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RECONCILING GAMEBIRD HUNTING AND BIODIVERSITY (REGHAB)
CONTRACT No: EVK2-CT-2000-200004
PROJECT COORDINATOR: Javier Viñuela Instituto de Investigación en Recursos Cinegéticos (IREC), Ronda de Toledo s/n, 13005-Ciudad Real, Spain
PROJECT WEBPAGE: www.uclm.es/irec/Reghab/inicio.html
(Report on Workpackage 2 – Deliverable no 7).
Impact of hunting management practices on biodiversity
Beatriz Arroyo1 & Pedro Beja2
January 2002
1 – Centre for Ecology and Hydrology, Banchory, Hill of Brathens, AB31 4BW, Scotland 2 – ERENA- Ordenamento e Gestão de Recursos Naturais, Lda., Av. Visconce Valmor 11, 3°; 1000-289 Lisboa, Portugal
INDEX
1. INTRODUCTION 1 Background 1 Aims 1 Scope and definitions 2
2. MANAGEMENT PRACTICES 3 Management for gamebird shooting 3 Dominant practices per country 3 Finland 3 France 5 UK 11 Spain 13 Portugal 15 Global patterns 17
3. EFFECT OF MAIN MANAGEMENT PRACTICES ON GAME AND NON-GAME SPECIES 3.1 PREDATOR CONTROL 18 Predation and gamebirds 18 Effect of predator control on small game species 19 Effect of predator control on species other than target game 24 Conclusions 26 3.2 RELEASES AND RESTOCKING 27 Impact of releases and restocking on the wild populations of target species 29 Species introductions 30 Genetic pollution 31 Sanitary problems 32 Conclusions 34 3.3 HABITAT MANAGEMENT 35 3.3.1 Introduction 35 3.3.2 Management in areas where the main land use is hunting: moorland 36 General background 36 Grouse shooting and preservation of moorland 36 Comparisons between managed and unmanaged moorland 37 Conclusions 38 3.3.3 Management in areas where the main land use is different from hunting: farmland 39 Field boundaries 41 Hedgerows 41 Herbaceous field margins 42 Sterile strips 44 Crops 44 Crop edges 44 Beetle banks 46 Undersown leys 47 Game crops 47 Fallows and set asides 48 Woodland 51 Supplementary resources 53 Provision of grain 53 Provision of open water 53 Habitat management to reduce predation pressure 54 Conclusions 54
4. CONCLUDING REMARKS 56
5. ACKNOWLEDGEMENTS 58
6. REFERENCES 59
1. INTRODUCTION
Background
Conflicts between gamebird hunters and nature conservationists arise, among other things, because of differences in perceptions about the relationship between hunting activities and wildlife conservation. On the one hand, there is a public belief that hunting and shooting may be detrimental for wildlife conservation, either directly if performed in an unsustainable way, or indirectly through the use of certain management practices (Escribano 2000). On the other hand, hunting managers spend money and effort to preserve game numbers and their habitats, and hunters thus claim that game conservation may play a role in wildlife conservation, both because whole ecosystems have been sometimes maintained through the income from hunting, and because some management practices have been judged positive to biodiversity in general (Potts 1992, Tapper 1999). However, it is not always clear what kinds and extent of game management actions may affect biodiversity in a positive or a negative manner. For instance, conservationists usually regard the control of predators to improve game numbers as a negative aspect of game management (Herranz 2001). However, high predation levels can also represent a threat to many non-game birds (Martin 1993, Newton 1993), particularly ground-nesting species in grassland ecosystems (Warner 1994). Predator control exerted due to game interests could therefore potentially benefit other species, (Suarez et al. 1993, Tapper 1999). Another example is the release of farm-reared birds to sustain or restore declining game bird numbers, which may have an important impact on ecosystems and on hunting pressure itself (Vargas & Duarte 2001).
Overall, therefore, game management has potentially a significant impact on biodiversity, and major conservation implications, by affecting habitat structure, food availability and intra- and inter-specific interactions on large tracts of land. Nevertheless, the effects of game management on animal communities are poorly documented, with most published information covering only a limited range of environmental and management situations. Because both conservationists and hunters are dependent on and demanding of natural resources, there are opportunities for them to become partners in wildlife management (Peyton 2000), and the integration of hunting management as a conservation tool is increasingly demanded (Taris 1997, Vargas & Duarte 2001). However, in order to evaluate how this partnership may be developed, a better understanding on the relationship between gamebird management and biodiversity is needed, to determine the best management practices to maximise benefits for both hunting and conservation purposes.
Aims
The objectives of this workpackage are to synthesise the available data on the effects of gamebird management and hunting on biodiversity. In particular, the objectives of WP2 are to review available information on the effect of habitat management and species management (reinforcement of gamebird populations, predator control) on gamebird populations and on species other than gamebirds.
Additionally, in order to evaluate whether quantitative estimates of the impact of hunting practices on biodiversity are currently possible at any scale, we review the available information on the extent of the implementation of the main management practices in each country/area. In particular, we review and discuss levels of implementation of different measures of habitat management, predator control, and restocking.
1
Scope and definitions
There are many problems associated with the evaluation of the effect of hunting management practices on biodiversity. First of all, biodiversity is a difficult concept to define. Despite the increasing attention received by the search for globally accepted measures for biodiversity (see Heywood 1995 for a review), it remains difficult to derive a precise scientific measure of biodiversity. In addition to the difficulty of defining biodiversity, there are problems in defining reference values and consistent measures for the different indicators, and defining spatial and temporal scales. Since our work is based on published literature, there is also the problem of biases associated with this: experimental work is rare, many studies contemplate only some animal or plant groups, and significant results are more likely to be published than non-significant ones.
Because of all these issues, we are aware that this review is only partial, and rather an initial approach to a complex issue. We have concentrated, for the purpose of this review, on work related to vertebrates (i.e., the effect of hunting practices on vertebrates other than gamebirds). Work related to other groups (insects or plants) has also been included when relevant, although research on published material about these groups has been less exhaustive.
2 2. MANAGEMENT PRACTICES
Management for gamebird shooting.
Gamebird management aims to maximise hunting yields, and ideally in a sustainable way. The first way to achieve this main aim is to manipulate limiting factors for the game species. These include nest and feeding cover, so habitat management is usually advised in management (Nadal 1989, Hudson & Newborn 1995, Vargas & Cardo 1996, Nadal 1997). Additionally, supplementary food or water have been used in certain cases (e.g. Segovia 1994, Hudson & Newborn 1995, Borralho et al. 1997) when they are limiting factors. Predation on nests or adults can be high in many game species, such as grouse (e.g. Angelstam et al. 1984, Redmond et al. 1982, Erikstad et al. 1982, Thirgood et al. 1998), partridges (Potts 1980, Bro et al. 2000, 2001) or wildfowl (Beauchamp et al. 1996, Sovada et al. 2001). Thus, predator control has been recommended in many cases to increase gamebird breeding success and thus the size of the harvestable populations (Charlez 1993, Borralho et al. 1997, Hudson & Newborn 1995, Romero 2001). Diseases and parasites may in cases limit populations, so sanitary control has also been used for game (e.g. Hudson & Newborn 1995, Bravo & Peris 1998).
Secondly, a way to maximize hunting yields is to artificially increase densities through restocking wild stocks with farm-reared birds. This has been increasingly done in recent decades, and particularly for farmland birds (e.g. Tapper 1999). Releases may have two different objectives, with a different time scale: reinforcement of breeding numbers (long term objectives), or increase of bird numbers during the hunting season (short-term objectives).
Finally, if gamebird numbers are to be maximized in a sustainable way, control of bags may be essential in order to avoid overshooting, so population management through control of bags is also a management option, and is extensively applied for some species in at least some countries.
We review the most common practices in the REGHAB participant countries (Finland, UK, France, Spain and Portugal), and summarize the available information about the extent with which these practices are implemented. Sources used include primarily information provided by hunting organizations, such as the ONCFS (in France), the GCT (in the UK), or FDC (Spain), completed when necessary with published information.
Dominant practices per country
Finland
Habitat management.
In Finland, the habitats are generally not managed for hunting purposes, i.e. there is no habitat management that is specifically aimed at preserving gamebird populations. However, modern forest and farming management instructions in Finland attempt to maintain or even increase biodiversity, and this includes taking actions that may have positive effects on gamebirds.
3 In forestry there is no habitat management that is specifically aimed at preserving gamebird populations (black grouse, capercaillie, hazel grouse or willow grouse), but rather the interests of gamebirds have been tried to put under the umbrella of maintenance of overall biodiversity in forest ecosystems. Forest planners and managers should be aware of habitat requirements of grouse species, but instructions and practice are often two completely different things. One problem in large-scale regional forest planning is that, especially in southern and central Finland, forests are mainly private. This means that forest planning is possible in small scale, but may be very difficult at a large (i.e. landscape level) scale. Finnish Forest Certification System (FFCS, also known as PEFC, Pan European Forest Certificate) aims at sustainable forestry, and this sustainability is assessed through 37 different criteria. PEFC has been criticized, and is generally not accepted by conservation organisations that have their own certification system (FSC, which is used e.g. in Sweden). Thus, it is virtually impossible to give any numbers (such as areas covered in hectares) of gamebird habitat management in forests.
On farmland, there are three main habitat management tools that have been used for gamebird (pheasant and grey partridge) protection: (i) fields specifically established for mammals (especially for moose) and gamebirds. These are usually small (less than 1 hectare) and are placed on fields that are not optimal for commercial harvesting. (ii) Creation of refuge sites (with higher vegetation especially in the middle of large farmland areas) and (iii) wider or broader field margins. There are practically two types of field margins: (1) “buffer zones”, which are 15 m wide uncultivated but managed belts between the field and large rivers or lakes. In 1999, the total area of these in Finland was 2230 hectares, which is about 0.11% of the total farmland area. (2) “buffer strips”, which are 3 m wide uncultivated but managed strips between the field and larger ditches. In 1999, the total area of these in Finland was 6240 hectares (0.3% of the total farmland area). These zones contain permanent vegetation (not bushes) and may thus be good for gamebirds and also for overall biodiversity (research of their significance is ongoing). The vegetation in these zones should be cut and collected away once per summer, and wrong timing of this activity may of course be harmful for animals living or breeding there. Organic production, covering at the moment about 140000 hectares in Finland, may be more beneficial to gamebirds than directed management in intensive farming areas.
Predator control
Predator control is a common practice in Finland for all gamebirds considered. Predators legally killed include foxes, mustelids, raccons and corvids, and culling can be intensive. For example, in 2000, numbers of predators killed included 50800 red foxes, 84700 raccoon dogs, 4300 stoats, 84300 American mink, 12800 pine martens, 7400 badger, 192300 hooded crows and 119200 magpies (FFGRI 2001).
Restocking
Pheasants and grey partridges are released in agricultural areas, but no exact records exist on the number of released birds, or on whether releases aim reinforcement of breeding numbers or short-term shooting interests only. A questionnaire sent to hunters allowed, however, to estimate that 15000 pheasants and 500 grey partridges are released annually (Mr Pentti Vikberg, from Finnish Hunters’ Central Organisation, comm. pers.).
4 Management of hunting bags
Until the late 1980s Finnish forest grouse were censused using the route method (Rajala 1974). In the late 1980’s this was replaced by the wildlife triangle census method (Lindèn et al. 1996). The total length of these equilateral triangle routes is 12 km. Censuses are carried out in Mid-August for gamebirds and again in mid-winter for snow-tracks of mammalian predators, hares and squirrels. Censuses are conducted by hunters, but the work is co- ordinated by Finnish Game and Fisheries Research Institute. Each year approx. 1500 triangles are censused. On the basis of August censuses, it is decided whether there is a need for hunting restrictions for any of the censused species in any area. Local hunting clubs can deny or restrict hunting of some species in their areas, e.g. such that killing of only one black grouse male or one capercaillie is allowed for each hunter.
Sanitary control
No sanitary control is carried out for gamebirds. Vaccination for rabies for raccon dogs (through vaccine-baited meals) was carried out in the early 1990’s.
Supplementary food and water
Some gamebird species (black grouse, pheasant and partridge) are fed with oats during winter in some areas, but the amount of food used for this purpose is not recorded. This winter feeding method is not supported and practiced by all hunters, because some think that feeding sites tend to attract avian predators, especially young goshawks.
