the vegetation of grouse moors, and consideration is given to these managements too. 3.2.2 Grazing-related effects Grazing regimes on grouse moors are expected to differ from those on moors where management for grouse shooting is not a major objective, with the likelihood that grazing management on grouse moors will be more sensitive to the need to maintain Calluna cover (Sotherton et al. 2009a). This may involve relatively low sheep stocking densities and deer numbers on grouse moors, and sometimes removal of sheep during autumn and winter when Calluna is particularly susceptible to grazing effects (Grant et al. 1987). Under such management, Calluna is less likely to be replaced by grass swards (Anderson & Yalden 1981, Bardgett et al. 1995). Moreover, distribution of grazing pressure will be influenced by rotational muirburn, which can act to distribute sheep and deer more evenly across a moor (because of the resultant distribution of nutritious new growth from burning), hence reducing the chance of localised concentration of grazing and consequent loss of Calluna (Grant & Hunter 1968, Phillips & Watson 1995). However, loss of Calluna is exacerbated by the interaction of heavy grazing with ‘poor’ burning practices (Thompson et al. 1995), . and high levels of atmospheric nitrogen deposition (Alonso et al. 2001, Terry et al. 2004, Edmondson et al. 2010), and outbreaks of infestation of Calluna by heather beetle Lochmaea suturalis can sometimes initiate transition from heather-dominated to grass-dominated vegetation covers (Berdowski 1987). This illustrates the challenges that may be involved in diagnosing the specific role of different grazing regimes on grouse moors and non-grouse moors. In Scotland, the reduced losses of heather moorland to grassland on grouse moors compared to other moors suggests that grazing regimes on grouse moors tend to differ from those on other moors in the ways described above (Robertson et al. 2001, 1.1). However, substantial losses of heather moorland to grassland have occurred on some grouse moors (Thirgood et al. 2000a), whilst past surveys of grouse moor estates in Scotland and northern England showed that over 80% kept sheep on the moor during the winter months (Hudson 1992). Therefore, at a broad level it seems that although a main effect of grazing management on grouse moors has been to

12

reduce losses of heather moorland, the full extent of differences in grazing regimes on grouse moors and other moors, and how these differences have varied with changes in the size of sheep and deer populations remains unclear. This lack of detailed information on the differences in grazing regime, combined with the fact that grazing and burning are closely interlinked and will interact in their impacts on vegetation (Hobbs & Gimingham 1987, MacDonald 2000, Tucker 2003), prevents a more detailed assessment of likely grazing-mediated effects of grouse moor management on moorland vegetation. 3.2.3. Factors influencing the response of moorland vegetation to fire Whilst the main aims of rotational muirburn are to ensure an abundance of young Calluna, create structural diversity and maintain Calluna dominance across the moor, the exact responses of moorland vegetation to burning are highly variable and affected by an array of factors (Fig 2). These are summarised in detail by Tucker (2003) and can be assigned to three main categories – i.e. pre-fire vegetation composition and age, fire severity (i.e. the extent to which a site is altered or disrupted by fire2

) and post-fire management regimes.

2 As defined in the Global Fire Monitoring Centre website: http://www.fire.uni-freiburg.de/literature/glossary.htm

13

Fig. 2. Key factors determining the outcome of vegetation re-establishment after a fire and their interrelationships, after Tucker (2003). Reprinted from English Nature Research Report (No. 550), Review of the impacts of heather and grassland burning in the uplands on soils, hydrology and biodiversity, with permission from Natural England. The characteristics of the vegetation that are important in this respect include the extent to which there are species that are fire-sensitive (i.e. with no means of surviving the fire or rapidly regenerating afterwards) or resistant, the extent to which species regenerate by vegetative means or from seed, and the amount of woody, flammable, material present in the sward. The presence and condition of the moss layer may also be important in influencing fire impact (Davies et al. 2010). Fire severity is in itself also determined by species composition and stand age (since this will affect fire temperature and intensity), along with such factors as wind speed, soil

14

and vegetation moisture levels, and method of burning, i.e. head-fire or back-fire. Head-fires involve burning with the wind, so creating faster, higher intensity fires, which consume less fuel, whilst back-fires burn against the wind, creating slower fires, where heat production is concentrated closer to the ground, so that more fuel is consumed. In relation to post-fire management, grazing is generally of greatest importance, with most moors having either wild mammalian herbivores (hares and deer) and/or domestic livestock present. As described above (3.2.2), the effects of burning and grazing on moorland vegetation are closely interlinked, because burning produces a source of young, nutritious, plant growth causing herbivores to concentrate their grazing on these areas (Hobbs & Gimingham 1987). Although the interactions between burning and grazing have not been quantified (MacDonald 2000, Stewart et al. 2004a, 2004b), it is clear that sufficiently high grazing levels on recently burnt areas will cause a shift from Calluna to grass dominance (Thompson et al. 1995), hence directing the succession away from the re-establishment of the pre-burn vegetation (Tucker 2003). Other managements, such as drainage, may also affect the post-fire succession, but little or no information appears to be available on their interactions with burning (Tucker 2003, Stewart et al. 2004a, 2004b). Additionally, site characteristics (e.g. soil type, topography), time of year and weather also affect vegetation response to fire, and interact with the above factors in producing this response. Overall, the wide range of factors and their interactions that may affect vegetation response to burning make it difficult to predict post-fire vegetation succession in many situations, with the exact direction of post-fire succession having the potential to be highly variable according to these different factors (Hobbs & Gimingham 1987). These sources of variation, along with issues of frequency and scale of fire, may often be overlooked when considering the effects of fire (MacDonald 2000), but they are of fundamental importance in drawing conclusions and in considering the extent to which the findings from particular studies can be generalised more widely. 3.2.4 Post-fire vegetation succession Fire affects the moorland vegetation community in numerous ways, including altering microclimates and soil nutrient balance, facilitating some species (e.g. by stimulating seed germination, or providing bare ground to colonise), and, perhaps most obviously, by killing or reducing species that are sensitive to fire (Mallik & Gimingham 1983, 1985, Mallik 1986, MacDonald 2000, Neimeyer et al. 2005, Harris et al. 2011). As expected for a habitat that has been created and is maintained to a large extent by fire, most of the characteristic and widespread plant species possess some form of fire survival mechanism (Hobbs et al. 1984). For example, the above ground buds of several species are protected (e.g. by sheathing leaves in the case of purple moor-grass, or by woody bark material and litter in the case of Calluna), whilst other species have underground rhizomes (e.g. blaeberry). A number of plant species (e.g. dwarf birch Betula nana, Lycopodium spp and Sphagnum spp) are often presumed to be susceptible to fire even though the evidence suggests that they may survive some fire events. According to MacDonald (2000), scrutiny of the available literature, together with detailed observation of fire events, indicates that such species may often survive vegetatively, following ‘normal’

15

management fires of relatively low severity. However, other species with shallow rooting systems, no rhizomes, creeping or low growing surface stems, and little capacity for re-sprouting from stem bases, are particularly vulnerable, even to fires of low severity – e.g. crowberry Empetrum nigrum and juniper Juniperus communis (MacDonald 2000). For several species that survive vegetatively, the extent of vegetative re-establishment declines with age (Hobbs & Gimingham 1984a, Davies et al. 2010, Fig. 3), partly as a consequence of a reduced ability of older plants to produce vegetative sprouts (Kayll & Gimingham 1965), and in the case of Calluna at least, also because there are fewer stems from which to regenerate in older, degenerate, stands (Miller & Miles 1970). Plants may also regenerate from seed sources following fire, and the soil seed bank will be of particular importance for species unable to survive fire vegetatively. Calluna is well known for forming a large, long-lived, seed bank (Hill & Stevens 1981), and whilst many other moorland species also form seed banks, they are rarely as long-lived (Hobbs et al. 1984). Studies on species-rich heathland in northeast Scotland found a marked reduction in the viable seeds of most species in stands older than 16 years (Mallik et al. 1984), and the average number of species germinating from soil samples declined from c.7 to c.2.5 with increasing stand age, up until c.30 years, after which there was an increase to c.4.5 in 40 year old stands (Hobbs et al. 1984). However, for Calluna and pill sedge Carex pilulifera, seed density increased with stand age, being highest in mature and degenerate stands of 25 – 40 year old Calluna, whilst the viable seeds of three other species (wavy hair-grass Deschampsia flexuosa, bell heather and slender St John’s-wort Hypericum pulchrum) increased in degenerate Calluna stands, following declines to low levels during building and mature stages (Mallik et al. 1984). For some species (including juniper) there is a short-lived seed bank only, or none at all (MacDonald 2000, Tucker 2003), and dispersal of seed into recent burns from surrounding areas may be limited (Mallik et al. 1984).

Fig. 3. The timing of critical life history events for major species present in species-rich heath in Scotland, after Tucker 2003, redrawn from Hobbs et al. 1984 (Journal of Ecology 72, pp 963-976).

16

Symbols: (o) time of disturbance; (p) time at which propagules are available on site; (m) time at which reproductive maturity is reached; (l) local loss from the community; (e) local extinction from the community. Calluna V represents persistence through vegetative regeneration; Calluna S represents persistence through seed germination. Reprinted from Tucker (2003), Review of the impacts of heather and grassland burning in the uplands on soils, hydrology and biodiversity. English Nature Research Report (No. 550). Reproduced with permission from Natural England and John Wiley & Sons Ltd. However, even where seed banks persist in older stands this may not guarantee successful or rapid regeneration. For example, Calluna seedling production one year after burning of Calluna heath overlying peaty soils was strongly influenced by post-fire substrate type, being extremely poor where live or dead pleurocarpous moss mats persisted after burning, somewhat higher where litter persisted and more than twenty times higher where only bare peat remained (Davies et al. 2010). Post-fire substrate type was, in turn, governed largely by stand age (moss mats being greatest in mature stands) and the moisture content of the moss/litter layer, as opposed to intrinsic fire characteristics and behaviour. Thus, post-fire Calluna regeneration is likely to be patchy in mature (and older) stands where there is little capacity for vegetative regeneration, leading to modified trajectories of vegetation succession, despite high seed densities (Davies et al. 2010). As a consequence of the above factors, the composition and age of the pre-fire vegetation are the main factors determining post-fire succession, provided;

(i) fire severity is within the limits often recommended for rotational muirburn (e.g. “quick, cool burns”, removing the dwarf shrub canopy but leaving behind some “stick” and not damaging the moss or litter layer or exposing bare soil surface – Defra 2007), or which may be considered ‘typical’ (e.g. as per the proxy measures of severity given by Davies et al. 2010), and (ii) grazing levels are not so high as to re-direct succession away from the pre-fire vegetation (Fig. 4).

The recommended muirburn practice is usually to burn during the building phase of Calluna and, under these circumstances, most regeneration will usually be by vegetative means, because vegetative re-establishment occurs and seedlings are out-competed by vegetative growth (Mallik & Gimingham 1983, Hobbs et al. 1984, Allchin et al. 1996, Davies et al. 2010). Under such conditions on species-rich heathland in northeast Scotland, the pattern of post-fire succession is for an initial rapid growth of grass and forb species, together with pioneer mosses and lichens, over the first two to three years (Hobbs et al. 1984). This is followed by a period of dominance by bell heather and bearberry Arctostaphylos uva-ursi, during which the abundance of grasses and herbs decline, with Calluna eventually reaching dominance after 8 – 10 years, at which time other species decline, pioneer mosses and lichens begin to disappear, and pleurocarpous mosses increase. By the time Calluna has reached the mature stage, only a few grass and herb species remain, but if left to enter the degenerate stage, gaps appear in the canopy, providing openings for recolonisation by some species (Hobbs et al. 1984). However, where muirburn management is particularly efficient and burning is controlled in such a way as to produce rapid vegetative regeneration, Gimingham (1995) notes that the pioneer phase may be virtually by-passed, with re-establishment of a young building phase stand one to two years after burning.

17

Where burning occurs in older, degenerate, Calluna stands the successional process is different (Hobbs & Gimingham 1984a, Fig. 4). This is because fewer species are present in degenerate stands, and many of these (including Calluna) are beyond the age at which they regenerate vegetatively (Fig. 3). Therefore, the re-establishment of many species (including Calluna) is from seed sources and is slower, and may be highly variable, according to the post-fire substrate (Davies et al. 2010). This may mean that, following burning, bare ground persists for longer, the phase of dominance by lichens, pioneer mosses and grasses is longer and, where present, aggressive rhizomatous species (e.g. blaeberry, cowberry Vaccinium vitis-idaea and bracken Pteridium aquilinum) prevent Calluna from becoming dominant (Hobbs & Gimingham 1984a).

Fig. 4. Examples demonstrating the effects of stand age at time of burning on species richness two years after fire (A) and initial Calluna dominance (B). Data from species–rich heathland in A, and from species-poor heathland in B. In B, P = pioneer (6 years), B = building (10 years), M1 = mature (15 years). Redrawn from Hobbs & Gimingham 1984. Reprinted from Journal of Ecology, Volume 72, Hobbs & Gimmingham, Studies on fire on Scottish Heathland communities: II Post fire vegetation development pp 585-610, 1984, with permission of John Wiley & Sons Ltd.

A

0

10

20

30

0 10 20 30 40

Stand age at time of fire (years)

Spe

cies

(num

ber)

B

18

The successional patterns described above refer to dry heath, where most detailed studies of the effects of fire on moorlands have been carried out. Less work has been undertaken on wet heath and blanket bog, with much of what is known on the effects of fire in these habitats derived from non-grouse-moor situations (e.g. Currall 1981, Hamilton 2000). As on dry heaths, the pre-fire vegetation community is likely to be the main determinant of post-fire succession, provided fire severity and grazing levels are sufficiently low (Tucker 2003). Calluna productivity and growth is lower than on dry heaths (Forrest & Smith 1975), so that it takes longer to attain dominance following burning (e.g. 11 – 15 years on blanket bog in the North Pennines and on wet heath on The Isle of Skye, northwest Scotland – Rawes & Hobbs 1979, Currall 1981, Hobbs 1984). Studies of wet heath on the Isle of Skye suggest three typical phases to post-fire development, with an initial graminoid phase, dominated by species with rapid recovery or ability to colonise bare ground (e.g. star sedge Carex echinata, common cottongrass Eriophorum angustifolium, heath rush Juncus squarrosus, mat grass Nardus stricta and Festuca spp). The second phase, which may last eight to 12 years, is characterised by dense growth of purple moor-grass or deer grass, causing a reduction in bare ground and colonization by new species, and elimination of shade-sensitive species. Cross-leaved heath often reaches a peak in abundance during the second phase, but Calluna and some other dwarf shrubs do not attain dominance until the third phase, typically at c.15 years, when the cover of graminoids declines and mosses develop under the canopy (Currall 1981). On blanket bogs, the Calluna growth cycles typical of heathland do not usually occur, because of layering from the burial of stems under Sphagnum, so that the age structure of the above ground stems is uneven and degeneracy does not occur (Rawes & Hobbs 1979, MacDonald et al. 1995). Longer burning rotations of c.20 years are required to maintain Calluna dominance in these habitats, with 10-year rotations causing a shift to dominance by hare’s-tail cottongrass Eriophorum vaginatum in blanket bogs (Currall 1981, Hobbs 1984). Burning may also cause a shift towards purple moor-grass dominance on wet heaths where this species is present (Ross et al. 2003). One regional-scale study has been undertaken on the effects of rotational muirburn on plant communities, using a chronosequence approach to determine changes in plant species cover in relation to time since burning (Harris et al. 2011). This work was carried out on predominantly blanket bog habitats (as determined by peat depth) in the Peak District, and demonstrated that Calluna was the only species to show an increasing response with time since burning, reaching 50% cover after c.12 – 13 years and > 90 % cover after c.30 years. All other species showed an increase immediately after burning but then decreased, with few species other than Calluna present in stands older than 20 – 25 years (Harris et al. 2011). The sites on which this study was undertaken had particularly low plant species richness, as is often typical of Peak District moorlands, whilst Sphagnum mosses (a key peatland group – 3.2.6) appeared to be rare or absent (Harris et al. 2011). These factors may influence the extent to which the findings apply to blanket bogs in other parts of the UK. On all moorland habitats, particularly severe fires are likely to produce different effects from those described above. Such fires may burn into the mor humus layer and are more likely to kill off the shoot bases of species such as Calluna or kill off

19

rhizomatous species, and destroy seed banks (Schimmel & Granström 1996, Tucker 2003). Ignition of peat soils may occur in some instances, causing wider damage. These effects sometimes arise during wildfires on moorlands (Maltby et al. 1990, Legg et al. 1992) but the extent to which they arise during rotational muirburn is unclear. Extended periods of drought and low soil or moss/litter moisture levels are likely to be particularly important in facilitating such severe fires (Davies et al. 2010), whilst the amount of woody (and hence highly combustible) material in the vegetation is of particular importance in determining the temperature of moorland fires. Highest fire temperatures occur in mature stands of c.30 year old Calluna, with temperatures in degenerate stands also being high but more variable, reflecting the patchier occurrence of woody material (Kenworthy 1963, Hobbs & Gimingham 1984b). Thus, in older stands, the effects of reduced species richness and lower potential for vegetative regeneration on post-fire succession may be compounded by factors such as the presence of more extensive moss mats and a greater likelihood of more severe fires. Unfortunately, little information is available on actual burning practises on grouse moors, with little known about factors such as typical fire severity, or actual rotation lengths in many areas. Certainly, in some areas rotational muirburn often results in the exposure of large amounts of bare ground (suggesting relatively severe fires – Yallop & Clutterbuck 2009), but (at least in terms of heather recovery) this is not always detrimental, because the moss or litter layer remaining after quick, cool, burns in mature and degenerate stands impairs seedling establishment (Davies et al. 2010). The extent to which rotation lengths diverge from the recommended practise across grouse moors is often unknown (1.2), and appears to be regionally (and presumably also temporally) variable. Thus, during the post-war period in the Grampian and Borders regions of Scotland an average of only 1 – 2% of heather was burnt per year (as opposed to an ideal of 6 – 10%), suggesting long rotations, although not all of this heather will have been under active grouse moor management, and these averaged figures presumably obscure much variation (Hester & Sydes 1992). More recent data from a wide range of moorlands in northern England suggest rotations range from 7 – 25 years on dry heath, 9 to 25 years on wet heath and 10 – 35 years on blanket bogs (Natural England, unpubl. data). These figures appear to be a reasonable match to the rotations generally recommended for these habitats, but there is considerable variation, with some areas of blanket bog being burnt on short rotations. Other data from northern England suggest a trend towards shorter rotations since the 1970s (Yallop et al. 2006). Reliable information on typical burning practises is critical to understanding the extent to which the patterns of succession described above apply to the actual situation on grouse moors. 3.2.5 Effects of rotational muirburn on species richness and diversity Moorlands tend to be habitats with inherently low plant species diversity compared to semi-natural, or natural, upland scrub and woodland, although those habitats that occur on calcareous soils (e.g. calcicolous grasslands, flushes and fens) have substantially greater species diversity than those occurring on acid soils (MacDonald 2000, Harris et al. 2011). At a broad level, dwarf-shrub heaths (probably the single most important habitat component of many grouse moors) tend to have higher species diversity and a greater number of “rare” species than several other moorland

20

habitats that occur on acid soils (MacDonald 2000), indicating that there may be benefits in this respect from land-uses that help to maintain this habitat. However, few, if any, data are available to allow a rigorous and direct assessment of how management for grouse shooting affects plant species diversity compared to other moorland land-uses (e.g. management for deer stalking or livestock production, or management abandonment). The information detailed above (3.2.4), suggests that relatively frequent and regular burning (i.e. 10 – 15 year rotations on many dry heaths) can be important to maintaining plant species diversity on dwarf-shrub heaths. This is because of the general decline in species diversity that occurs as stands age and Calluna becomes increasingly dominant (up until the degenerate phase) and also because of the patchwork of pioneer and building phase stands (with differing species compositions), which is produced as a consequence of muirburn management. However, these potential benefits to species diversity may not occur if the severity of fires on grouse moors causes elimination or substantial reductions in the abundance of certain species, or if rotations are longer than recommended, so that many vegetation stands are burnt in the mature phase rather than the building phase, with the consequent effects on post-fire succession. Where management enables rapid re-establishment of Calluna (essentially omitting the pioneer phase – 3.2.4) low diversity will result (Gimingham 1995). The extent to which rotational muirburn regimes retain stands of degenerate Calluna will also be important, as some moss and lichen species are restricted to such stands, so that retention of some areas of degenerate Calluna will add to overall diversity (MacDonald 2000). In addition to these factors, it is apparent that the effects of rotational muirburn on blanket bogs (and to a lesser extent wet heaths also) are little studied, so that there is greater uncertainty on the effects on plant species diversity in these habitats. A review of quantitative studies considering the effects of burning on dry dwarf-shrub heath in Britain and Ireland identified five studies, comprising 13 different data sets, which included appropriate controls or unburnt comparators and which were suitable for inclusion in a meta-analysis (Stewart et al. 2004a, 2005). Overall, this analysis detected no significant effect of burning on either species richness or diversity (as measured by Simpson’s index, to account for the equitability of species’ abundances), with some studies detecting increases in richness and diversity in burnt stands (Welch et al. 1994, Stevenson et al. 1996) and others decreases (Elliot 1953, Hobbs & Gimingham 1984a). However, the analyses did detect an effect of stand age at the time of burning on the response of species diversity (and to a lesser extent species richness) to burning, with post-burn diversity decreasing as stand age at the time of burning increased, as expected from existing knowledge of post-fire succession on heathlands - see 3.2.4 (Stewart et al. 2004a, 2005). Although the review by Stewart et al. (2004a, 2005) provides a valuable assessment of the quantitative evidence available on the effects of burning on plant species diversity on dry dwarf-shrub heath, there are considerable limitations to this evidence base. Thus, the geographical distribution of the studies is restricted, with all but one of the datasets derived from the eastern Highlands of Scotland, whilst the different studies generally provide insufficient information to enable potentially

21

important effects of site conditions (e.g. soil moisture levels) or other managements (e.g. grazing regimes) to be taken into account. Perhaps most importantly, four of the five studies (including 11 of the datasets) are short-term and consider the effects of a single burning event only. The sole exception used a combination of palaeoecological methods and aerial photographs to determine the fire histories at five sites in Tayside, south-eastern Highlands (Stevenson et al. 1996). Here plant species composition on stands with a history of serial burning (burnt two to three times in the previous 40 - 50 years) was compared with that on nearby stands not subject to burning during the past 70 or more years. Species richness was greater on each of the three stands that had been burnt recently (3 - c.4 years previously) than on the associated unburnt stands (with a mean of 7.9 species per 0.25 m2 quadrat compared to 5.3 across the three sites). However, it was lower on both stands that had been burnt c.19 – 20 years previously (now in the mature phase of Calluna growth), than on the associated unburnt stands (with a mean of 4.9 species per 0.25 m2 quadrat compared to 5.8 across the two sites), hinting that serial burning may eliminate some species characteristic of the later stages of stand development. A study of lichen species diversity on dry dwarf-shrub heath, undertaken at Mar Lodge in the Scottish Highlands, examined spatial variation in species number across muirburn patches burnt 1 - 18 years previously, as well as in a sample of unburnt areas. This demonstrated that the number of species present in burnt stands was low immediately after burning but increased to a maximum 10 - 15 years after burning, declining thereafter (Davies & Legg 2008). Species richness was highly variable in unburnt areas (being affected by the density of vegetation in such stands, with pleurocarpous mosses replacing lichens under very dense vegetation) but, overall, was lower than in areas burnt 10 - 15 years previously. However, lichen species composition also varied with time since burning, with the numbers of corticolous (i.e. bark-living) species being highest in unburnt stands, so that diversity across a moor is likely to be maximised by having stands of a range of ages, including some areas set aside from burning (Davies & Legg 2008). Few data are available to assess the effects of burning on plant species diversity on blanket bogs. Extensive studies from across five moors in the Peak District demonstrate that rotational muirburn is important in maintaining plant species diversity on blanket bogs in this region of the UK, with few species other than Calluna present in stands of > 30 years of age (Harris et al. 2011, 3.2.4). However, the particularly low plant species richness on the Peak District blanket bogs (including the near absence of Sphagnum mosses from many parts) (Tallis 1997) may mean that these findings are not applicable to blanket bogs elsewhere in the UK (3.2.4). Elsewhere, there were no apparent differences in vascular plant species richness between unburnt plots and plots burnt 1 - 2 and 7 - 8 years previously in County Antrim, Northern Ireland (McFerran et al. 1995). This was also the case for the different treatments on the Hard Hill3

3 For further details of the experimental set up at Moor House, see figure prepared by John Adamson (CEH)

plots at Moor House North Pennines, 18 – 19 years after first burning; these plots comprising four replicates (each 10 x 30 m in area) of 10 and 20 year burning rotations and of plots unburnt since first burn, each

http://www.ecn.ac.uk/IMAGES/SITES/MOO%20BURNING/Layout.pdf)

22

under regimes of ‘normal’ and no sheep grazing (Rawes & Hobbs 1979). However, subsequent comparisons of the two different burning rotations under the sheep grazing regimes, at 23 - 25 years after first burning, indicate a tendency for greater numbers of lichen species to occur in the 20-year rotation treatment (Hobbs 1984). Data from unreplicated management trials undertaken elsewhere at Moor House indicate that burning followed by relatively heavy sheep grazing resulted in a general decline in plant species richness after seven years, relative to the changes observed under a heavy grazing regime without burning (Rawes & Hobbs 1979). This appeared to result primarily from a decline in bryophyte species on the burnt plot as well as an increase in vascular plant species on the unburnt plot. The findings from the studies detailed above describe the effects of burning on plant species richness, or diversity, at the stand scale – comparing burnt and unburnt ‘patches’ (essentially equivalent to α-diversity – Whittaker 1972). However, when considering the overall effects of muirburn on plant species diversity, it is also necessary to consider effects at larger scales and account for the variation produced by creating a mosaic of different successional stages, and hence different species compositions (essentially equivalent to β-diversity – Whittaker 1972). Therefore, if rotational muirburn at least maintains plant species diversity at the stand scale, then it should produce an overall increase in species diversity across a moor. This is illustrated by the Mar Lodge study of lichen species richness where both the total number of species present and species composition vary with time since burning, so that increases in species richness are greater at this larger scale (Davies 2001). Although no studies appear to have measured the effects of rotational muirburn on overall plant species richness or diversity at such larger scales, the material reviewed here suggests that, whilst vegetation effects may be highly variable, in many instances this management should promote plant species richness and diversity. This conclusion is difficult to reconcile with the widely reported perception that rotational muirburn creates low plant species diversity and Calluna monocultures (Gimingham 1995, Davies et al. 2008, Yallop et al. 2009). This discrepancy could arise if such perceptions tend to be derived from observations of sites with inherently low species diversity (irrespective of burning management), or if the effects of rotational muirburn are often confounded with other managements causing low species diversity (e.g. drainage and heavy grazing), when making such observations (as implied by MacDonald 2000). However, it is also possible that the findings from the quantitative studies reviewed here on the effects of rotational muirburn on vegetation have limited applicability to actual grouse moors, with many of the relevant studies deriving from a restricted geographical area and being based upon burning that is undertaken for experimental purposes, rather than by gamekeepers managing active grouse moors. Insufficient information is available to determine which, if any, of these explanations is most likely. Furthermore, information on the extent to which degenerate stands of Calluna generally occur on grouse moors (likely to contribute to overall plant species diversity on the moor) is lacking. 3.2.6 Effects of rotational muirburn on habitat condition Although species richness and diversity are important aspects of biodiversity, some of the most important measures of moorland plant biodiversity are those concerned

23

with the status of individual species or assemblages that are considered to be of high conservation or functional importance, or which are invasive. Such species or assemblages may be used to monitor the state of, and changes in, habitat ‘quality’, and hence may provide measures of the ecological condition of moorland habitats. Calluna and other dwarf shrubs are the main diagnostic species of dwarf shrub heathland, whilst they are also typical on blanket bog habitat. Clearly, rotational muirburn will generally act to promote Calluna cover because it is designed to do so. Exceptions may occur where post-fire grazing regimes are too high, fires are too frequent, or old Calluna stands are burnt. These situations do arise on some grouse moors but, although the extent to which this happens is unclear, this management regime appears to have been more successful than others in maintaining Calluna cover (Robertson et al. 2001). Several other dwarf shrub species are also likely to benefit from rotational muirburn, or are at least maintained under such management regimes, although crowberry is susceptible to fires (Table 1). Amongst species of conservation priority that are associated with moorland habitats, juniper (a UK BAP species) is notable for having no fire survival mechanism (3.2.4, Table 1), so that it is likely to be vulnerable to muirburn (Hobbs et al. 1984). After most fires, regeneration of juniper will need to be from seed, with seedling establishment unlikely where vegetative re-growth of Calluna is rapid, whilst the species’ poor seed dispersal limits its ability to colonise burnt areas where colonisation by heather is by seed, and therefore, slower (Hobbs et al. 1984). However, the effects of muirburn may be complex, even for such a fire-sensitive species, and the fact that the species often requires disturbed ground for regeneration may mean that muirburn could, in some circumstances, provide opportunities as well as threats (Thomas et al. 2007). Thus, no clear correlation has been demonstrated between burning management and juniper occurrence, whilst in the Scottish Highlands the species may be most abundant in areas where management by burning is common (MacDonald, pers comm., after Tucker 2003). The recent decline of juniper in England is attributed more to heavy grazing pressure and consequent low recruitment and establishment of seedlings, than to effects of muirburn (Drewitt & Manley 1997). On blanket bogs, and to a lesser extent wet heaths, the effects of rotational muirburn on Sphagnum mosses are critical from the perspective of habitat condition and function; Sphagnum being major peat forming species, which are particularly effective at holding water and absorbing nutrients deposited on the bog surface (Verhoeven & Liefveld 1997, Rydin et al. 2006). Sphagnum mosses have often been considered particularly susceptible to detrimental effects of burning, either through direct effects of fire damage, or through subsequent effects of exposure to drying of the peat as a consequence of fire (Pearsall 1950, Ratcliffe 1964, Rowell 1990). However, other observational evidence suggests that burning sometimes has little impact on Sphagnum (Daniels 1991, Barkman 1992), whilst MacDonald (2000) considers that it may be relatively well able to survive “fairly low intensity” management fires vegetatively. For example, in the boreal peatlands of Canada, complete vegetation recovery occurred within a few decades of peat surface fires, which were probably more severe than most moorland management fires in the UK, given that Sphagnum