France
Habitat management
Many small-scale habitat management actions have been implemented for upland gamebirds. However, the extent of practices is relatively small and they usually concern experimental settings over a reduced number of sites. For example, almost no habitat management is carried out for rock ptarmigan or hazel grouse within hunted areas. Within protected areas, some management occurs for these two species, sometimes in an experimental way. Measures taken, for example, include control in ptarmigan habitats of grazing by sheep. In hazel grouse habitats, measures to favour food plants include thinning of dense forests, creation of small clearings and manipulation of cattle grazing practices (e.g. Leonard 1993, 2000, Ellison et al. 1994, Parc Naturel du Haut Jura 1997). However, management usually concerns less than 5% of the surface. For the black grouse, several interventions have taken place in hunted and protected areas (maintenance of breeding habitats by control of invasive shrubs, changing grazing regimes; Magnani 1993, Cornut & Dubost 1998 Jouglet et al. 1999, Parc National des Ecrins 1999, Novoa et al. 2001, Anthelme et al. 2001). In the Alps such shrub control has been carried out on ca.1000 ha, which also represents a small percentage of the surface favourable for the species. For rock partridge and Pyrenean grey partridge, specific management actions (controlled burning and grazing, removal of brush) have been implemented only in a small number of experimental sites, representing a low percentage of the species’ distribution range (Novoa et al. 1990, FDCI 1997, Blanchemain 1997, Novoa et al. 1998, Novoa & Landry 1998, Curtet 2000, Michallet & Toigo 2001). For capercaillie, forest management guidelines exist for the whole distribution range, and are either
5 compulsory (in Vosgues) or facultative. In the latter case, they are increasingly implemented in Jura, Cévennes and Pyrénées.
In contrast, because the abundance and breeding success of gamebirds in agricultural habitats is so strongly correlated with farming practices and the structure of the farming habitat (see Bro 1998 for a review, Bro et al. 2000), management of habitat in agricultural areas for improving grey partridge or red-legged partridge numbers has been implemented in many areas, in the form of contracts with farmers. Measures implemented have been described in several hunting books (e.g. Koch 1992, Pasquet 1995, Chantelat & Lorgnier du Mesnil 1995), and include planting of cover crops, management of set-asides, management of field margins, etc. An inquiry carried out in 1997-98 on 485 managed territories (Mayot 1999) showed that “wildlife” set-asides (JEFS, set-asides in which cover is chosen and managed to favour gamebirds and other wildlife) were mentioned in 32% of them, but there was an important geographical variation in its implementation. Average density was 0.22 to 1.4 ha of JEFS / 100 ha of farmland. Other game crops are less common than JEFS, and their importance also varied between departments. Average density was 0.23 ha / 100 ha (Mayot 1999).
Before releasing birds to reinforce populations, habitat management (and predator control) are advised to increase the success probability (Havet & Biadi 1990, Mayot et al. 1993) but these measures are not frequently applied or at low intensity.
There are no statistics to quantify habitat management carried out when making releases for shooting interests only, but they are in any case less extensive than for wild populations. No (or very little) habitat management is carried out specifically for increasing numbers of passerines or pigeons for hunting purposes.
Releases
In France, the purpose of releases differs between programmes: sometimes, releases aim to reinforce breeding populations, and these are mainly carried out in spring/summer. At other times, releases are made just before the hunting season (in autumn), to increase temporarily numbers for shooting.
Releases of pen-reared birds are particularly common for gamebirds typical of agricultural habitats (pheasant, partridges or quail), for both reinforcement of breeding numbers and for autumn increase of shooting numbers.
The number of releases of partridge has increased strongly in the last decades: Yeatman (1976) estimated at ca. 400000 the number of released red-legged partridges between 1970 and 1975. Pinet (1984) estimated 800000 10 years later, and in 1995, the estimate was ca. 2.5 million birds (Tupigny 1996). In that year, total number of partridges released was estimated (from numbers reared) at 4472000 (Tupigny 1996). In terms of grey partridges, the two types of releases (reinforcement of breeding numbers and shooting interests) occur in France but no statistics exist to quantify their relative importance. Furthermore, the situation varies regionally. In some Departments or some hunting estates, releases are forbidden: wild populations are managed to adjust the hunting bag to the local density and demography of the grey partridge. In other departments or territories, releases occur either to reinforce breeding numbers or to increase the hunting bag. A recent inquiry indicates that, in 1998, release of grey partridges occurred in 17% of communes (Reitz in press). Red-legged partridges are often released to preserve the wild grey partridge (and avoid a genetic pollution) while
6 maintaining partridge hunting. Because of this, red-legged partridges are released sometimes in areas outside their normal breeding range. According to Pinet (1984), releases occurred in 37 departments in the 1980s. In 1998, releases occurred in 73 departments (Reitz in press), which represents 50% more than the area of natural presence. An unpublished inquiry by the ONCFS and IMPCF in the Mediterranean area from 1995 showed that, from a sample of 50000 reared red-legged partridges, 43% were released in summer, 12% in spring, and the other 45% in the autumn. From the 333 communes sampled, 47% made shooting releases, 55% reinforcement releases in summer, and only 34% made spring reinforcement releases.
Releases of pheasant have also increased in recent times. Up to recent times, the pheasant was an artificially maintained species. However, there is nowadays a change in attitude, and a wish to maintain natural populations. Some of the releases occurring nowadays respond to that aim, so some releases nowadays are aimed to reinforce breeding numbers, rather than shooting numbers. An inquiry performed in 1995 estimated to up to 10 million pheasants reared in France, against 5-7 million in the 1980s (Tupigny 1996). Mayot and Biadi (1986) reported 6-10 million pheasants reared in the 1980s, and quoted the figure of 12 – 15 million pheasants released for shooting in the 1990s. The comparison between the number of pheasants killed by hunting and the number of released birds attests that pheasant hunting is artificial in France. Releases mainly occur during the hunting season for shooting interests. However, the practice of summer releases is becoming more common, to obtain birds of better quality. Only 21 % of releases are currently for reinforcement of breeding numbers (Mayot 1996).
There were increasing (and uncontrolled) numbers of quails released between the 1960’s and the 1990’s, then they probably decreased. Given that 75% of the annual hunting bag in the South-West were japanese birds, it has been estimated that 400000 - 500000 domestic quails were released in the 1980’s (ONC), although this figure might be over-estimated. A precise inquiry (FDC Haute-Garonne, 1986) indicated that 4000 Japanese quails were released in August over an area of 64000 ha of stubble (ca. 0.06 bird/ha). Applying the same rate to the 5 million ha of cereals of the SAU leads to a maximal estimation of 300000 birds. The majority of these birds were released just before the opening of hunting, in late August. Nevertheless, there was also a number of observations of japanese quail released in June (thousands in the department of Maine-et-Loire in 1985). Nowadays, release of japanese quail is strictly prohibited because of not being a native species.
Releases for other gamebirds are very rare or non-existent. When they have been carried out, they responded to conservation, not to hunting, interests, and to respond to the symbolic value of a given species in certain regions. For example, to reintroduce capercaillie in the Cévennes National Park, 598 pen-reared individuals were released between 1978 and 1994. The programme stopped after establishment of the population. For the Pyrenean grey partridge, releases are prohibited in the most departments. Summer releases are allowed in two departments to reinforce existing populations, but are restricted to a few places and concern only small numbers of birds (440 in 2000). No captive-reared rock partridge, black grouse or rock ptarmigan have been released in France, but a small number of pen-reared hazel grouse (<50) may have been set free.
Predator control
Species legally considered as “harmful/pest” include 12 mammals (weasel, pine marten, beech marten, american mink, polecat, koipu, muskrat, racoon, racoon dog, red fox, rabbit,
7 wildboar) and six birds (rook, magpie, carrion crow, jay, spotted starling, woodpigeon) at the national level, although this list may be modified at the Departmental scale by the authorities. The list is fixed after consultation between hunting representatives, farmers and conservationists (Conseil National / Départemental de la Chasse et de la Faune Sauvage). Categorisation of a species as “pest” is based on the economic impact on human activities (agriculture, livestock, fishing), health risks and protection of wildlife. Hunting interests are not taken into account as such. It is usually the owner who should implement the culling, and the hunters who are also farmers use this right. Currently, right of culling is extended to non- owners and estate managers for hunting interests, if authorised by the owners. After mid 1980s, trappers must be licensed, they follow a training course, and all traps must be identified and licensed.
No predator control is carried out with the objective of increasing numbers of hazel grouse, black grouse, rock ptarmigan or rock partridge. Predator control carried out with the objective of increasing numbers of capercaillie is globally slight, and irregularly distributed (no planning at the regional level).
In contrast, predator control is commonly implemented to increase numbers of gamebirds in agricultural areas (quail, grey partridge, red-legged partridge, pheasant). It occurred in > 95% of managed territories sampled (Mayot 1999). In relation to the intensity of culling there was an important geographical variation (Mayot 1999), but in a scale from 0 (nil) to 3 (strong), predator control was described as weak (lower than 2 on average) in only 26% of Departments.
Management of hunting bags
In France, hunting plans (quotas for hunting bags based on breeding density and breeding success) are not always compulsory, but most species have some limits of hunting bags, particularly those of mountain habitats. For all mountain gamebirds, when there is no shooting plan, bags are limited by the number of days when hunting is allowed, season bag limits, closed areas, and by the protection of hens for the 2 sexually dimorphic grouse (black grouse and capercaillie). Moreover since 1998, for all forest and upland gamebirds, the hunters are obliged by a national law to declare the number of birds shot.
Management of hunting bags of agricultural species (pheasants, partridges and quail) are not compulsory by law, so they are usually less intense or lacking. The latter is associated with the problem of shooting releases, as no hunting plans occur when pen-reared birds are released for shooting purposes only. However, this is changing since the mid 1980s for the grey partridge, as implementation of hunting plans depends on the will of hunters to manage their populations in a sustainable way. In areas with hunting plans, hunters must tag harvested birds to facilitate control of numbers shot.
Finally, management of territories by protecting some areas is common. On semi-private hunting territories (“A.C.C.A.”, “A.I.C.A.” or “société de chasse”), at least 10% of the area should be protected or is often protected (i.e. hunting is forbidden). The latter also results in some limitation of numbers shot by limiting the area where hunting is allowed.
We summarise below the situation for each species.
8 Shooting of capercaillie has been banned since 1967 in the Haute Savoie (French Alps), and since 1973 and 1974 in the Departments of the mountains of the Jura and the Vosges, due to the strong decline of populations. In these three mountain chains, the capercaillie was classified as a protected species in 1985, when law n° 76-629 on nature protection was enacted. Hunting is forbidden in the Cévennes, where the capercaillie was reintroduced. In the Pyrenees only male capercaillie can be shot (females are protected) and shooting regulations (quota numbers) vary according to Department. (Since 1998, the hunters must declare the number of birds shot). Additionally, during open season, hunting is allowed only 3 days per week and, within the Departments open to hunting, shooting of capercaillie is also banned in such areas as National Parks, reserves, and many localities of low population density. In Pyrenees-Orientales and Haute-Garonne, hunting plans are based on reproductive success estimates calculated on reference sites.
A hunting plan (quota) exists for Pyrenean grey partridge in some hunting associations of the Pyrénées-Orientales since 1997. A hunting plan also exists since 1999 in the national forests administered by the National Bureau of Forestry (ONF) in the Department of the Ariège. Pointing dogs are used to determine indices of reproductive success, which are taken into consideration for setting quotas. In some Departments, there is no daily or season bag limit (Ariège, Pyrenees Atlantiques). In the Aude, the Hautes-Pyrenees and the Pyrénées- Orientales, the daily bag limit is 2 partridges. In the Haute-Garonne, the season limit is 6 partridges. Shooting is banned in such areas as National Parks, reserves, and some localities of low population density in the Pyrenees Atlantiques. During open season, hunting is allowed only a maximum of 3 days per week in the Pyrenees.