24

apparently had to re-colonise bare peat (Kuhry 1994). It is also possible that a lack of burning may lead to declines in Sphagnum abundance in some situations, as a result of the accumulation of standing dead material from grasses and sedges (Chapman & Rose 1991). Studies in northwest Scotland found that Sphagnum recovery following burning was variable, but often considerable within two years after burning, with indications that fire may encourage Sphagnum in some cases (Hamilton 2000). The extent of damage to Sphagnum in these studies increased with the amount of overlying Calluna, presumably because this created hotter and more severe fires. However, these studies were conducted in an area with little or no grouse moor management where burning is undertaken primarily to improve grazing for livestock, and where management fires usually extend over larger areas than on grouse moors. Furthermore, these findings derive from single fire events, rather than from the cumulative effects of rotational muirburn. Data from the Moor House Hard Hills plots (3.2.5) indicate that, 23 - 25 years after first burning, Sphagnum cover was higher in the 10-year burning rotations under sheep grazing than in the 20-year rotations under sheep grazing in three of the four replicates of these treatments, no Sphagnum being recorded in the other replicate (Hobbs 1984). Whilst this might suggest benefits to Sphagnum from short rotation burning, the data presented do not allow the actual trends in Sphagnum cover to be determined on the different treatments over the 23 – 25 years. In an unreplicated management trial at Moor House, Sphagnum cover declined equally over a seven year period on an area where burning was followed by relatively heavy grazing as on an area where heavy grazing was applied without any burning (Rawes & Hobbs 1979). Rotational muirburn is also important in terms of its effects on invasive species, or on the spread of dominating, non-dwarf shrub species. Most obviously, rotational muirburn may be important in maintaining areas of moorland free from scrub and woodland, where grazing levels are low enough to allow the establishment and growth of seedlings (Gimingham 1995). Whether such effects represent costs or benefits to overall biodiversity is debatable, and is likely to vary between sites according to the biodiversity value of the moorland and woodland habitats in question, as well as the conservation priorities that are in place. However, muirburn is clearly important in maintaining moorland habitats in such situations. Inappropriate burning practices have been suggested as one of several causes of the bracken (Pteridium aquilinum) expansion on moorlands (Drewitt & Manley 1997). However, the extent to which rotational muirburn, as opposed to wildfires or burning for livestock production, is likely to contribute to any such effect is unknown, whilst evidence that the establishment of new bracken plants is encouraged by burning appears to be lacking (MacDonald 2000). Furthermore, management that maintains a vigorous cover of competing species (as rotational muirburn aims to do) tends to limit, rather than encourage, the spread of bracken, with one long-term study in the Quantock Hills, southwest England, demonstrating that dwarf-shrub heath was more likely to have been lost to bracken if it was not burnt between 1938 and 1987 than if it was burnt at least once during that period (Ninnes 1995). There is, however, strong evidence that burning can facilitate the

25

spread of purple moor-grass; a highly competitive, tussock-forming, deciduous grass that is resistant to fire and which has increased considerably in distribution and abundance across western European heathlands and moorlands in recent decades (Hansen 1976, Berendse et al. 1994, Chambers et al. 1999). Purple moor-grass regenerates rapidly following fire and may become dominant on wet heaths, with burning often leading to increases in its cover and eventual dominance, particularly if undertaken frequently or under wet conditions which prevent removal of the accumulated grass litter (Currall 1981, Ross et al. 2003). The extent to which rotational muirburn for red grouse management contributes to such detrimental habitat change is unknown and will depend upon the exact situations in which burning is being undertaken and the nature of the burning itself. One previous review has attempted to assess the effects of burning on the overall condition of moorland habitats. This was concerned with blanket bogs in Britain and Ireland, attempting to interpret the findings of relevant studies in terms of favourable condition criteria (JNCC 2004). Only eight studies, comprising 11 different datasets, were identified which provided quantitative assessments of effects on plant species composition and included appropriate controls or unburnt comparators (Stewart et al. 2004b, 2005). In contrast to the analogous review undertaken on burning and species diversity on dry dwarf-shrub heath (see 3.2.5), the data in this review were of insufficient quality to enable a meta-analysis, so that they were synthesised by ‘vote counting’ (Stewart et al. 2004a, 2004b, 2005). Overall, burning was found to: promote the dominance of a few species or a switch in dominance from ericoids to graminoids; increase the area of bare ground; result in decreased abundance of key species; result in minor changes in floristic composition. Of the 11 datasets, three provided evidence of degradation, five evidence of contrasting effects and three showed no evidence of degradation. Limiting comparisons to the higher quality datasets (i.e. randomized controlled trials as opposed to site comparisons) provided greater evidence for degradation (two of these five indicating degradation, and three showing contrasting effects). Across the full 11 datasets, evidence of degradation was essentially limited to those examining recently burnt areas (1 – 7 years), with the three datasets examining longer post-burn periods providing little evidence of degradation (one providing contrasting evidence and two providing no evidence). In addition, further analyses of data from the replicated experiment at the Hard Hill plots (included in the comparisons of Stewart et al.) have been undertaken, since the completion of this review. These show that 50 years after all plots were subjected to an initial burn, the above ground live biomass of dwarf shrubs and bryophytes was reduced on the 10-year rotation relative to the unburnt control (by 51% and 92%, respectively), whilst graminoid biomass increased on the 10-year rotation (by 88% relative to the unburnt control) (Ward et al. 2007). Stewart et al. (2004b, 2005) conclude that the weight of evidence indicates that burning either degrades blanket bog or has contrasting effects, and that, pending further research, burning on blanket bog and wet heath should normally be avoided to achieve or maintain favourable condition. However, the authors recognise that circumspection is required given the small sample size, variable timescales and problems in interpreting favourable condition. Additionally, four of the 11 datasets used in the review actually concerned wet heath and not blanket bog habitats, with

26

two datasets only considering rotational burning, and six deriving from the North Pennines and none from Scotland (where the largest expanses of blanket bog in the UK occur). Therefore, whilst this review represents the best quantitative evidence currently available on the effects of burning on the overall condition of blanket bogs, there are substantial limitations in the extent to which it can be assumed to reliably reflect effects of rotational muirburn on blanket bogs, and in the extent to which findings can be assumed as representative of blanket bogs throughout the UK. A summary of the impacts of burning management on some key moorland plant species is provided in Table 1.

27

Table 1. A summary of the impacts of burning management on key moorland plant species and taxa, after Tucker (2003). Reprinted from Tucker (2003), Review of the impacts of heather and grassland burning in the uplands on soils, hydrology and biodiversity. English Nature Research Report (No. 550). Reproduced with permission from Natural England and John Wiley & Sons Ltd.

28

3.2.7 Effects of drainage and drain blocking on moorland vegetation In past decades, large areas of blanket bog and wet heath were drained using open ditches (commonly known as grips) cut in rows, usually parallel to the contours (Stewart & Lance 1983). Hill farmers aiming to improve the ground for livestock grazing undertook much of this, with drainage in some areas beginning in the late 1940s and substantial agricultural grants available to facilitate such work by the 1970s and 1980s (Bowers 1985, I. Condliffe, pers. comm.). However, drainage was also undertaken for red grouse management, with 55% of a sample of grouse moor owners in northern England reporting they had drained part of their moors (Hudson 1984). Subsequently, potential benefits of this management to grouse via increased Calluna quality and cover have been considered to be outweighed by detrimental effects on wet areas, which may provide young red grouse chicks with important invertebrate food sources (Hudson 1986, Park et al. 2001). Some grouse moors may now be more likely to promote drain blocking (Hudson 2008), with several actively involved in conducting such management, although its adoption by grouse moors probably remains patchy and concentrated in parts of northern England, where peat restoration projects have overlapped most with grouse moors (Waddell 2008, Armstrong et al. 2009, Jonczyk et al. 2009). Two studies examining the effects of drainage on moorland vegetation on blanket bog and wet heath sites in the North Pennines found that detectable effects were relatively localised and, as expected, most marked downslope of drains (Coulson et al. 1990, Stewart & Lance 1991). Overall, drainage reduced the cover of species dependent on high water tables, notably cottongrass and Sphagnum spp, and tended to increase the cover of those with affinities to drier heathland (e.g. cloudberry Rubus chamaemorus). These effects were associated with a lowering of the water table, and relatively few effects were detectable beyond 4 m from the drain. Temporal data

29

from an experimental set of drains dug at 15 m intervals to a depth of 0.5 m (Conway & Millar 1960), demonstrated minor changes in Calluna cover only, with cover downslope of the ditch peaking after eight years but declining subsequently so that no statistically significant effects were detectable after 19 years (Stewart & Lance 1991). At this site, Sphagnum capillifolium (the most abundant Sphagnum species at the site) had been virtually eliminated from the drain edges at 19 years and had declined markedly in quadrats as far as midway to the next drain downslope, an effect that was still maintained 25 years later. Across a wider sample of drains, no statistically significant differences in Calluna cover could be found when comparing cover upslope and downslope of the drains, but the current year’s growth was significantly greater downslope, an effect that extended midway (10 m) to the next drain (Stewart & Lance 1991). In the second study, effects of drainage on plant species composition were more marked on drier, low altitude (c.450 m asl), sites on relatively shallow peat (250 – 270 mm) than on wetter, high altitude (1100 – 1500 m asl), blanket bog sites on deep peat (>1000 mm), demonstrating that effects vary according to site conditions (e.g. climate and topography), as well as in relation to the density of drains themselves (Coulson et al. 1990). At these sites, the main recorded change was a decrease in Calluna cover, and increase in grass cover, within 5 m downslope of the drain, whilst the nutrient content (nitrogen and phosphorus) of Calluna was greater 1.5 m downslope of the drain than at 1.5 m upslope of the drain on the low altitude sites. Importantly, Sphagnum cover was not recorded separately from that of other mosses in this study, and more detailed recording of plant species and taxa may have revealed greater effects of drainage. Wide-scale surveys at Lake Vyrnwy, Wales, found evidence for declines in the cover of deer grass, purple moor-grass and both cottongrass species up to 15 m from the drains, although declines in Sphagnum cover were more limited and localised (Wilson et al. 2010a). Relatively few studies have yet been undertaken on the effects of drain blocking on plant species composition on UK blanket bogs. Changes in the cover and frequency of plant species six years after blocking at a single drain in the Flow Country in northeast Scotland showed few differences to the changes recorded over the same period at an adjacent unblocked drain (Bottin 2002). However, other wider-scale studies in the Flow Country, comparing vegetation conditions along gradients from the drain edge up to 15 m out from the edge at blocked and unblocked drains, found evidence that blocking reduces the cover of vegetation characteristic of drier conditions and of peat degradation, whilst increasing the cover of peat forming species (Bellamy et al. 2011). Such effects were not ubiquitous at blocked drains, with indications that it may take several years for such effects to manifest. Other studies at Lake Vyrnwy, which compare changes in vegetation following drain blocking, obtained similar findings, with increases in the cover of Sphagnum mosses and other wet tolerant species following blocking, whilst species characteristic of drier conditions declined (Wilson et al. 2010a). 3.2.8 Summary and synthesis of effects on vegetation Grouse moor management has major effects on moorland vegetation. In the past grouse moors have been important in preventing greater losses of heather moorland to afforestation and conversion to grassland and, at a broad level, there is little doubt that their management regimes are more effective in conserving

30

dwarf shrub dominated habitats than are those associated with livestock farming and deer stalking. However, the more detailed effects of grouse moor management on vegetation, and the extent to which individual managements represent costs or benefits to the biodiversity value of moorland vegetation, are more difficult to determine. Rotational muirburn is the management of most importance in determining the vegetation on grouse moors, whilst drainage and drain blocking will also have direct effects on vegetation. The maintenance of Calluna cover on grouse moors is likely to be sensitive to grazing regimes and their interaction with factors outside immediate management control such as atmospheric nitrogen deposition and heather beetle outbreaks , but a detailed assessment of grazing-mediated effects of grouse moor management on moorland vegetation is difficult because information on differences in grazing regimes between grouse moors and other moors is limited, and grazing and burning effects are likely to interact. Rotational muirburn aims to maintain and promote Calluna dominance across a moor, although the exact response of moorland vegetation to burning is variable and determined by an array of factors, including vegetation composition and age, fire severity and post-fire management regimes. Findings from studies that have examined the effects of muirburn on plant species richness and diversity on dwarf shrub heath tend to suggest that rotational muirburn will produce an increase in these attributes. However, these studies have considerable limitations (e.g. restricted geographical distribution and poor coverage of blanket bog habitats) and may not be truly representative of the conditions that occur on managed grouse moors, especially in terms of severity of burning and burn rotation length. At the same time, the contradictory, but frequently reported, perception that rotational muirburn creates low species diversity and Calluna monocultures may arise if such perceptions are often derived from observations of sites with inherently low diversity (irrespective of management), or if the effects of rotational muirburn are often confounded with other managements causing low diversity (e.g. drainage and heavy grazing). Management regimes that combine ‘sensitive’ rotational muirburn (e.g. appropriate rotation lengths and fires that are not so severe as to burn into soils) with the retention of some stands of degenerate Calluna seem likely to produce maximum plant species richness and diversity on dwarf shrub heath, but the extent to which such regimes occur on grouse moors is unknown. A different situation may arise on many blanket bog habitats because the Calluna growth cycles typical of heathland do not generally occur (due to layering of stems), creating an uneven age structure where a decline in plant species diversity with age of vegetation (as observed in heathland) seems unlikely. In terms of the effects of rotational muirburn on plant species or taxa that are of particular conservation or functional importance on moorland habitats, the evidence base is again limited. For example, even for juniper (a species of high conservation importance, which has no fire survival mechanism), the full nature of the relationship with muirburn appears to be poorly understood, whilst studies demonstrating how rotational muirburn affects Sphagnum mosses (a key component of blanket bogs), and how this varies according to variation in muirburn regimes on grouse moors, are lacking. In relation to blanket bog habitats, insufficient data are available to enable a rigorous assessment of how

31

rotational muirburn affects their overall condition, although the available data indicate that detrimental effects are apparent mainly on more recently burnt areas, with little evidence of detrimental impacts in studies that were undertaken seven or more years after burning. Few studies are available on the effects of either drainage or of drain blocking on moorland vegetation. The latter management has been undertaken widely on blanket bogs in recent years, including on grouse moors, although the full extent of this activity on grouse moors is unclear. The most detailed study of drainage effects demonstrated reductions in the cover of plant species typical of wet conditions (e.g. cottongrass species and Sphagnum spp), whilst recent findings from a small number of studies on drain blocking suggest some reversal of these effects at least, although responses may be variable and inconsistent. Determining the effects of drain blocking on the vegetation of grouse moors, and the ways in which this management interacts with rotational muirburn to affect vegetation, would be of considerable value. It is clear that fundamental gaps in knowledge remain concerning the effects of grouse moor management on moorland vegetation, and on the extent to which these effects confer costs and benefits to biodiversity. Knowledge of the effects of the main managements is inadequate and, even where such knowledge exists, information on the interactions between different managements and on exactly how these managements are applied across grouse moors (and the variation in this – e.g. in terms of scale and intensity) is lacking. These factors greatly limit our ability to assess the overall effects of grouse moor management on moorland vegetation. To address these knowledge gaps, valuable topics for further research include: • Assessment of variation in plant species richness and diversity, and the

occurrence of key attributes of habitat condition, between grouse moors and other moors in relation to muirburn intensity and frequency.

• Measurement of differences in grazing regimes between grouse moors and other moors.

• Quantification of differences in grazing regimes between grouse moors and other moors.

• Quantification of burning regimes on grouse moors across different regions and moorland habitats, particularly in terms of rotation-lengths, fire severity and the extent to which older Calluna stands are retained.

• Determining the response of some key plant species and taxa to rotational muirburn regimes, and to variation in critical attributes of muirburn management (e.g. burn frequency and fire severity).

• Determining the extent to which drain blocking is undertaken on grouse moors and assessment of the effects of drain blocking and interactions between this and rotational muirburn.

32

3.3 Moorland Invertebrates 3.3.1 Managements of relevance In the context of grouse moors, rotational muirburn is the management that is likely to be of greatest importance in influencing the abundance and diversity of invertebrate species (as for moorland vegetation). The main focus of this section is therefore on the effects of rotational muirburn, with additional consideration given to the effects of drainage and drain blocking. The relationships between grouse moor management and grazing regime are also likely to be of considerable importance to moorland invertebrates, with variation in grazing regimes known to have substantial effects on invertebrate abundance and diversity (Dennis et al. 1997, 2001, 2008, Gardner et al. 1997). The likelihood that grazing regimes on grouse moors favour the maintenance of Calluna cover compared to those on other moors is likely to be a major effect, and will benefit some invertebrate groups (e.g. Lepidoptera – Haysom & Coulson 1998), but not others (e.g. hemipteran species richness is higher in moorland swards that have lost Calluna – Littlewood et al. 2006). However, as for vegetation, determining the detailed effects of grazing regimes on grouse moors on moorland invertebrates is difficult because of the lack of detailed information on how grazing regimes on grouse moors differ from those on other moors, together with the complex interactions that occur between rotational muirburn and grazing regime (3.2.2). Additionally, disentangling effects of rotational muirburn from those of grazing regime is problematic because both directly affect the composition and structure of the vegetation upon which moorland invertebrates depend, so that their effects may be confounding in some ways. 3.3.2 Moorland invertebrate communities and the main factors affecting diversity and abundance In terms of overall numbers, studies in northern England indicate that enchytraeid worms, springtails (Collembola) and mites (Acari) make up the vast bulk of invertebrates across a wide range of moorland habitats. Although the contribution of these invertebrates to the total invertebrate standing crop (biomass) is much less, the enchytraeids in particular still contribute substantially in this respect, especially on blanket bogs where they may account for over half of the total standing crop (Coulson 1988). Overall, worms (enchytraeids and earthworms (Lumbricina) combined) may comprise over half of the invertebrate standing crop in many situations, particularly on grasslands overlying base-rich mineral soils, where earthworm densities may be more than an order of magnitude greater than on other moorland habitats (Coulson & Butterfield 1985, Coulson 1988). In terms of the insects, true flies (Diptera) and lepidopterans tend to make major contributions to the standing crop. True flies, particularly craneflies (Tipulidae) are most numerous on blanket bogs and other peat-based habitats, being estimated to comprise 20% of the standing crop on such habitats in one study, compared to 4% on dry heath (Coulson 1988). In contrast, lepidopterans (which occur at lower densities overall) were estimated to comprise approximately one third of the invertebrate standing crop on dry heath, compared to just 1% on blanket bogs (Coulson 1988). Beetles (Coleoptera) and bugs (Hemiptera) are other common insect groups on moorlands (with the latter group sometimes particularly numerous), whilst spiders (Araneae) and harvestmen

33

(Opiliones) are also important in terms of the above-ground invertebrate community (Coulson 1988, Dennis et al. 2008). These studies from northern England indicated that the total invertebrate standing crop is substantially greater on upland limestone grasslands than on other moorland habitats, due largely to the abundance of earthworms in this habitat, with blanket bog having a greater standing crop than dry heath (Coulson 1988). However, numbers of macro-invertebrates (i.e. excluding enchytraeids, springtails and mites) were greatest on blanket bogs, due to the higher numbers of bugs and craneflies on this habitat. Limestone grasslands held a greater number of invertebrate species than other moorland habitats, at least amongst the macro-invertebrates, although highest diversity was found in areas with a mix of blanket bog and dry heath (Coulson & Butterfield 1985, Coulson 1988). This reflected the fact that (as also found in several other studies – Gardner 1991, Usher 1992) species assemblages differed between these broad habitat-types, so that greater interspersion of the habitats tended to increase overall species diversity. These broad comparisons provide insight into some of the main differences in the diversity and abundance of invertebrate species on different moorland habitats, as well as indicating the importance of factors such as soil nutrient status and habitat variability in determining invertebrate communities and species diversity. In addition to these, a wide range of other environmental and habitat factors affect the abundance and diversity of different invertebrate groups on moorland, notably altitude and rainfall (Coulson 1988), soil wetness and soil organic content (Cherrill & Rushton 1993, Sanderson et al. 1995, Gardner et al. 1997), topography (Downie et al. 1995), vegetation structure (Downie et al. 1995, Gardner et al. 1997) and vegetation composition (Cherril & Rushton 1993, Sanderson et al. 1995). Therefore, there will be a number of different ways in which managements associated with grouse moors affect moorland invertebrates, most obviously and directly through changes in the structure and composition of the vegetation, but also by affecting plant age and nutrient status, and soil wetness and nutrient status. Furthermore, grouse moor management may also affect the freshwater invertebrates of moorland streams and rivers, with recent studies at Moor House indicating that such streams can have a relatively high species richness of freshwater insects (Brown et al. 2009). 3.3.3 Effects of rotational muirburn on invertebrate abundance and diversity Rotational muirburn will affect invertebrate abundance and diversity by several different means, with the immediate effects being destructive, through direct mortality and reductions in food availability in the burnt area. Marked declines in the abundance of insect species in the immediate and short term period (hours to 1 – 2 months) after fires is a widespread occurrence across many habitats, and in a few instances such declines have been shown to be sufficient to result in local population extinction (Swengel 2001). For many invertebrates on moorland, re-colonisation of burnt areas following rotational muirburn is likely to be relatively rapid, as burning is usually carried out in small patch sizes (1.2), and many of the invertebrates on moorland are highly mobile, making threats to population viability unlikely (Usher & Smart 1988, Gardner & Usher 1989, Glaves & Haycock 2005). Thus, even for some phytophagous insects, such as leafhoppers (Cicadellidae), movement of relatively

34

large numbers into and across recently burnt areas can occur within three months of burning, despite the fact that the removal of Calluna and other vegetation will represent a loss of habitat (Gardner & Usher 1989). However, this situation may be different for species that are associated mainly, or solely, with degenerate Calluna stands, which may be a rare feature on some grouse moors (Tucker 2003). One of the main ways in which rotational muirburn affects invertebrate diversity and abundance is via the creation of a patchwork of vegetation stands in different growth phases, which increases the structural variation of vegetation across an area of moorland. This is expected to increase invertebrate diversity, because the species assemblages of several groups differ in relation to variation in Calluna growth phase, and in the composition and structure of moorland vegetation. Studies of ground beetles (Carabidae) and spiders on moorland in the North York Moors have shown that the development of vegetation cover from sparse or short to well developed is a major factor separating the distribution of the different species, with different species assemblages occurring on burnt or cut moorland than on areas where the vegetation is well developed (Gardner 1991, Usher 1992). Earlier studies in the same area demonstrated that numbers of spider and harvestman species (as determined from pitfall trap catches) in the first summer after burning were higher near to the edge ( ≤ 5 m) of 35 m wide burnt strips than in adjacent unburnt vegetation, although lowest species richness occurred in the central areas of the burnt strips (Usher & Smart 1988). Analogous studies of ground beetles found less variation in species richness in relation to burning, with similar numbers of species occurring in these different areas (Gardner & Usher 1989). Importantly, individual species differed in the extent of their association (as determined from numbers caught) with the burnt and unburnt vegetation, and within the burnt strips with distance from the edge, so providing strong indications that the overall effect of rotational muirburn is to increase the species diversity of these invertebrate groups on moorland. Amongst the Lepidoptera, the diversity of species present as larvae on dry dwarf shrub heath managed by rotational muirburn increased with the height of Calluna, which, in turn, was positively correlated with green shoot density and flower density (Haysom & Coulson 1998). The overall increase in species diversity with Calluna height in this study was due largely to the presence of uncommon moth species in taller Calluna, and other studies have found no significant differences between building, mature and degenerate Calluna phases in the actual number of lepidopteran species present as larvae (Fielding 1992). Instead, changes in the contribution of common species to the community in different height zones was the main factor causing greater species diversity in taller Calluna, in the Haysom & Coulson study (1998). This suggested that the presence of different aged stands is important in maintaining overall species diversity. Rotational muirburn can also benefit lepidopteran diversity by helping to maintain blaeberry in the presence of Calluna (3.2.4, Table 1); blaeberry being another important host species for larvae, and often supporting different species to Calluna (MacDonald & Haysom 1997). However, to benefit lepidopteran species diversity overall, rotational muirburn would have to be carried out in such a way that it maintained a substantial proportion of tall Calluna (> 20 cm), including the retention of unburnt stands, and short burning rotations (e.g. 10-year) are likely to be detrimental (MacDonald & Haysom 1997).

35

Studies on dry dwarf shrub heath in northeast Scotland found that the number of different invertebrate families present in stands of different Calluna growth phases was greatest in pioneer stands (with 52), and lowest in mature stands (41), although diversity (as measured by William’s index) was highest in degenerate stands, and lowest in both building and mature stands (Gimingham 1985). Greater numbers of families (55) were found in areas of unmanaged Calluna, where there was a heterogeneous mix of Calluna bushes of different ages. These data on numbers and diversity of families were determined from sampling with suction apparatus (Johnson et al. 1957), and so sampled invertebrates on the vegetation and ground surface. Additional sampling of aerial insects and of invertebrates in litter in this study revealed that the diversity of these components of the invertebrate fauna also appeared to be highest in the unmanaged areas, whilst amongst the managed stands highest diversity of litter invertebrates occurred in the degenerate phase. However, it is not clear whether areas had been left unmanaged because of intrinsic differences (e.g. in soil type) that might have affected invertebrate populations, whilst no details were given on the extent of variation in diversity measures between different stands in the same growth phase or between different unmanaged areas. An assessment of how diversity was likely to vary between managed and unmanaged areas at a larger scale (where managed areas would comprise a patchwork of even-aged stands in different growth phases - i.e. equivalent to β-diversity – 3.2.5) would have been valuable in this study. The studies from the North York Moors found that the overall catches of certain invertebrate groups differed between the burnt and unburnt sample areas. Thus, ants (Formicidae) were more numerous immediately along the edge of the burnt and unburnt vegetation, whilst catches of leafhoppers increased in the burnt areas compared to the unburnt vegetation, as was the case for several of the most abundant ground beetle species (Gardner & Usher 1989). However, these findings do not necessarily demonstrate that burning increases the actual density of these invertebrate groups, because catches from pitfall traps are affected by the activity of the animals and by vegetation structure, as well as by their abundance (Greenslade 1964). In particular, the higher numbers of leafhoppers caught in burnt areas was probably due to different activity patterns in the burnt and unburnt vegetation, with insects in the unburnt vegetation feeding in the Calluna canopy, making them less likely to be trapped (Gardner & Usher 1989). Other studies using pitfall traps on dry dwarf shrub heath in northeast Scotland found that total numbers of invertebrates caught were greatest in stands in the pioneer and degenerate Calluna phases, with lowest catches of most groups in the building phase, but these findings are subject to the same potential bias (Barclay-Estrup 1972). For some invertebrate groups at least, sampling with suction apparatus is likely to produce more reliable estimates of abundance (Southwood & Henderson 2000), although few such studies appear to have been undertaken on moorlands. In Gimingham’s (1985) study in northeast Scotland (see above), invertebrates that feed upon the unlignified green shoots or which are sap feeders (e.g. bugs, weevils (Curculionidae) and some lepidopteran larvae) were most abundant in the building and (in some cases) mature phases, where their food supply was likely to be greatest.

36

By contrast, ground feeders and dwellers, such as springtails, ground beetles and some spiders were amongst the taxa that were most abundant in the pioneer phase. These findings support the view that pitfall trap catches underestimate the abundance of some groups (e.g. leafhoppers) in unburnt Calluna, whilst reflecting real differences in the abundance of other groups (e.g. ground beetles) between burnt and unburnt Calluna. Lepidopterans comprise a major component of the invertebrate community on dry dwarf shrub heath (3.3.2), and densities of their larvae increase with Calluna height. Studies based upon sweep net sampling (and verified using jarring of vegetation with Berlese-Tullgren extraction), demonstrated an approximate doubling of larval density with every 20 cm increase in Calluna height, with this relationship being similar across the four major sub-groups examined – i.e. macrolepidoptera, microlepidoptera, geometrids and noctuids (Haysom & Coulson 1998). In addition to the initial destructive effects of fire, and the effects produced by creating a patchwork of different Calluna growth phases, rotational muirburn could affect overall invertebrate diversity and abundance by several other means. These include changes in nutrient status and moisture levels. The release of nutrients after burning is suggested as a factor that could cause a temporary increase, at least, in invertebrate abundance (Tucker 2003), although any long-term decline in nutrient status as a consequence of burning may have the opposite effect. Detrimental effects could arise if rotational muirburn causes habitats to become drier (e.g. via reductions in Sphagnum mosses, particularly on wet heath and blanket bog – 3.2.6). Any such effects could be of considerable importance because of their likely impact on groups such as craneflies, which comprise a major part of the invertebrate community on the wetter moorland habitats (3.3.2) and for which certain of the species are susceptible to desiccation (Coulson 1988). Furthermore, site wetness is often as, or more, important than vegetation structure in separating the distribution of different species amongst such groups as ground beetles and spiders (Gardner 1991, Usher 1992), so that the loss of wet areas would reduce diversity. However, strong evidence that rotational muirburn consistently causes moorland habitats to become drier appears to be lacking (Worrall et al. 2007a, 3.2.6, 4.3.2). The findings described above relate primarily to macro-invertebrates that have at least part of their life-cycle on or above the ground surface. Few studies appear to have been carried out on the effects of burning on the invertebrates of moorland litter and soils, or of moorland streams. Increases in microbial numbers may occur after burning on shallow and deep peat, possibly because of increases in pH and nutrient availability (Maltby & Edwards 1984). However, other studies point to detrimental effects of burning. Substantial reductions in the total densities of the micro- and meso-fauna (particularly litter-dwelling springtails and enchytraeid worms) were found one year after management burning compared to mature Calluna in the North York Moors (Brown 1986), whilst other studies suggest that rotational muirburn reduces diversity of litter-dwelling invertebrates on dry heath (Gimingham 1985, see above). Effects of non-management fires may be substantial, with one study finding reduced densities and diversity of soil fauna eight years after the fire (Aked 1984), but such fires are likely to be of greater severity than management fires.