A hunting plan for rock ptarmigan exists in the Department of the Pyrénées-Orientales since 1990. Because of the low densities and periodic disappearance of ptarmigan from edges of the distribution range in this Department, the quota has always been zero. In 2001, for the first time a shooting quota (91 birds) was established in the Department of the Hautes-Alpes. A hunting plan also exists since 1999 in the national forests administered by the National Bureau of Forestry (ONF) in the Department of Ariège. In some Departments, there is no daily or season bag limit (Haute-Savoie, Savoie). In the other Departments, the daily bag limit is 1 to 2 rock ptarmigan. In the Haute-Garonne, the season limit is 3 rock ptarmigan. Within the Departments open to hunting, shooting of rock ptarmigan is banned in such areas as National Parks, Reserves, and some localities of low population density. During open season, hunting is allowed only 1 to 4 days per week in the Alps and a maximum of 3 days per week in the Pyrenees. In summary, where there is no shooting plan, bags are limited by the number of days of hunting, daily or season bag limits and closed areas. In the Alps, about 69% of the area of potential habitat of rock ptarmigan is open to hunting. Within the area open to hunting, about 9% is controlled by private hunting interests, the rest being open to public hunting. No similar data are available for the Pyrenees.
A hunting plan for hazel grouse exists in the Department of the Jura since 1994, in the Ain since 1995 and in the Hautes-Alpes since 2001. Because of low densities, the quota has always been zero. In 2001, for the first time a quota (total of 3 birds!) was established in the Department of the Hautes-Alpes. In the other Departments open to hunting (Haute-Savoie, Savoie, Isère), there is no daily bag limit. Within the Departments open to hunting, shooting is banned in such areas as National Parks, reserves, and some localities of low population density. During open season, hunting is usually allowed only 4 days per week. In summary, where there is no shooting plan, bags are limited by the number of days of hunting, daily bag limits, closed areas and method of hunting (use of whistle banned). In terms of geographic
9 limitations, the exact percentage of the geographic range of hazel grouse in France open to hunting cannot be calculated, but a rough estimate would be 20%.
Hunting plans (quotas) for rock partridge exist in 3 Departments: in the Isère since 1987 (Sibut et al. 1998), in the Savoie since 1993 and in the Alpes-Maritimes since 2000. Pointing dogs are used to determine indices of reproductive success, which are taken into consideration for setting quotas. In two Departments (Hautes-Alpes and Alpes de Haute-Provence before 1997), bags are limited by the number of hunting days, a daily bag limit of 2 birds per hunter and the areas closed to hunting (National Parks and Reserves). Shooting of rock partridge has been closed in the Haute-Savoie since 1974 ans in the Drôme since 1980.
For black grouse, at the national level, only hunting of males is authorised. Additionally, some additional limitation of number of birds killed and/or number of hunting days may be implemented at the department level. Hunting plans are based on reproductive success obtained with pointing dogs on 35 reference areas distributed over the Alps, the trend in numbers of cocks counted on 32 reference areas, changes in occupied range (results of national inquiries on status by commune), and the vulnerability of populations to shooting. Finally, small, isolated populations on the edge of the distribution range are usually not hunted.
For quail, the number of hunting days per week was reduced in the 1970’s, but today there is no pattern of game management concerning the hunting bag.
For pheasant, where/when reared birds are released, hunting bags are not limited. On the contrary, in the case of wild populations, bags can be limited through different measures, undertaken where most hunters are favourable to such limitation to manage their wild population. Quotas may be controlled by limiting the number of birds harvested (quota defined according to reproductive success), or by limiting the number of hunting days. The latter is less frequently applied than the quota. Additionally, qualitative bags (shooting 2 cocks for 1 hen) are sometimes implemented for protecting females. This rule is often applied in increasing populations and not in stable populations where no particular rule is applied. On other occasions, hunting is oriented towards males: the existence of harems and the sex-ratio in a non-hunted population (1.5 cock : 1 hen, Biadi & Mayot 1990) justifies this practice. Indeed, it is assumed that the proportion of cocks shot is roughly equivalent to the proportion of cocks that would not have reproduced anyway because not all cocks are territorial. Finally, limitation of bags also occurs by protecting some areas. On semi-private hunting territories, at least 10% of the area is protected. The impact of such a measure depends upon the total area protected, its spatial configuration, and the habitat type(s) in this protected area.
Population management of the grey partridge has greatly changed recently: almost non- existent 20 years ago, partridge populations are nowadays managed on ca. 50% of communes (limitation of hunting, collective game management through membership to “GIC” – Reitz, in press). Bags can be limited by imposing a harvest quota or by limiting the number of hunting days. These measures are only undertaken where most hunters are favourable to such limitation to manage their wild population. Limiting numbers shot by quotas occurs in 12.5% of communes, whereas limiting number of hunting days occurs in 25% of communes. On many areas, hunters are allowed to hunt grey partridge fewer than 10 days a year. A combination of methods exists in some areas in southern France. Apart from these management types (which have a legal basis), hunters can limit their bags by defining their own rules for their specific hunting area (bag limits per season, per day, per hunter-day, non-
10 hunted parts of the area, etc.). However, in other areas, there are no particular rules of shooting limitation. In grey partridges, when shooting birds are released, the wild population can be protected against overshooting through a quota estimated for the wild birds (and thus marking released birds is needed to identify them). Additionally, when there are releases for reinforcing breeding numbers, hunting is usually prohibited for several years.
For red-legged partridge, there is no legal obligation to implement management of hunting bags. However, according to Reitz (in press), limitation of the number of hunting days is a common practice in many areas, and limitation of numbers harvested is frequent in the Centre and Provence-Alpes-Côte d’Azur regions. These two measures apply to 23% of populations of red-legged partridge.
Sanitary control
Sanitary control is implemented in pens for pen-reared gamebirds. Otherwise, no sanitary control is carried out for wild birds.
Supplementary food or water
No artificial feeding is carried out for mountain galliformes (hazel grouse, rock ptarmigan, black grouse, rock partridge, capercaillie or Pyrenean grey partridge).
In contrast, provision of grain is implemented commonly for grey partridges on managed areas, sometimes in combination with protection against raptors (in the form of a net, or a net with a bush). Water is sometimes provided together with grain. In a national inquiry concerning 485 managed territories in 1997-1998, supplementary food was implemented in 99% of territories, and was the only management taken in 26% of them (Mayot 1999). Provision of grain is also associated with releases of pheasants and partridges. Artificial water points are also a common management technique for pheasants and red-legged partridges.
UK
Habitat management
The most extensive habitat management for gamebird hunting in the UK, in terms of land area, concerns management of upland moorland for red grouse shooting (hunting). Current estimation of moorland in England and Wales is 7790 and 6360 km2 respectively (Bargett et al. 1995), whereas in Scotland it is estimated to cover 31000 km2 (Tapper 1999). It is estimated that 26% of that area is actively managed for driven grouse shooting in the 1980s, as opposed to 54% in the 1940s (Mackey et al. 1998). Heather moorland in recent decades has declined, changing to pine plantations or intensive grazing areas. At least in Scotland, loss of heather cover has been less in areas with a continuing interest in grouse management (Barton & Robertson 1997), so grouse management is associated with habitat preservation.
Habitat management within grouse moors is also intensive, and and primarily involves the rotational burning of heather patches (muirburn) (this occurs in 90% of 49 grouse estates sampled, GCT unpublished data). In areas where it is not suitable to burn, heather is cut. This was done in 16% of estates sampled. Burning or cutting creates a heather mosaic benefitial to grouse (Hudson & Newborn 1995). On average, each estate burned 95 ha of heather a year.
11 Bracken (Pteridium aquilinum) is also controlled on moors (spraying from the air), as it can replace heather, and additionally it provides a good habitat for ticks that spread Louping ill (Hudson 1986). 64% of 61 estates reported conducting bracken control, with an average area of 35 ha per estate (GCT unpubl. report). Drainage in areas may also be blocked, because the amount of invertebrates is higher in wetter areas (Hudson & Newborn 1995). 16% of 61 estates reported drain blockage (GCT unpubl. report). Additionally, numbers of sheep are also controlled in grouse moors to prevent excessive grazing of heather. 12% of 49 estates reported a reduction in sheep densities (GCT unpubl. report). Finally, there may be actions of heather restoration if needed. 28% of 49 estates reported reseeding heather with an average of 32.6 ha of moorland reseeded over the last 10 years per estate (GCT unpubl. report).
Management for other upland gamebirds for the purpose of shooting is less intensive or non existent. (There are, however, management programmes for declining species – e.g. Black grouse recovery projects in Wales etc). Black grouse benefits from some of the management carried out for red grouse. No habitat management is carried out with the specific purpose of increasing numbers of rock ptarmigan.
In the lowlands, habitat management directed to favour grey partridge and pheasant also occurs frequently. Habitat management in agricultural lowlands include planting of cover crops (either as a block or mixing of cover crops with main crops). From a sample of estates in Norfolk and Northumberland (n = 145), 74% were found to plant game cover crops. The area of game cover was about 4% of the tilled land within these estates (GCT unpubl. report). Additionally, the creation of conservation headlands (areas around a crop which are not sprayed with herbicides or pesticides, therefore promoting invertebrates and weeds for game and potentially other species) was implemented first in the UK and they are currently used in many estates (although no current quantitative data exist to evaluate the extent of this practice). Furthermore, set asides may also be managed for wildlife (by planting cover beneficial to wild birds). The uptake of such an option by a land owner may be influenced by whether there are shooting interests on the land, but no quantitative data exists for this aspect.
Finally, there is also management of woodland (cutting/maintaining rides, coppicing, woodland regeneration) for shooting interests of pheasant and capercaillie. Maintenance of old woodland was more common in areas with shooting interests, as well as planting new woodlands for game, particularly in more recent years (Duckworth et al. 1999). Similarly, another study showed that far more woodland creation and management takes place on land used for pheasant releases than elsewhere. However, in areas where shooting was mainly based on released pheasants rather than wild birds, management of crop and crop boundaries showed few consistent differences between game and non-game sites (Hinsley 1999), reflecting the lower intensity of those practices in such cases.
Releases
Captive rearing and releases of red-legged partridges for hunting purposes started in the 1960s in response to the decline of grey partridge populations (to maintain a partridge shoot despite the decline of the grey), and then increased rapidly. In the 1970s, releases of chukar and red- legged/chukar hybrids started. The latter was prohibited in 1992. An estimated two million red-legged partridges are currently released every year (Tapper 1999). Grey partridges are also released in increasing numbers, although at a much lower scale than red-legged partridges (Tapper 1992).
12 Pheasants are also released in big numbers. Releases had been carried out throughout the whole of the 20th century, but numbers have increased dramatically in recent decades, and currently around 20 million pheasants are estimated to be released each year (Tapper 1999).
Predator control
Predator control is intensive both in the uplands and in the lowlands. It aims primarily at controlling red fox and corvids (crows, rooks, jackdaws, jays and magpies), but also stoat, weasel, mink, hedgehog, polecat, feral cats, grey squirrels and brown rats. Game keepers are employed on most estates, and one of their main jobs (together with habitat management) is to control predators.
Management of hunting bags
In grouse estates, control of hunting bags is very strict. Control is exercised through limiting the number of hunting days, which is calculated according to breeding densities and breeding success (calculated annually with pointer dogs). Number of lagomorphs shot in each estate is also regulated. In fact, populations of lagomorphs are actively controlled not only for sport, but also to avoid damage to heather by excessive eating of young heather shoots, and to avoid attracting buzzards or foxes that may put the grouse under increased predation pressure.
Unlike grouse, hunting bags of pheasants or red-legged partridges are not controlled (or to a lower extent). Grey partridges are only rarely shot nowadays.
Sanitary control
Medicated grit (grit with a coating of an Anthelmintic drug, which reduces worm burdens) is provisioned in most red grouse estates. No sanitary measures are taken for other upland species, or in the lowlands.
Supplementary food or water
Non-medicated grit is sometimes provided in some red grouse estates, although no information exists on the exact number of estates or the quantity put out. Supplementary food (grain) is also provided in the lowlands for both partridges and pheasants, but no quantitative data on the extent was available.
Spain
Management for gamebird hunting in Spain is relatively common. Nevertheless, it is globally difficult to quantify or even get accurate descriptive information about the type and extent of practices implemented, because there is no tradition in the hunting societies in Spain to gather even basic information about game densities or economic benefits, least on other variables (Vargas 1997).
Habitat management
After the strong decline of partridge populations at the end of the 1980s, stronger emphasis on habitat management in agricultural areas has occurred in Spain, and preferred practices have
13 been described in many hunting journals and publications (e.g. Segovia 1994, APROCA 1996, Vargas & Cardo 1996, Nadal 1997). Habitat management for partridges include the maintenance of traditional dry cereal regimes (with rotational set-aside every three years, i.e. each field is cultivated only once in three years), reduction of pesticides and fertilisers, maintenance of bushy areas and field margins, maintenance of olive trees-vineyard mosaic, planting cover under olive trees, etc.
Habitat management for other species is much less intensive or non-existent. Management for Pyrenean grey partridge is mainly based on the maintenance of traditional livestock regimes, but the application of these practices for hunting purposes is rare. No particular management is carried out for other gamebirds.
Releases
Numbers of partridges released have increased in recent times. Currently, around 3,000,000 red-legged partridges are reared and released every year in Spain. Farm rearing of partridges has become an important business in the 1990s. Some hybrids red-legged – rock partridge have been released. Releases are carried out at different times of the year. An inquiry in Andalucia showed that there was a big disparity between hunting societies in relation to the preferred time for releases (Vargas & Roman 1996). It is however frequent to release birds at the end of the hunting season, at least in Andalucia (Vargas 1997). Records of releases are not very accurate. However, data from a sample of hunting areas in Castilla-León showed an average of 170 release points per year, mainly of partridge and rabbit (APROCA unpubl. report).
Pheasant hunting is not so popular as in other northern countries, so fewer birds are released here than in other countries. However, records or releases are also very scanty. In the above- mentioned sample, only 3.3% of all individuals released in Castilla-León were pheasants.
Predator control
This is a very common practice in the whole country, mainly in partridge hunting areas. Many hunters believe that the main problem for game is predators (Vargas & Roman 1996). Predators legally controlled include red fox, feral dogs, magpies, carrion crows, snakes, rats. Keepers employed for that purpose often carry out predator control, but additional culling by hunters in associative hunting areas is also carried out.
Management of hunting bags
Many private estates have managers that manage hunting bags based on breeding data and hunting pressure. However, in most areas, hunting societies impose quotas by reducing numbers shot and number of hunting days, but without a clear evaluation of the shooting surplus (Vargas 1997). The need to develop and implement detailed management plans for hunting bags is increasingly demanded by hunting organisations (APROCA 1996, Montoya Oliver 2000).
Additionally, some hunting associations are implementing the establishment of a protected area within the estate to protect game (Vargas 1997), although no quantitative data was found about the extent of this measure.
14 Sanitary control
Sanitary control is carried out in pens for partridges and pheasants. No sanitary control is implemented for wild gamebirds. Otherwise, sanitary control is carried out for rabbits, by vaccinating against haemorragic disease (Bravo & Peris 1998).
Supplementary food or water
Provision of water places and grain for partridges is a very common practice both in private estates and in associative hunting estates.
Portugal
In Portugal, the legal hunting framework has changed considerably since the early 1970s, thus making it very variable the extent and kinds of game management carried out in the country. The most recent legal framework dates from September 2000, and this will probably induce further changes to the game management system.
Until very recently, there were broadly two types of hunting regimes: a general public hunting regime, with no management other than the limitation of the number of hunting days and the daily bag per hunter; a special hunting regime, where management practices were usually carried out, and where managers have potential control over the number of birds hunted (Borralho et al. 1996, Borralho et al. 1997, Borralho et al. 2000). In the present legal context, the general hunting regime was discontinued, assuming that management is always needed to keep the hunting activities sustainable. In consequence, there are currently four categories of hunting areas in the country: National Hunting Areas (NHA), managed by a government agency and open to the general hunters on a restrictive basis and given the payment of an access fee; Municipal Hunting Areas (MHA), managed by a range of local, public or private organisations, and open to the general hunters given the payment of an access fee; Tourist Hunting Areas (THA), managed by a private company on a commercial basis; Associative Hunting Areas (AHA), managed by an association of hunters and open to its members and invited persons, provided there is no payment of access fees. Overall, these areas occupy about 35% (3,140,000ha) of the country: 0.5% - NHA; 2.8% - MHA; 10.8% - THA; 21.4% - AHA. There are still large areas occupied by the general hunting regime, but these are being converted to other hunting categories; it is expected that they will disappear in the medium term.
For the creation of any of the extant categories of hunting areas, it is necessary the approval by the governmental hunting agency of management and annual exploitation plans. These should detail how and where the management actions will be carried out, what game species will be hunted and how the populations will be exploited and monitored. However, it is very difficult to get quantitative information about the extent of these practices at a national level, as hunters are not forced to report on these activities at a national level, so no centralised information exists on these issues. Results of a recent inquiry based on questionnaires sent to hunters (Cipriano 1999) showed that 46% of hunters participate in management actions in the AHA to where they belong, These hunters spent overall 80 hr per year in these activities, 98% of which were non reimbursed.
15
Habitat management
In Portugal, particularly in AHA and THA, there is some of habitat management techniques to improve the conditions for small game. This management is targeted primarily at increasing the carrying capacity for the red-legged partridge. One of the main management activities is the creation of game crops, intended primarily for supplementary feeding during periods of food scarcity. These, however, may also be useful in providing cover and breeding habitat. Game crops include the plantation of small patches with sorghum, sunflower or leguminous across the game estates (e.g. Borralho et al. 2000). Another frequent habitat management practice is the burning of bramble (Rubus spp.) thickets, in order to reduce the potential cover to mammalian carnivores. In normal conditions, only an estimated 12% of hunters participate in habitat management activities in AHA (Cipriano 1999).
Releases
The release of red-legged partridges is one of the most common game management actions carried out in Portuguese hunting areas (Castro Pereira et al. 1996, Capelo & Castro Pereira 1996, Castro Pereira et al. 1998, Carvalho et al. 1998). Release of pheasants and quails has also been carried out in the past few years, though to a much lesser extent than the partridge. Releases of quails have included both the common (Coturnix coturnix) and the Japanese (C. japonica) species. Currently, Portuguese legislation only allows the restocking with autochthonous species or sub-species. An estimated 24% of hunters in AHA declared to have been involved in release activities (Cipriano 1999), but there is currently no estimate of the number of released birds.
Predator control
Predators legally controlled in Portugal include Egyptian mongooses, red foxes, feral cats and dogs, magpies, jays and carrion crows. Data from inquiries allowed an estimate of 0.74 mongooses captured each year per km2 in Portuguese game estates (SD = 1.47, range 0-14.5, n = 1408) (Borralho et. al 1996). An estimated 20% of hunters have declared to participate in predator control in AHA (Cipriano 1999).
Management of hunting bags
For resident game birds (red-legged partridge and pheasant), the annual exploitation plans for each hunting area should define clearly the numbers of each species that can be shot each year. Under the general hunting regime and for migratory species, the hunting bag is lightly controlled through the limitation of the number of hunting days per week and by a daily limit to the numbers of each species that may be shot per hunter. Weekly, there may be at most three hunting days for migratory species in special hunting areas during the hunting season, whereas under the general regime hunting activities can only be carried out on Thursdays, Sundays and Bank Holidays. For game birds the daily hunting bags are set annually by the government hunting agency; in 2001 these are one pheasant, three partridges and ten quails per hunter per day.
16 Sanitary control
To the best of our knowledge, there is no regular sanitary control for gamebirds in Portuguese hunting areas.
Supplementary food and water
Provision of grain was implemented by 19% of hunters in associative areas (Cipriano 1999), and is a common activity for partridges and pheasants. Additionally, in southern and dry areas like Portugal, distribution of red-legged partridges is clearly associated with water distribution (Borralho et al. 1998), so provision of water points is also a common practice.
Global patterns
Despite the differences observed between countries, management of habitats and of gamebirds is relatively common in all countries considered. Habitat management at large scale only occurs in grouse moors in Great Britain, and in some areas in Spain where private estates are managed for red-legged partridges. Small scale habitat management for gamebird hunting interests is particularly common in farmland areas, for partridges and pheasants. Predator control is common in all countries, and is exerted for all species that are commonly hunted. Only for species that are hunted in small quantities (mountain gamebirds in the UK, France or Spain), predator control is not carried out.
Overall, and with the exception of areas where the primary land use is hunting (such as grouse moors), management of habitats and predators is more common and intense for farmland species, those that are currently declining at a higher rate (see WP1). In contrast, management of bags is currently more common for upland galliformes, although it is becoming more common for farmland gamebirds, particularly in UK and France.
Information lacking
Quantitative information of the extent of management practices is particularly lacking in Spain, Portugal and Finland. Of these three countries, this lack of information is particularly important for the Iberian Peninsula, given that the overall intensity of management is higher there than in Finland.
17 3- EFFECT OF MAIN MANAGEMENT PRACTICES ON GAME AND NON-GAME SPECIES
3.1 PREDATOR CONTROL
Predator control is a highly debated practice. Some conservationist groups have expressed ethical and biological reasons against this practice (e.g. Messmer et al. 1999), whereas hunters consider it essential to maintain game numbers (e.g. Reynolds et al. 1988), and even to maintain healthy ecosystems (Tapper 1999). Among scientists there is not a consensual opinion (Marcstrom 1995, Swam 1995, Coté & Sutherland 1997), and very different positions exist, partly dependant on the complexities in the predator-prey relationships, but also because of the socio-economical and ethical complexities of this issue (Herranz 2001).
This debate is particularly relevant nowadays given that, on the one hand, conservation and ethical considerations are more important now than several decades ago and, on the other hand, there is a perceived and sometimes proved increase of certain predators in recent decades (Herranz 2001, Tapper 1999), which has lead hunters to overtly claim that predator control is more important now than ever (e.g. Vargas & Roman 1996). However, at least in some cases, predator control of some species is carried out without the knowledge on whether predators culled have or not an influence on the target game (Herranz 2001 for Jackdaw or dormouse).
Finally, given the overall impact of human management on ecosystems, and the changes occurring in predator guilds in recent times, predator control is sometimes claimed as important for conservation purposes (Suarez et al. 1993). Therefore, predator control may be beneficial in some systems, and this has been claimed by hunters (Reynolds & Tapper 1996). On the other hand, predator control has been considered as a factor destabilising predator guilds, and thus detrimental for conservation (Moral Castro 1999).
In an extreme, one of the sources of conflict is that predator control may sometimes affect species that are legally protected, such as birds of prey. For example, in Europe, heavy persecution due to conflicts with hunting interests resulted in severe declines or the extirpation of a number of raptor species from several countries (Newton 1979). The issue of the relationship between illegal control of raptors is detailed in the report for Workpackage 3 (Mañosa 2002). We present here a brief review of the known and potential effects of legal predator control on different population parameters on target and non-target species. We concentrate mainly on studies referring particularly to gamebirds, although studies related to other birds (pigeons or passerines) or lagomorphs are included when relevant, given that hunting gamebirds is usually associated with hunting of these species (see report of Workpackage 1, Viñuela 2002).
Predation and gamebirds
Predation is found to be the main cause of nest failure and mortality of adults in most gamebirds in all REGHAB countries. In France, mammalian and avian predators are the main cause of nest and adult mortality in hazel grouse (Bernard-Laurent et al. 1994, Léonard & Montadert 2000), capercaillie (Menoni et al. 1991), ptarmigan (Miquet & Deana 2000), grey partridges (Reitz et al. 1993, 1999; Bro 1998, Bro et al. 2001), Pyrenean grey partridges (Novoa 1998) and black grouse (Caizergues 1997, Bernard-Laurent 1994). Predation is also reported to be the main cause of mortality of released pheasants, mainly by ground carnivores
18 (in particular the red fox) and in some areas by raptor species (Havet & Biadi 1990, Mayot & Brouillard 1993, Mayot et al. 1993, 1997). Similar results have been found in Fennoscandia for tetraonids (e.g. Angelstam et al. 1984, Marcstrom et al. 1988). In the UK, predation is also the most important cause of nest failure and adult mortality in grouse (Thirgood et al. 1998) or partridges (Potts 1980). In Spain, predation by generalist predators (mainly foxes and crows, but also other species) is also important in partridges (Yanes & Suarez 1996, Herranz 2001).