37

One study only appears to have investigated effects of rotational muirburn on the invertebrate fauna of moorland streams, with only initial, preliminary, findings reported as yet (Brown et al. 2009). These indicate that neither the overall abundance, nor the taxonomic richness4

, of aquatic invertebrate larvae differed significantly between three 2nd order streams draining North Pennines’ catchments managed by rotational muirburn and 2nd order streams at Moor House, where there is no rotational muirburn (except on small-scale experimental plots – 3.2.5). Although significantly lower abundances of some stonefly (Plecoptera) and mayfly (Ephemeroptera) species (and associated higher abundances of some of the more widespread and commoner true flies) were found on the managed sites, these effects could conceivably be due to confounding differences between Moor House and the managed sites (e.g. in terms of grazing regime and vegetation type).

3.3.4 Effects of rotational muirburn on species and taxa of conservation importance Few studies have been undertaken on individual invertebrate species of conservation importance on moorlands, but there is evidence that rotational muirburn may benefit some rare and important species. Some nationally rare species of ground beetles and spiders are associated with the open conditions of recently cut or burnt moorland (Usher 1992), whilst amongst four broad habitats on a grouse moor in southern Scotland, dry heather moorland managed by rotational muirburn was ranked highest for nationally rare and scarce ground beetles (Eyre et al. 2003). In this latter study, the only nationally scarce plant bug species recorded also occurred on the managed Calluna habitat, but rare or scarce rove beetles (Staphylinidae) were not associated with this habitat, whilst there were no rare or scarce spiders (the other invertebrate group studied) recorded. Overall, streamside habitats (particularly sediment) were of greatest importance for nationally rare or scarce species of these invertebrate groups, with unmanaged wet heather moorland also holding some rare and scarce species. Thus, the presence of both managed and unmanaged habitats together was considered important in contributing to the conservation importance of the moor for invertebrates (Eyre et al. 2003). Low intensity patch burning, as will occur under some rotational muirburn regimes, is positively correlated with the occurrence of the large heath butterfly Coenonympha tullia; a UK BAP species that has undergone marked declines (Dennis & Eales 1997). Blanket bogs provide the main habitat for this species in the UK, and it is predominantly found on wet sites with surface water, where there is an abundance of hair’s-tail cottongrass (an important larval food plant) and cross-leaved heath (an important nectar source for adults), preferably in close association (Dennis & Eales 1999). Rotational muirburn is likely to increase the cover of both hair’s-tail cottongrass and cross-leaved heath on blanket bogs (3.2.4). By contrast, high intensity burning, or low intensity burning across a whole site at one time, is negatively correlated with the occurrence of the large heath butterfly (Dennis & Eales 1997). Presumably, negative effects of such burning practices could arise through declines in Sphagnum abundance or damage to peat, causing subsequent drying of the habitat.

4 Defined as the number of mayfly (Ephemeroptera), stonefly (Plecoptera), caddisfly (Trichoptera) and beetle species plus the number of genera/families of other groups.

38

Detrimental effects of rotational muirburn on invertebrate species of conservation importance are likely to occur where burning regimes fail to retain sufficient areas of degenerate phase or unmanaged Calluna, because of the likelihood of such habitats holding rare or scarce species (Haysom & Coulson 1998, Eyre et al. 2003, Tucker 2003). Insufficient information appears to exist to allow any rigorous assessment of what might constitute a sufficient area of such habitat for ensuring that conservation interests are maximised. 3.3.5 Effects of drainage and drain blocking Few studies have been undertaken on the effects of drainage on moorland invertebrates. In the study by Coulson et al. (1990), the most marked effects appeared to occur on a low altitude blanket bog site, with few detectable effects at a high altitude site (i.e. analogous to effects on vegetation in that study – 3.2.7). At the low altitude site, the numbers of bugs, true-flies, click beetles (Elateridae) and rove beetles caught in pitfall traps placed 1.5 m downslope of drains were higher than in traps placed 1.5 m upslope of drains, whilst there were no differences in the numbers of ground beetles, spiders, harvestmen or enchytraeid worms caught. These differences may have been due to the reduction in Calluna cover immediately downslope of drains (3.2.7) causing subsequent changes in invertebrate abundance, although the uncertainty over the interpretation of pitfall trap catches has to be borne in mind (3.3.3). At the high altitude site, the only effect was a reduction in the catches of enchytraeids downslope of drains, a finding that was attributed to the lowering of the water table on the downslope side, combined with the susceptibility of enchytraeids to desiccation (Coulson et al. 1990). There are few studies assessing the effects of drain blocking on terrestrial moorland invertebrates. A recent study (Carroll et al 2011) found that the abundance of craneflies, a major prey item for breeding birds, is positively related to soil moisture - cranefly abundance was consistently low where the peat was dry and generally higher (though variable) in wet areas. Blocking drains increased soil moisture and cranfly abundance though the strength and significance of effects varied between years (Carroll et al. 2011). A study into the impact of drainage and drain blocking on the invertebrates of moorland streams found no significant differences in the overall abundance or taxonomic richness of aquatic invertebrate larvae between streams on, three sites with drainage in place, three sites where drains had been blocked and at Moor House, where drains are absent (Brown et al. 2009). However, the abundance of some species of stonefly and mayfly were significantly lower on the drained sites than on either the Moor House sites or the sites with blocked drains. As with the analogous investigations into muirburn effects on freshwater invertebrates (3.3.3), it is conceivable that these differences are due to confounding site effects, but in this case the similar abundance of these species on sites at Moor House and on sites with blocked drains provides stronger evidence for an effect of drainage. Such differences may be due to the increased suspended sediment associated with drained catchments. More recently, the authors of the above study reported a positive effect of drain blocking. Following drain blocking, the water quality of stream ecosystems was improved such that they sustained broadly similar macroinvertebrate

39

communities to streams in catchments where there was no peat drainage (Ramchunder et al. 2012). The authors note that this benefit is largely going unnoticed by the majority of peatland restoration projects. 3.3.6 Summary and synthesis of effects on invertebrates The effects of grouse moor management on moorland invertebrates are poorly understood. One study examining the occurrence of some invertebrate groups in areas managed as grouse moor compared to unmanaged areas provides some evidence for grouse moor management increasing the conservation value of the invertebrate community, but it also indicates that benefits will be greatest where managed and unmanaged areas occur together on a moor. Rotational muirburn on dry dwarf shrub heath increases the diversity and abundance of some invertebrate groups (notably ground beetles and spiders) and appears to have the potential to increase the overall diversity of terrestrial macro-invertebrates at least, through the resultant increase in the structural diversity of the vegetation. Whether or not this potential benefit is actually realised on grouse moors will depend on the exact nature of the muirburn regime, with regimes that prevent the development of unmanaged or degenerate Calluna stands reducing the diversity and abundance of some groups (notably lepidopteran larvae) and being likely to have detrimental effects overall. It is also unclear whether such potential benefits arise on wet heath or blanket bog, although a carefully managed, ‘sensitive’, rotational muirburn regime on such habitats is expected to increase the abundance of at least one species of conservation importance (the large heath butterfly). Some limited evidence suggests that muirburn may reduce the abundance of soil invertebrates but does not affect the overall abundance or taxonomic richness of aquatic invertebrates in moorland streams, although it may reduce the abundance of some mayfly and stonefly species. Given the importance of soil moisture as a determinant of invertebrate diversity and abundance on moorlands, drainage is likely to have had detrimental effects on both of these aspects, whilst drain blocking may be expected to be beneficial. A study in the North Pennines found a positive effect of drain blocking on water quality and associated aquatic invertebrates. Limited information is available on the effects of grouse moor management on invertebrates. Furthermore, as for vegetation, an understanding of the full effects of grouse moor management on invertebrates is greatly restricted by a lack of knowledge on the ways in which these managements are applied across grouse moors, and the ways in which they interact to affect invertebrate communities. Research that examines variation in invertebrate diversity and abundance between grouse moors and other moors, and in relation to the intensity of grouse moor management is a priority, as is similar work on individual species or taxa of particular conservation importance. Further work to improve the understanding of the effects of rotational muirburn and drain blocking on invertebrate abundance and diversity is also important.

40

3.4 Moorland birds 3.4.1 Assessing the effects of grouse moor management on the moorland bird community The UK’s upland bird community is widely recognised as being of international importance, comprising unique species assemblages and including populations that are of national and international conservation importance (Thompson et al. 1995). Within the wider UK upland bird community, moorlands provide a major breeding habitat for several of the species of conservation importance, with the breeding populations of some virtually confined to moorlands (e.g. merlin Falco columabrius, golden plover Pluvialis apricaria, dunlin Calidris alpina, greenshank Tringa nebularia, ring ouzel Turdus torquatus). Populations of several of the bird species associated with moorlands are currently declining, both at the UK level and, in some cases, more widely across their European range (Burfield & Bommel 2004, Sim et al. 2005, Pearce-Higgins et al. 2009), with several species listed as UK BAP5

species (e.g. black grouse, curlew Numenius arquata and ring ouzel).

In many respects, the evidence base for assessing the effects of grouse moor management is greater for birds than for other taxa. Uniquely, comparative studies have examined how population densities of some species vary in relation to the intensity of grouse moor management, whilst there have been large-scale field experiments to determine the effects of particular managements on populations, and detailed studies of the ecology of some species on grouse moors (Redpath & Thirgood 1997, Tharme et al. 2001, Fletcher et al. 2010). In terms of the different managements associated with grouse moors, several are likely to affect bird populations, with legal predator control, the illegal control of birds of prey and rotational muirburn likely to be of most importance (and having been subject to the most detailed study). As for vegetation and invertebrates, grazing regimes affect moorland bird populations and the role of grouse moors in influencing grazing regimes will be important. As the bird species most closely associated with Calluna, red grouse clearly benefit from the lower grazing pressures likely on many grouse moors compared to other moorlands (Miller et al. 1966, Hudson 1992). Several other species are also associated with Calluna cover to varying degrees (e.g. merlin, hen harrier and ring ouzel), but none are as dependent on extensive cover as are red grouse (Pearce-Higgins et al. 2009). Because grouse moors have been an important means by which Calluna cover has been retained on moorlands, it seems likely that they have benefited such species in this respect. However, the full extent to which the grazing regimes present on grouse moors benefit such species compared to grazing regimes on other moorlands is difficult to determine. Once again, this is in large part due to the lack of detailed information on how grazing regimes on grouse moors differ from those on other moors, together with the complex interactions that occur between rotational muirburn and grazing regime. Additionally, as for invertebrates, disentangling effects of rotational muirburn from those of grazing regime can be problematic because both directly affect the composition and structure of the vegetation upon which moorland birds depend, so that their effects on bird populations may often be confounding.

5 See – www.ukbap.org.uk/PrioritySpecies.aspx?group=1

41

The following assessment of the effects of grouse moor management focuses primarily (but not exclusively) on those species that; (i) are most strongly associated with moorland habitats; (ii) have breeding ranges that overlap with the occurrence of (the main areas of) grouse moors; and (iii) for which moorland management regimes are expected to have major impacts (Pearce-Higgins et al. 2009, Table 2). Red grouse themselves are not considered in detail for the reasons discussed in section 1.3., although they clearly fulfil the three criteria listed above, and are an integral component of the moorland bird community (as well as being a UK BAP species). 3.4.2 Effects of rotational muirburn Rotational muirburn may affect moorland birds in two main ways – i.e. by direct destruction of nests and through modification of habitat (including the structural characteristics and the availability and condition of both plant and invertebrate food sources). The available data suggest limited overlap only between the date on which muirburn generally has to cease (1.2) and the time of clutch initiation of many ground nesting birds on moorland (Glaves & Haycock 2005, Moss et al. 2005). Amongst the ground nesting species that show greatest overlap in this respect, several select nest-sites in short vegetation (notably golden plover, lapwing Vanellus vanellus and curlew - Whittingham et al. 2002, Robson 1998), where burning is unlikely to occur on grouse moors, but others select nest-sites in tall Calluna, where muirburn is likely. Short-eared owls and stonechats Saxicola torquata may be amongst the most vulnerable species in this respect, given that a substantial proportion of nesting attempts are initiated before 15 April (estimated as 16 - 32% and 28 – 40%, respectively), and that both species frequently nest in tall Calluna (Roberts & Bowman 1986, Moss et al. 2005, Pearce-Higgins & Grant 2006). Because hen harriers also select stands of tall Calluna for nesting, they too may be vulnerable to nest losses from muirburn. However, below 450 m (where the majority nest), this is only likely to be important where the burning season is extended to 30 April (few pairs nest before 15 April, but up to c.25% are estimated to have started clutches by 30 April) (Redpath et al. 1998, Moss et al. 2005). Overall there appears to be little direct evidence of nest losses from muirburn amongst moorland birds, but such evidence may be difficult to obtain and detailed nest survival studies of short eared owls and stonechats on grouse moors (the two species, other than red grouse, considered most vulnerable to such effects) are lacking. Several moorland bird species require access to short, open, vegetation, for purposes such as nesting (golden plover – Whittingham et al. 2002) and foraging (ring ouzel – Burfield 2002), whilst breeding densities of some species are positively correlated with the extent of short, open, vegetation (e.g. golden plover and skylark Alauda arvensis – Pearce-Higgins & Grant 2006). Where grazing is absent or at low levels, muirburn is the main way in which short, open, vegetation is created and maintained on many moorlands, and it is therefore of potential importance to those species requiring access to such vegetation. At the same time, as indicated above, several moorland species require access to tall dense vegetation (often Calluna), usually for the purposes of providing suitable nest-sites (e.g. black grouse, hen harrier and ring ouzel – Picozzi & Hepburn 1986, Redpath et al. 1998, Burfield 2002), with the densities of some species positively correlated with the extent of tall Calluna (Pearce-Higgins & Grant 2006). For these species it is conceivable that rotational muirburn

42

could be detrimental, if burning regimes were particularly intensive and reduced the area of tall Calluna to such an extent that nest-site availability limited breeding densities. However, several of the same species may benefit from other consequences of rotational muirburn (see below). Amongst the waders, rotational muirburn appears to benefit several species by creating suitable nesting (and possibly chick-rearing) habitat. Both golden plover and curlew select recently burnt areas for nesting on heather moorland (Whittingham et al. 2002, Robson 1998), as presumably do lapwings6

, given their strong preference for breeding in short, open, vegetation (O’Brien 2002). Such benefits do not necessarily translate into higher breeding densities of these species, however. Suitable nesting sites may occur in other situations on moorland (e.g. golden plover are often most abundant on blanket bog, where there may be little muirburn – Ratcliffe 1976), whilst small-scale patch burning may offer no advantages over more extensive burning (as undertaken for sheep or deer management) in this respect. Evidence of effects of rotational muirburn on wader breeding densities derive mainly from studies identifying associations with grouse moor management, although because they are correlative these studies have, at best, limited ability to disentangle the effects of rotational muirburn from those of other attributes and managements associated with grouse moors (Haworth & Thompson 1990, Tharme et al. 2001, Daplyn & Ewald 2006). The most comprehensive study of this type, undertaken across extensive areas of the Scottish and English uplands, demonstrated that the breeding densities of golden plover, lapwing and curlew (but not snipe Gallinago gallinago) were higher on heather moorland managed for grouse shooting than on other heather moorland, even after accounting for potentially confounding habitat effects (Tharme et al. 2001). Several different indices of grouse moor management intensity were measured in this study, including measures of muirburn. The analyses provided strongest evidence for positive effects of muirburn on golden plover, with their densities (but not those of curlew or lapwing) being correlated with measures of burning, after adjustment for other habitat effects. However, for all three of these species, the differences in density between grouse moors and other moors remained after accounting for burning, suggesting that other managements (most likely predator control – 3.4.3) were more important in causing the higher densities on grouse moors.

A second study, confined to the Peak District moorlands, also attempted to assess the role of grouse moor management in determining breeding bird densities (and changes in density between 1990 and 2004), and to distinguish the effects of muirburn from other managements associated with grouse moors (Daplyn & Ewald 2006). Amongst the waders, the analyses in this second study indicated overall benefits of grouse moor management for golden plover and dunlin. Considering muirburn specifically, this study found some evidence for positive effects on golden plover, lapwing and curlew, but for negative effects on dunlin, with conflicting findings for snipe. However, the ability to account for any associated habitat effects was limited in this study, since it relied upon broad habitat measures derived from 1988, which pre-dated even the earlier bird survey data (Daplyn & Ewald 2006). Furthermore, other analyses of the same bird data in relation to more detailed, 6 Whilst lapwings do sometimes occur on grouse moors, the species is primarily found in the uplands on enclosed rough pastures and meadows (Gibbons et al. 1993).

43

contemporary, habitat measures (but less detailed grouse moor management measures), found less evidence for effects of grouse moor management on bird abundance (Pearce-Higgins et al. 2006). Thus, the findings of these analyses on the effects of grouse moor management, and specifically muirburn, on the Peak District wader populations are inconclusive. In contrast to some of the main moorland wader species, rotational muirburn appears to have detrimental effects on meadow pipit Anthus pratensis populations. Tharme et al. (2001) found that meadow pipit densities were lower on heather moorland managed for grouse shooting than on other heather moorland, with densities (adjusted for other habitat effects) negatively correlated with one of the measures of burning used in this study. Additionally, in a second study, meadow pipit densities were also negatively correlated with the extent of muirburn, both across a sample of grouse moors and on a series of plots within a single moor (Smith et al. 2001). The causes of this relationship are unclear. More generally, on moorlands, meadow pipit densities are related to the ratio of graminoid to dwarf shrub (mainly Calluna) vegetation, with highest densities where dwarf shrub cover is c.40% (Smith et al. 2001, Pearce-Higgins & Grant 2006). Densities may also be lower where short (< 15 cm) vegetation is more extensive (Buchanan et al. 2007). Thus, the negative relationship between meadow pipit densities and muirburn could conceivably arise from confounding relationships between the extent of muirburn and grass-Calluna ratios, or more simply through the resultant increase in short vegetation. Less evidence exists for effects of rotational muirburn on other moorland passerines. Tharme et al. (2001) found no differences in the densities of skylarks, wheatears Oenanthe oenanthe and whinchats Saxicola rubetra between heather moorland managed for grouse shooting and other heather moorland, after accounting for other habitat differences, whilst densities of these species were not correlated with muirburn measures, after adjustment for other habitat effects. The analyses of the Peak District data by Daplyn & Ewald (2006) found somewhat contrasting findings to this, with evidence of negative effects of burning on skylark, reed bunting Emberiza schoeniclus, wheatear, whinchat and twite Carduelis flavirostris, and of positive effects on ring ouzel. However, as for the waders, these findings were generally not supported by other analyses of the same bird data which used more detailed, contemporary habitat measures (Pearce-Higgins et al. 2006). Several moorland passerines, including ring ouzel and twite, select nest-sites in relatively, tall, dense vegetation, often Calluna (Brown et al. 1995, Burfield 2002, Wilkinson & Wilson 2010). By reducing the extent of tall Calluna, it is conceivable that rotational muirburn could cause nest-site availability to become limiting for these species, but this would probably require particularly intensive muirburn regimes, which may be rare on many, but not all, grouse moors (Hester & Sydes 1992, Rhodes 1996, MacDonald 2000, but see Yallop et al. 2006). Furthermore, other vegetation can sometimes provide alternative nest-sites (e.g. bracken for twite – Raine 2006), whilst ring ouzels select nest-sites on steep slopes and in ghylls (Sim et al. 2007a) where burning should be less likely (Defra 2007).

44

There appears to be no evidence for clear positive or negative associations between muirburn and black grouse abundance (Tharme et al. 2001), and potentially it may have both beneficial and detrimental effects on this species. As for red grouse, benefits of rotational muirburn may arise from the greater nutritional content of the resultant young heather shoots (Savory 1978), heather being a major food plant during the late autumn and winter (Picozzi & Hepburn 1986). Furthermore, burning on blanket bogs will increase the cover of hair’s-tail cottongrass (3.2.4), with cottongrass inflorescences being a protein-rich food source that the hens often exploit prior to egg-laying (Trinder 1975, Starling-Westerberg 2001). However, negative effects may arise through rotational muirburn preventing the development of scrub and open woodland at the moorland edge (attributes that would probably improve overall habitat conditions for black grouse - Pearce-Higgins et al. 2007, Watson & Moss 2008), or by intensive muirburn regimes removing potential nesting habitat; black grouse often selecting nest-sites in patches of tall, dense, Calluna (Picozzi & Hepburn 1986, Grant & Dawson 2005). Thus, whether muirburn confers benefits or disadvantages to this species is likely to depend upon the situation and locations in which it is carried out. Of the main predatory birds breeding on moorland, hen harrier, merlin and short-eared owl all select tall Calluna for nesting (Bibby 1986, Roberts & Bowman 1986, Redpath et al. 1998). Muirburn regimes that retain little or no degenerate Calluna could reduce breeding densities by limiting nest-site availability, which may be more likely than for passerine species selecting tall Calluna nest-sites, because more extensive stands of taller Calluna may be required (Redpath et al. 1998). The temporary loss of merlin breeding territories was attributed to increased burning on game estates in one study in northeast Scotland (Rebecca & Cosnette 2003), whilst the use of burning to remove tall heather at potential or historical harrier nest sites has been reported in several instances (B. Etheridge, T. Melling, S. Garnett, pers. comm.). Meadow pipits are usually the major prey species of both hen harriers and merlin on moorland, particularly during the nestling period, with inter-site variation in harrier breeding densities determined by pipit density, in the absence of illegal persecution (Newton et al. 1984, Bibby 1987, Redpath & Thirgood 1999). Despite this, prey capture rates for harriers are higher on moorland managed for grouse than on other moorlands (Redpath et al. 2002a), as are brood size at fledging amongst successful broods (Etheridge et al. 1997). Higher prey capture rates on grouse moors may be due to higher densities of other important prey (e.g. red grouse chicks), but seem more likely to be due to other factors, such as differences in habitat or prey behaviour, which make prey more accessible on grouse moors (e.g. harriers selectively hunt along edges of muirburn patches – Redpath 1992). Field voles are also important prey for harriers, and the main prey for short-eared owls (Village 1987, Redpath et al. 2002b). There appear to be no studies examining the effects of rotational muirburn on vole densities, but these are highest in grassy habitats with little or no grazing (Hewson 1982, Evans et al. 2006, Wheeler 2008). Consequently, rotational muirburn seems unlikely to benefit vole abundance because it will promote Calluna cover in most situations, and Hope et al. (1996) suggested that vole increases were most likely where reductions in sheep and red deer grazing impacts were associated with absent or infrequent burning. However, on blanket bog benefits of burning for voles could arise via the resultant increase in hair’s-tail cottongrass,

45

provided grazing levels are low (Wheeler et al. 2008). Rotational muirburn benefits both golden eagle and peregrine falcon Falco peregrinus (the two other raptors most commonly associated with grouse moors), via effects on important prey species. Red grouse and mountain hares are the preferred prey of golden eagles in Scotland (Watson et al. 1993), and rotational muirburn contributes to the higher densities of both on moorlands managed for red grouse (Picozzi 1968, Miller et al. 1970, Stoddart & Hewson 1984). In the absence of illegal persecution, golden eagle breeding success tends to be higher in grouse moor areas due to the abundance of these preferred prey species (Watson et al. 1992). Similarly, red grouse are a major prey species of moorland nesting peregrines (Ratcliffe 1993, Redpath & Thirgood 1999). Few studies have investigated the effects of rotational muirburn on overall bird species richness or diversity. Diversity (as measured by Simpson’s index) was positively correlated with the extent of muirburn across a sample of grouse moors in Scotland and northern England in one study, although there was no relationship with species richness (Smith et al. 2001). The effect of muirburn on species diversity in this study was relatively weak (accounting for 9% only of the variation in diversity), whilst there was also a statistically significant increase in diversity from west to east across the sample of grouse moors. Given that grouse moor management tends to be most intensive on drier eastern moors (Hudson 1992, Watson & Moss 2008), this effect may have been confounded with that of muirburn, making it difficult to assess the true effect of muirburn on bird species diversity in this study, but no details on the level of inter-correlation between these variables is given. Where rotational muirburn prevents the development of scrub (and ultimately woodland) at the moorland edge then it will presumably lead to a reduction in the overall bird species richness and diversity, at least at the scale of the individual moorland holding. 3.4.3 Effects of legal predator control The effect of predation on bird populations has been the subject of considerable study, with several comprehensive reviews undertaken (Newton 1998, Côté & Sutherland 1997, Gibbons et al. 2007, Holt et al. 2008, Roos et al. 2012). These have concluded that predation can limit bird populations, with populations of ground nesting wildfowl, gamebirds, waders and gulls appearing to be most susceptible to limitation by predation. Furthermore, the control of multiple predators by gamekeepers tends to be more effective in increasing the abundance of prey species than is the control of single predators, at least in part because this reduces the chances of compensatory predation by other predators (Reynolds & Tapper 1996, Holt et al. 2008). Thus, the control of generalist predators such as foxes, mustelids and crows on grouse moors may be expected to increase the breeding success and population sizes of at least some of the bird species associated with moorlands, most notably gamebirds and waders. The extent to which there is evidence to support this position for UK moorlands, and for the species of interest, is considered below. For black grouse, predation is usually the main proximate cause of nest failure and adult mortality, often accounting for > 80% and > 70% of failures and mortality at these stages, respectively (Ellison 1978, Storaas et al. 1982, Angelstam 1984, Picozzi & Hepburn 1986, Caizergues & Ellison 1997, Warren & Baines 2002). This can be the

46

case even where predator control is intensive, with, for example, predation (largely by stoats Mustela erminea) accounting for most nest failures and adult deaths on and around grouse moors in the North Pennines (Warren & Baines 2002, Baines et al. 2007). Some of the strongest evidence for the effects of predation on black grouse populations is from studies involving the removal of foxes and pine martens (Martes martes) in a ‘switch-over’ experiment carried out on two islands in the Baltic, where breeding success and population size increased under the predator control regime (Marcström et al. 1988). Elsewhere, increases in black grouse abundance in both Norway and Sweden were recorded following declines in fox densities resulting from an epidemic of sarcoptic mange, with subsequent declines in black grouse abundance when fox populations recovered (Lindstrom et al. 1994, Smedshaug et al. 1999). In Finland, positive effects of controlling mammalian predators were detected on breeding success, but not on population size, whilst large-scale correlative studies detected negative effects of mammalian predator densities on breeding success, but generally failed to detect effects on either black grouse density or population trends (Kurki et al. 1997, Kauhala et al. 2000, Kauhala & Helle 2002). In Britain, the removal of crows in an experimental study at Abernethy Forest resulted in higher breeding success, but only during years of low June rainfall; breeding success being low when June rainfall was high, irrespective of whether or not crows were removed (Summers et al. 2004). Although this study did not examine effects on population density, annual changes in the numbers of male black grouse at Abernethy in spring are directly related to the previous year’s breeding success, suggesting that population level effects of crow removal were likely (Grant et al. 2009). Although foxes were also removed in the Abernethy study, there was no detectable effect of removal on their abundance (Summers et al. 2004). Studies of the effects of habitat management on black grouse across a series of sites in Wales indicate that a population response was limited to those sites where predator control was undertaken (and where foxes and crows were less abundant), with predator control being associated with higher breeding success (RSPB unpublished). However, other correlative studies in Britain have failed to detect effects of gamekeeper presence or grouse moor management intensity on either black grouse breeding success or density (Baines 1996, Tharme et al. 2001). Several studies indicate higher breeding success of some moorland waders where predator control is undertaken, and suggest that this results in higher breeding densities. Thus, findings from an experimental study, carried out on four separate moorland sites and involving a switch-over in predator control status, demonstrate that the removal of foxes, mustelids and crows results in higher breeding success amongst lapwing, golden plover and curlew, with 3.5 times as many pairs producing fledged young when predator control is undertaken (Fletcher et al. 2010). The levels of breeding success recorded in the absence of predator control in this study were probably insufficient to maintain population stability, with less than 20% of lapwing, golden plover and curlew pairs estimated to have fledged young on plots without predator control, and whilst breeding numbers of these waders increased with predator control in place, they declined without it. Furthermore, as detailed above (3.4.2), in an extensive correlative study the densities of all three of these species were one and a half to almost three times higher on heather moorland managed for grouse shooting than on other heather moorland, after accounting for potentially

47

confounding habitat effects (Tharme et al. 2001). These differences in density remained after accounting for differences in the extent of burning as well, whilst densities of both lapwing and golden plover (adjusted for other habitat effects) were negatively related to crow abundance (a likely index of predator control intensity), suggesting that predator control was important in causing the higher densities on grouse moors (Tharme et al. 2001). Other studies have documented declines in numbers of lapwing, golden plover and curlew following the cessation of gamekeeping on moors (Parr 1992, 1993, Baines et al. 2008). In these cases, the declines have been associated with increases in numbers of generalist predators, such as crows and foxes and, in one study, with a marked reduction in golden plover breeding success (resulting from increased nest predation), which appeared sufficient to account for the observed population decline (Parr 1992, Harding et al. 1994). However, in these cases there have also been confounding effects, which could conceivably have caused the observed declines (e.g. increased overwinter losses of adult golden plover or changes in vegetation condition - Parr 1992, Baines et al. 2008). Another large-scale predator control experiment, undertaken on lowland wet grassland sites in this instance, found that fox and crow removal had no consistent overall effect on the breeding success or population trends of lapwings, but that it did increase lapwing breeding success on those sites with highest fox and crow densities (Bolton et al. 2007). Additionally, studies across a range of upland farmland sites found that predation rates on lapwing nests and declines in their breeding numbers were greater where foxes were most abundant (O’Brien 2001). For curlew, high rates of nest predation have been identified as the likely cause of population declines in several instances (in the UK and elsewhere in Europe), with mammalian predators (particularly foxes) being most important (Mulder & Swaan 1988, Grant et al. 1999, Valkama & Currie 1999, Grimm 2005), so providing further indications that this species benefits from predator control. Overall, there is considerable evidence that legal predator control, as undertaken on grouse moors, does benefit populations of lapwing, golden plover and curlew. However, such benefits may not be ubiquitous; two studies undertaken on grouse moors in the North Pennines found that levels of golden plover and curlew breeding success were probably too low to maintain population stability, due largely to high predation rates on nests and chicks by stoats (Whittingham 1996, Robson 1998). For other moorland wader species, there is less evidence for population level benefits arising as a consequence of predator control. Thus, the experimental manipulation of predator control had no effect on snipe abundance (breeding success not being measured), and their densities were not related to the intensity of grouse moor management on heather moorlands (Tharme et al. 2001, Fletcher et al. 2010). At Langholm Moor in southern Scotland, snipe numbers increased following the cessation of gamekeeping (Baines et al. 2008). Data on dunlin and greenshank responses to predator control are lacking, although elsewhere declines in dunlin populations have been attributed to high predation rates on nests by crows and foxes (Jönsson 1991).

48

Perhaps unsurprisingly, given the above findings, there is some evidence of correlations between measures of predator and breeding wader populations at regional and national spatial scales. For example, Amar et al. (2011) found that lapwing declines were greatest where crows are most abundant and grouse moor management less intensive across multiple upland regions. However, these relationships were not found for curlew, dunlin or snipe, whilst golden plover, declines were greater in areas of intensive grouse moor management. These outcomes suggest that grouse moor management practices are not the only factors driving breeding wader population changes in the uplands, with proximity to forestry plantations also associated with greater declines for snipe and golden plover in the above study (Amar et al. 2011). Three of the predatory bird species associated with moorlands are largely, or entirely, ground nesting - i.e. hen harrier, merlin and short-eared owl. Given the relative vulnerability of ground nesting species to predation, then breeding success or population densities of these species may also be expected to benefit from legal predator control. However, little evidence currently exists to support this. Following the cessation of gamekeeping at Langholm Moor, harriers declined to low numbers (Baines et al. 2008), but the reasons for this decline are unclear and could include either an increase in predation of nests by foxes or a resumption of illegal persecution, or both (Ratcliffe 2007, Sotherton et al. 2009a). Data from across much of the hen harrier range in Scotland show that breeding productivity on moorland not managed for grouse shooting was no lower than on two grouse moors where there was no illegal persecution, whilst nest success did not differ between islands without foxes and areas within the fox range (Green & Etheridge 1999). In Wales, nest predation by foxes accounts for a high proportion of harrier breeding failures but, despite this, failures from predation did not differ between areas with and without gamekeepers, whilst the rate of predation on nests had no significant effect on annual variation in breeding productivity (Whitfield et al. 2008). Experimental removal of hooded crows from harrier territories on Orkney (where foxes are absent), also failed to improve hatching or breeding success (Amar & Redpath 2002), despite crows being important egg and chick predators of harriers in this area (Picozzi 1984). Merlin and short-eared owl nests may also be predated by foxes (Lockie 1955, Williams & Parr 1995), although there appears to be no evidence to suggest that such predation limits breeding densities. Overall, there are no indications that merlin are any more abundant on heather moorland managed for grouse shooting than on other heather moorland (Tharme et al. 2001). There appears to be little evidence for effects of legal predator control on the breeding densities of moorland passerines, except for carrion and hooded crow, which have reduced densities on grouse moors because they are the targets of control (Baines 1996, Tharme at al 2001, Fletcher et al. 2010). Predator control may increase meadow pipit breeding success, but it does not appear to result in higher abundance of this species or of skylark (Tharme et al. 2001, Fletcher et al. 2010). Indeed meadow pipit abundance is lower on grouse moors than other heather moorland (after accounting for potentially confounding habitat effects), probably as a consequence of greater muirburn levels (3.4.2) (Tharme et al. 2001, Smith et al. 2001). For two other passerine species considered in the extensive correlative study (wheatear and

49

whinchat), there were no significant differences in density between grouse moors and other moors, whilst ring ouzel nest survival rates did not differ between two study areas with different intensities of gamekeeping (Tharme et al. 2001, Burfield 2002). 3.4.4 Effects of illegal persecution of predatory birds Several species of predatory bird breeding on moorlands are subject to illegal killing, nest destruction and disturbance because of their documented or reputed impact in reducing the shootable surplus of red grouse (Thirgood et al. 2000b, Park et al. 2008). Such activity has the potential to limit population densities and restrict the geographical range of these species, and when in the past it was more widespread and indiscriminate it has been responsible for reducing populations of some species to low levels and even extinction in Britain and Ireland (Newton 1979). In Scotland at least, the illegal use of poisoned baits (a commonly used method of killing predatory birds) is greater on grouse moors than elsewhere in the uplands, with reported poisoning incidents being four to seven times higher than expected (on the basis of the amount of land) in areas where ‘strip’ muirburn (as produced by burning rotations on grouse moors) was recorded (Whitfield et al. 2003). The known and more widely perceived impact of hen harriers on red grouse has brought the species into direct conflict with grouse moor interests. Following the near extinction of the hen harrier in the early 20th century, the species began to re-occupy the UK mainland with the recovery coincident with a decline in grouse moors and game-keeping intensity, associated with the First and Second World War and post-war afforestation (Watson 1977, Hudson 1992). Persecution on grouse moors reduces hen harrier productivity and survival substantially, and limits population size and range (Etheridge et al. 1997, Whitfield et al. 2008, Anderson et al. 2009, Fielding et al. 2011). Recent population trends demonstrate declines in hen harrier abundance in those regions of Scotland where grouse moor management is most intensive (Fielding et al. 2011), whilst the species remains absent from most of the English uplands (Sim et al. 2007b). The situation is particularly severe on land managed for ‘driven’ grouse shooting (as opposed to ‘walked-up’ shooting), with such land estimated to be capable of supporting 500 successful breeding pairs of hen harrier but possibly holding as few as five successful pairs in 2008 (Redpath et al. 2010). It is also likely that persecution on grouse moors limits golden eagle populations, with non-occupation of apparently suitable territories and occupation by sub-adult birds associated with the occurrence of grouse moors in the Scottish Highlands (Whitfield et al. 2004, 2007). Peregrine breeding success is also lower on grouse moors, with numbers and breeding success on English grouse moors limited by persecution, where grouse moors appear to be ‘sink’ populations (Amar et al. 2012). However, in the case of the peregrine current persecution levels appear to have relatively little affect in restricting their range at a national scale (Gibbons et al. 1993, Thirgood et al. 2000b, Hardey et al. 2003). Persecution has been reported at merlin nests and, although often abundant on grouse moors, breeding numbers at Langholm Moor increased after the cessation of persecution (Rebecca & Cosnette 2003, Amar et al. 2008). Previous studies demonstrate that the breeding distributions

50

of both buzzard Buteo buteo and raven Corvus corax may be limited by grouse moors (possibly as a result of persecution), but it is unclear whether this is still true given the continued increases in the abundance of both species (Gibbons et al. 1995, Park et al. 2005, Sim et al. 2005, Risely et al 2010). The red kite Milvus milvus population in north Scotland also appears to be affected by persecution on grouse moors. Although kites generally breed on lower ground, their breeding areas in this region lie close to grouse moors and dispersing first and second-year birds are often poisoned in such areas, with this source of mortality limiting the population growth rate in north Scotland (Smart et al. 2010). In addition to the more obvious direct effects of illegal persecution, outlined above, there may be indirect effects on the populations of other moorland birds. Raptors (and owls) may predate the adults, juveniles and chicks of a wide range of moorland bird species, whilst ravens are often predators of their eggs and chicks (Newton et al. 1984, Ratcliffe 1993, 1997, Graham et al. 1995, Watson 1997, Redpath & Thirgood 1999). Given the apparent effect of legal predator control on the breeding densities of several bird species on grouse moors (3.4.3), then it is conceivable that persecution will have similar effects on some prey species (particularly given the extent to which populations of some predatory birds are reduced). Many of the studies undertaken to assess the specific impact of such predatory birds on their prey populations have focussed on gamebirds, with some demonstrating likely effects of raptors in limiting prey populations (Nielsen 1999, Valkama et al. 2005, Bro et al. 2006). In the UK, the most detailed study of this type demonstrated that predation by hen harriers and peregrines limited red grouse populations at Langholm Moor in southwest Scotland, following cessation of persecution, although the extent to which this finding applies to other moors elsewhere in the UK remains unclear (Thirgood et al. 2000a, 2000c, Park et al. 2008). Such detailed studies of the effects of predatory birds on the populations of moorland bird species other than red grouse are lacking, and little information exists from which to assess the likely effects of persecution on these other species. Studies at Langholm Moor demonstrated that harrier predation appeared to limit numbers of breeding meadow pipit, and possibly skylark, but analyses of population trend data provided no evidence for effects of increased raptor numbers on the abundance of breeding lapwing, golden plover or curlew (Amar et al. 2008, Baines et al. 2008). Correlations across a large number of plots, widely distributed across the British uplands, showed no statistically significant effects of raven abundance on either the abundance of, or the changes in the abundance of, five wader species (i.e. golden plover, lapwing, dunlin, snipe and curlew) over a c.10 – 20 year period up to either 2000 or 2002 (Amar et al. 2010). However, correlations between the change in raven abundance and the change in wader abundance during this period suggested possible negative effects of ravens on both lapwing and curlew, with relationships showing relatively large effect sizes and being close to statistical significance for these species. Raptors (particularly goshawk Accipiter gentilis and peregrine) are frequently the main proximate cause of mortality of black grouse (see 3.4.3), and predation by goshawks may be sufficient to drive black grouse population cycles in parts of northern Europe, where alternative prey are scarce (Tornberg et al. 2005). However, in the UK, increases in black grouse populations have been recorded in

51

situations with high levels of reported raptor predation (Lindley et al. 2003, Bowker et al. 2007, RSPB, unpubl. data), so that the extent to which illegal persecution on and around grouse moors affects black grouse populations is unclear. 3.4.5 Effects of other managements Of the other activities associated with grouse moors, managements concerned with moorland drainage and with controlling the abundance of sheep ticks Ixodes ricinus may be of greatest relevance to moorland birds (other than red grouse). Moorland drainage reduces the cover of plant species typical of high water tables (3.2.7). Although such effects may be relatively localised around the drains themselves, they have the potential to be detrimental to moorland bird populations, given that several species select vegetation-types typical of wet conditions (particularly during chick-rearing) and that breeding abundance may be correlated with the extent of such vegetation (Parr & Watson 1988, Baines et al. 1996, Whittingham et al. 2001, Pearce-Higgins & Grant 2006, Hoodless et al. 2007). Selection for vegetation typical of wet conditions is often linked to a greater abundance of invertebrate prey in such areas (Baines et al. 1996, Park et al. 2001, Whittingham et al. 2001), and drainage may reduce the abundance of some key invertebrate prey (e.g. craneflies) through the lowering of the water table (Buchanan et al. 2006, Pearce-Higgins et al. 2010). However, available data are insufficient to provide a reliable assessment of the likely effects of drainage on overall invertebrate abundance (3.3.5). In addition, reductions in cottongrass cover associated with drainage (3.2.7), may have more direct detrimental effects on black (and possibly red) grouse, because the inflorescences are an important, protein rich, food source for hens prior to the onset of egg-laying (Hudson 1992, Starling-Westerberg 2001). Conversely, drain blocking may have beneficial effects on moorland bird species, because this management tends to raise water tables, reducing the cover of vegetation typical of drier conditions (Worrall et al. 2007b, 3.2.7), with recent studies suggesting that it may also result in increased densities of cranefly larvae, which are important prey for some moorland bird species (Buchanan et al. 2006, Carroll et al. 2011, 3.3.5). The parasitic nematode worm reduces the breeding productivity and survival of red grouse, and is an important factor in contributing to population cycles in this species (Hudson 1992, Seivwright et al. 2005). In recent times, grouse moor managers have sought to overcome the effects of the parasitic worms both through the provision of medicated grit and by catching and directly dosing infected birds (Newborn & Foster 2002). The former approach entails applying an active anthelmintic chemical agent (flubendazole) directly onto grit, with the grit dispensed via grit boxes distributed across the moor. Red grouse receive the medicated treatment when they ingest the grit to aid the physical digestion of heather. Once ingested, the grit passes through the grouse and is released back into the environment. Data are lacking on the extent to which medicated grit is applied on grouse moors, and the trends in rates of application, but anecdotal accounts suggest that improved formulations of medicated grits aremay now being applied at a landscape scale in some areas. The possible side-effects of introducing powerful and long-lasting chemicals into moorland ecosystems are unstudied (Watson & Moss 2008).

52

Red grouse are susceptible to the louping ill virus, which may cause high levels of chick mortality in the species and reduce population size (Hudson 1992). Sheep ticks act as the vector for louping ill and so measures may be taken to control sheep ticks on grouse moors where louping ill is prevalent. The most widespread control measure is probably the treatment of sheep flocks with acaricides (chemicals that kill ticks or prevent their feeding) and, sometimes, vaccination to reduce the prevalence of louping ill in the flock (Newborn 2004, Laurenson et al. 2007). However, mountain hares and deer may also be important hosts for sheep ticks and this has led to the culling of mountain hares on some grouse moors in the Scottish Highlands (Laurenson et al. 2003, Newey 2010). These different managements may have a number of effects on moorland birds, other than red grouse. Whether or not other ground nesting birds on moorland are susceptible to tick borne diseases such as louping ill, is not known, but high levels of infestation with ticks have been found amongst wader chicks in some studies, with instances of associated mortality in curlew chicks (Grant et al. 1999, Newborn et al. 2009). Thus, some species may benefit from measures to control tick numbers on grouse moors. The culling of mountain hares, on the other hand, is likely to have detrimental effects on golden eagles, given that they are important prey for this species (3.4.2) (Watson & Moss 2008). 3.4.6 Summary and synthesis of effects on birds Moorlands provide breeding habitat for a bird assemblage of high conservation importance, which is particularly notable for its raptor, grouse and wader species. This assemblage includes several species that are current conservation priorities in the UK. The available evidence demonstrates that grouse moor management has major effects on the populations of several of these species. Through their role in maintaining Calluna cover and in creating areas of open vegetation by rotational muirburn, grouse moors are important in providing key habitat elements for some moorland bird species (e.g. nest-sites for some waders, notably golden plover, and suitable foraging habitat for several raptors), although rotational muirburn may also have detrimental effects on the habitat conditions of some species (e.g. by reducing, or eliminating, extensive stands of tall Calluna). However, the most striking effects of grouse moor management on bird populations result from the marked contrasts between the negative effects on several species of predatory bird (resulting largely from illegal persecution) and the beneficial effects on several species of wader and grouse (resulting mainly from control of predators such as corvids, foxes and mustelids, but probably also rotational muirburn – Table 2). Negative effects on predatory birds are sufficient to limit population densities and distribution of some species, whilst grouse moor management results in higher breeding success and breeding densities of some waders, and possibly black grouse. Both the illegal persecution of predatory birds and legal predator control regimes extend beyond the managed heather moorland on some grouse moors and on to moorland edge habitats. Thus, species more strongly associated with moorland edge habitats may also be affected by these managements (e.g. detrimental effects on red kites in north Scotland and beneficial effects on lapwings on rough grassland and in-bye habitats). The full importance of grouse

53

moors in maintaining (or slowing declines of) populations of certain species that are declining nationally and across much of their European range (e.g. black grouse, lapwing and curlew) is not known, but it is conceivable that they are critical in this respect. Broad-scale correlative analyses suggest that this may be the case for lapwing, but not for other waders (Amar et al. 2011). In terms of further research on grouse moors and birds, the key priority should be to establish the role of this management in determining population trends of these species at regional and national scales. Despite the fact that the evidence base for assessing the effects of grouse moor management is greater for birds than for other taxa, for some predatory species (e.g. short-eared owls) and waders (e.g. dunlin), and for most of the moorland passerines, the overall impact of grouse moor management on their populations is poorly understood (Table 2). This includes rapidly declining species, such as ring ouzel. More widely, it seems likely that grouse moor management generally limits overall bird species diversity by restricting scrub and woodland habitats (Thompson et al. 1997), and it would be valuable to establish the extent to which such habitats could be developed on moorland edge areas, without compromising the status of moorland bird species of conservation importance. Research on the importance of other managements that may be associated with grouse moors (e.g. drain blocking) is also warranted.

54

Table 2. Summary of likely effects of grouse moor management on moorland bird species. The effect of the main individual managements is assessed, along with the likely overall net effect of grouse moors.

Rotational muirburn

Legal predator control

Illegal persecution of

predatory birds

Net effect

Black grouse (+/-) Difficult to assess whether it will create net benefit (via increased young nutritious heather and, on blanket bog, cottongrass) or cost (via prevention of scrub/woodland development and loss of potential nests-sites). Net effect likely to depend upon situation in which muirburn occurs.

+ Several studies demonstrate increased breeding success and/or densities.

(+) Predation by raptors often main proximate cause of adult mortality. Some studies indicate possible effects on density.

+/= Evidence for population-level effects of predator control on black grouse suggest a positive effect overall, but studies have failed to detect associations between black grouse densities and grouse moor management or gamekeeper presence.

Hen harrier + Although muirburn reduces the densities of a key prey species (meadow pipit) and may reduce nest-site availability, prey capture rates are higher where muirburn occurs, and (possibly as a consequence) brood sizes at fledging of successful nests are larger on grouse moors.

(+) As a ground nester, expected to be vulnerable to predation but little evidence of major effects on breeding success or density.

- Persecution reduces breeding success, densities, survival and range.

- Persecution is the main factor affecting population size and range.

Golden eagle + Increases densities of important prey species (red grouse and mountain hare).

= Predation unlikely to have major effects on this species.

- Persecution reduces densities, survival and range.

- Persecution is the main factor affecting population size and range.

55

Merlin ? Reduces the densities of a key prey species (meadow pipit) and may reduce nest-site availability, but could increase prey availability (as for hen harrier).

(+) As a ground nester, expected to be vulnerable to predation but little evidence of major effects on breeding success or density.

? Extent and effects of persecution are unknown.

= Available data provide no evidence for effects.

Peregrine + Increases densities of an important prey species (red grouse).

= Predation unlikely to have major effects on this species.

- Persecution reduces breeding success and densities.

- Effects of persecution on overall population size less marked than for some other raptors, but is probably still sufficient to confer a net cost.

Golden plover

+ Creates suitable nesting habitat, and more generally maintains short vegetation selected by this species.

+ Causes higher breeding success and densities.

= Available data provide no evidence for effects.

+ Benefits from managements produce higher densities on grouse moors.

Dunlin ? Lack of suitable data from which to assess effects.

(+) No strong evidence of effects, but may be vulnerable to predation impacts.

? Insufficient data to assess effects.

? Insufficient knowledge to assess overall effects.

Snipe = Available data provide no evidence for effects.

(+) No strong evidence of effects, but (as a wader species) they may be vulnerable to predation impacts.

= Available data provide no evidence for effects.

= Experimental and correlative studies suggest little effect of grouse moor management on density.

Curlew + Creates areas selected as nesting habitat.

+ Causes higher breeding success and densities.

= Available data provide little evidence for effects.

+ Benefits from managements produce higher densities on grouse moors.

Greenshank ? Lack of data from which to assess effects.

(+) No strong evidence of effects, but (as a wader species) they may be vulnerable

? Insufficient data to assess effects.

? Insufficient knowledge to assess overall effects.

56

to predation impacts.

Short-eared owl

(-) May cause nest loss and reduce nest-site availability, but little evidence for effects.

(+) As a ground nester, expected to be vulnerable to predation but little evidence of major effects on breeding success or density.

? Extent and effects of persecution are unknown.

? Insufficient knowledge to assess overall effects.

Skylark = Available data provide little evidence for effects

(+) No effect on density. No data on breeding success effects, but benefits may be expected given effects on meadow pipits (the other abundant, widespread, moorland passerine).

+ Evidence of increase in abundance when raptor numbers reduced.

= Experimental and correlative studies suggest little effect of grouse moor management on density.

Meadow pipit

- Associated with reduced densities.

+ Increases nesting success, but no detectable effect on density.

+ Evidence of increase in abundance when raptor numbers reduced.

- Densities lower on grouse moors.

Whinchat ? Lack of suitable data from which to assess effects.

? Lack of suitable data from which to assess effects.

? Lack of data from which to assess effects.

= Correlative studies suggest little effect of grouse moor management on density.

Stonechat (-) May cause nest loss and reduce nest-site availability, but little evidence for effects.

? Lack of data from which to assess effects.

? Lack of data from which to assess effects.

? Insufficient knowledge to assess overall effects.

Wheatear ? Lack of suitable data from which to assess effects.

? Lack of suitable data from which to assess effects.

? Lack of data from which to assess effects.

= Correlative studies suggest little effect of grouse moor management on density.

57

Ring ouzel (-) Could reduce nest-site availability

= No evidence for effects on breeding success or density.

? Lack of data from which to assess effects.

? Insufficient knowledge to assess overall effects.

Raven (+) Possible benefits via increased grouse and wader densities, eggs and chicks of which may provide important food source

(+) Possible benefits via increased grouse and wader densities, eggs and chicks of which may provide important food source

- Persecution reduces densities and range.

- Effects of persecution on overall population size less marked than in the past but are probably still sufficient to confer a net cost.

Twite (-) May reduce nest-site availability, but little evidence for effects.

? Lack of suitable data from which to assess effects.

? Lack of data from which to assess effects.

? Insufficient knowledge to assess overall effects.

+ or -: Quantitative evidence of direct effect on species in terms of population density, breeding success or foraging success. =: Quantitative evidence is available but has failed to detect positive or negative effects. (+) or (-): Lack of direct evidence for effects (as defined for + and - above), but possible direction of effect given knowledge of nesting requirements, food supply and predation rates. ?: Lack of evidence for assessing effects.

58

4. Wider Environmental Effects: Grouse Moor Management, Carbon and Water 4.1 Grouse moors and wider ecosystem services Traditionally the main land-uses in the UK uplands, and specifically on moorlands, have been those of livestock farming, forestry for timber production, game management and, to lesser extents, recreation and biodiversity conservation, with the associated land managements predominating throughout the uplands (Bonn et al. 2009). In recent years, however, increased recognition has been given to the importance of UK moorlands as C stores, major sources of drinking water and as determinants of water flows and, hence, downstream flood risk (Smith et al. 2007, Holden et al. 2007a). Thus, moorlands are viewed as being critical in providing these key ecosystem services (Bonn et al. 2009). The moorland habitats of main interest in these respects are those overlying peat soils; primarily blanket bogs, but also wet heaths and some acid grassland (subsequently referred to as blanket peats in this section). The UK may hold 10 – 15% of the global resource of such blanket peat habitats (Lindsay et al. 1988, Milne & Brown 1997, Dunn & Freeman 2011), with northern peatlands globally estimated to hold c.20 – 40% of global soil C stocks (equivalent to 25 - > 50% of all atmospheric C), despite representing just 2% of global land area (Gorham 1991, Schlesinger 1991). Estimates of UK peatland C stores vary greatly but they are clearly large and likely to be over 3000 million tonnes, of which more than 70% occurs in Scotland (Cannell et al. 1999, Milne & Brown 1997, Smith et al. 2007). Additionally, the uplands provide 70% of the UK’s water supply, with colouration and the occurrence of particulate and dissolved organic carbon (POC and DOC – section 4.2) from peat soils being major factors affecting water quality, and incurring high financial costs for treatment (Orr 2008). As well as causing low aesthetic quality, DOC increases the potential for biological contamination and, when water is chlorinated, can lead to production of potentially carcinogenic tri-halomethanes7

, whose concentration in drinking water is limited by law in the UK (Hsu et al. 2001, Clay et al. 2009).

Although peatlands have represented major long-term C sinks they are highly sensitive to change, with even relatively small imbalances between the production and decay of C potentially causing them to shift from C sinks to sources (Bragg & Tallis 2001, Laiho 2006). Given the size of their C store, this could have considerable implications for climate change, as recognized by the Intergovernmental Panel on Climate Change (IPCC) in calling for the maintenance and enhancement of global C sinks and stores (Watson et al. 2000). Such concerns may be well founded. Evidence is emerging that climate warming may increase carbon dioxide (CO2) emissions from northern peatlands, hence creating positive feedback to climate change (Dorrepaal et al. 2009), whilst increases in riverine DOC from peat soils across the northern hemisphere (including the UK – 4.2) could represent the depletion of peatland C stores, in addition to detrimentally affecting water quality (Freeman et al. 2001, Hejzlar et al. 2003, Evans et al. 2005, Skjelkvåle et al. 2005, Monteith et al. 2007, Worrall & Burt 2007). Furthermore, data on soil C stores from England and Wales suggest that they have declined between the late 1970s and the early years of this

7 See - http://www.saskh2o.ca/PDF-WaterCommittee/DissolvedOrganicCarbon.pdf

59

century, with rates of loss an order of magnitude greater in peat than in other soils (Bellamy et al. 2005). Threats to the persistence of peatland C stores are not restricted to climatic change. Land management and other anthropogenic influences (e.g. air pollution – Tallis 1987) also affect the ability of peatlands to accumulate and store C, and may affect the vulnerability of C stores to climate change (Bragg & Tallis 2001, Holden et al. 2007a, Worrall & Evans 2009, Dunn & Freeman 2011). Any such management related effects are particularly relevant to the UK’s blanket peats, as these are amongst the most heavily managed of northern peatlands. Within the UK there has been considerable debate on the potential effects of different land managements on peatland C stores, with grouse moor management (and particularly muirburn) sometimes being a focus of such debate (Pearce 2006, Yallop et al. 2009). Despite this, recent reviews note the lack of scientific evidence on the environmental effects of some of the key management practices on blanket peats (Tucker 2003, Ramchunder et al. 2009). Worrall et al. (2010) conclude that no single management or restoration intervention is universally positive for greenhouse gases and note a paucity of data on changes in carbon budgets before and after management interventions. The most recent review of the heather and grass burning regulations in England and Wales found that a lack of scientific data meant it was difficult to provide evidence to support any major changes to the code (Glaves & Haycock 2005). 4.2 Peatland soil structure and carbon flows The nature of peat soils is relevant to understanding the different C flows in peatlands, and the effects of land management on their C balance and store. Soils in an intact peatland are often considered to comprise an upper oxic layer (the acrotelm) and a lower, permanently saturated, anoxic layer (the catotelm) (Fig. 5). The acrotelm is ‘active’ in that it contains live, growing, plant material, abundant peat-forming aerobic bacteria, and has a high rate of water movement through it due to the larger pore spaces of less decomposed material (Ingram 1978, Holden 2009a). Within the acrotelm, the water table fluctuates, and plant material undergoes a progressive degree of breakdown and decomposition with depth, leading to peat formation. Towards the base of the acrotelm and within the catotelm, there is little water movement, and the catotelm comprises older more decomposed peat. The roots of some species (e.g. cottongrass species) can, however, penetrate the catotelm, creating small pockets where conditions are less anaerobic (Clymo 1992). The extent to which peat soils adhere to this generalised model does, however, vary, and where the lowest position of the water table differs between wet and dry years the distinction between acrotelm and catotelm may be less clear (Rydin & Jeglum 2006).