Because of that, predation is believed by hunters to be a cause of decline of small game (Vargas & Roman 1996) or a limiting factor for gamebirds (Herranz 2001). However, predation may have a positive effect on gamebirds, because predators may have a sanitary effect on prey, if predation is directed primarily towards ill animals, and there is selective capture of weaker or more accessible individuals (Delibes 1980, Temple 1987, Hudson et al. 1992, Fernandez Llario & Hidalgo 1995, Moller & Erritzoe 2000). In any case, the effect of predation as a limiting factor for gamebirds has been rarely evaluated, because it is difficult todisentangle their effects from that of other factors driving gamebird populations. In some cases, studies have suggested that, indeed, predation may play a role in population limitation. A wider description of the impact of predation as a limiting factor is detailed in the report for Workpackage 3 (Mañosa 2002).
The negative impact of predators on gamebirds may be limited by destructive methods (predator control, aiming to limit the number of predators), or else by indirect methods aiming not to limit the number of predators, but their impact, for example manipulation of habitat to reduce predation risk. In this section, we concentrate on the effect of predator control per se, because predator control for hunting purposes is very common in all REGHAB countries, at least for some species (see section 3 for details). A more detail review on indirect methods to reduce predation risk will be carried out within workpackage 4.
Predator control is implemented under the assumption that reducing the density of predators is going to reduce predation rates on both nestlings and adults, thus increasing breeding success and the shooting surplus, and reducing adult mortality and thus increasing breeding densities of game. We review subsequently the results of studies attempting to address this issue.
Effect of predator control on small game species
Some studies have shown a correlation between the mortality rate of gamebirds and estimates of the abundance of predators (e.g. Bro et al. 2001). This has been used to imply that limiting the density of predators would result in reducing mortality rate of game. Whereas this might be true, correlative data like that does not take into account potential differences between areas in ecological variables, which may affect simultaneously predator abundance, game abundance, and game vulnerability to predators. Experimental work can, in contrast, test whether removing predators has any effect on game population dynamics (breeding success or density).
Previous reviews of experimental predator control studies showed that predator control often increases breeding success of small game, and thus the size of the autumn (harvestable) population in many cases, although it is less clear whether predator control affects breeding density (see reviews in Reynolds et al. 1988, Stahl & Migot 1993, Newton 1993). Similar results were also found by Coté & Sutherland (1997), who performed a meta-analysis on the results of 20 experimental studies on different species (gamebirds and others). According to
19 their analysis, predator removal had a large, positive and significant effect on hatching success of the target species. Removal of predators also had a large and significant positive effect on post-breeding population size, but the effect on breeding population sizes of target birds was not statistically significant, according to the meta-analysis.
These previous reviews combined studies carried out on all different types of game (lagomorphs, wildfowl, galliforms, passerines) as well as non game species. We present here a summary of the results of experimental and semi-experimental studies (the latter includes comparisons between areas where predator control is carried out or not, even if culling was not done with an experimental purpose; and comparisons for a given area between years when predators had been decimated by illness, or killed because of rabies) of predator control on gamebird species (Table 1). We concentrate mainly on studies referring to European species, but we have included American studies when referring to the same species as here. Studies related to lagomorphs are only included when relevant, given that gamebird hunting is usually associated with hunting of these species (see report of workpackage 1), and that predator control is usually implemented for increasing numbers of both game types (see section 3).
For gamebirds, 75% of 17 studies found some positive results of predator removal on reproductive success (2 of them found positive results only some of the years of the study, but not others), and an equivalent proportion of them found that predator control resulted in increased autumn numbers (thus bags); 43% of 14 studies found positive results on breeding densities (Table 1). The pattern did not differ markedly between species (Table 2); the effect of predator control on breeding densities was more marked for agricultural species than for upland species, but there were very few studies carried out with partridges.
Table 2. Effect of predator control on population parameters of galliformes species, according to species habitat (proportion of studies that find a positive effect, sample size in brackets). Upland galliformes are Black Grouse, Capercaillie, Hazel Grouse, Willow Grouse and red grouse; agricultural species are grey partridge, red-legged partridge and pheasant.
Reproductive success Bags / Autumn Breeding densities population size Agricultural 80% (5) 75% (8) 66% (3) Upland 75% (12) 72% (11) 36% (11)
The above-mentioned studies were all carried out with wild populations. Few studies have been carried out with the effect of predator control on the survival of released pen-reared birds, although it is well known that they suffer high rates of predation in the days following release (see section 3.2), and predator control is usually recommended in release campaigns (see section 2). An experimental study of predator control did not report any significant difference in survival rates of hand-reared pheasants released in spring-summer (Mayot et al. 1993).
Similar patterns were found for lagomorphs, although there exist fewer studies on the effect of predator removal on hare or rabbit populations, particularly in southern Europe. Predator removal was usually associated with an increase in numbers in the autumn, but only 50% of 10 studies found an effect on breeding numbers. The effects are apparently more marked for hares, for which most studies found positive effects (Marcstrom et al. 1989, Jensen 1970, Trautman et al. 1974, Lindstrom et al. 1994, Smedshaug et al. 1999; but see Guthery & Beasom 1977, Kauhala et al. 1999). In rabbits, results are more contrasting (Trautman et al. 1974, Guthery & Beasom 1977, Newsome et al. 1989).
20 Table 1. Experimental (or semi-experimental) studies on the influence of predator control on population parameters of European gamebird species. Species Increased nest Increased Increased breeding Predators culled Reference success postbreeding numbers numbers Black Grouse (YES) Carnivores crows Baines 1991 Black Grouse NO NO Carnivores crows Baines 1996 Black Grouse YES YES NO Carnivores Kauhala et al. 2000 Black Grouse YES YES YES Carnivores Marcstrom et al. 1988 Black Grouse YES NO NO Crow Parker 1984 Black Grouse YES Fox Lindstrom et al. 1994 Black Grouse YES Fox Smedshaug et al. 1999 Black Grouse NO Fox, marten Ellison 1996 Capercaillie NO NO NO Carnivores Kauhala et al. 2000 Capercaillie YES YES YES Carnivores Marcstrom et al. 1988 Capercaillie YES Fox Smedshaug et al. 1999 Hazel grouse YES YES NO Carnivores Kauhala et al. 2000 Red grouse YES Carnivores crows Tharne et al. 2001 Willow grouse YES Crows Erikstad et al.1982 Willow grouse YES YES NO Carnivores Kauhala et al. 2000 Willow grouse (YES) NO NO Crows Parker 1984 Willow grouse YES Fox Smedshaug et al. 1999 Pheasant YES NO NO Carnivores crows Chesness et al. 1968 Pheasant YES Fox Jensen 1970 Pheasant YES Carnivores crows Trautman et al. 1974 Pheasant NO Carnivores crows Trautman et al. 1974 Pheasant NO Crows Goransson & Loman 1982 Grey Partridge YES Fox Jensen 1970 Grey Partridge YES YES YES Carnivores crows Tapper et al. 1996 Grey Partridge YES Carnivores Frank 1970 Redleg partridge YES YES YES Carnivores crows Potts 1980 Redleg partridge YES Carnivores crows Ricci et al. 1990
21
22
Factors influencing the efficacy of predator control
Despite the general trends seen above, no positive effect of predator control on the breeding success or abundance of game has been found in many studies. The general lack of effect of predator control on breeding density suggests that predation on adults is compensatory in many cases, not additive. Additionally, the efficacy of predator control on breeding success (and/or breeding density) depends on many variables. The relationship between predation and interacting factors is complex, and we are still far from having enough knowledge to manage populations in all circumstances. Among others, the following factors have been found to influence this relationship:
Habitat structure. Habitat is one of the most important variables affecting the relationship between prey and predator and thus, potentially, affecting the effect of predator removal on prey mortality. Habitat characteristics affect predation risk in many gamebirds (Angelstam 1986, Brittas & Willebrand 1991). For example, partridge nest predation depends on vegetation cover (Rands 1987, Potts 1980), vegetation structure (Yanes & Suarez 1995), or the spatial structure of hedges (Ricci et al. 1990). The predation risk on breeding female partridges is also related to the frequentation of particular habitat features (Bro 1998, Reitz & Mayot in press). The presence of a copse, a bush or a cover such as a fodder crop, rape-seed or set-aside is likely to decrease the predation risk by raptors. Partridges killed by mustelids were near a copse and had many lanes in their home range (Reitz & Mayot in press). Additionally, surveys of Buzzard (Buteo buteo) predation at pens of released Pheasants (Phasianus colchicus) showed pens with heavy predation had little shrub cover, much deciduous canopy and few Pheasants relative to the size of pen (Kenward et al. 2001).
A stronger effect of predator removal has been found in degraded areas, i.e., when habitat is degraded and thus vulnerability of the prey is higher because of lack of proper habitat and concealment (Chesness et al. 1968, Stahl & Migot 1993, Sovada et al. 2001). This may explain the apparently higher effect of predator control on species of agricultural habitats (Table 2).
The intensity of culling (e.g. area culled, number of trappers or number of traps per surface unit). Given the mobility of certain predators, for example foxes, and the number of floaters in a normal population (Reynolds et al. 1993), culling may not be reflected on reduced number of predators in a given area, and regional suppression of predators may be only achieved through many local control efforts within a region (Reynolds 2000). Stahl & Migot (1993) found that the effect of predator removal was more marked when the surface where culling took place was big (11/13 studies in their review) than when it was small (only 1/9). Experiments in France (Drillon 1995) suggest that the intensity of predator trapping necessary to have a positive effect on capercaillie populations is very costly, because it should be done over large areas to be effective.
The diversity of predators culled. Removing one predator may be without effect on the prey if removal is followed by compensatory predation by other predators within the guild. Stahl and Migot (1993) in their review found that the effect of predator removal was more marked if several main predators were controlled (only one out of six studies found a positive effect of predator removal if only one predator was controlled, as opposed to 8/10 when all predators were killed). Other studies have found the same effect (Parker 1984, Greenwood
23 1986, Norrdahl & Korpimaki 1995). In contrast, Coté & Sutherland (1997) did not find heterogeneity among studies in relation to whether all or only one predator had been removed.
Intraguild relationships. Some predators may act on the other predators of the community, either eliminating them from the area through territorial exclusion, or other mechanisms (e.g. intraguild predation), so eliminating some predators may make overall predation pressure on prey higher or at least not different, because abundance of other predators will change as a result of the elimination of one of the predators (Palomares et al. 1995, 1996, 1998, Sovada et al. 1995).
Game abundance and trends. Predator control exerted in declining populations do not necessarily stop the decline (Beauchamp et al. 1996), and may even be entirely ineffective (Parr 1983, Coté & Sutherland 1997).
Abundance of alternative prey. Many studies have found that the effect of predator removal on galliformes depended on the abundance of voles (game being secondary prey for many predators), and are only visible in years of low vole abundance (Parker 1984, Marcstrom et al. 1988, Kauhala et al. 2000). Food abundance for game. The effect of predator removal may only be found in years of overall poor food supply for the game, the only ones when predation may be limiting (Baines 1991). In conditions of food scarcity, game may need to move further distances to find it, and/or body condition may be poorer. Additionally, in those conditions, vigilance may be reduced. All these factors may increase predation risk when food is in short supply.
In summary, predator control seems to affect the size of the harvestable population of small game, but not necessarily the size of the breeding population. The efficacy of predator control programmes is likely to vary according to population and ecological variables. To maximise efficiency, predator control has to be carried out in combination with habitat manipulation, and/or be very intensive, culling all potential predators, and over a large area, and particularly in years or conditions when predation is highest (when food conditions for the predators are low, and when habitats are highly degraded).
Effects of predator control on species other than target game
The effect that predator control exerted for game management may have on other species is difficult to assess with the available evidence. Much fewer studies had been carried out that try to assess such an impact, and the only experimental work designed to such effect (The Otterburn experiment, Hoodless 2000) has started only recently. All other studies are either correlative or based in indirect evidence.