60

Fig. 5. Simplified representation of carbon flow and peat formation in peatland. Encircled symbols represent gases and dashed arrows show microbial processes; NPP (net primary production), DOC (dissolved organic carbon), POC (particulate organic carbon), DIC (dissolved inorganic carbon), DCO2 (dissolved carbon dioxide). Adapted from Rydin and Jeglum (2006). Reprinted from The Biology of Peatlands (Series – Biology of Habitats) (Rydin & Jeglum) Page 249 (Figure. 12.4), (2006), by permission of Oxford University Press. Sequestration from the atmosphere is in the form of CO2, as a consequence of plant photosynthesis for primary productivity, and this has to exceed losses of CO2 from plant and soil respiration if peat is to accumulate. Thus, the accumulation of organic matter is dependent upon the rate of decomposition of dead plant material as well as the rate of plant productivity. The waterlogged and acidic conditions that are characteristic of blanket peat habitats are critical to peat formation because they inhibit decomposition, but plant species composition is also important. In particular, the dead material of Sphagnum mosses has a low decomposability, resulting from the production of phenolics that inhibit microbial decomposition, and the creation of acidic conditions that retards the decay of litter of both Sphagnum and co-occurring plants (Verhoeven & Liefveld 1997, Børsheim et al. 2001, Rydin et al. 2006). Other gaseous losses of C occur from methane (CH4) production resulting from the breakdown of organic matter under anaerobic conditions, with release to the atmosphere either directly as CH4 or, following oxidation, as CO2 (Fig. 5). The composition of the above ground vegetation can have a marked effect on CH4 release, with the aerenchyma system of some graminoids, such as cottongrass species and deer grass, acting as a conduit for direct CH4 release (Greenup et al. 2000, Marinier et al. 2004). Sphagnum mosses may again be important in this context, with suggestions that they may inhibit nitrogen uptake in decomposer bacteria and slow the production of CH4, whilst they may also provide an effective oxidising zone as CH4 moves towards the surface, hence reducing emissions (Lindsay 2010). Although CH4 fluxes in peatlands are generally small compared to those of CO2, they are of disproportionate importance in terms climate change implications. This is because the global warming potential (GWP) of CH4 is estimated to be 62 and 23 times greater than for CO2 over periods of 20 and 100 years, respectively, with the difference over time due to the shorter lifespan of CH4 in the atmosphere (IPCC 2001). Therefore, it is

DOC in rainfall

+ POC, DIC, DCO2

61

possible for a peatland to be a C sink, and yet still have a net warming effect on climate (Baird et al. 2009). The other C fluxes are non-gaseous and often fluvial, with DOC and POC being of main concern, whilst dissolved CO2 and dissolved inorganic C (DIC) are minor components (Billet et al. 2004, Worrall & Evans 2009). DOC includes both the colloidal fraction and organic C in true solution in waters leaving the peatland, being separated from POC by the size of the particles (although this distinction is essentially arbitrary, being defined by the size of filter used – Giller & Malmqvist 1998). Although small inputs of DOC occur via rainfall, fluvial fluxes are essentially unidirectional, representing losses of C from the system (Worrall & Evans 2009). Many previous studies of peatland C balance have focussed solely on the direct exchange of C between the atmosphere and the land surface, and have failed to account for fluvial fluxes. However, although the fluxes of DOC and POC may be substantially smaller than gross gaseous fluxes (but see below), the fact that they are unidirectional means that they can be of similar magnitude to the net exchange of CO2 (Smith et al. 2007). Their inclusion in calculations of C balance can have a substantial effect, and can be sufficient to turn a catchment from an apparent C sink to a substantial C source (Worrall et al. 2003, Billett et al. 2004). The inclusion of fluvial fluxes is therefore critical to assessing the overall C balance of peatlands, but less certainty appears to exist over the implications of the observed increases in DOC (4.1) for peatlands and their continued ‘stability’. Humic substances, such as DOC, are complex and their origins and mechanisms of production are unclear, but DOC may derive from plant litter, root exudates, microbial biomass or soil organic matter (Smith et al. 2007, Lindsay 2010). Increases in DOC could indicate a destabilisation of ‘old’ soil C (if the DOC derives from the catotelm), but could also arise from an accelerated throughput of ‘new’ C that has been assimilated relatively recently from the atmosphere (Smith et al. 2007). Studies, based upon measurements of radiocarbon (14C) to estimate DOC age yield mixed findings in this respect, although recent work from North Wales suggests that most DOC exported from organic soils is ‘new’ (i.e. post 1950s), especially at high flows, when the majority of DOC export occurs (Smith et al. 2007). The implication of these findings is that the observed increases in DOC from organic soils are likely to be due either to an increased production of ‘new’ DOC from recent plant material or to a decreasing retention of this DOC within the soil profile, and that they are unlikely to indicate large-scale peatland destabilisation. Whilst this issue is of clear importance in assessing the implications of increasing DOC for peatland maintenance and climate change, it is less relevant to the problems of maintaining potable water quality. In terms of the causes of the observed increases in DOC from peat soils (4.1), large-scale environmental processes, as opposed to local management factors, are likely to be the main drivers, given the wide geographical occurrence of these trends across habitats ranging from moorland and plantation forest in the UK, to unmanaged mixed forests in North America (Smith et al. 2007). Possible causes for these increases include decreases in acid deposition (soil solution DOC being highly sensitive to changes in acidity and ionic strength) and elevated levels of atmospheric CO2, which

62

may accelerate production of labile litter and root exudates, whilst a combination of rising air temperatures and associated drying of peat may make a partial contribution (Smith et al. 2007). However, recent studies show that, in the South Pennines, increases in rotational muirburn on blanket peats are stronger correlates of the observed marked increases in DOC export than are declines in sulphate deposition or increases in temperature, suggesting that in some situations local management factors (and specifically increases in rotational muirburn) may be important drivers of increased DOC export (Clutterbuck & Yallop 2010, Yallop et al. 2010, 4.3.2). Losses of POC from peatlands may occur as a result of the action of wind as well as water, with production of POC primarily controlled by vegetation cover and the physical erosion of the peat surface being the primary source of POC (Holden et al. 2007b, Smith et al. 2007). Thus, in most UK studies, DOC tends to be the major component of fluvial fluxes with POC a minor component only, but these findings have been derived from relatively intact peatlands (Hope et al. 1997, Dawson et al. 2002). However, POC accounted for c.53% of the fluvial flux at Moor House in the North Pennines, and 80% in studies from the more severely eroding peatlands of the South Pennines (Pawson et al. 2008, Worrall & Evans 2009), and it may represent the single most important flux in UK blanket peats suffering from severe erosion (Holden et al. 2007b). Losses of POC also occur via soil pipes (subsurface channels that transport water, sediment and solutes). Although soil pipes occur naturally in blanket peats, management causing desiccation may increase their prevalence, and hence the extent of subsurface erosion (Holden 2006). The extent of piping in blanket peats is correlated with the occurrence of Calluna, and although it is difficult to disentangle cause and effect (both being associated with drier conditions), field studies and laboratory experiments indicate that Calluna can increase pipe frequency. This may be as a consequence of the woody stem structure opening up preferential flow paths which allow pipe network development (Holden 2005a). Debate remains over the ultimate fate of POC, and to what extent it represents a loss of C to the atmosphere, as opposed to becoming buried in anoxic conditions in lakes or reservoirs. However, recent work appears to suggests that a large proportion of peatland POC could be rapidly altered to gaseous and dissolved forms in the fluvial system, so that C is made available in a climatically active form (Pawson et al. 2008). 4.3 Effects of grouse moor management on wider environmental issues 4.3.1 Geographical overlap of grouse moors with soil carbon stocks Fundamental to considering the importance of grouse moor management in affecting soil C store and fluxes is the extent to which areas with grouse moor management coincide with the main areas of soil C. Comparing the estimated distribution of soil C with an index of grouse moor management intensity (based upon the occurrence of ‘strip’ burns, taken to represent rotational muirburn – Anderson et al. 2009), shows considerable overlap at a broad scale (Fig. 6). However, the vast majority of Britain’s soil C stores occur in upland blanket peats, which are part of the wider upland environment within which grouse moors occur. Within the uplands, areas of overlap are limited, with some of the main areas of deep peat having little or no grouse moor

63

management (e.g. Shetland, the far north of the Scottish mainland, the western Highlands and Islands). Areas of most overlap include parts of the central Highlands and Southern Uplands, and the North and South Pennines, and within these areas finer-scaled information is required to assess the full extent of overlap. Additionally, grouse moors coincide with many areas overlying peat that occurs to lesser depths, but which will nonetheless make important contributions to the overall soil C store. Some areas of overlap are also sources of potable water for large human populations (e.g. The Peak District).

64

(A)

(B) Fig. 6. The distribution of soil carbon (A) and the intensity of ‘strip’ burning in Britain (B). Estimated soil carbon distribution is from Milne & Brown (1997) and uses data from the Soil Survey of England and Wales and the National Soils Database for Scotland, in combination with the 1990 ITE Land Cover Map (Barr et al. 1993) and Forest Authority bulk density data for Scottish peats. The index of ‘strip’ burning is from Anderson et al. (2009) and measures the extent of ‘strip’ burning at a 10 km resolution from mainly 2005 – 2006 satellite images. This measure is used by Anderson et al. (2009) to generate an index of grouse moor management intensity. Note, Shetland is omitted from map B but no grouse moor management occurs there. Figure (A) is reprinted from Journal of Environmental Management, Volume 49, Milne & Brown, Carbon in the Vegetation and Soils of Great Britain, pp 413-433, 1997, with permission of Elsevier; Figure (B) is reprinted from Biological Conservation, Volume 142 (3), Anderson et al., Using distribution models to test alternative hypotheses about a species’ environmental limits and recovery prospects, pp488-499, 2008, with permission from Elsevier.

65

4.3.2 Effects of rotational muirburn on carbon store and water quality Based upon the above summary (4.2), it is apparent that potential effects of rotational muirburn on both C store and water quality could arise via multiple routes. Plant community structure, soil temperature, water table position and the chemical composition of plant tissues and peat are considered to be the principal controls on the C cycle in peatlands (Blodau 2002), and it is likely that burning will affect some or all of these. Thus, burning has marked effects on plant species composition on blanket peat habitats (3.2.4) and, in this respect, the effects on Sphagnum mosses are likely to be of particular importance due to their role in facilitating peat accumulation (4.2). If it is assumed that Sphagnum generally does recover rapidly following low severity fires (3.2.6), then the impacts on peat accumulation will be determined largely by fire severity (and presumably also frequency of rotation). Short burning rotations on blanket peat habitats also tend to favour dominance by hair’s-tail cottongrass, or sometimes purple moor-grass, over Calluna (3.2.4), and are therefore expected to cause an increase in CH4 release, through the ‘methane shunt’ effects of their aerenchyma tissues (4.2). The extent to which rotational muirburn may influence soil piping, and resultant subsurface erosion, appears not to have been investigated, but in this instance short rotations might be beneficial, given the apparent positive association between Calluna cover and piping (short rotations on blanket bog favouring dominance by hair’s-tail cottongrass – 3.2.4). Through the loss of the vegetation canopy, burning will also remove shading and increase the exposure of the peat surface to solar irradiation. This may lead to higher soil temperatures and an increase in microbial activity, so increasing CO2 release, and perhaps also DOC via drying of peat (Kim & Tanaka 2003, Smith et al. 2007). Mitchell & McDonald (1992) conclude that severe peat drying, as a result of erosion, ditching, burning or drought, greatly increases the accumulation of decomposition products. Dried peat, particularly bare peat, takes much longer to re-wet thereby prolonging the time for decomposition, so that when saturation is achieved, subsequent runoff will carry a very high load of water discolouring organic solutes. Additionally, where ignition of the peat occurs, water repellent waxes may be produced, further affecting the moisture content of the peat (Clymo 1983), but such effects are more typical of severe wildfires and the extent to which they apply to rotational muirburn regimes is unknown. In contrast to effects that may cause drying out of the peat, studies on the Hard Hill plots at Moor House (3.2.5) have found that water tables are highest under the most frequent burning rotation (10 years) and lowest on unburnt plots, at least during the spring and summer months (Worrall et al. 2007a). This effect presumably results from deeper rooting and greater evapo-transpiration rates of the mature Calluna on the unburnt plots, again indicating the importance of effects that are mediated via changes in plant community. Other sources of C loss from burning may result from direct combustion of peat deposits (and hence release of C to the atmosphere) during the fire and from subsequent erosion. However, again, such effects are most characteristic of severe wildfires (Maltby et al. 1990), with post fire erosion affected mainly by the extent to which the peat surface is exposed and the time taken for re-vegetation to occur (Imeson 1971). Thus, under carefully controlled burning regimes on dry dwarf shrub

66

heath, Kinako & Gimingham (1980) showed that erosion increased for c.18 – 20 months after burning, after which time vegetation was well established. On podzols and peaty gleys in the North York Moors, post-fire erosion rates were slow where the ground surface comprised burnt Calluna remains, but up to 10 times higher where burning had been more severe and had exposed the peaty or mineral subsoil (Imeson 1971). However, in both of these studies post-fire erosion was measured by changes in the level of the soil surface, which may produce overestimates if there is long-term peat compaction and shrinkage following burning (Legg et al. 2010), the extent of which may vary between fires of different severity. A further factor that may influence the effects of muirburn on C budgets is that some C is produced during the fire in the form of black C, much of which is charcoal. Some of this at least may be extremely resistant to decomposition and may represent an important C sink (Farage et al. 2009, Clay & Worrall 2011). The production of black C during managed muirburn has not been estimated, although on one moor, where c.14% of the above-ground biomass survived a wildfire, 4.3% of the combusted biomass was estimated to be converted to black C (Clay & Worrall 2011). In other habitats, < 3% of the C consumed by the fire is generally converted to black C, and in well-managed, relatively low intensity burns it might be expected to be a small amount (Legg et al. 2010). This component has not been considered in studies assessing the effects of muirburn on C budgets. Clearly, the potential effects of rotational muirburn on C stores and balance, and on DOC production, are complex, with different effects operating in different directions and the magnitude of these effects likely to vary according to a wide range of factors (e.g. fire severity, length of burning rotation and initial vegetation condition). Given this, determining the overall impact of rotational muirburn is likely to be problematic, particularly because relatively few studies have addressed effects of rotational muirburn specifically, as opposed to effects of burning more generally. The evidence available from those studies that have considered the effects of rotational muirburn is considered below. Much of the research that is specific to rotational muirburn has been conducted on the Hard Hill plots at Moor House, where a long-term experiment provides the opportunity to compare 10 and 20-year burning rotations with unburnt vegetation (3.2.5). Recent work examined how the C stores and fluxes (CO2, CH4 and DOC) on the 10–year rotations differed from those on unburnt controls, 50 years after the plots had been established and when all had been subjected to an initial burn (Ward et al. 2007). In relation to C stores, differences were found only in the above ground vegetation, and in the upper peat horizons (including living plant roots and associated mycorrhizal fungi). In both cases, burning was associated with reduced C stores compared to the unburnt controls (by 56% and 60%, respectively). However, within the organic horizons (sampled to a depth of 1 m), which held over 99% of C stores, there were no differences between burnt and unburnt plots (Ward et al. 2007). Overall, these findings suggested an estimated total C loss attributable to burning of 25.5 g C m-2 y-1. However, an earlier study, undertaken on the same plots three years into the burning cycle, indicated losses of 73 g C m-2 y-1 from burning, suggesting some of the losses of C might be made up with time since burning, although the

67

methods of measurement also differed between the two studies (Garnett et al. 2001). In terms of CO2 fluxes, burning was associated with increases in the gross fluxes of both photosynthesis and respiration, indicating accelerated C processing rates under rotational muirburn regimes (Ward et al. 2007). Overall, photosynthesis increased at greater rates than respiration, making burnt plots a greater net CO2 sink than unburnt plots. Burning was also associated with a small (overall 12%), but significant, reduction in CH4 release, despite the greater graminoid, and specifically hair’s-tail cottongrass, cover under the 10-year burning rotation (and despite the fact that water tables are reported to be highest under the 10-year rotation plots, at least during summer months – see above). Soil water DOC (from depths of 10 and 50 cm) did not differ in relation to burning. The reduction in C stores in the upper peat horizon under burning appears to be difficult to reconcile with the greater CO2 sequestration and reduced CH4 release recorded on the burnt plots in this study. However, burning may have been associated with greater C losses from either erosion and/or direct combustion of peat, whilst the measurements of gaseous flux (and of DOC) are a snapshot at the end of the burning cycle only, and may not represent the overall flux during the full 10 years of the rotation. Another study, measuring C losses from the combustion of the vegetation during muirburn, suggested that muirburn causes minor losses from the system only (Farage et al. 2009). However, data from this study were collected from just two individual burns, which appeared to produce anomalously low estimates of the amount of material combusted during burning (Legg et al. 2010). Furthermore, estimates of other (substantially larger) components of the overall C budget in this study were taken from the published literature but, in some cases at least, from habitats that are not directly comparable with upland heaths and bogs (e.g. soil respiration rates relating to lowland heaths in southern England, where both climate and soil types are markedly different). A further attempt to assess the effects of rotational muirburn on the overall C budget of blanket bogs combined data from a wide range of disparate studies measuring at least one of the different relevant fluxes, but including few studies that measured all main fluxes (Worrall & Bell 2009). This suggested a negative impact of muirburn, but the assessment is dependent upon major assumptions concerning the relative importance of the different fluxes to the overall budget (being based on data from a single site), and incorporates data that are unlikely to reliably reflect the effects of rotational muirburn on blanket bogs (e.g. from non-UK sites and wildfires). Other studies of DOC flux in relation to burning have also been carried out on the Hard Hill plots, considering the 20-year rotation, as well as the 10-year rotation and unburnt controls, and examining soil water and runoff water at different stages in the burning cycle (Worrall et al. 2007a, Clay et al. 2009). The longer-term data from these studies cover almost three complete years, and show no detectable differences between the different burning treatments in either soil water or runoff water DOC at the end of the burning cycle (Clay et al. 2009). However, effects on water quality at the end of the burning cycle were detected, with the 10-year rotation having significantly higher absorbance values (a measure of water colour) in runoff water than other treatments, although absorbance values for soil water were significantly lower in the 20-year rotation than in the other treatments. In the weeks immediately

68

following burning, both absorbance and DOC showed peaks in both soil and runoff water, and DOC was slightly (but not significantly) higher in soil water in the year after burning, whilst in runoff water it decreased relative to the unburnt control in the year following burning. Overall, these studies suggested that burning in itself did not lead to dramatic increases in soil or runoff water DOC (Clay et al. 2009). A potential limitation of the studies carried out on the Hard Hill plots is that the burning regimes on these plots may not accurately reflect those typically found on actual grouse moors, so affecting the wider applicability of the findings. Thus, because they occur on a national nature reserve, burning may be carried out more sensitively on the Hard Hill plots than is typical on many grouse moors, with a greater likelihood of cooler, less severe, fires and less chance of fires damaging the peat surface. Certainly, it will be considerably more difficult to achieve the burning requirements on a grouse moor, and gamekeepers managing grouse moors seem likely to face greater time constraints and work pressures and be less likely to ensure the same level of care when burning on blanket peats. Such differences could be of fundamental importance in determining the overall effects of burning on C stores and DOC levels. Additionally, the studies on the Hard Hill plots inevitably relate to a much smaller scale than that of the moorland holding or the catchment. At larger spatial scales, correlational studies have examined the relationships between rotational muirburn and DOC concentration on grouse moors (Yallop & Clutterbuck 2009,Yallop et al. 2009). These studies were conducted primarily on 50 small catchments (0.3 – 3 km2 in area), located within three distinct areas of the Pennines and one area of the North York Moors. Habitat and soil type were measured within each small catchment, together with the extent of different categories of burnt vegetation (as measured from 25 cm resolution aerial photographs). Amongst these different measurements, DOC concentration in the catchment outflow streams was most highly correlated with the amount of recent burning on deep peat (recent burns defined as those without visible Calluna regeneration) (Yallop & Clutterbuck 2009). Correlations with the amount of recent burning on other soil types, or with overall burning, were either non-significant or considerably less significant, highlighting that the apparent strong effects of rotational muirburn on DOC concentration were associated with burning on deep peat, as opposed to burning on moorland more generally. However, the amount of new management burning (and associated area of bare peat) on deep peat was also highly positively correlated with the overall extent of deep peat, which is itself a known determinant of DOC concentration (Aitkenhead et al. 1999). Based upon the data presented by Yallop & Clutterbuck (2009), the effects of recent burning on DOC concentration appear to remain statistically significant even after first accounting for the extent of deep peat. This work therefore provides evidence that rotational muirburn (at least on deep peat) contributes to increases in DOC. Further work in the South Pennines also suggests a temporal association between DOC export and rotational muirburn, with a doubling of DOC export in three blanket peat catchments being associated with increases in rotational muirburn over this period, whilst other studies indicate that catchments with little or no burn management on blanket peat do not show such increase in DOC export (Clutterbuck & Yallop 2010, Yallop et al. 2010). A further notable finding from Yallop & Clutterbuck (2009) is that field

69

measurements from a sample of burns classed as recent in their study areas revealed an average cover for bare ground of c.85%. Based upon previous studies, burning that exposed this level of bare ground could potentially lead to relatively high rates of peat erosion (Imeson 1971). A study of changes in water colour (DOC) in the headwaters of the River Nidd, North Yorkshire, found no evidence that burning had an effect on water colour in the upland sub-catchments studied (Chapman et al. 2010). The conclusions of this study were questioned in a critique of the methods used (Yallop et al. 2011). The effect of drainage and burning on the release of DOC from deep peat soils and the associated impacts on water colour continues to be the subject of much debate (Yallop et al. 2011, Chapman et al. 2011). Holden et al. (2011) conclude that whilst there is no clear cut message from research that has taken place to date at different scales, the balance of evidence suggests that moorland burning results in increased colour in raw water. Holden et al. (2011) also note that other water quality variables may deteriorate with burning. These finding are consistent with the results of a recent study of drinking water colour between 1995-2006, undertaken across 18 sites in Yorkshire, which identified rotational muirburn and vegetation type (particularly heather) as the two most statistically significant variables influencing water colour (Grayson et al. 2012). In addition to the direct effects of rotational muirburn on the various C fluxes and overall C accumulation, a further factor that has to be considered in assessing the full effect of rotational muirburn on C stores and DOC release is the extent to which this management reduces the risk of wildfire. Rotational muirburn prevents (or at least reduces) the build up of litter, and the development of extensive stands of mature and degenerate Calluna, which have high fuel loads and may be highly combustible, particularly during extended dry periods causing low soil or moss/litter moisture levels (Davies et al. 2008, 2010). Under such conditions, severe wildfires can occur, with a greater risk of killing off Sphagnum and exposing bare peat for longer, due to the low levels of vegetative regeneration (3.2.4). Furthermore, such wildfires may result in the ignition of the peat deposits, and may burn over large areas (e.g. > 5000 ha – BBC 2007) for several days, resulting in bare ground that may take decades to recover (Maltby et al. 1990, Davies et al. 2008). Whether or not the detrimental effects from wildfires outweigh those that may result from rotational muirburn will depend primarily upon the extent to which rotational muirburn actually causes detrimental effects, the risk of severe wildfires occurring, and the extent to which these risks are reduced by having rotational muirburn regimes in place. These are largely unknown, but it seems likely that they will vary considerably between regions, and possibly also between moors within a region (e.g. climate and visitor pressure are two factors affecting the risk of wildfire – McMorrow et al. 2009). The potential importance of wildfire as a factor causing degradation of blanket peats and, hence reduction of C stores, and the possible role of rotational muirburn in preventing such events is illustrated by an example from the Forest of Bowland. Here a series of dry years in the early 1900s and decline in moorland management due to a shortage of gamekeepers after the First World War, combined with an exceptional summer drought in 1921, are believed to have precipitated a catastrophic burn, leading to the onset of summit-type erosion and loss of Sphagnum species from many areas (Mackay & Tallis 1996).

70

Despite the breadth and number of studies undertaken on this subject, and the extent to which managed burning occurs across the uplands, two recent JNCC reviews concluded that there is surprisingly little published information on the impact of rotational burning on complete greenhouse gas budgets and no complete studies of more intensively burnt areas or areas subject to wildfire (Evans et al. 2011, Worrall et al. 2011a). 4.3.3 Effects of drainage and drain blocking on carbon stocks and water quality Moorland drainage causes a lowering of the water table and an associated change in plant community structure (3.2.7). Lowering of the water table leads to desiccation of the peat, with a subsequent increase in aerobic microbial activity and hence CO2 release (4.2). Although earlier studies indicate that effects on the water table tended to be restricted to within a few metres of the drains (Stewart & Lance 1991, 3.2.7), more recent work demonstrates cumulative effects, particularly where drains are close together, with more widespread lowering of the water table (Holden et al. 2006). In terms of changes in plant species composition, the resulting decline in the cover of Sphagnum spp (3.2.7) is likely to substantially reduce the prospects for peat accumulation (4.2), and likely increases in Calluna cover (but see Coulson et al. 1990, 3.2.7) may cause increased soil piping and subsequent losses of POC from the system (4.2). Conversely, benefits to C balance could arise through reductions in the cover of cottongrass species, which combined with the lowering of the water table, may reduce CH4 emissions (Marinier et al. 2004, Strack et al. 2004, 4.2). Drainage may also reduce the potential for DOC retention within the soil (Holden et al. 2004), with several studies finding higher levels of DOC and/or water colour on drained slopes or in drained catchments than for undrained comparators (Mitchell & MacDonald 1995, Wallage et al. 2006, Gibson et al. 2009). However, such effects may not occur consistently across all catchment types, whilst there may also be seasonal interactions with drainage effects on DOC concentration, so that increases might occur at some times of year but decreases at other times (Hughes et al. 1997, Holden et al. 2007a). The extensive correlative study on DOC concentration undertaken by Yallop et al. (2009) also considered the presence of drainage within their catchments, as well as the extent of burning (4.3.2). No effect of drainage was detected, leading to the conclusion that drainage had little direct influence on DOC concentration in their study areas. However, drained catchments were virtually limited to one of the four study areas, where 11 of the 15 catchments were drained, so that drainage may have been confounded by other characteristics that differed between this study area and the other three study areas. Drains on blanket peats also represent a major source of erosion (and hence POC) in many situations. The exposure of bare peat, and the resulting desiccation, will increase the susceptibility to wind and rain erosion, with evidence from a range of sites showing drains to be major sources of fine sediment to the stream system (Holden et al. 2007a, 2007c). The drainage ditches may also be subject to severe scouring, widening and deepening, often by several metres (Holden 2009a). However, these effects are not ubiquitous and factors such as topography may have a major influence on drain erosion. Thus, across four blanket peat sites located in northern England and southern and northern Scotland, natural infilling of drains

71

was frequent on gentle slopes (< 4o), with drains on slopes < 2o rarely eroded, but with drains on steeper slopes (> 4o) rarely infilled (Holden et al. 2007c). The extent of overhanging vegetation also affected the re-vegetation of drains. Such effects may account for the observation that many drained peatlands show no visual evidence of erosion (Smith et al. 2007). Drainage also leads to increased export of POC via effects on soil pipes. Measurements from a wide range of blanket peat sites across Britain show that drained sites have significantly higher amounts of soil piping than those that have not been drained, and that soil pipe density and pipe diameter increase with the age of the drainage network (Holden 2005b, 2006). Thus, soil pipe density is predicted to have doubled 35 years after installation of the drains and such effects can continue even 80 years after drainage. These effects were predicted to cause an exponential increase with time since installation in POC export from soil pipes (Holden 2006). Drain blocking raises the local water table (Worrall et al. 2007b, Armstrong et al. 2010) and so might be expected to reverse the effects of drainage on C stores and DOC export. However, whilst too few studies have yet been undertaken to assess the full effects of this management, it seems likely that such a view is overly simplistic. Studies on the response in terms of plant species composition are only now becoming available and still derive from a restricted range of sites (3.2.7), whilst the few UK studies examining gaseous C flux in relation to this management are still in progress, or have been relatively small scale and have failed to detect statistically significant effects of management (Gray 2006, Baird et al. 2009). Studies of drained peatland sites elsewhere have found that rewetting is associated with increases in CH4 emissions, although in some (but not all) cases such increases appear insufficient to negate the GWP benefits of restoration resulting from the changes to net CO2 exchange (Tuittila et al. 2000, Waddington & Day 2007, Baird et al. 2009). These findings derive from the restoration of cutover raised bogs and may not be entirely applicable to drain blocking on blanket bogs, but where blocking creates open water and wet areas supporting plants such as common cottongrass and bogbean (Menyanthes trifoliata)this is likely to result in increased CH4 emissions (Baird et al. 2009). Whilst such an effect is unlikely to negate benefits to overall C balance, it may negate any benefits in terms of GWP, due to the greater GWP of CH4 compared to CO2 (4.2). However, in the longer-term, the overall net effect of blocking on CH4 emissions may depend upon the extent to which increases in species such as common cottongrass and bogbean are balanced by increases in Sphagnum cover, which may act to reduce CH4 emissions (4.2). To date, UK studies on drain blocking in blanket peats have focused primarily on the effects on DOC. These have sometimes produced mixed results, suggesting a potentially complex situation in which the effects of blocking may not be consistent or readily predictable, although some of this inconsistency may result from differences in exactly what has been measured, as well as in the length of time since blocking took place. Thus, whilst some studies consider effects on DOC concentration only, others have considered the overall DOC yield, which is a product of both concentration and discharge rates, and is ultimately of most relevance.

72

Of studies focused on DOC concentration, a comparison of undrained, drained and blocked sites within a single catchment in northern England found that DOC concentration and absorbance were lowest at the blocked site (and highest at the drained site). These differences were statistically significant, with DOC concentration and absorbance being 60% and 70% lower, respectively, on the blocked site compared to the drained site (Wallage et al. 2006). The composition of DOC at the blocked site differed from that at the two other sites, suggesting that drain blocking may also have altered the DOC production and/or transportation processes. However, this study lacked any data from the pre-blocking period, so that differences could have been due to confounding site effects (although the three sites were located within 1 km of each other and on comparable slopes, soil types and vegetation). In another study, examining short-term effects only, DOC concentration increased in the first 10 months following blocking, relative to pre-blocking concentrations and concentrations in unblocked drains, although no associated catchment-scale change in river water colour was detected (Worrall et al. 2007b). A wide-scale survey, encompassing 320 drains (including 266 which were blocked) on 32 sites in northern England and northern Scotland, found that DOC concentrations were on average 28% lower in blocked drains with flowing water than in unblocked drains with flowing water, whilst there was no difference in DOC concentrations between blocked and unblocked drains with standing water (drains with standing water having higher concentrations than those with flowing water) (Armstrong et al. 2010). Although the difference between the blocked and unblocked sites with flowing water was not statistically significant (P < 0.10), at all but one of the 60% of sites where blocked drains had lower DOC concentrations than unblocked drains, blocking had occurred at least seven years previously. At the remaining 40% of sites, blocking had occurred within the past three years, suggesting that in many instances the effects of blocking on DOC may be detectable only after several years. Furthermore, a multivariate analysis of these data, incorporating a range of site and environmental effects, detected a statistically significant effect of blocking on DOC concentration, although no details are given on the magnitude of this effect (Armstrong et al. 2010). Two studies considering effects on DOC yield suggest benefits of drain blocking. In the first, there was evidence of a significant reduction in DOC export (of 39% on average), resulting mainly from an effect of drain blocking in decreasing the flow from the drain, with only a minimal reduction in DOC concentration following blocking (Gibson et al. 2009). A second study examined effects at a landscape scale, comparing DOC concentrations and yields between blocked and unblocked drains and in relation to pre-blocking measures (Wilson et al. 2011a, b). This demonstrated that blocking resulted in a marked decline in DOC yield in streams compared to the unblocked situation, despite a lesser increase in yield in the drains following blocking and no strong effect of blocking on DOC concentration. Similarly, POC yield in streams declined following blocking, although there were no marked effects on yield in drains or in POC concentration (Wilson et al. 2011a, b). Thus, benefits of blocking arose mainly from reduced water flows from the drains and a change in the chemistry of the DOC. A limitation of this study is that it considers effects only up to c.25 months after blocking.

73

4.3.4 Effects of grouse moor management on water flows In recent years a number of major flood events have occurred in the UK with substantial socio-economic costs, and attention has been given to understanding the role of upland land management on downstream water flows and associated flood risk. Many grouse moors lie in the headwaters of some of our major river systems, so that some of this attention focuses on how these areas are managed. Whilst the majority of grouse moors in the English and Scottish uplands often include areas of blanket bog, few comprise extensive areas of blanket bog in favourable condition. This is largely because over generations, moorland managers (grouse moor managers and farmers) have sought to ‘improve’ upland heath and bog habitats for shooting and grazing through burning and drainage. Whilst fire has been a key management tool going back over hundreds of years, the cutting of drains is relatively recent and increased greatly in the 1970s and 1980s (1.1 and 3.2.7). Until recently, blanket bog was thought to function rather like a sponge – capturing and holding rainwater and then releasing it slowly. As such, blanket bogs were thought to reduce the effect of floods at times of high rainfall and sustain base flows during times of low rainfall (Holden & Burt 2003). O’Connell et al. (2004) reviewed the impacts of rural land use and management on flood generation and concluded that while there is substantial evidence that changes in land use and management practices affect runoff generation at the local scale, the effects are complex. Here we review recent literature on the impact of grouse moor management (e.g. drainage, burning) and more recent habitat restoration (e.g. re-wetting) on water storage and water flows. A number of environmental changes in blanket peatlands (e.g. drainage, erosion) have been linked to increased flood peaks downstream (Grayson et al. 2010). Drainage results in an immediate lowering of the water table adjacent to the drain (3.2.7, Holden et al. 2004, Holden et al 2006, Wilson et al 2010b), with one study finding long-term differences in the hydrological response of drained catchments some 50 years after the drains were cut (Holden et al. 2006). A number of older studies found that, in the short term, drainage in the uplands might actually increase the storage available for storm water by lowering water tables and increasing the antecedent soil moisture storage capacity (e.g. Burke 1975, Robinson 1986). A more recent review by Holden et al. (2004) found reports of drainage having positive and negative impacts on flood peak, suggesting that the results depend on a complex interplay of rainfall intensity with such factors as hydraulic conductivity, topography, drainage density and soil moisture capacity, which are in turn specific to the site, and location on the hillslope. Until relatively recently, it has been difficult to show if ditch blocking has any impact on the flood peak downstream. Following drain blocking, the volume of water flowing along drainage channels may be significantly reduced particularly where escape routes have been created to allow water to flow with the natural slope (Holden 2009b). A number of studies have found that the installation of dams in drains raises the water table and slows down the water discharged through the drainage network (Worrall et al. 2007b, Wilson et al. 2010b). At Lake Vyrnwy in Wales, the stability of the water table changed after drain blocking with extreme

74

fluctuations in the summer months removed and fluctuations in the autumn declining considerably (Wilson et al. 2010b, 2011b). The authors note that although high water tables result in lower storage potential, at times of high rainfall this did not necessarily translate into greater flows in response to restoration. Indeed the study at Vyrnwy found that restoration achieved apparently contradictory results of both raising water tables (albeit slightly) and decreasing the severity of peak flows. The authors suggest that this may be due to the reduction in connectivity of the drainage network caused by grip blocking and an increase in importance of overland flow. This shift in process has been modelled by Lane et al. (2003) who found diffuse overland and subsurface flow was up to two orders of magnitude slower than that delivered by grips. The movement of water across a bog surface typically depends on slope, water depth and surface vegetation. Where land management results in the loss of or a change to the surface vegetation, the overland flow of water may change. A study in the Upper Wharfe catchment found that overland flow velocities (movement of water across the surface) were slowest over a bog surface dominated by Sphagnum and significantly faster over areas dominated by bare peat. Recovering areas, dominated by bog cotton exhibited an intermediate effect (Holden et al. 2008). Thus, a blanket peatland landscape dominated by Sphagnum may be expected to release overland flow more slowly than the same peatland dominated by bare peat or bog cotton (Grayson et al. 2010). Ultimately, our ability to more precisely predict the impacts of grouse moor management (including peatland restoration works) on water flows is challenged by the problem of disentangling multiple and interacting management activities (e.g. grazing, burning, drainage, restoration), and carefully designed, large-scale, experiments may be required to reach firm conclusions on impacts. 4.3.5 Summary and synthesis of wider environmental effects Upland blanket peats represent important C stores, holding the vast majority of the UK’s soil C, whilst moorlands are also major sources of drinking water. In recent years, there has been increasing concern over the effects of land-use (including grouse moor management) on both C stores and water quality. Rotational muirburn is of greatest importance in determining how grouse moors affect C stores and water quality, given its fundamental role to grouse moor management and its potential to affect vegetation, soils and hydrology. The effects of rotational muirburn on these aspects are likely to be complex, with different effects operating in different directions and with the magnitude of these effects likely to vary according to a wide range of factors. These include many of the same factors that are fundamental to determining the effects of rotational muirburn on vegetation, and on which the existing knowledge base is inadequate (3.2.8). In terms of effects on C stores and water quality, much may depend upon the severity of fires on grouse moors and the rotation lengths applied, and hence there may be considerable geographical and temporal variation in the nature and magnitude of effects.

75

Very little evidence is available from which to assess the impact of rotational muirburn on overall C stores, and few studies have tackled this topic in detail. Findings from one detailed study suggest annual losses of total soil C were small, with burning also associated with a greater net uptake of CO2. However, these findings were obtained from a long-term experimental site, and it is unclear how representative this is of burning practices on real grouse moors. Similarly, findings from this same site suggest that rotational muirburn has little effect on soil or runoff water DOC levels, but other extensive, correlative, studies provide strong evidence that rotational muirburn on deep peat soils causes marked increases in DOC at the catchment scale. In addition to the direct effects, rotational muirburn may confer benefits indirectly by reducing the risks of severe wildfires, which would have greater detrimental impacts on C stores and DOC export. Whether or not the detrimental effects of wildfires outweigh those that may result from rotational muirburn is likely to vary regionally, according to the risk of such wildfires. Furthermore, it is important to bear in mind that grouse moors are rare or absent from many of the UK’s most extensive areas of blanket peats, so that other factors are likely to be of greater importance in affecting peatland C stores at the national scale. However, where intensive burning regimes (short rotations) are undertaken on blanket peat habitats, detrimental effects may be inevitable (Yallop et al. 2006). Past drainage is likely to have caused increased losses of soil C via both POC and DOC, and drain blocking may offer a means of reversing, or at least reducing, these effects. The evidence to date suggests that drain blocking does reduce DOC concentrations in some instances but that the response is variable and inconsistent, whilst the restoration of higher, stable water tables that reduces water flows in the drains can be important in causing a reduction in the overall yields of both DOC and POC. As for rotational muirburn, there are multiple ways in which drain blocking might affect overall C stores, as well as the Global Warming Potential of peatlands, but little research on these further aspects has yet been undertaken. There are fundamental knowledge gaps concerning the effects of grouse moor management on C stores and DOC export. Currently, insufficient evidence is available to allow conclusions to be reached on whether rotational muirburn causes major detrimental effects to overall C stores, after accounting for benefits that may accrue from reducing risks of wildfires. However, there is growing evidence that intensive rotational muirburn (involving short rotations) on deep peat causes marked increases in DOC export. Given the wide range of potential ways in which rotational muirburn could affect soil C stores, it is clear that further, wider-scale, research on these topics is urgently required. In this respect, there is some overlap with requirements for assessing effects on vegetation, because an improved understanding of actual burning regimes on grouse moors and on responses of key plant taxa is one aspect of this. An improved understanding of the full effects of drain blocking on overall C stores (and GWP) is also required, as is research to determine the factors affecting

76

the efficacy of drain blocking in reducing DOC export. Drain blocking may offer a means for grouse moors to produce environmental benefits and determining how this management interacts with rotational muirburn in affecting C stores and DOC may prove to be a valuable avenue for investigation. Finally, a more detailed and accurate assessment of the extent to which grouse moors overlay deep peat soils is essential to determining the extent of potential impacts at a national scale, and for focusing attention on areas of greatest concern. Many grouse moors are located in the headwaters of major river systems. The way in which grouse moors are managed may play a role in moderating the effects of major rainfall events on downstream flooding. Drainage has been found to have both positive and negative impacts on flood peak, suggesting that the results depend on a complex interplay of factors. However, following drain blocking, the volume of water flowing along drainage channel outlets may be significantly reduced. Drain blocking raises the water table and slows down the water discharged through the drainage network. This may be because blocking drains reduces the connectivity between drains and allows the water to escape from the drain network and flow overland with the natural gradient. The movement of water across a bog surface typically depends on slope, water depth and surface vegetation. Water moves more rapidly over smooth surfaces (e.g. bare peat) and more slowly over rough surfaces (e.g. moss dominated). Whilst the precise way in which grouse moors are managed may be important in moderating water flows, it is difficult to disentangle the multiple and interacting effects of grazing, burning, drainage and habitat restoration without conducting further research at a range of scales. 5. Acknowledgements The review was commissioned by the RSPB’s Uplands Working Group. We are grateful to colleagues, especially Jeremy Wilson, for reading through earlier drafts of this review and providing constructive comments and advice. We are grateful to a number of grouse moor owners, managers and keepers and others with an interest in upland land management who have openly shared their knowledge with us, thereby helping us complete this review. Thanks also to Matt Davies and Alistair Hamilton for providing access to unpublished analyses, and to Angus MacDonald for supplying a copy of an unpublished report. Finally, we are grateful to the RSPB Library (Elizabeth Allen, Ian Dawson and Lynn Giddings) for helping source key references.

77

6. References

Aitkenhead, J.A., Hope, D. & Billet M.F. 1999. The relationship between dissolved organic carbon in stream water and soil organic carbon pools at different spatial scales. Hydrological Processes, 13, 1289–302.

Aked, T.J. 1984. Soil invertebrate fauna of moorland habitats and their relationship to fire damaged areas. In Moorland Management. North York Moors National Park, Helmsley.

Allchin, E.A., Putwain, P.D., Mortimer, A.M. & Webb, N.R. 1996. Burning heathland for management: fire temperatures and vegetative regeneration. Aspects of Applied Biology, 44, 407-412.

Alonso, I., Hartley, S.E. & Thurlow, M. 2001. Competition between heather and grasses on Scottish moorlands: interacting effects of nutrient enrichment and grazing regime. Journal of Vegetation Science, 12, 249-260.

Amar, A., Court, I.R., Davison, M., Grimshaw, T., Pickford, T. & Raw, D. 2012. Linking nest histories, remotely sensed land use data and wildlife crime records to explore the impact of grouse moor management on peregrine falcon populations. Biological Conservation, 145, 86-94

Amar, A. & Redpath, S.M. 2002. Determining the cause of the hen harrier decline on the Orkney Islands: an experimental test of two hypotheses. Animal Conservation, 5, 21-28.

Amar, A., Grant, M., Buchanan, G., Sim, I., Wilson, J., Pearce-Higgins, J.W. & Redpath, S. 2011. Exploring the relationships between wader declines and current land use in the British uplands. Bird Study, 58, 13-26.

Amar, A., Redpath, S., Sim, I. & Buchanan, G. 2010. Spatial and temporal associations between recovering populations of common raven Corvus corax and British upland wader populations. Journal of Applied Ecology, 47, 253–262.

Amar, A., Thirgood, S., Pearce-Higgins, J.W. & Redpath, S. 2008. The impact of raptors on the abundance of upland passerines and waders. Oikos, 117, 1143-1152.

Anderson, B.J., Arroyo, B.E., Collingham, Y.C., Etheridge, B., Fernandez-De-Simon, J., Gillings, S., Gregory, R.D., Leckie, F.M., Sim, I.M.W., Thomas, C.D., Travis, J. & Redpath, S.M. 2009. Using distribution models to test alternative hypotheses about a species’ environmental limits and recovery prospects. Biological Conservation, 142, 488-499.

Anderson, P. & Yalden, D.W. 1981. Increased sheep numbers and the loss of heather moorland in the Peak District, England. Biological Conservation, 20, 195-213.

Angelstam, P. 1984. Sexual and seasonal differences in mortality of the black grouse Tetrao tetrix in boreal Sweden. Ornis Scandinavica, 15, 123-124.

Armstrong, A., Holden, J., Kay, P., Brian, F., Foulger, M., Gledhill, S., McDonald, A. & Walker, A. 2010. The impact of peatland drain-blocking on dissolved organic carbon loss: results from a national survey. Journal of Hydrology, 381, 112-120.

Armstrong, A., Holden, J., Kay, P., Foulger, M., Gledhill, S., McDonald, A.T. & Walker, A. 2009. Drain-blocking techniques on blanket peat: A framework for best practice. Journal of Environmental Management, 90, 3512-3519.

Bain, C.G., Bonn, A., Stoneman, R., Chapman, S., Coupar, A., Evans, M., Gearey, B., Howat, M., Joosten, H., Keenleyside, C., Labadz, J., Lindsay, R., Littlewood,

78

N., Lunt, P., Miller, C.J., Moxry, A., Orr, H., Reed, M., Smith, P., Swales, V., Thompson, D.B.A., Thompson, P.S., Van de Noort, R., Wilson, J.D. & Worrall, F. 2011. IUCN UK Commission of Inquiry on Peatlands. IUCN UK Peatland Programme, Edinburgh

Baines, D. 1996. The implications of grazing and predator management on the habitats and breeding success of black grouse Tetrao tetrix. Journal of Applied Ecology, 33, 54-62.

Baines, D., Redpath, S., Richardson, M. & Thirgood, S. 2008. The direct and indirect effects of predation by hen harriers Circus cyaneus on trends in breeding birds on a Scottish grouse moor. Ibis, 150, 27-36.

Baines, D., Warren, P. & Richardson, M. 2007. Variations in the vital rates of black grouse (Tetrao tetrix) in the United Kingdom. Wildlife Biology, 13 (Suppl 1), 109-116.

Baines, D., Wilson, I.A. & Beeley, G. 1996. Timing of breeding in black grouse Tetrao tetrix and capercaillie Tetrao urogallus and distribution of insect food for the chicks. Ibis, 138, 181-187.

Baird, A., Holden, J. & Chapman, P. 2009. Literature Review of Evidence on Emissions of Methane in Peatlands – Defra Project SP0574. Unpublished report to Defra

Ball, D.F., Radford, G.L. & Williams, W.M. 1983. A Land Characteristic Data Bank for Great Britain. Institute for Ecology Occasional Paper 13. Institute for Ecology, Bangor.

Barclay-Estrup, P. 1972. Arthropod populations in a heathland as related to cyclical changes in the vegetation. Entomologist’s Monthly Magazine, 109, 79-84.

Bardgett, R.D., Marsden, J.H. & Howard, D.C. 1995. The extent and condition of heather on moorland in the uplands of England and Wales. Biological Conservation, 71, 155-161.

Barkman, J.J. 1992. Plant communities and synecology of bogs and heath pools in the Netherlands. In: Verhoeven, J.T.A. (ed) Fens and Bogs in the Netherlands, pp. 173-235. Kluwer Academic Publishers.

Barr, C.J, Bunce, R.G.H., Clarke, R.T., Fuller, R.M., Furze, M.T., Gillespie, M.K., Groom, G.B., Hallam, C.J., Hornung, M., Howard, D.C. & Ness, M.J. 1993. Countryside Survey 1990, Main Report. Department of the Environment, London.

BBC. 2007. Forest park grass blaze put out. http://news.bbc.co.uk/1/hi/scotland/south_of_scotland/6562687.stm

Bellamy, P.H., Loveland, P.J., Bradley, R.I., Lark, R.M. & Kirk, G.J.D. 2005. Carbon losses from all soils across England and Wales 1978-2003. Nature, 437, 245-248.

Bellamy, P.E., Stephen, L., Maclean, I.S. & Grant, M.C. 2011. Response of blanket bog vegetation to drain blocking. Applied Vegetation Science, 15, 129-135.

Berdowski, J.J.M. 1987. Transition from heathland to grassland initiated by the heather beetle. Vegetatio, 72, 167-173.

Berendse, F., Schmitz, M. & De Visser, W. 1994. Experimental manipulation of succession in heathland ecosystems. Oecologia, 100, 38-44.

Bibby, C.J. (1987). Foods of breeding merlins Falco columbarius in Wales. Bird Study, 34, 64-70.

Bibby, C.J. 1986. Merlins in Wales: Site occupancy and breeding in relation to vegetation. Journal of Applied Ecology, 23, 1-12.

79

Billet, M.F., Palmer, S.M., Hope, D., Deacon, C., Storeton-West, R., Hargreaves, K.J., Flechard, C. & Fowler, D. 2004. Linking land-atmosphere-stream carbon fluxes in a lowland peatland system. Global Biogeochemical Cycles, 18 (1), GB1024.

Blodau, C. 2002. Carbon cycling in peatlands – a review of processes and controls. Environmental Reviews, 10, 111-134.

Bolton, M., Tyler, G., Smith, K. & Bamford, R. 2007. The impact of predator control on lapwing Vanellus vanellus breeding success on wet grassland nature reserves. Journal of Applied Ecology, 44: 534–544.

Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds). 2009. Drivers of Environmental Change in Uplands. Routledge, Oxon.

Børsheim, K.Y., Christensen, B.E. & Painter, T.J. 2001. Preservation of fish by embedment in Sphagnum moss, peat or holocellulose: experimental proof of the oxopolysaccharidic nature of the preservative substance and of its antimicrobial and tanning action. Innovative Food Science and Emerging Technologies, 2, 63-74.

Bottin, G. 2002. West Cross Lochs Hill Drain Restoration Monitoring, RSPB Forsinard Reserve, Sutherland, Scotland. Unpublished RSPB report. RSPB, Sandy.

Bowers, J.K. 1985. British agricultural policy since the Second World War. Agricultural History Review, 33, 66-76.

Bowker, G., Bowker, C. & Baines, D. 2007. Survival rates and causes of mortality in black grouse Tetrao tetrix at Lake Vyrnwy, North Wales, UK. Wildlife Biology, 13, 231-237.

Bragg, O.M. & Tallis, J.H. 2001. The sensitivity of peat-covered upland landscapes. Catena, 42, 345-360.

Bro, E., Arroyo, B. & Migot, P. 2006. Conflict between grey partridge Perdix perdix hunting and hen harrier Circus cyaneus protection in France: a review. Wildlife Biology, 12, 233–247.

Brown, A.F. & Bainbridge, I.P. 1995. Grouse moors and upland breeding birds. In: Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp.51-66. HMSO, Edinburgh.

Brown, A.F., Crick, H.Q.P. & Stillman, R.A. 1995. The distribution, numbers and breeding ecology of twite Acanthis flavirostris in the South Pennines of England. Bird Study, 42, 107-121.

Brown, L., Holden, J., Ramchunder, S. & Langton, R. 2009. Understanding Moorland Aquatic Insect Ecology to Inform Biodiversity Conservation and Sustainable Land Management. Unpublished report to North Pennines AONB Peatscapes Partnership.

Brown, R.W. 1986. The Effect of Controlled Burning and Season on Moorland Soil Fauna. 2nd Report on the Moorland Management Programme. North York Moors National Park, Helmsley.

Buchanan, G.M., Grant, M.C., Sanderson, R.A. & Pearce-Higgins, J.W. 2006. The contribution of invertebrate taxa to moorland bird diets and the potential implications of land-use management. Ibis, 148, 615-628.

Buchanan, G.M., Pearce-Higgins, J.W. & Grant, M.C. 2007. Testing the generality of moorland bird habitat-associations to inform conservation management. Bird modelling: Bird-habitat associations. In: Environmentally Sustainable & Economically Viable Grazing Systems for Restoration and Maintenance of Heather

80

Moorland: England and Wales – Final Report, Defra Project BD1228. Unpublished report to Defra.

Bunce, R.G.H. & Barr, C.J. 1988. The extent of land under different management regimes in the uplands and the potential for change. In: Usher, M.B. & Thompson, D.B.A. (eds) Ecological Change in the Uplands. Blackwell, Oxford.

Burfield, I. & van Bommel, F. 2004. Birds in Europe: Population Estimates, Trends and Conservation Status. BirdLife International, Cambridge.

Burfield, I.J. 2002. The Breeding Ecology and Conservation of the Ring Ouzel Turdus torquatus in Britain. Unpublished PhD thesis, University of Cambridge.

Burke, W. 1975. Effect of drainage on the hydrology of blanket bogs. Irish Journal of Agricultural Research, 14, 145-162.

Caizergues, A. & Ellison, L. 1997. Survival of black grouse Tetrao tetrix in the French Alps. Wildlife Biology, 3, 177-186.

Cannell, M.G.R., Milne R, Hargreaves K.J., Brown, T.A.W., Cruikshank, M.M., Bradley, R.I., Spencer, T., Hope, D., Billet, M.F., Adger, W.N. & Subak, S. 1999. National inventories of terrestrial carbon sources and sinks: the UK experience. Climatic Change, 42, 505-530.

Carroll, M.J., Dennis, P., Pearce-Higgins, J.W. & Thomas, C.D. 2011. Maintaining northern peatland ecosystems in a changing climate: effects of soil moisture, drainage and drain blocking on craneflies. Global Change Biology, http://onlinelibrary.wiley.com/doi/10.1111/j.1365-2486.2011.02416.x/abstract

Chambers, F.M., Mauquoy, D. & Todd, P.A. 1999. Recent rise to dominance of Molinia caerulea in environmentally sensitive areas: new perspectives from palaeolecological data. Journal of Applied Ecology, 36, 719-733.

Chapman, P.J., McDonald, A.T., Tyson, R., Palmer, S.M., Mitchell, G. & Irvine, B. 2010. Changes in water colour between 1986 and 2006 in the headwaters of the River Nidd, Yorkshire, UK. Biogeochemistry, 101, 281-294.

Chapman, P.J., Palmer, S.M., Irvine, B.J., Mitchell, G. & McDonald, A.T. 2011. A response to ‘Changes in water colour between 1986 and 2006 in the headwaters of the River Nidd, Yorkshire, UK: a critique of methodological approaches and measurement of burning management’ by Yallop et al. Biogeochemistry (2 November 2011), pp. 1-5, doi:10.1007/s10533-011-9665-0

Chapman, S.B. & Rose, R.J. 1991. Changes in the vegetation at Coom Rigg Moss National Nature Reserve within the period 1958-86. Journal of Applied Ecology, 28, 140-153.

Cherrill, A.J. & Rushton, S.P. 1993. The Auchenorhyncha of an unimproved moorland in northern England. Ecological Entomology, 18, 95-103.

Clay, G.D. & Worrall, F. 2011. Charcoal production in a UK wildfire – how important is it? Journal of Environmental Management, 92, 676-682.

Clay, G.D., Worrall, F. & Fraser, E.D.G. 2009. Effects of managed burning upon dissolved organic carbon (DOC) in soil water and runoff water following a managed burn of a UK blanket bog. Journal of Hydrology, 367, 41-51.

Clutterbuck, B. & Yallop, A.R. 2010. Land management as a factor controlling dissolved organic carbon release from upland peat soils 2: Changes in DOC productivity over four decades. Science of the Total Environment, 408, 6179-6191.

Clymo, R.S. 1983. Peat. In: Gore, A.J.P. (ed) Mires, Swamp, Fen and Moor. General Studies. Ecosystems of the World 4a, pp. 159-224. Elsevier Scientific, Amsterdam.

81

Clymo, R.S. 1992. Models of peat growth. Suo, 43, 127-136. Conway, V.M. & Millar, A. 1960. The hydrology of small peat-covered catchments in

the northern Pennines. Journal of the Institute of Water Engineers, 14, 415-424. Côté I.M. & Sutherland, W.J. 1997. The effectiveness of removing predators to protect

bird populations. Conservation Biology,11, 395–405. Coulson, J.C. & Butterfield, J.E.L. 1985. The invertebrate communities of peat and

upland grasslands in the north of England and some conservation implications. Biological Conservation, 34, 197-225.

Coulson, J.C. 1988. The structure and importance of invertebrate communities on peatlands and moorlands, and effects of environmental and management changes. In: Usher, M.B. & Thompson, D.B.A. (eds) Ecological Change in the Uplands. Blackwell, Oxford.

Coulson, J.C., Butterfield, J.E.L. & Henderson, E. 1990. The effect of open drainage ditches on the plant and invertebrate communities of moorland and on the decomposition of peat. Journal of Applied Ecology, 27, 549-561.

Crowle, A. & McCormack, F. 2009. Condition of upland terrestrial habitats. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 209-227. Routledge, Oxon.

Currall, J.E.P. 1981. Some Effects of Management by Fire on Wet Heath Vegetation in Western Scotland. Unpublished PhD thesis, University of Aberdeen.

Daniels, R.E. 1991. Management of Sphagnum on lowland heaths. In: Auld, M.H.D., Pickess, B.P. & Burgess, N.D. (eds) Proceedings of Heathlands Conference II. History and Management of Southern Lowland Heaths. RSPB, Sandy.

Daplyn, J. & Ewald, J. 2006. Birds, Burning and Grouse Moor Management. Unpublished report to Moors for the Future.

Davies, G.M. & Legg, C.J. 2008. The effect of traditional management burning on lichen diversity. Applied Vegetation Science, 11, 529-538.

Davies, G.M. 2001. The Impact of Muirburning on Lichen Diversity. Unpublished MSc thesis, University of Edinburgh.

Davies, G.M., Gray, A., Hamilton, A. & Legg, C. 2008. The future of fire management in the British uplands. International Journal of Biodiversity Science and Management, 4, 1-21.

Davies, G.M., Smith, A.A., MacDonald, A.J., Bakker, J.D. & Legg, C.J. 2010. Fire intensity, fire severity and ecosystem response in heathlands: factors affecting the regeneration of Calluna vulgaris. Journal of Applied Ecology, 47, 356-365.

Dawson, J.J.C., Billet, M.F., Neal, C. & Hill, S. 2002. A comparison of particulate, dissolved and gaseous carbon in two contrasting upland streams in the UK. Journal of Hydrology, 257, 226-246.

Defra. 2007. The Heather and Grass Burning Code. 2007 Version. www.naturalengland.gov.uk

Dennis, P., Skartveit, J., McCracken, D.I., Pakeman, R.J., Beaton, K., Kunaver, A. & Evans, D.M. 2008. The effects of livestock grazing on foliar arthropods associated with bird diet in upland grasslands of Scotland. Journal of Applied Ecology, 45, 279-287.

Dennis, P., Young, M.R. & Bentley, C. 2001. The effects of varied grazing management on epigeal spiders, harvestmen and pseudoscorpions of Nardus stricta grassland in upland Scotland. Agriculture, Ecosystems and Environment, 86, 39–57.

82

Dennis, P., Young, M.R., Howard, C.L. & Gordon, I.J. 1997. The response of epigeal beetles (Col. Carabidae, Staphylinidae) to varied grazing regimes on upland Nardus stricta grasslands. Journal of Applied Ecology, 34, 433–443.

Dennis, R.L.H. & Eales, H.T. 1997. Patch occupancy in Coenonympha tullia (Müller, 1764) (Lepidoptera: Satyrinae): habitat quality matters as much as patch size and isolation. Journal of Insect Conservation, 1, 167-176.

Dennis, R.L.H. & Eales, H.T. 1999. Probability of site occupancy in the large heath butterfly Coenonympha tullia determined from geographical and ecological data. Biological Conservation, 87, 295-301.

Dorrepaal, E., Toet, S., van Logtestijn, R.S.P., Swart, E., van de Weg, M.J., Callaghan, T.V. & Aerts, R. 2009. Carbon respiration from subsurface peat accelerated by climate warming in the subarctic. Nature, 460, 616-619.

Downie, I.S., Butterfield, J.E.L. & Coulson, J.C. 1995. Habitat preferences of sub-montane spiders in northern England. Ecography, 18, 51-61.

Drewitt, A.L. & Manley, V.J. 1997. The vegetation of the mountains and moorlands of England. English Nature Research Report no. 218. English Nature, Peterborough.

Dunn, C. & Freeman, C. 2011. Peatlands: our greatest source of carbon credits? Carbon Management, 2, 289-301

Edmondson, J.L., Carroll, J.A., Price, E.A.C. & Caporn, S.J.M. 2010. Bio-indicators of nitrogen pollution in heather moorland. Science of the Total Environment, 408, 6202-6209.