Predator control could potentially have a direct impact on other species, if those species are also limited by predation (and by the same predators). In their meta-analyses, Coté & Sutherland (1997) did not find any heterogeneity in the effect of predator control between game and non-game species, suggesting that the positive effects of predator control on breeding success could be shared by other species. However, the lack of effect of predator control on breeding densities (both for game and non game species) also suggested that such a benefit would be limited. In contrast, positive effects of predator control had been found in some studies. For example, Suarez et al. (1993) found that passerine nests in a nature reserve suffered high mortality rates due to predation (mainly by foxes and dogs). There was, in contrast, higher success in a close-by area where there was a hunting reserve where feral dogs
24 and occasionally foxes were killed. The latter suggests that lack of control of nest predators after declaration of the reserve may had resulted in an increased and unsustainable predation rate for passerines. Similarly, the hatching success of 5 out of 6 wader species in Scotland increased significantly after crow and gulls were controlled (Parr 1993), although breeding numbers over the three years of the study did not increase. Tharme et al. (2001) found that density of golden plover, lapwing and curlew were higher on grouse moors where predators where culled than on moors where no predator control was implemented, while meadow pipit, skylark, whinchat and crows were less abundant (although the latter was probably due to habitat rather than predator management). Stoate and Szczur (1994) found that hatching success of some songbirds (particularly those having low values of hatching success in control conditions) increased as a result of corvid removal. In contrast, other studies have found no beneficial effect. Baines (1996) found that neither density nor breeding success of Black grouse differed between keepered and non keepered moors. Finally, Green and Etheridge (1999) analysed whether nest success of hen harriers varied between keepered and non keepered moorland (as removal of foxes could potentially increase breeding success of that ground-nesting raptor). No significant effect was found, once controlling for differences related to human interference, and they suggested that even a supposed beneficial effect of the control of foxes and other predators on hen harrier nest success, its effect on the population trend of hen harriers would be small relative to the large negative effect of persecution.
Additionally, given that predators may be related between them through intraguild predation or competitive exclusion, predator control may have an indirect impact on other species by altering the structure of the predator guild. Usually, removal of top predators may produce a release in numbers of mesopredators. Therefore, predator control may be positive if the unmanaged structure of the predator guild is biased towards mesopredators, and these are the ones culled for hunting purposes. For exemple, Travaini et al. (1997) observed that diversity and evenness of carnivore community decreased over 6 years in a protected area (where no culling of foxes was carried out) as compared with a less protected area (managed for rabbit and partridge, and where continuous culling of foxes was carried out, including illegal control methods such as poison or snares). This result was apparently due to a disproportionate increase of red fox in the protected area, and it was suggested that overabundance of red foxes could thus on a long term basis negatively affect endangered specialist species such as lynx.
Alternatively, if predators affecting non-target prey are species potentially limited by culled predators, removal of those top predators may be detrimental to non-target prey because of meso-predator release. For example, removal of foxes in some areas may produce an increase of rodent numbers, which in itself may have high impact on nest predation of passerines or small ground birds. A higher failure of partridge nests in areas with fox removal has been found in central Spain (J. Vinuela com. pers.). Dion et al. (1999) evaluated whether predator control for waterfowl could have a negative impact on grassland songbirds nesting in the same habitat. Management for wetlands include control of fox, striped skunk, raccoon and badger, some of which predate on rodents (nest predators of passerines), but who are also predators of passerines. In their study, removal of those predators did not have a significant effect on the nest success of song-birds: the relative effect of medium-size mammal predators decreased in removal areas, but the effect of bird and small mammal predators increased.
25 Conclusions
Predation is an important factor of nest and adult mortality in gamebirds. Relatively little is known, however, about the regulating or limiting effect of predation on wild populations. Overall, predator control increases significantly breeding success of gamebirds and hence the size of the harvestable population. The effect of predator control on breeding densities is less consistent. The efficacy of such practice may vary enormously according to an array of ecological and population variables. In many instances, predator control programmes have to be intensive and large-scale to achieve an effect on game populations.
The effect of predator control on species other than gamebirds is little studied. Both positive and negative effects may be expected, and the relative importance of both would depend on the type and extent of control exerted. However, no studies up to now have shown negative effects of predator control on other species, and available information for positive effects is inconclusive.
In summary, predator control increases the economic benefits of game, but it is unclear whether it is always critical for the exercise of hunting (as breeding densities are not always affected, although see Redpath & Thirgood 1999, Thirgood et al. 2000). Whether such a management technique is also beneficial for other species is even less clear, although it has the potential to be (at least for species that are limited by the same set of predators as game, or for maintaining balanced and diverse predator guilds).
26 3.2 RELEASES AND RESTOCKING
The introduction into hunting grounds of captive-reared or, less frequently, wild-caught individuals, with an intention to increase the number of birds available for shooting, it is one of the most widely used gamebird management techniques. Although it is not possible to estimate with certainty the extent of this practice, it is likely that at least several million gamebirds are introduced each year in Europe (see section 2, and Table 1). These introductions correspond to at least two very distinctive practices, though there is a range of variants and intermediate stages:
Releases are defined herein as the introduction of birds shortly before the shooting period, with the sole objective of short-term increasing numbers to boost the hunting yield. This technique is used when the harvests are much higher than the supply provided by the wild populations, thus the natural stocks need to be supplemented to sustain the shooting interests. A typical example is the pheasant (Phasianus colchicus) shooting in the UK and France, where the breeding populations cannot sustain the hunting bags (Robertson 1990, E. Bro pers comm.).
Restocking refers to bird introductions aimed at reinforcing or re-establishing a local breeding population. This technique is frequently used in association with habitat management, to revert declines or to speed up population recoveries (e.g., Brun & Aubineau 1989, Carvalho et al. 1998). In these circumstances, the hunting yield remains limited by the survival and fecundity of the breeding population, which are conditional on other management actions, such as predator control or habitat improvement. Additionally, in some cases, restocking is associated with a ban of hunting for several years, as is the case in France (see section 2). Restocking has been used also to establish gamebird populations outside their normal geographic ranges (Holechek 1981, Robertson 1996) and in conservation programs for declining or threatened populations (Putaala & Hissa 1998, Svedarsky et al. 2000).
A number of rearing and release methods have been developed to introduce gamebirds into hunting areas, aimed essentially at enhancing the survival and breeding productivity of the birds (Anon. 1983, 1994). The main problem to be overcome is the low survival of hand- reared birds, which seems related primarily to high predation levels, largely associated with their poor behavioural, morphological and physiological capacity to live in the wild (Robertson & Hill 1992, Schroth 1991, Leif 1994, Capelo & Castro Pereira 1996, Carvalho et al. 1998, Putaala et al. in press). These limitations are believed to cause also the low reproductive output of released birds (Hill & Robertson 1988, Putaala & Hissa 1998)
A large proportion of the high mortality rates of hand-reared birds may be assigned to their inadequate anti-predator behaviour (Thomas 1987, Havet & Biadi 1990, Dowell 1990, Mayot et al. 1993, Antilla, Putaala & Hissa 1995). Also, hand-reared gallinaceous birds tend to have shorter intestinal tracts and lighter gizzards than wild birds, which result from being fed on low-fibre, commercial diets (Moss 1972, Hanssen 1979, Paganin & Meneguz 1992, Putaala & Hissa 1995, Liukkonen-Anttila et al. 2000, Millán et al. 2001). Greek partridges (Alectoris rufa) with larger intestines survive longer after release (Paganin et al. 1993) and reproduce better (Bergero et al. 1995). Moreover, sudden switches from commercial to natural diets, which are relatively indigestible, low in nutrient quality and containing plant secondary compounds, may render released birds more susceptible to starvation and predation (Liukkonen-Anttila et al. 1999, 2001). Predators may also find hand-reared birds easier to catch due to their limited flying ability and endurance (Robertson et al. 1993, Putaala et al.
27 1997), caused by modifications of the musculature and heart associated with the restriction of movements in aviary conditions (Majewska et al. 1979, Putaala & Hissa 1995, Mäkinen et al. 1997, Pyörnilä et al. 1998, Liukkonen-Anttila et al. 2000). Besides the low quality of hand- reared birds, the high stocking densities of vulnerable prey should also contribute to increase the losses, by prompting functional and numerical responses of the predators (Kenward 1977, Harradine et al. 1997, Kenward et al. 2001).
Table 1. Examples of gamebird restocking or release operations in Europe. The figures refer to annual numbers, except where indicated otherwise. When there were different figures available for one species in a given country, the most recent estimate was used. Species Numbers. Year- Country References 1 Red-legged partridge Alectoris rufa » 2.5 millions France Tupigny (1996) unknown Portugal Carvalho et al. (1998) » 3 millions Spain Gortázar et al. (2000) » 2 millions UK Tapper (1999)
Alectoris rufa ´ A.graeca unknown Spain Negro et al. (2001)
Chukar Partridge Alectoris chukar 100,000 – Cyprus Hadjisterkotis 1998 150,000
Grey partridge Perdix perdix 500 Finland P. Vikberg, from Finnish Hunters’ Central Organisation, pers. comm. » 2 millions France Tupigny (1996) unknown UK Tapper (1992)
Japanese quail Coturnix c. japonica 300.0001 France Guyomarc’h & Boutin (pers. comm.) unknown Portugal Rafael et al. (1996) unknown Spain Rodriguez-Teijeiro (1993)
Ring-necked pheasant Phasianus 15,000 Finland P. Vikberg, from Finnish Hunters’ colchicus Central Organisation, comm. pers.). 10-15 millions France Mayot & Biadi (2000) unknown Portugal Faria & Torrinha (1996) » 20 millions UK Tapper (1999)
Reeve’s pheasant Syrmaticus reveesi 10,000 – 50,000 France Alain Roobrouck pers. comm. 1 Numbers released during the 1980’s; releases of Japanese quails are currently illegal in France.
The contribution of hand-reared birds to the breeding population is frequently small, due to their low survival rates in the few weeks after release (Havet & Biadi 1990, Brittas et al. 1992, Hill & Robertson 1998, Putaala & Hissa 1998). There is also some evidence that even those hand-reared birds surviving to the breeding season may have a much lower breeding success than wild birds. In pheasants, Hill & Robertson (1998) found that hand-reared males were less able to establish territories and hand-reared females were more vulnerable to predation than their wild counterparts. In addition, it is possible that captive conditions cause long-term physiological stress on birds, lowering their fertility (Hissa et al. 1983).
28 Biological effects of release and restocking operations result from the introduction of large number of these hand-reared, low-quality birds. In this review these potential effects were grouped into four broad categories, which include the demographic impacts on the wild populations of the released/restocked species, introduction of exotic species, genetic pollution and sanitary problems. A previous review on this subject was carried out by Dowell (1992).
Impact of releases and restocking on the wild populations of target species
Detailed quantitative assessments of the effects of releases and restocking on the demography of wild gamebird populations are generally lacking. Although these operations usually result in short-term increase of population densities, it is far from clear whether they contribute significantly to the breeding population. Indeed, some studies have documented or suggested negative effects on the wild stocks (Robertson 1990, Robertson & Dowell 1990, Robertson & Hill 1992). Conversely, others have related recoveries of declining populations in association with successful restocking programs (Carvalho et al. 1998). The following factors have been suggested to contribute for a depressing effect of releases/restocking on wild breeding populations (Hudson & Rands 1988, Robertson 1990, Robertson & Dowell 1990, Schroth 1991, Robertson & Hill 1992, Hodder & Bullock 1997, Putaala & Hissa 1998):
· Competition with wild counterparts for optimum habitats. The productive potential of the best breeding habitats may be under exploited if occupied by released birds.
· Released birds increase population density, which may lead to greater density- dependent rates of mortality.
· In monogamous birds, the pairing of wild and released birds may reduce the breeding potential of the former.
· Genetic deterioration of recipient, small wild populations in large-scale releases.
· Increased prevalence of diseases and parasites.
· Overshooting of the wild population.
· Reliance on hand-rearing birds may lead to neglect of other management options benefiting the wild population. Releasing does not provide as much incentive for landowners to manage their estates in a way sympathetic to conservation in general.