Elliott, R.J., 1953. The Effects of Burning on Heather Moors of the South Pennines. Unpublished PhD thesis, University of Sheffield.

Ellison, L.N. 1978. Black grouse population characteristics on a hunted and three unhunted areas in the French Alps. Proceedings of the Woodland Grouse Symposium.

Etheridge, B., Summers, R.W. & Green, R.E. 1997. The effects of illegal killing and destruction of nests by humans on the population dynamics of the hen harrier Circus cyaneus in Scotland. Journal of Applied Ecology, 34, 1081-1105.

Evans, C.D., Monteith, D.T. & Cooper, D.M. 2005. Long-term increases in surface water dissolved organic carbon: Observations, possible causes and environmental impacts. Environmental Pollution, 137, 55-71.

Evans, C., Worrall, F., Holden, J., Chapman, P., Smith, P. & Artz, R. 2011. A programme to assess evidence gaps in greenhouse gas and carbon fluxes from UK peatlands. JNCC Report No. 443. JNCC, Peterborough.

Evans, D.M., Redpath, S.M., Evans, S.A., Elston, D.A., Gardner, C.J., Dennis, P. & Pakeman, R. 2006. Low intensity, mixed livestock grazing improves the breeding abundance of a common insectivorous passerine. Biology Letters, 2, 636-638.

Eyre, M.D., Luff, M.L. & Woodward, J.C. 2003. Grouse moor management: habitat and conservation implications for invertebrates in southern Scotland. Journal of Insect Conservation, 7, 21-32.

Farage, P., Ball, A., McGenity, T.J., Whitby, C. & Pretty, J. 2009. Burning management and carbon sequestration of upland heather moorland in the UK. Australian Journal of Soil Research, 47, 351-361.

83

Fielding, A., Haworth, P., Whitfield, P., McLeod, D. & Riley, H. 2011. A Conservation Framework for Hen Harriers in the United Kingdom. JNCC Report No: 441. JNCC, Peterborough.

Fielding, C.A. 1992. Aspects of the Ecology of the Lepidoptera Associated with Heather Calluna vulgaris. Unpublished PhD thesis, University of Durham.

Fletcher, K., Aebischer, N.J., Baines, D., Foster, R. & Hoodless, A.N. 2010. Changes in breeding success and abundance of ground-nesting moorland birds in relation to the experimental deployment of legal predator control. Journal of Applied Ecology, 47, 263-272.

Forrest, G.J. & Smith, R.A.H. 1975. The productivity of a range of blanket bog vegetation types in the Northern Pennines. Journal of Ecology, 63, 173-199.

Fraser of Allander Institute. 2010. An Economic Study of Scottish Grouse Moors. Game & Wildlife Conservation Trust, Perth.

Freeman, C., Evans, C.D., Monteith, D.T., Reynolds, B., Fenner, N. 2001. Export of organic carbon from peat soils. Nature, 412, 785.

Fuller, R.J. & Gough, S.J. 1999. Changes in sheep numbers in Britain: implications for bird populations. Biological Conservation, 91, 73-89.

Game Conservancy Trust. 2001. Nervous ticks: the growing concerns over louping ill in the upland. The Game Conservancy Trust Review of 2001. Game Conservancy Trust, Fordingbridge.

Gardner , S.M., Hartley, S.E., Davies, A. & Palmer, S.C.F. 1997. Carabid communities on heather moorlands in northeast Scotland: The consequences of grazing pressure for community diversity. Biological Conservation, 81, 275-286.

Gardner, S.M. & Usher, M.B. 1989. Insect abundance on burned and cut upland Calluna heath. The Entomologist, 108, 147-157.

Gardner, S.M. 1991. Ground beetle (Coleoptera: Carabidae) communities on upland heath and their association with heathland flora. Journal of Biogeography, 18, 281-289.

Garnett M.H., Ineson P., Stevenson A.C. & Howard D.M. 2001. Terrestrial organic carbon storage in a British moorland. Global Change Biology, 7, 375–388.

Gibbons, D., Gates, S., Green, R., Fuller, R.J. & Fuller, R.M. 1995. Buzzard Buteo buteo and ravens Corvus corax in the uplands of Britain: limits to distribution and abundance. Ibis, 137, S75-S84.

Gibbons, D.W., Reid, J.B., & Chapman, R.A. 1993. The New Atlas of Breeding Birds in Britain and Ireland. Poyser, London.

Gibbons, D.W., Amar, A., Anderson, G.Q.A., Bolton, M., Bradbury, R.B., Eaton, M.A., Evans, A.D., Grant, M.C., Gregory, R.D., Hilton, G.M., Hirons, G.J.M., Hughes, J., Johnstone, I., Newbery, P., Peach, W.J., Ratcliffe, N., Smith, K.W., Summers, R.W., Walton, P. & Wilson, J.D. 2007. The predation of wild birds in the UK: A review of its conservation impact and management. RSPB Research Report Number 23. RSPB, Sandy.

Gibson, H.S., Worrall, F., Burt, T.P. & Adamson, J.K. 2009. DOC budgets of drained peat catchments: implications for DOC production in peat soils. Hydrological Processes, 23, 1901-1911.

Giller, P.S. & Malmqvist, B. 1998. The Biology of Streams and Rivers. Oxford University Press, Oxford.

Gimingham, C.H. 1975. An Introduction to Heathland Ecology. Oliver & Boyd, Edinburgh.

84

Gimingham, C.H. 1985. Age-related interactions between Calluna vulgaris and phytophagous insects. Oikos, 44, 12-16.

Gimingham, C.H. 1995. Heaths and moorland: an overview of ecological change. In: Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp.9-19. HMSO, Edinburgh.

Glaves, D.J. & Haycock, N.E. (eds) 2005. Defra Review of the Heather and Grass Burning Regualtions and Code: Science Panel Assessment of the Effects of Burning on Biodiversity, Soils and Hydrology. Unpublished report to Defra.

Gorham, E. 1991. Northern peatlands: role in the carbon cycle and probable responses to climatic warming. Ecological Applications, 1, 182-195.

Graham, I.M., Redpath, S.M. & Thirgood, S.J. 1995. The diet and breeding density of common buzzards Buteo buteo in relation to indices of prey abundance. Bird Study, 42, 165-173.

Grant, M. & Dawson, B. 2005. Black grouse habitat requirements in forested environments: Implications for conservation management. In: Proceedings of The 3rd European Black Grouse Conference.

Grant, M.C., Cowie, N., Donald, C., Dugan, D., Johnstone, I., Lindley, P., Moncreiff, R., Pearce-Higgins, J.W., Thorpe, R. & Tomes, D. 2009. Black grouse response to dedicated conservation management. Folia Zoologica, 58, 195-206.

Grant, M.C., Orsman, C., Easton, J., Lodge, C., Smith, M., Thompson, G., Rodwell, S. & Moore, N. 1999. Breeding success and causes of breeding failure of curlew Numenius phaeopus in Northern Ireland. Journal of Applied Ecology, 36, 59-74.

Grant, S.A. & Hunter, R.F. 1968. Interactions of grazing and burning on heather moors and their implications in heather management. Grass and Forage Science, 23, 285-293.

Grant, S.A., Suckling, S.A., Smith, H.K., Torvell, L., Forbes, T.D.A. & Hodgson, J. 1987. Comparative studies of diet selection by sheep and cattle: blanket bog and heather moor. Journal of Ecology, 75, 947-960.

Gray, A. 2006. The Influence of Management on the Vegetation and Carbon Fluxes of Blanket Bog. Unpublished PhD thesis, University of Edinburgh.

Grayson, R., Holden, J. & Rose, R. 2010. Long-term change in storm hydrographs in response to peatland vegetation change. Journal of Hydrology, 389, 336-343.

Grayson, R., Kay, P., Foulger, M. & Gledhill, S. 2012. A GIS based MCE model for identifying water colour generation potential in UK upland drinking water supply catchments. Journal of Hydrology, 420-421, 37-45.

Green, R.E. & Etheridge, B. 1999. Breeding success of the hen harrier Circus cyaneus in relation to the distribution of grouse moors and the red fox Vulpes vulpes. Journal of Applied Ecology, 36, 472-483.

Greenslade, P.J.M. 1964. Pitfall trapping as a method for studying populations of Carabidae (Coleoptera). Journal of Animal Ecology, 33, 301-310.

Greenup, A.L., Bradford, M.A., McNamara, N.P., Ineson, P. & Lee, J.A. 2000. The role of Eriophorum vaginatum in CH4 flux from an ombrotrophic peatland. Plant and Soil, 227, 265-272.

Grimm, M. 2005. Bestandsentwicklung und gefährdungsursachen des groβen brachvogels Numenius arquata in den Belziger Landschaftwiesen (Brandenburg). Vogelwelt, 126, 333-340.

Haines-Young, R.H., Barr, C.J., Black, H.I.J., Briggs, D.J., Bunce, R.G.H., Clarke, R.T., Cooper, A., Dawson, F.H., Firbank, L.G., Fuller, R.M., Furse, M.T., Gillespie,

85

M.K., Hill, R., Hornung, M., Howard, D.C., McCann, T., Morecroft, M.D., Petit, S., Sier, A.R.J., Smart, S.M., Smith, G.M., Stott, A.P., Stuart R.C. & Watkins J.W. 2000. Accounting for Nature: Assessing Habitats in the UK Countryside. DETR, London.

Hamilton, A. 2000. The Characteristics and Effects of Management Fire on Blanket-bog Vegetation in North-west Scotland. Unpublished PhD thesis, University of Edinburgh.

Hansen, K. 1976. Ecological studies in Danish heath vegetation. Dansk Botanisk Archiv, 31, 7-116.

Hardey, J., Rollie, C.J. & Stirling-Aird, P.K. 2003. Variation in breeding success of inland peregrine falcon (Falco peregrinus) in three regions of Scotland 1991-2000. In: Thompson, D.B.A., Redpath, S.M., Fielding, A.H., Marquiss, M. & Galbraith, C.A. (eds) Birds of Prey in a Changing Environment, pp.99-109. The Stationery Office, Edinburgh.

Harding, N.J., Green, R.E. & Summers, R.W. 1994. The Effects of Future Changes in Land Use on Upland Birds in Britain. Unpublished RSPB report. RSPB, Sandy.

Harris, M.P.K, Allen, K.A., McAllister, H.A., Eyre, G., Le Duc, M.G. & Marrs, R.H. 2011. Factors affecting moorland plant communities and component species in relation to prescribed burning. Journal of Applied Ecology, 48, 1411-1421.

Haworth, P.F. & Thompson, D.B.A. 1990. Factors associated with the breeding distribution of upland birds in the South Pennines, England. Journal of Applied Ecology, 27, 562–577.

Haysom, K.A. & Coulson, J.C. 1998. The Lepidoptera fauna associated with Calluna vulgaris: effects of plant architecture on abundance and diversity. Ecological Entomology, 23, 377-385.

Hejzlar, J., Dubrovsky, M., Buchtele, J. & Ruzicka, M. 2003. The apparent and potential effects of climate change on the inferred concentration of dissolved organic matter in a temperate stream (the Malse River, South Bohemia). Science of the Total Environment, 310, 143–152.

Hester, A.J. & Sydes, C. 1992. Changes in burning of Scottish heather moorland since the 1940s from aerial photographs. Biological Conservation, 60, 25-30.

Hewson, R. 1982. The effect upon field vole (Microtus agrestis) habitat on removing sheep from moorland in west Scotland. Journal of Zoology, 197, 304-307.

Hill, M.O. & Stevens, P.A. 1981. The density of viable seed in soils of forest plantations in upland Britain. Journal of Ecology, 69, 693-709.

Hobbs, R.J. & Gimingham, C.H. 1984a. Studies on fire in Scottish heathland communities. II. Post-fire vegetation development. Journal of Ecology, 72, 585-610.

Hobbs, R.J. & Gimingham, C.H. 1984b. Studies of fire in Scottish heathland communities. I. Fire characteristics. Journal of Ecology, 72, 223-240.

Hobbs, R.J. & Gimingham, C.H. 1987. Vegetation, fire and herbivore interactions in heathland. Advances in Ecological Research, 16, 87-173.

Hobbs, R.J. 1984. Length of burning rotation and community composition in high-level Calluna-Eriophorum bog in N England. Vegetatio, 57, 129-136.

Hobbs, R.J., Mallik, A.U. & Gimingham, C.H. 1984. Studies on fire in Scottish heathland communities. III. Vital attributes of the species. Journal of Ecology, 72, 963-76.

86

Holden, J. & Burt, T, P. 2003. Runoff production in blanket peat covered catchments. Water Resources Research, 39 (7), 1191.

Holden, J. 2005a. Piping and woody plants in peatlands: cause or effect? Water Resources Research, 41, W06009.

Holden, J. 2005b. Controls of soil pipe frequency in upland blanket peat. Journal of Geophysical Research, 110, F01002.

Holden, J. 2006. Sediment and particulate carbon removal by pipe erosion increase over time in blanket peatlands as a consequence of land drainage. Journal of Geophysical Research, 111, F02010.

Holden, J. 2009a. Upland hydrology. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 113-134. Routledge, Oxon.

Holden, J. 2009b. A Grip-blocking Review. Unpublished report to Environment Agency.

Holden, J., Chapman, P.J. & Labadz, J.C. 2004. Artificial drainage of peatlands: Hydrological and hydrochemical process and wetland restoration. Progress in Physical Geography, 28, 95-123.

Holden, J., Chapman, P.J., Palmer, S.M., Kay, P. & Grayson, R. 2011. A review of moorland burning impacts on raw water quality with a focus on water colour. Final report to Yorkshire Water Services, Project B4897, University of Leeds, Leeds.

Holden, J., Chapman, P.J., Evans, M.G., Hubacek, K., Kay, P. & Warburton, J. 2007b. Vulnerability of Organic Soils in England and Wales – Final Technical Report, Defra Project SP0532. Unpublished report to Defra.

Holden, J., Evans, M.G., Burt, T.P. & Horton, M. 2006. Impact of land drainage on peatland hydrology. Journal of Environmental Quality, 35, 1764-1778.

Holden, J., Gascoign, M. & Bosanko, N.R. 2007c. Erosion and natural revegetation associated with surface land drains in upland peatlands. Earth Surface Processes and Landforms, 32, 1547-1557.

Holden, J., Kirby, M.J., Lane, S.N., Milledge, D.G., Brookes, C.J., Holden, V. & McDonald, A.T. 2008. Overland flow velocity and roughness properties in peatlands. Water Resources Research, 44, 11.

Holden, J., Shotbolt, L., Bonn, A., Burt, T.P., Chapman, P.J., Dougill, A.J., Fraser, E.D.G., Hubacek, K., Irvine, B., Kirkby, M.J., Reed, M.S., Prell, C., Stagl, S., Stringer, L.C., Turner, A. & Worrall, F. 2007a. Environmental change in moorland landscapes. Earth-Science Reviews, 82, 75-100.

Holden, J., Wallage, Z.E., Lane, S.N. & McDonald, A.T. 2011. Water table dynamics in undisturbed, drained and restored blanket peat. Journal of Hydrology, 402, 103-114

Holt, A.,R., Davies, Z.G., Tyler, C. & Staddon, S. 2008. Meta-analysis of the effects of predation on animal prey abundance: evidence from UK vertebrates. PLoS ONE, 3, e2400.

Hoodless, A.N., Ewald, J.A. & Baines, D. 2007. Habitat use and diet of common snipe Gallinago gallinago breeding on moorland in northern England. Bird Study, 54, 182-191.

Hope, D., Picozzi, N., Catt, D.C. & Moss, R. 1996. Effects of reducing sheep grazing in the Scottish Highlands. Journal of Range Management, 49, 301-310.

87

Hope, D., Billett, M.F. and Cresser, M.S. 1997. Exports of organic carbon in two river systems in NE Scotland. Journal of Hydrology, 193: 61-82.

Hsu, C.H., Jeng, W.L., Chang, R.M., Chien, L.C. & Han, B.C. 2001. Estimation of potential lifetime cancer risks for trihalomethanes from consuming chlorinated drinking water in Taiwan. Environmental Research, 85, 77–82.

Hudson, D. 2008. Grouse Shooting. Quiller Publishing, Shrewsbury. Hudson, P.J. 1984. Some effects of sheep management on heather moorlands in the

north of England. In: Jenkins, D. (ed) The Impact of Agriculture on Wildlife and Semi-natural Habitats, pp. 143-9. Institute of Terrestrial Ecology Symposium, no. 13.

Hudson, P.J. 1986. The Red Grouse, The Biology and Management of a Wild Gamebird. Game Conservancy, Fordingbridge.

Hudson, P.J. 1992. Grouse in Space and Time. Game Conservancy Ltd., Fordingbridge. Hudson, P.J. 1995. Ecological trends and grouse management in upland Britain. In:

Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp. 282-293. HMSO, Edinburgh.

Hughes, S., Reynolds, B., Hudson, J.A., & Freeman, C. 1997. Effects of summer drought on peat soil solution chemistry in an acid gully mire. Hydrology and Earth System Sciences, 1, 661-669.

Huntings Surveys 1986. Monitoring Landscape Change. Vol. 1. Main Report. Huntings Surveys, Borehamwood.

Imeson, A.C. 1971. Heather burning and soil erosion on the North Yorkshire Moors. Journal of Applied Ecology, 8, 537-542.

Ingram, H.A.P. 1978. Soil layers in mires: function and terminology. Journal of Soil Science, 29, 224-227.

IPCC. 2001. Climate Change 2001: The Scientific Basis. Cambridge University Press, Cambridge.

Jenkins, D., Watson, A. & Miller, G.R. 1963. Population studies on red grouse, Lagopus lagopus scoticus (Lath.), in north-east Scotland. Journal of Animal Ecology, 32, 153-161.

JNCC. 2004. Common Standards Monitoring Guidance for Upland Habitats. Joint Nature Conservancy Council, Peterborough.

Johnson, C.G., Southwood, T.R.E. & Entwhistle, H.M. 1957. A new method of extracting arthropods and molluscs from grassland and herbage with a suction apparatus. Bulletin of Entomological Research, 48, 211-218.

Jonczyk, J., Rimmer, D. & Quinn, P. 2009. Geltsdale grip block monitoring report. Unpublished report to North Pennines AONB Peatscapes Partnership.

Jönsson, P.E. 1991. Reproduction and survival in a declining population of the southern Dunlin Calidris alpine schinzii. Wader Study Group Bulletin, 61 (Suppl.), 56-68.

Kauhala, K. & Helle, P. 2002. The impact of predator abundance on grouse populations in Finland – a study based on wildlife monitoring counts. Ornis Fennica, 79, 14-25.

Kauhala, K., Helle, P. & Helle, E. 2000. Predator control and the density and reproductive success of grouse populations in Finland. Ecography, 23, 161-168.

Kayll, A.J. & Gimingham, C.H. 1965. Vegetative regeneration of Calluna vulgaris after fire. Journal of Ecology, 53, 729-734.

Kenworthy, J.B. 1963. Temperatures in heather burning. Nature, 200, 1226.

88

Kim, Y. & Tanaka, N. 2003. Effect of forest fire on the fluxes of CO2, CH4 and N2O in boreal forest soils, interior Alaska. Journal of Geophysical Research-Atmospheres, 108(D1), 8154.

Kinako, P.D.S. & Gimingham, C.H. 1980. Heather burning and soil erosion on upland heaths in Scotland. Journal of Environmental Management, 10, 277-284.

Kuhry, P. 1994. The role of fire in the development of Sphagnum-dominated peatlands in western boreal Canada. Journal of Applied Ecology, 82, 899-910.

Kurki, S., Helle, P., Lindén, H. & Nikula, A. 1997. Breeding success of black grouse and capercaillie in relation to mammalian predator densities on two spatial scales. Oikos, 79, 301-310.

Laiho, R. 2006. Decomposition in peatlands: Reconciling seemingly contrasting results on the impacts of lowered water levels. Soil Biology & Biochemistry, 38, 2011-2024.

Lane, SN. Brookes, CJ. Hardy, RJ. Holden, J. James, TD. Kirby, MJ. McDonald, AT. Tayefi, V. Yu, D. 2003. Land management, flooding and environmental risk: new approaches to a very old question. Proc. Of National CIWEM Conference.

Laurenson, M.K., McKendrick, I.J., Reid, H.W., Challenor, R. & Mathewson, G.K. 2007. Prevalence, spatial distribution and the effect of control measures on louping-ill virus in the Forest of Bowland, Lancashire. Epidemiology and Infection, 135, 963-973.

Laurenson, M.K., Norman, R.A., Gilbert, L., Reid, H.W. & Hudson, P.J. 2003. Identifying disease reservoirs in complex systems: mountain hares as reservoirs of ticks and louping-ill virus, pathogens of red grouse. Journal of Animal Ecology, 72, 177-185.

Legg, C., Davies, G.M. & Gray, A. 2010. Comment on: ‘Burning management and carbon sequestration of upland heather moorland in the UK’. Australian Journal of Soil Research, 48, 100-103.

Legg, C.J., Maltby, E. & Proctor, M.C.F. 1992. The ecology of severe moorland fire on the North York Moors: seed distribution and seedling establishment of Calluna vulgaris. Journal of Ecology, 80, 737-752.

Lindley P., Johnstone I. & Thorpe R. 2003. The status of black grouse in Wales in 2002 and evidence for population recovery. Welsh Birds, 3, 318–329.

Lindsay, R. 2010. Peatbogs and Carbon: A Critical Synthesis. Unpublished report to RSPB Scotland.

Lindsay, R.A., Charman, D.J., Everingham, F., O’Reilly, R.M., Palmer, M.A., Rowell, T.A. & Stroud, D.A. 1988. The Flow Country: The Peatlands of Caithness and Sutherland. Nature Conservancy Council, Peterborough.

Lindström, E.R., Andrén, H., Angelstam, P., Cederlund, G., Hörnfeldt, B., Jäderberg, L., Lemnell, P-A., Martinsson, B., Sköld, K. & Swenson, J.E. 1994. Disease reveals the predator: sarcoptic mange, red fox predation, and prey populations. Ecology, 75, 1042-1049.

Littlewood, N.A., Pakeman, R.J. & Woodin, S.J. 2006. The response of plant and insect assemblages to the loss of Calluna vulgaris from upland vegetation. Biological Conservation, 128, 335-345.

Lockie, J. D. 1955. The breeding habits and food of short-eared owls after a vole plague. Bird Study, 2, 53-69.

Lovat, Lord 1911. The Grouse in Health and in Disease. Smith, Elder & Co., London.

89

MacDonald, A. & Haysom, K. 1997. Heather moorland management for Lepidoptera. Scottish Natural Heritage Information & Advisory Note no. 78.

MacDonald, A. 2000. Effects of Grouse Moor Management on Moorland Plant Diversity. Unpublished report, Grouse moor management & biodiversity workshop, Napier University, Edinburgh.

MacDonald, A.J., Kirkpatrick, A.H., Hester, A.J. & Sydes, C. 1995. Regeneration by natural layering of heather (Calluna vulgaris): frequency and characteristics in upland Britain. Journal of Ecology, 32, 85-99.

Mackay, A.W. & Tallis, J.H. 1996. Summit-type blanket mire erosion in the Forest of Bowland, Lancashire, UK: Predisposing factors and implications for conservation. Biological Conservation, 76, 31-44.

Mackey, E.C., Shewry, M.C. & Tudor, G.J. 1998. Land Cover Change Scotland - From the 1940s to the 1980s. The Stationery Office, Edinburgh.

Mallik, A.U. & Gimingham, C.H. 1983. Regeneration of heathland plants following burning. Vegetatio, 53, 45-58.

Mallik, A.U. & Gimingham, C.H. 1985. Ecological Effects of Heather Burning – II Effects on seed germination and vegetative regeneration. Journal of Ecology, 73, 633-644.

Mallik, A.U. 1986. Near-ground micro-climate of burned and unburned Calluna heathland. Journal of Environmental Management, 23, 157-171.

Mallik, A.U., Hobbs, R.J. & Legg, C.J. 1984. Seed dynamics in Calluna-Arctostaphylos heath in north-eastern Scotland. Journal of Ecology, 72, 855-871.

Maltby, E., Legg, C.J. & Proctor, M.C.F. 1990. The ecology of severe moorland fire on the North York Moors: effects of the 1976 fires, and subsequent surface vegetation development. Journal of Ecology, 78, 490-518.

Maltby, T. & Edwards, T. 1984. Microbiological response to heather burning. In: Moorland Management. North York Moors National Park, Helmsley.

Marcström, V., Kenward, R.E. & Engren, E. 1988. The impact of predation on boreal tetraonids during vole cycles: an experimental study. Journal of Animal Ecology, 57, 859-872.

Marinier, M., Glatzel, S. & Moore, T.R. 2004. The role of cotton-grass (Eriophorum vaginatum) in the exchange of CO2 and CH4 at two restored peatlands, eastern Canada. Ecoscience, 11, 141-149.

McFerran, D.M., McAdam, J.H. & Montgomery, W.I. 1995. The impact of burning and grazing on heathland plants and invertebrates in County Antrim. Proceedings of the Royal Irish Academy, 95B, 1-17.

McMorrow, J., Lindley, S., Aylen, J., Cavan, G., Alberton, K. & Boys, D. 2009. Moorland wildfire risk, visitors and climate change: patterns, prevention and policy. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 404-431. Routledge, Oxon.

Miller, G.R. & Miles, J. 1970. Regeneration of heather (Calluna vulgaris (L.) Hull) at different ages and seasons in north-east Scotland. Journal of Applied Ecology, 7, 51-60.

Miller, G.R., Jenkins, D. & Watson, A. 1966. Heather performance and red grouse populations. I. Visual estimates of heather performance. Journal of Applied Ecology, 3, 313-326.

Miller, G.R., Jenkins, D. & Watson, A. 1970. Responses of red grouse populations to experimental improvement of their food. In: Watson, A. (ed) Animal

90

Populations in Relation to Their Food Resources, pp. 323-325. Blackwell Scientific Publications, Oxford.

Milne, R. & Brown, T.A. 1997. Carbon in the vegetation and soils of Great Britain. Journal of Environmental Management, 49, 413-433.

Mitchell, G. & McDonald, A.T. 1992. Discolouration of water by peat following induced drought and rainfall simulation. Water Research, 26, 321-326

Mitchell, G. & McDonald, A.T. 1995. Catchment characterization as a tool for upland water quality management. Journal of Environmental Management, 44, 83-95.

Monteith, D.T., Stoddard, J.L., Evans, C.D., de Wit, H.A., Forsius, M., Hogasen, T., Wilander, A., Skjelkvale, B.L., Jeffries, D.S., Vuorenmaa, J., Keller, B., Kopacek, J. & Vesely, J. 2007. Dissolved organic carbon trends resulting from changes in atmospheric chemistry. Nature, 450, 537-541.

Moss, D., Joys, A.C., Clark, J.A., Kirby, A., Smith, A., Baines, D. & Crick, H.Q.P. 2005. Timing of breeding of moorland birds. BTO Research Report no. 362. BTO, Thetford.

Muirburn Working Party. 1977. A Guide to Good Muirburn Practice. HMSO, Edinburgh.

Mulder, J.L. & Swaan, A.H. 1988. De Vos in Het Noord-hollands Duinreservaat Deel 5: de Wulpenpopulatie. Rijk-instituut voor Natuurbeheer, Arnhem.

Natural England 2009. Mapping values: the vital nature of our uplands – an atlas linking environment and people. Natural England, Peterborough.

Neimeyer, T., Niemeyer, M., Mohamed, A., Fottner, S. & Härdtle, W. 2005. Impact of prescribed burning on the nutrient balance of heathlands with particular reference to nitrogen and phosphorus. Applied Vegetation Science, 8, 183-192.

Newborn, D. 2004. Tick treatment for sheep keeps louping ill at bay. The Game Conservancy Trust Review of 2003. Game Conservancy Trust, Fordingbridge.

Newborn, D. & Baines, D. 2011. Enhanced control of sheep ticks in upland sheep flocks: repercussions for red grouse co-hosts. Medical and Veterinary Entomology, 26, 63-69.

Newborn, D., Fletcher, K.L., Beeston, R. & Baines, D. 2009. Occurrence of sheep ticks on moorland wader chicks. Bird Study, 56, 401-404.

Newborn, D. & Foster, R. 2002. Control of parasite burdens in wild red grouse Lagopus lagopus scoticus through the indirect application of anthelminthics. Journal of Applied Ecology, 39, 909-914.

Newey, S.J., Gilbert, L. & Harrison, A. 2010. Insufficient evidence for culling mountain hares for tick and louping-ill virus (LIV) control. Knowledge Scotland, Science Policy Connections Online. Research Briefing, Biodiversity.

Newton, I. 1979. Population Ecology of Raptors. Poyser, Berkamsted. Newton, I. 1998. Population Limitation in Birds. Academic Press, London. Newton, I., Meek, E.R. & Little, B. 1984. Breeding season foods of merlins Falco

columbarius in Northumbria. Bird Study, 31, 49-56. Nielsen, Ó.K. 1999. Gyrfalcon predation on ptarmigan: numerical and functional

responses. Journal of Animal Ecology, 68, 1034-1050. Ninnes, R.B. 1995. Bracken, heath and burning on the Quantock Hills, England. In:

Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp. 194-199. HMSO, Edinburgh.

91

O’Brien, M. 2001. Factors Affecting Breeding Wader Populations on Upland Enclosed Farmland in Northern Britain. Unpublished PhD thesis, University of Edinburgh.

O’Brien, M. 2002. The relationship between field occupancy rates by breeding lapwing and habitat management on upland farmland in northern Britain. Aspects of Applied Biology, 67, 85-92.