Unfortunately, the evidence available to support the above suggestions is meagre and in some cases contradictory. Robertson & Dowell (1990) provided evidence that releases of grey and pure red-legged partridges had no obvious detrimental effects on the wild populations. They also suggested that releasing pheasants may have had some depressive effects on the wild stocks but this did not appear to be endangering them. In contrast, releasing chukar or hybrid partridge led to the virtual extinction of the wild stock. In general, there was no evidence that releasing birds had led to increases in the production of wild game for any species or halter their declines. Using a computer simulation approach, Robertson & Hill (1992) have shown that as the number of pheasants released increases, the productivity of the breeding population steadily declines to reach a lower equilibrium point than that attained in the absence of hand rearing. These authors concluded that there was a tendency for hand-rearing to become self- perpetuating: after the release of birds the productivity declines as does the incentive for wild
29 bird management; this in turn leads to an increased reliance on hand-rearing if the bags are to be maintained.
Other studies, though not showing any deleterious effects of releases and restocking on the wild populations, concluded that due to poor survival and reproduction output of hand-reared birds, restocking operations might have little value in attempts to boost populations of the grey (Putaala & Hissa 1998) and the red-legged partridges (Gortázar & Villafuerte 2000). This contrasts with the results by Carvalho et al. (1998), who reported increased summer densities of red-legged partridges one year after a restocking operation, suggesting that it helped increasing the breeding population. However, the restocking was accompanied with other management actions (stop of hunting, predator control, provision of grain and water points) and the relative contributions of wild and hand-reared birds to population growth were not estimated, thus making it impossible to single out the effects of restocking from that of other confounding factors.
The above studies were all carried out in farmland habitats, and there is very little data from releases/restocking carried elsewhere. However, in a study carried out in Scottish moorlands, Price (1994) found that the breeding population of wild grouse Lagopus lagopus declined during a release experiment, although there was no direct evidence that this was caused by the release.
In summary, it seems reasonable to infer that releases and restocking may be effective in helping gamebird populations to sustain high shooting pressure, though they may have at best limited value to increase the wild breeding population. Moreover, the effects may be highly variable, depending on the techniques used, the quality of the birds and other management actions taken (predator control and habitat management). It is certain, however, that releases and restocking may at least in some circumstances contribute for depressing wild gamebird populations and, through neglect, may decrease their habitat quality.
Species introductions
Gamebird introductions outside their normal geographic range are generalised worldwide and, in cases such as the ring-necked pheasant, have been carried out for a very long time (e.g. Holechek 1981). This may have the biological consequences normally associated with species introductions, including competition with and eventual replacement of native species (Hodder & Bullock 1997). Surprisingly, however, there is very little understanding of their effects on local ecosystems and biodiversity. Nevertheless, none of the dramatic negative impacts of species introductions has hitherto been attributed to gamebirds, which suggest that their ecological impacts may be subtle and difficult to detect.
The main negative effect of gamebird introductions seem to occur when a species or sub- species is introduced to the range of a closely related taxa, which may lead to the extinction or genetic deterioration of the wild stock (see details in the sections Impacts on the wild population and Genetic pollution). When birds are released in sympatry with closely related species, hybridisation is likely and this may produce unfit hybrids and depress the average fitness of the local natural populations (Randi & Bernard-Laurent 1998). Similar effects may occur when exotic sub-species or artificial hybrids are released.
These introductions are primarily caused by some species being much widely available or easier to hold in captivity than their close relatives. Examples include: Japanese quails
30 Coturnix japonica released within the range of the Common quail Coturnix coturnix in at least Portugal (Rafael et al. 1996), Spain (Rodriguez-Teijeiro et al. 1993) and France (Saint- Jalme et al. 1987).
In summary, we could find no evidence for ecosystem or community-level effects resulting from the introduction of exotic gamebird species. However, it is possible that some of these may exist, though they may be subtle and difficult to detect. Population-level effects are more likely, with some documented cases related primarily to interference with demographic processes, genetic pollution and the spread of diseases and parasites (see corresponding sections).
Genetic pollution
Often, the use of non-native or even non-local stocks in species translocations are seen as negative, because of their potential to cause the contamination of local genetic stocks, genetic erosion, or spread of unsuitable genotypes (Black 1991, Hodder & Bullock 1997). There is evidence that mixing between genetically divergent populations has the potential to introduce locally maladaptive traits (e.g. Templeton 1986). Conversely, the use of non-local individuals could be beneficial due to increases in genetic diversity of a species within an area, particularly if adverse effects of inbreeding or genetic bottlenecks can be reduced (Hodder & Bullock 1997). Most frequently, however, it is considered that the artificial mixing of populations, and in particular the restocking of local populations with birds originating from different areas, should be avoided as they contribute for mixing formerly isolated genetic profiles (Lucchini & Randi, 1998, Randi et al. 1998).
Gamebird populations frequently show a distinctive geographic structuring in genetic sub- populations, which may or may not have some kind of formal taxonomic recognition (Kark et al. 1999, Lucchini & Randi 1998, Randi et al. 1998, Liukkonen-Anttila 2001). For instance, the genetic diversity of chukar partridges (Alectoris chukar) in Israel varies significantly over short geographical distances, in association to a steep environmental gradient (Kark et al. 1999). This variation may reflect to some extent the adaptation to local environmental conditions, though evidence to support this view is still scarce (Kark et al. 1999). Other gamebirds for which there is evidence for marked phylogeographical structure, independently of taxonomic recognition, include the rock (Alectoris graeca) and the grey partridges (Liukkonen-Anttila 2001).
Despite this information on the genetic structure of gamebird populations, there is much less data on the spread of alien genes or maladaptive traits in wild populations following releases, or the loss of genetic diversity. However, some evidence suggests that genetic consequences may be prevented to some extent by the poor breeding output of the released hand-reared birds. Genetic studies of the grey-partridge in Europe using mtDNA, have documented the presence of two well-defined lineages, which originated from the separation of eastern and western populations during the Pleistocene (Liukkonen-Anttila 2001). These two lineages coexist in countries such as Finland, Sweden and Ireland, probably as a consequence of restocking operations. Nevertheless, Liukkonen-Anttila (2001) refer that “Despite the numerous releases of grey partridges, amazingly few [genetic] marks of these could be seen in Continental Europe”. One possible explanation to this pattern was attributed to the Haldane’s rule, which supports the reduced viability or even sterility of the hybrid female offspring of two different subspecies (Liukkonen-Anttila 2001). Although the Haldane’s rule mechanism may effectively prevent the spread of alien genes into established wild stocks, it may also
31 contribute to the negative impacts of restocking on the breeding populations, by decreasing productivity. This may be illustrated by the strong negative effects of releasing chukar or hybrid partridges on the populations of red-legged partridges (Robertson & Dowell 1990). A study by Randi & Bernard-Laurent (1998, 1999) on the population genetic consequences of natural hybridisation between red-legged and rock partridges supports these views. They found that the diffusion of hybrid genomes was much lower than expected on the basis of neutral diffusion models, suggesting that this could be due to natural selection against hybrid genotypes.
The above mechanisms would suggest that there should be limited spread of maladaptive traits resulting from genetic pollution. In opposition to this view, it was suggested that hybridisation between native European and released Japanese quails (sometimes regarded as sub-species), may be leading in France to widespread genetic pollution resulting in the loss of migratory behaviour (Dérégnaucourt 2000). Releases of Japanese quails were carried out in France, Spain and Portugal, in response to population declines of common quails. Cross- breeding experiments between the two quails showed that it is easy to obtain the hybrid combinations F1, F2 as well as backcrosses, and that the fertility of the hybrid pairs and the hatchability of their eggs do not differ from that of pure European quails (Dérégnaucourt, Guyomarc’h & Aebischer in press). There is evidence that domestic Japanese quails may breed under natural conditions (Nichols, Robinson & Cheng 1992), and thus hybridization may occur in the wild (Dérégnaucourt, Guyomarc’h & Aebischer in press). Domestic quails have lost their migratory behaviour, and few hybrids show migratory restleness in captivity (Dérégnaucourt 2000). From these observations, it has been speculated that the massive release of domestic quails may result in the replacement of native migratory populations by hybrid sedentary populations (Dérégnaucourt 2000). This hypothesis remains tentative, as there seems to be no hard evidence to support it; this would require a fielld evaluation of the genetic composition of quail populations in areas where large scale releases have taken place.
The loss of genetic diversity due to restocking operations is probably best illustrated by the presumed extinction of pure forms of the subspecies P.p.italica, which seems to have resulted from the intensive release of hand-reared grey partridges in Italy (Matteucci 1988, Montagna et al. 1990). This sub-species, together with P.p.hispaniensis, is listed as endangered in the Annex I of the European Union (EU) Birds Directive.
In summary, it is likely that releases/restocking may result in the mix of formerly isolated genetic units, causing the introduction of alien genes into wild, breeding populations. However, it is uncertain the extent to which these genes will actually spread, due to the low contribution of hand-reared birds to the breeding population and to the natural selection against the hybrid forms. There is still a possibility, however, that in some circumstances this genetic pollution may result in the spread of maladaptive traits and loss of genetic diversity.
Sanitary problems
The potential spread of diseases and parasites in the wild is considered an important factor that should be accounted for in species translocation programs (Viggers, Lindenmayer & Spratt 1993). This problem may be particularly serious in the case of hand-reared gamebirds, as the artificial environment of aviaries and the high stocking densities multiply the risk for the spread of parasites and infectious diseases (Beer 1998, Anon. 1994, Pennycott & Duncan 1999, Pennycott 2000). Introduction of diseases and parasites in the breeding population may increase mortality or reduce the reproductive output of the wild birds (Hudson 1986,
32 Woodburn 1995). Parasitism may increase the risk for predation, since predators may selectively prey upon birds that carry heavy parasite burdens. This effect has been reported for red grouse Lagopus lagopus scoticus (Hudson et al. 1992), pheasants Phasianus colchicus (Tompkins et al. 2000), capercaillie Tetrao urogallus, black grouse Tetrao terix and hazel grouse Bonasia bonasia (Rätti et al. 1999). In addition to posing a threat to other wild birds, infected populations of wild birds also form a reservoir of infection for domestic livestock, companion animals and even people (Hoodless et al. 1998, Pennycott et al. 1998).
The best evidence for releases/restocking interfering with wild populations through diseases is provided by research carried out in England on the parasite-mediated competition between pheasant and grey partridges (Tompkins & Hudson 1999, Tompkins et al. 1999, 2000a,b, 2001). Pheasants are maintained in the British countryside at high densities through increasing numbers of released birds (Tapper 1999), whereas the grey partridge is a declining species, probably in association with the intensification of agriculture and increased predation pressure (Potts 1986, Sotherton 1998). The pheasant is a reservoir host for the caecal worm Heterakis gallinarum, infection by which may be highly detrimental to the partridge, but not nearly as much to the pheasant. There is evidence that H. gallinarum cannot persist within partridge populations without the presence of alternative host species, and that infection from pheasants largely determine the worm burdens of partridges in the wild. Based on the best available evidence, a 2-host shared macroparasite model has shown that the observed impact of H.gallinarum may cause the exclusion of grey partridges when pheasants are present, supporting the hypothesis that apparent competition with pheasants may have played a role in grey partridge declines observed over the past 50 years (Tompkins et al. 2000b).
Another system in which gamebird releases may play a role, is the transmission dynamics of the Lyme disease spirochaete, Borrelia burgdorferi, mediated by the relationship between the tick, Ixodes ricinus, with its hosts. Pheasants in lowland woods in the UK are heavily infested with ticks and can be competent amplifying hosts for B. burdorferi sensu lato (Kurtenbach et al. 1998a). As a result, the millions of pheasants released annually to satisfy game shooting interests represent a large potential host reservoir for ticks and spirochetes, and thus they may be important in maintaining the transmission of Lyme disease (Hoodless et al. 1998). It is not known, however, whether this affects other wildlife populations, though the Lyme disease is highly deleterious to humans. The presence of large numbers of pheasants may actually reduce the potential risk of Lyme disease to people (Hoodless et al. 1998). Pheasants selectively infect nymphs feeding on them with a genospecies of spirochete (B. garinii) that causes neuroborreliosis in humans but is not maintained by mammalian hosts, and filter the genospecies (B.burgdorferi sensu stricto) maintained by mammals out of the system (Kurtenbach et al. 1998b). This ensures a high infection prevalence only in adult questing ticks, which are more conspicuous and less numerous than nymphs (Hoodless et al. 1998).