O’Connell, P.E., Beven, K.J., Carney, J.N., Clements, R.O., Ewen, J., Fowler, G.L., Harris, J., Hollis, Morris, J., O’Donnell, G.M., Packman, J.C., Parkin, A., Quinn, P.F., Rose, S.C., Shepherd, M. & Tellier, S. 2004. Review of Impacts of Rural Land Use and Management on Flood Generation. R&D Technical Report FD2114/TR. Environment Agency, Bristol.

Orr, H.G., Wilby, R.L., McKenzie Hedger, M. & Brown, I. 2008. Climate change in the uplands: a UK perspective on safeguarding regulatory ecosystem services. Climate Research, 37, 77-98.

Park, K.J., Calladine, J.R., Graham, K.E., Stephenson, C.M. & Wernham, C.V. 2005. The Impacts of Predatory Birds on Waders, Songbirds, Gamebirds and Fisheries Interests. Unpublished report to Scotland’s Moorland Forum (available at - http://www.moorlandforum.org.uk/documents/PBRFinal.pdf).

Park, K.J., Graham, K.E., Calladine, J. & Wernham, C.W. 2008. Impacts of birds of prey on gamebirds in the UK: a review. Ibis, 150 (suppl. 1), 9-26.

Park, K.J., Robertson, P.A., Campbell, S.T., Foster, R., Russell, Z.M., Newborn, D. & Hudson, P.J. 2001. The role of invertebrates in the diet, growth and survival of red grouse (Lagopus lagopus scoticus) chicks. Journal of Zoology, 254, 137-145.

Parr, R. & Watson, A. 1988. Habitat preferences of black grouse on moorland-dominated ground in north-east Scotland. Ardea, 76, 175-180.

Parr, R. 1992. The decline to extinction of a population of golden plover in north-east Scotland. Ornis Scandinavica, 23, 152–158.

Parr, R. 1993. Nest predation and numbers of golden plovers Pluvialis apricaria and other moorland waders. Bird Study, 40, 223–231.

Pawson, R.R., Lord, D.R., Evans, M.G. & Allott, T.E.H. 2008. Fluvial organic carbon flux from an eroding peatland catchment, southern Pennines, UK. Hydrological Earth System Sciences, 12, 625-634.

Pearce, F. 2006. Grouse-shooting popularity boosts global warming. New Scientist. Pearce-Higgins, J.W. & Grant, M.C. 2006. Relationships between bird abundance and

structure of moorland vegetation. Bird Study, 53, 112-125. Pearce-Higgins, J.W., Beale, C.M., Wilson, J. & Bonn, A. 2006. Analysis of Moorland

Breeding Bird Distribution and Change in the Peak District. Unpublished report to Moors for the Future (report no. 11).

Pearce-Higgins, J.W., Dennis, P., Whittingham, M.J. & Yalden, D.W. 2010. Impacts of climate on prey abundance account for fluctuations in a population of a northern wader at the southern edge of its range. Global Change Biology, 16, 12-23.

Pearce-Higgins, J.W., Grant, M.C., Beale, C.M., Buchanan, G.M. & Sim, I.M.W. 2009. International importance and drivers of change of upland bird populations. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 209-227. Routledge, Oxon.

92

Pearce-Higgins, J.W., Grant, M.C., Robinson, M.C. & Haysom, S.L. 2007. The role of forest maturation in causing the decline of black grouse Tetrao tetrix. Ibis, 149, 143-155.

Pearsall, W.H. 1950. Mountains and Moorlands. Collins, London. Phillips, J. & Watson, A. 1995. Key requirements for management of heather

moorland: now and for the future. In: Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp. 344-361. HMSO, Edinburgh.

Picozzi, N. & Hepburn, L.V. 1986. A study of black grouse in north-east Scotland. Proceedings of the 3rd International Grouse Symposium, 462-480.

Picozzi, N. 1968. Grouse bags in relation to the management and geology of heather moor. Journal of Applied Ecology, 5, 483-488.

Picozzi, N. 1984. Breeding biology of polygynous hen harriers Circus c. cyaneus in Orkney. Ornis Scandinavica, 15, 1–10.

Rackham, O. 1986. The History of the Countryside. Dent & Sons, London. Raine, A. F. 2006. The Breeding Ecology of Twite Carduelis flavirostris and the Effects of

Upland Agricultural Intensification. Unpublished PhD thesis, University of East Anglia.

Ramchunder, S.J., Brown, L.E. & Holden, J. 2009. Environmental effects of drainage, drain-blocking and prescribed vegetation burning in UK upland peatlands. Progress in Physical Geography, 33, 49-79.

Ramchunder, S.J., Brown, L.E. & Holden, J. 2012. Catchment-scale peatland restoration benefits stream ecosystem biodiversity. Journal of Applied Ecology, 49, 182-191.

Ratcliffe, D.A. 1964. Mires and bogs. In: Burnett, J. H. (ed) TheVegetation of Scotland, pp. 426-478. Oliver & Boyd, Edinburgh.

Ratcliffe, D.A. 1976. Observations on the breeding of the golden plover in Great Britain. Bird Study, 23, 63-116.

Ratcliffe, D.A. 1993. The Peregrine Falcon (2nd edition). Poyser, London. Ratcliffe, D.A. 1997. The Raven. Poyser, London. Ratcliffe, D.A. 2007. Galloway and the Borders. New Naturalist, Collins, London. Rawes, M. & Hobbs, R. 1979. Management of semi-natural blanket bog in the

northern Pennines. Journal of Ecology, 67, 789-807. Rebecca, G.W. & Cosnette, B.L. 2003. Long-term monitoring of breeding merlin (Falco

columbarius) in Aberdeenshire, north-east Scotland 1980-2000. In: Thompson, D.B.A., Redpath, S.M., Fielding, A.H., Marquiss, M. & Galbraith, C.A. (eds) Birds of Prey in a Changing Environment, pp. 183-197. The Stationery Office, Edinburgh.

Redpath, S., Amar, A., Madders, M., Leckie, F. & Thirgood, S. 2002a. Hen harrier foraging success in relation to land use in Scotland. Animal Conservation, 5, 113-118.

Redpath, S., Madders, M., Donnelly, E., Anderson, B., Thirgood, S., Martin, A. & McLeod, D. 1998. Nest site selection by hen harriers in Scotland. Bird Study, 45, 51-61.

Redpath, S.M. & Thirgood, S.J. 1997. Birds of Prey and Red Grouse. Stationery Office, London.

93

Redpath, S.M. & Thirgood, S.J. 1999. Numerical and functional responses in generalist predators: hen harriers and peregrines on Scottish grouse moors. Journal of Animal Ecology, 68, 879-892.

Redpath, S.M. 1992. Behavioural interactions between hen harriers and their moorland prey. Ornis Scandinavica, 23, 73-80.

Redpath, S.M., Amar, A., Smith, A., Thompson, D.B.A. & Thirgood, S. 2010. People and nature in conflict: can we reconcile hen harrier conservation and game management? In: Baxter, J. & Galbraith, C.A. (eds) Species Management: Challenges and Solutions for the 21st Century, pp. 335-350. The Stationery Office, Edinburgh.

Redpath, S.M., Thirgood, S.J. & Clarke, R. 2002b. Field vole Microtus agrestis abundance and hen harrier Circus cyaneus diet and breeding in Scotland. Ibis, 144 (1) on-line), E33-E38.

Reynolds, J.C. & Tapper, S.C. 1996. Control of mammalian predators in game management and conservation. Mammal Review, 26, 127-156.

Rhodes, A.N. 1996. Moorland Fire History From Microscopic Charcoal in Soils and Lake Sediments. Unpublished PhD thesis, University of Newcastle-upon-Tyne.

Risely, K., Baillie, S.R., Eaton, M.A., Joys, A.C., Musgrove, A.J., Noble, D.G., Renwick, A.R. & Wright, L.J. 2010. The Breeding Bird Survey 2009. BTO, Thetford.

Roberts, J.L. & Bowman, N. 1986. Diet and ecology of short-eared owls Asio flammeus breeding on heather moor. Bird Study, 33, 12-17.

Robertson, P.A., Park, K.J. & Barton, A.F. 2001. Loss of heather moorland in the Scottish uplands: the role of red grouse management. Wildlife Biology, 7, 37-42.

Robinson, M. 1986. Changes in catchment runoff following drainage and afforestation. Journal of Hydrology, 86, 71-84.

Robson, G. 1998. The Breeding Ecology of Curlew Numenius arquata on North Pennine Moorland. Unpublished PhD thesis, University of Sunderland.

Roos, S., Smart, J., & Gibbons, D.W. 2012. Review of the predation of wild birds in the UK: An updated review (2007-2011) of its conservation impact and management. RSPB, Sandy.

Rosenburgh, A. & Marrs, R. 2010. The Heather Beetle: a review. Unpublished report to the Heather Trust (available from www.heathertrust.org.uk).

Ross, S., Adamson, H. & Moon, A. 2003. Evaluating management techniques for controlling Molinia caerulea and enhancing Calluna vulgaris on upland wet heathland in northern England, UK. Agriculture, Ecosystems & Environment, 97, 39-49.

Rowell, T.A. 1990. Management of peatlands for conservation. British Wildlife, 1, 144-156.

Rydin, H. & Jeglum, J.K. 2006. The Biology of Peatlands. Oxford University Press, Oxford.

Rydin, H., Gunnarsson,U. & Sundberg, S. 2006. The role of Sphagnum in peatland development and persistence. In: Wieder, R.K. & Vitt, D.H. (eds) Boreal Peatland Ecosystems, pp. 47-65. Springer-Verlag, Berlin.

Sanderson, R.A., Rushton, S.P., Cherrill, A.J. & Byrne, J.P. 1995. Soil, vegetation and space: an analysis of their effects on the invertebrate communities of a moorland in northeast England. Journal of Applied Ecology, 32, 506-518.

Savory, C.J. 1978. Food consumption of red grouse in relation to the age and productivity of heather. Journal of Animal Ecology, 47, 269-282.

94

Schimmel, J. & Granström, A. 1996. Fire severity and vegetation response in the boreal Swedish forest. Ecology, 77, 1436-1450.

Schlesinger, W.H. 1991. Biogeochemistry: An Analysis of Global Change. Academic Press, San Diego.

Scottish Natural Heritage. 2005. Constructed Tracks in the Scottish Uplands. Scottish Natural Heritage (available at – http://www.snh.gov.uk/docs/A308736.pdf).

Scottish Natural Heritage. 2010. Condition of Designated Sites. Scottish Natural Heritage (available at - http://www.snh.gov.uk/docs/B686627.pdf)

Seivwright, L.J., Redpath, S.M., Mougeot, F., Leckie, F. & Hudson, P.J. 2005. Interactions between intrinsic and extrinsic mechanisms in a cyclic species: testosterone increases parasite infection in red grouse. Proceedings of the Royal Society, 272, 2299-2304.

Sim, I.M.W., Burfield, I.J., Grant, M.C., Pearce-Higgins, J.W. & Brooke, M de L. 2007a. The role of habitat composition in determining breeding site occupancy in a declining ring ouzel Turdus torquatus population. Ibis, 149, 374-385.

Sim, I.M.W., Dillon, I.A., Eaton, M.A., Etheridge, B., Lindley, P., Riley, H., Saunders, R., Sharpe, C. & Tickner, M. 2007b. Status of the hen harrier Circus cyaneus in the UK and Isle of Man in 2004, and a comparison with the 1998/89 and 1998 surveys. Bird Study, 54, 256-267.

Sim, I.M.W., Gregory, R.D., Hancock, M.H. & Brown, A.F. 2005. Recent changes in the abundance of British upland breeding birds. Bird Study, 52, 261-275.

Skjelkvåle, B.L., Stoddard, J., Jeffries, D., Tørseth, K., Høgåsen, T., Bowman, J., Mannio, J., Monteith, D., Mosello, R., Rogora, M., Rzychon, D., Vesely, J., Wieting, J., Wilander, A. & Worsztynowicz, A. 2005. Regional scale evidence for improvements in surface water chemistry 1990-2001. Environmental Pollution, 137, 165-176.

Smart, J., Amar, A., Sim, I.M.W., Etheridge, B., Cameron, D., Christie, G. & Wilson, J.D. 2010. Illegal killing slows population recovery of a re-introduced raptor of high conservation concern – the red kite Milvus milvus. Biological Conservation, 143, 1278-1286.

Smedshaug, C.A., Selaes, V., Lund, S.E. & Sonerud, G.A. 1999. The effect of a natural reduction of red fox Vulpes vulpes on small game hunting bags in Norway. Wildlife Biology, 5, 157-166.

Smith, A.A., Redpath, S.M., Campbell, S.T. & Thirgood, S.J. 2001. Meadow pipits, red grouse and the habitat characteristics of managed grouse moors. Journal of Applied Ecology, 38, 390-400.

Smith, P., Smith, J.U., Flynn, H., Killham, K., Rangel_Castro, I., Foereid, B., Aitkenhead, M., Chapman, S., Towers, W., Bell, J., Lumsdon, D., Milne, R., Thomson, A., Simmons, I., Skiba, U., Reynolds, B., Evans, C., Frogbrook, Z., Bradley, I., Whitmore, A. & Falloon, P. 2007. ECOSSE: Estimating Carbon in Organic Soils – Sequestration and Emissions. Final Report. Scottish Executive Environment and Rural affairs Department.

Sotherton, N., May, R., Ewald, J., Fletcher, K. & Newborn, D. 2009a. Managing uplands for game and sporting interests: an industry perspective. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 209-227. Routledge, Oxon.

95

Sotherton, N., Tapper, S. & Smith, A. 2009b. Hen harriers and red grouse: economic aspects of red grouse shooting and the implications for moorland conservation. Journal of Applied Ecology, 46, 955-960.

Southwood, T.R.E. & Henderson, P.A. 2000. Ecological Methods (3rd edition). Blackwell, Oxford.

Staines, B.W., Balharry, R. & Welch, D. 1995. Moorland management and impacts of red deer. In: Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp. 294-308. HMSO, Edinburgh.

Starling-Westerberg, A. 2001. The habitat use and diet of black grouse Tetrao tetrix in the Pennine hills of northern England. Bird Study, 48, 76-89.

Stevenson, A.C. & Birks, H.J.B. 1995. Heaths and moorland: long-term ecological changes, and interactions with climate and people. In: Thompson, D.B.A., Hester, A.J. & Usher, M.B. (eds) Heaths and Moorlands: Cultural Landscapes, pp. 224-239. HMSO, Edinburgh.

Stevenson, A.C., Rhodes, A.N., Kirkpatrick, A.H. & MacDonald, A.J. 1996. The determination of fire histories and an assessment of their effects on moorland soils and vegetation. Scottish Natural Heritage Research, Survey and Monitoring Report no. 16. Scottish Natural Heritage, Battleby.

Stewart, A.J.A & Lance, A.N. 1991. Effects of moor-draining on the hydrology and vegetation of northern Pennine blanket bog. Journal of Applied Ecology, 28, 1105-1117.

Stewart, A.J.A. & Lance, A.N. 1983. Moor-draining: a review of impacts on land use. Journal of Environmental Management, 17, 81-99.

Stewart, G.B., Coles, C.F. & Pullin, A.S. 2004a. Does Burning of UK Sub-montane, Dry Dwarf-shrub Heath Maintain Vegetation Diversity? Systematic Review no. 2. Unpublished report, Centre for Evidence-Based Conservation.

Stewart, G.B., Coles, C.F. & Pullin, A.S. 2004b. Does Burning Degrade Blanket Bog? Systematic Review no. 1. Unpublished report, Centre for Evidence-Based Conservation.

Stewart, G.B., Coles, C.F. & Pullin, A.S. 2005. Applying evidence-based practice in conservation management: lessons from the first systematic review and dissemination projects. Biological Conservation, 126, 270-278.

Stoddart, D.M. & Hewson, R. 1984. Mountain hare, Lepus timidus, bags and moor management. Journal of Zoology, 204, 563-565.

Storaas, T., Wegge, P. & Larsen, B.B. 1982. Nest predation among capercaillie and black grouse as affected by habitat location and cover. Proceedings of the 2nd International Grouse Symposium, 131

Strack, M., Waddington, J.M. & Tuitttila, E-S. 2004. Effect of water table drawdown on northern peatland methane dynamics: Implications for climate change. Global Biogeochmical Cycles, 18, GB4003.

Summers, R.W., Green, R.E., Proctor, R., Dugan, D., Lambie, D., Moncreiff, R., Moss, R. & Baines, D. 2004. An experimental study of the effects of predation on the breeding productivity of capercaillie and black grouse. Journal of Applied Ecology, 41, 513-525.

Swengel, A.B. 2001. A literature review of insect responses to fire, compared to other conservation managements of open habitats. Biodiversity and Conservation, 10, 1141-1169.

96

Tallis, J.H. 1987. Fire and flood at Holme Moss: erosion processes in an upland blanket mire. Journal of Ecology, 75, 1099-1129.

Tallis, J.H. 1997. The southern Pennine experience: an overview of blanket mire degradation. In: Tallis, J.H., Meade, R. & Hulme, P.D. (eds.) Blanket Mire Degradation - Causes, Consequences and Challenges, pp. 7-15. Macaulay Land Use Research Institute, Aberdeen on behalf of the Mires Research Group.

Tapper, S. 1992. Game Heritage. An Ecological Review from Shooting and Gamekeeping Records. Game Conservancy Ltd., Fordingbridge.

Terry, A.C., Ashmore, M.R., Power, S.A., Allchin, E.A. & Heil, G.W. 2004. Modelling the impacts of atmospheric nitrogen deposition on Calluna-dominated ecosystems in the UK. Journal of Applied Ecology, 41, 897-909.

Tharme, A.P., Green, R.E., Baines, D., Bainbridge, I.P. & O’Brien, M. 2001. The effect of management for red grouse shooting on the population density of breeding birds on heather-dominated moorland. Journal of Applied Ecology, 38, 439-457.

Thirgood, S.J., Redpath, S.M., Haydon, D.T., Rothery, P., Newton, I. & Hudson, P.J. 2000a. Habitat loss and raptor predation: disentangling long- and short-term causes of red grouse declines. Proceedings of the Royal Society of London B, 267, 651-656.

Thirgood, S., Redpath, S., Newton, I. & Hudson, P. 2000b. Raptors and Red grouse: Conservation Conflicts and Management Solutions. Conservation Biology, 14, 95-104.

Thirgood, S.T., Redpath, S.M., Rothery, P. & Aebischer, N.J. 2000c. Raptor predation and population limitation in red grouse. Journal of Animal Ecology, 69, 504-516.

Thomas, P.A., El-Barghathi, M. & Polwart, A. 2007. Biological flora of the British Isles: Juniperus communis L. Journal of Ecology, 95, 1404-1440.

Thompson, D.B.A., Gillings, S., Galbraith, C.A., Redpath, S.M. & Drewitt, J. 1997. The contribution of game management to biodiversity: a review of the importance of grouse moors for upland birds. In: Fleming, L.V., Newton, A.C., Vickery, J.A. & Usher, M.B. (eds) Biodiversity in Scotland: Status, Trends and Initiatives, pp. 198–212. The Stationery Office, Edinburgh.

Thompson, D.B.A., MacDonald, A.J., Marsden, H. & Galbraith, C.A. 1995. Upland heather moorland in Great Britain: A review of international importance, vegetation change and some objectives for nature conservation. Biological Conservation, 71, 163-178.

Thompson, P.S., Amar, A., Hoccum, D.G., Knott, J. & Wilson, J.D. 2009. Resolving the conflict between driven-grouse shooting and conservation of hen harriers. Journal of Applied Ecology, 46, 950-954.

Tornberg, R., Korpima¨ki, E., Jungell, S. and Reif, V. 2005. Delayed numerical response of goshawks to population fluctuations of forest grouse. Oikos, 111, 408-415.

Trinder, N. 1975. The Feeding Value of Moorland Plants. ADAS Science Service 1975, 176-187.

Tucker, G. 2003. Review of the impacts of heather and grassland burning in the uplands on soils, hydrology and biodiversity. English Nature Research Report no. 550. English Nature, Peterborough.

Tuittila, E-S., Komulainen, V-M., Vasander, H., Nykänen, H., Martikainen, P.J. & Laine, J. 2000. Methane dynamics of a restored cut-away peatland. Global Change Biology, 6, 569-581.

97

Usher, M.B. & Smart, L.M. 1988. Recolonization of burnt and cut heathland in the North York Moors by arachnids. Naturalist, 113, 103-111.

Usher, M.B. 1992. Management and diversity of arthropods in Calluna heathland. Biodiversity and Conservation, 1, 63-79.

Valkama, J. & Currie, D. 1999. Low productivity of curlews Numenius arquata on farmland in southern Finland: causes and consequences. Ornis Fennica, 76, 65-70.

Valkama, J., Korpimäki, E., Arroyo, B., Beja, P., Bretagnolle, V., Bro, E., Kenward, R., Mañosa, S., Redpath, S.M., Thirgood, S. & Viñuela, J. 2005. Birds of prey as limiting factors of gamebird populations in Europe: a review. Biological Review, 80, 171–203.

Verhoeven, J.T.A. & Liefveld, W.M. 1997. The ecological significance of organochemical compounds in Sphagnum. Acta Botanica Neerlandica, 46, 117-130.

Village, A. 1987. The Kestrel. Poyser, Calton. Waddell, L. 2008. Plugging our carbon sinks. Shooting Times & Country Magazine,

18 Sept, 27-28. Waddington, J.M. & Day, S.M. 2007. Methane emissions from a peatland following

restoration. Journal of Geophysical Research, 112, G03018. Wallage, Z.E., Holden, J. & McDonald, A.T. 2006. Drain blocking: An effective

treatment for reducing dissolved organic loss and water discolouration in a drained peatland. Science of the Total Environment, 367, 811-821.

Ward, S.E., Bardgett, R.D., McNamara, N.P., Adamson, J.K. & Ostle, N.J. 2007. Long-term consequences of grazing and burning on northern peatland carbon dynamics. Ecosystems, 10, 1069-1083.

Warren, P. & Baines, D. 2002. Dispersal, survival and causes of mortality in black grouse Tetrao tetrix in northern England. Wildlife Biology, 8, 129-135.

Watson, A. 2011. Red Grouse and Ptarmigan research in Scotland. Scottish Birds, 31, 230-235.

Watson, A. & Moss, R. 2008. Grouse. New Naturalist, Collins, London. Watson, D. 1977. The Hen Harrier. Poyser, Berkhamsted. Watson, J. 1997. The Golden Eagle. Poyser, London. Watson, J., Leitch, A.F. & Rae, S.R. 1993. The diet of golden eagles Aquila chrysaetos in

Scotland. Ibis, 135, 387-393. Watson, J., Rae, S.R. & Stillman, R. 1992. Nesting density and breeding success of

golden eagles in relation to food supply in Scotland. Journal of Animal Ecology, 61, 543-550.

Watson, R.T., Noble, I.R., Bolin, B., Ravindranath, N.H., Verardo, D.J., & Dokken, D.J. (eds) 2000. Land Use, Land-Use Change, and Forestry: IPCC Special Report. Cambridge University Press, Cambridge.

Welch, D., Scott, D., Moss, R., & Bayfield, N.G. 1994. Ecology of Blaeberry and its Management in British Moorlands. Institute of Terrestrial Ecology, Grange-over-Sands.

Wheeler, P. 2008. Effects of sheep grazing on abundance and predators of field vole (Microtus agrestis) in upland Britain. Agriculture, Ecosystems & Environment, 123, 49-55.

98

Whitfield, D.P., Fielding, A.H. & Whitehead, S. 2008. Long-term increase in the fecundity of hen harriers in Wales is explained by reduced human interference and warmer weather. Animal Conservation, 11, 144-152.

Whitfield, D.P., Fielding, A.H., McLeod, D.R.A. & Haworth, P.F. 2004. The effects of persecution on age of breeding and territory occupation in golden eagles in Scotland. Biological Conservation, 118, 249-259.

Whitfield, D.P., Fielding, A.H., Mcleod, D.R.A., Morton, K., Stirling-Aird, P. & Eaton, M. 2007. Factors constraining the distribution of golden eagles Aquila chrysaetos in Scotland. Bird Study, 54, 199-211.

Whitfield, D.P., McLeod, D.R.A., Watson, J., Fielding, A.H. & Haworth, P.F. 2003. The association of grouse moor in Scotland with the illegal use of poisons to control predators. Biological Conservation, 114, 157-163.

Whittaker, R.H. 1972. Evolution and measurement of species diversity. Taxon, 21, 213–251.

Whittingham, M.J. 1996. Habitat Requirements of Breeding Golden Plover Pluvialis apricaria. Unpublished PhD thesis, University of Sunderland.

Whittingham, M.J., Percival, S.M. & Brown, A.F. 2001. Habitat selection by golden plover Pluvialis apricaria chicks. Basic and Applied Ecology, 2, 177-191.

Whittingham, M.J., Percival, S.M. & Brown, A.F. 2002. Nest-site selection by golden plover: why do shorebirds avoid nesting on slopes? Journal of Avian Biology, 33, 184-190.

Wilkinson, N.I. & Wilson, J.D. 2010. Breeding ecology of the Twite Carduelis flavirostris in a crofting landscape. Bird Study 57, 142-155.

Williams, I.T. & Parr, S.J. 1995. Breeding merlins Falco columbarius in Wales in 1993. Welsh Birds, 1, 14-20.

Williams, J.M. (ed) 2006. Common Standards Monitoring for Designated Sites: First Six Year Report 2006. Joint Nature Conservation Committee, Peterborough. http://www.jncc.gov.uk/page-3520.

Wilson, L., Wilson, J., Holden, J., Johnstone, I., Armstrong, A. & Morris, M. 2011a. Ditch blocking, water chemistry and organic carbon flux: evidence that blanket bog restoration reduces soil erosion and carbon loss. Science of the Total Environment, 409, 2010-2018.

Wilson, L., Wilson, J., Holden, J., Johnstone, I., Armstrong, A. & Morris, M. 2011b. The impact of drain blocking on an upland blanket bog during storm and drought events, and the importance of sampling-scale. Journal of Hydrology, 404, 198-208.

Wilson, L., Wilson, J., Holden, J., Johnstone, I., Armstrong, A. & Morris, M. 2010b. Recovery of water tables in Welsh blanket bog after drain blocking: Discharge rates, time scales and the influence of local conditions. Journal of Hydrology, 391, 377-386.

Wilson, L., Wilson, J.M. & Johnstone, I. 2010a. The effect of blanket bog drainage on habitat condition and on sheep grazing, evidence from a Welsh upland bog. Biological Conservation, 144, 193-201.

Worrall, F. & Bell, M. 2009. Best Practice for Managing Soil Organic Matter in Agriculture: Managing SOM in ‘Upland’ Agriculture - Final Report, Defra Project SP08016. Unpublished report to Defra.

Worrall, F. & Burt, T.P. 2007. The flux of dissolved organic carbon from UK rivers. Global Biogeochemical Cycles, 21, GB1013.

99

Worrall, F. & Evans, M.G. 2009. The carbon budget of upland peat soils. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 93-112. Routledge, Oxon.

Worrall, F., Armstrong, A. & Adamson, J.K. 2007a. The effects of burning and sheep-grazing on water table depth and soil water quality in a upland peat. Journal of Hydrology, 339, 1-14.

Worrall, F., Armstrong, A. & Holden, J. 2007b. Short-term impact of peat drain-blocking on water colour, dissolved organic carbon concentration, and water table depth. Journal of Hydrology, 337, 315-325.

Worrall, F., Bell, M.J. & Bhogal, A. 2010. Assessing the probability of carbon and greenhouse gas benefit from the management of peat soils. Science of the Total Environment, 408, 2657-2666.

Worrall, F., Reed, M., Warburton, J. & Burt, T. 2003. Carbon budget for a British upland peat catchment. Science of the Total Environment, 312, 133-146.

Worrall, F., Chapman, P., Holden, J., Evans, C., Artz, R., Smith, P. & Grayson, R. 2011a. A review of current evidence on carbon fluxes and greenhouse gas emissions from UK peatlands. JNCC Report No. 442. JNCC, Pterborough.

Worrall, F., Rowson, J.G., Evans, M.G., Pawson, R, Daniels, S. & Bonn, A. 2011b. Carbon fluxes from eroding peatlands – the carbon benefit of revegetation following wildfire. Earth Surface Processes and Landforms, 36, 1487-1498

Yallop, A.R. & Clutterbuck, B. 2009. Land management as a factor controlling dissolved organic carbon release from upland peat soils 1: Spatial variation in DOC productivity. Science of the Total Environment, 407, 3803-3813.

Yallop, A.R., Clutterbuck, B. & Thacker, J.I. 2009. Burning issues: The history and ecology of managed fires in the uplands. In: Bonn, A., Allott, T., Hubacek, K. & Stewart, J. (eds) Drivers of Environmental Changes in Uplands, pp. 209-227. Routledge, Oxon.

Yallop, A.R., Clutterbuck, B. & Thacker, J. 2010. Increases in humic dissolved organic carbon export from upland peat catchments: the role of temperature, declining sulphur deposition and changes in land management. Climate Research, 45, 43-56.

Yallop, A.R., Clutterbuck, B. & Thacker, J.I. 2011. Changes in water colour between 1986 and 2006 in the headwaters of the River Nidd, Yorkshire, UK: a critique of methodological approaches and measurements of burning management. Biogeochemistry, DOI: 10.1007/s10533-011-9628-5

Yallop, A.R., Thacker, J.I., Thomas, G., Stephens, M., Clutterbuck, B., Brewer, T. & Sannier, C.A.D. 2006. The extent and intensity of management burning in the English uplands. Journal of Applied Ecology, 43, 1138-1148.