In summary, it seems that releases/restocking of hand-reared gamebirds have strong potential to spread diseases and parasites into wild populations, both because the rearing conditions are favourable to the spread of infectious diseases and because the high stocking densities maintained artificially in the wild create the conditions for them to become important host reservoirs. These parasites and infectious diseases may then be transmitted to other wild populations, eventually causing decreased fitness through reduced survival and reproductive output. Research on these subjects is still limited, but it may well be that the sanitary problems associated with releases/restocking operations are more widespread than has hitherto been found.
33 Conclusions
Overall, it seems that releases/restocking have a positive effect on harvestable population of target game species, but not necessarily on breeding populations. This is particularly evident in the case of releases with the sole objective of providing birds for shooting. In general, however, releases/restocking seem to have at best neutral effects on the receiving population, and may even contribute for depressing it. There are some exceptions, in cases where restocking is accompanied by other management actions, that together may revert population declines or to boost population recoveries.
The loss of genetic diversity may be one of the most obvious effects of releases/restocking on biodiversity. By breaking the isolation among distinctive population units, release/restocking may lead to the loss of adaptive alleles and to the spread of maladaptive traits. There is however limited information about the extent of these processes in the wild. Indeed, some studies have suggested that the spread of alien alleles may be lower than it might be expected, due to the small contribution of introduced birds to the breeding population and to the natural selection against hybrids.
The introduction of diseases and parasites may be an important mechanism through which releases/restocking can negatively affect other wildlife populations. Yet, we could find only a single case in which this effect is supported by hard scientific evidence. Effects may well be widespread, however, at least in areas where high stocking densities are maintained in the wild by the massive release of hand-reared birds. A close sanitary scrutiny of these populations should be sought.
We could find no study documenting ecosystem or community-level impacts of gamebird release/restocking. This does not prove that the effects are absent, but it suggests that these should either be subtle or hard to demonstrate. It should be particularly interesting to study how releases/restocking affects assemblages of carnivore and birds of prey, and in turn how it induces predator – gamebird conflicts and the kinds and extent of predator control.
In summary, it may be concluded that releases/restocking tend to be associated with intensive forms of hunting, and not to game management based on the sustainable exploitation of breeding populations. The exceptions are few and primarily related to the recovery of populations depressed by overshooting or low habitat quality, when associated with other management actions. In general, releases/restocking may have several negative consequences for wildlife populations and genetic diversity, though very few of the effects suggested have been supported by hard scientific evidence. Future research is needed, and it should probably be focused on ecosystem and community-level processes.
34 3.3 HABITAT MANAGEMENT
3.3.1 - INTRODUCTION
Management of habitats for the benefit of game and hunting interests is a common practice to a greater or lesser extent in all REGHAB countries (see section 2). Because habitat beneficial for game is potentially beneficial for many other species, enhancing habitat for game is potentially positive for biodiversity. The evaluation of the effect of habitat management on biodiversity is however complex. First, because the relationship between habitat and plant and fauna communities is complex. Second, because the overall effect that management may have on communities will depend on the overall extent to which the habitats are managed.
The evaluation of the relationship between habitat management for hunting and biodiversity may be assessed in two different contexts: when habitat is managed in areas where the main land use is hunting, and when habitat management is implemented in areas where the main land use is different from hunting.
As seen in section 2, the first scenario only happens (in the REGHAB countries) in certain moorland areas in Scotland and Northern England managed for red grouse shooting, and in some dehesa areas in Spain managed for red-legged partridges. In these cases, habitat is managed at large scale (at the landscape level), and management practices include a large degree of preservation and maintenance of the habitat, rather than modification of it (section 2). The influence of management on biodiversity in these cases may thus be evaluated at two different levels: first, comparing similar habitats managed and unmanaged for hunting; secondly, comparing these habitats with what would be the alternative land use if hunting would not be the main economic resource in these areas.
The second scenario concerns measures taken in e.g. farmland areas and in commercially managed forest. In such cases, the overall surface concerned by management is relatively small at the landscape level (see section 2), but potentially the effect of such management may be strong, particularly if the habitats are very degraded (as it is the case of intensive farmland in many countries).
For the purpose of the review, we concentrate in two case studies. We evaluate the relationship between management of grouse moors and biodiversity as an example of the first scenario, and the influence of habitat management for gamebirds in farmland on biodiversity. The choice of these two case studies is based on the amount of information existing (moorland habitats have been studied in further detail than Spanish dehesas, for example) and on the relevance of the habitats (farmland area is an extremely important habitat for conservation of many declining species, Tucker and Heath 1994).
35 3.3.2 - MANAGEMENT IN AREAS WHERE THE MAIN LAND USE IS HUNTING: GROUSE MOORS.
General background
Upland habitats cover some 30-35% of the British land surface and support many populations of breeding birds of international conservation importance, such as golden plover, dunlin and greenshank (Thompson et al. 1995, Tapper 1999).
Among upland habitats, heather moorland constitutes the most important habitat. One of the major uses of heather moorland is red grouse shooting. The main alternatives to grouse moor management on upland moors are intensive grazing (by sheep or deer), afforestation or a system of nature reserves (Tapper 1999).
Management techniques in grouse moorlands are directed, first, to maintain their habitat and, second, to control some of the other limiting factors in red grouse populations, such as predators and parasites (see section 2). In terms of habitat management, practices include mainly rotational burning of heather to prevent tree regeneration, and secondarily heather planting or cutting, controlled grazing and bracken removal (Hudson & Newborn 1995).
Leaving apart the economics of the issue (reviewed in workpackage 1), and the potential positive effects of predator control (reviewed in section 3.1), is the management of such ecosystem positive for other species? It has been stated that grouse moors are associated with more dunlin and golden plovers, that meadow pipits were also observed more frequently on moors where the number of grouse shot was high, and mountain hares seem to do best in areas managed for red grouse (Hudson 1992). In contrast, Thompson et al. (1997) considered that many moorland swathes are remarkably uniform and dull and with an apparently limited range of breeding birds, and consider debatable to what extent grouse moor management actually enhances nature conservation of moorlands.
Despite the attention that this habitat and the issues related to it have arose in recent years (e.g. May 1997, Mead 1997, Thirgood & Redpath 1997, Thirgood et al. 1999), relatively little research has been done to specifically assess the latter. We summarise here the results of the main studies relating the relationship between management for grouse and biodiversity. The majority of the studies evaluating the impact of management for grouse have compared distribution and/or abundance of species between areas managed or unmanaged for grouse. Because of that, it is difficult to separate the relative effect of each practice separately, or even the effect of habitat management only from other practices such as predator control.
Grouse shooting and preservation of moorland
Heather moorland has declined in recent years (Thompson & MacDonald 1995): around 20% of upland heather moorland present in the UK in the 1940s has changed under afforestation, agricultural reclamation, high grazing pressures and bracken invasion. Of that remaining, 70% is estimated to be at risk of change (Thompson & Macdonald 1995). However, the rate of loss has been lowest where heather moorland is managed for grouse shooting (Barton & Robertson 1997). In the 1940, around 54% of heather ground in Scotland was actively managed for driven grouse shooting. By the 1980’s, this figure had fallen to 26% (Barton & Robertson 1997).
36 Grouse shooting may thus be positive for the maintenance of an ecosystem. In that sense, Tapper (1999) indicated that the main benefits from grouse shooting for conservation come indirectly from the management of heather, namely the reduction of overgrazing, afforestation and invasion of bracken. No studies have specifically evaluated the relative biodiversity value of heather moorland versus forests, for example. However, heather moorland is a habitat of international importance (Thompson et al. 1995), so the cessation in grouse moor management and a consequent shift to other land uses would be negative (Thompson et al. 1988, Hudson 1992, Usher & Thompson 1993).
Comparisons between moorland managed and unmanaged for grouse.
Haworth and Thompson (1990) found that golden plover, curlew, redshank, short-eared owls, and to some extent merlins, were more frequent in upland areas managed by gamekeepers, although the significance of the association was weak compared with the influence of topography. Areas managed for red grouse and subject to intensive gamekeeping were favoured apparently because: 1) rotational burning of Calluna provides suitable nesting habitat; 2) areas of Calluna and Empetrum much favoured for nesting are largely kept free from disturbance; and 3) potential predators were controlled.
In contrast to that, Thompson et al (1997) found that in the Scottish Highlands, more bird species were more widely distributed in squares with little or no grouse moor than in squares with much grouse moor, but in Southern Scotland, England and Wales the converse was true. They suggested that species richness and diversity would increase over upland areas in the absence of burning, if scrub and woodland developed in open mosaics, but the abundance of some key moorland birds would be greatly reduced, and the actual economic cost of this whole-scale regeneration would be considerable.
Another study compared harrier hunting success in moors managed and unmanaged for grouse (Redpath et al. in press). They found a significant difference in capture rates, with harriers on managed grouse moors capturing prey at a higher rate than elsewhere, supporting the idea that prey were more available on grouse moors. However, they found no difference in strike rates between the habitats, suggesting that prey were not more abundant on grouse moors, only more available maybe due to changes in vegetation structure, prey behaviour or predator behaviour.
Tharme et al (2001) compared densities of different bird species in moorland areas managed and not managed for grouse. According to their analyses, densities of breeding golden plover and lapwing were 5 times higher and those of curlew twice as high on grouse moors as on other moors, while meadow pipit, skylark, whinchat, and carrion/hooded crow were less abundant on grouse moors. The increase in wader densities was apparently related to predator control, whereas they found evidence of a positive effect of heather burning on the density of golden plover and a negative effect on meadow pipit. Their results provide correlative evidence that moorland management benefits some breeding bird species and disbenefit others.
The same mixed effects of grouse management are observed when looking at the control of grazing levels. Sheep reduction on a moor improves heather cover and height (Hudson & Newborn 1995), and Baines (1996) found that densities and breeding success of black grouse were highest on lightly grazed moors. Low levels of grazing were associated with taller vegetation, which may provide more cover for nesting and more invertebrates for chicks. In
37 contrast, increased sheep numbers increase the proportion of grass vs heather in the landscapes. Meadow pipits and field voles prefer upland grassy heath habitats to stands of pure heather (Smith et al. 2001), and thus Hen harriers tend to have higher breeding densities on moors where there were more meadow pipits as well as field voles (Redpath & Thirgood 1999). Low grazing and reduced amount of grass are also likely to reduce the vole-dependent species, such as short-eared owl, kestrel and barn owl and weasel. Likewise, reduced number of meadow pipits could also negatively affect merlins and cuckoos (Potts 1998).
In summary, rotational burning seems to help some species, especially Golden Plover, by creating a mosaic of habitats and edges for feeding and nesting/hiding in, and by providing a high diversity of arthropods. The effect of limiting grazing levels is mixed. Low levels of grazing may benefit moors by maintaining the sward height, controlling invasion of trees, etc; high levels of grazing remove beneficial plant species and effect heather cover detrimentally, but may be beneficial for maintaining high grass levels (and the species dependant on grass areas).
Conclusions
Overall, management of moorland for grouse is beneficial for some species, and detrimental for others. The overall biodiversity value of management will thus depend on conservation priorities.
In general, management for grouse seems not to be optimal for biodiversity, but the shift to other land uses would have even more adverse consequences. In that sense, maintenance of this habitat through shooting interests is likely to be positive for overall biodiversity.
38 3.3.3. MANAGEMENT IN AREAS WHERE THE MAIN LAND USE IS NOT HUNTING: FARMLAND
In present-day, intensive agricultural landscapes there is little room for wildlife, and thus many species formerly associated with extensive farming regimes are currently suffering severe declines, particularly in North America and Europe (Freemark 1995, McLaughlin & Mineau 1995, Pain, Hill & McCracken 1997, Pain & Pienkowski 1997). In this context, gamebird management has been claimed to provide the incentive and the economic basis for landowners, farmers and estate managers to manage land under their stewardship in a way more sympathetic to nature conservation in general (e.g. Piddington 1980, Sotherton, Boatman & Robertson 1992, Robertson 1992). This claim is based on the idea that management actions taken for gamebirds are also valuable for other key groups of organisms, which in turn are generally regarded as surrogates for the general health of the agro- ecosystem (e.g. wildflowers, butterflies and songbirds).
Habitat management for game birds in farmland habitats has developed over the past few decades, primarily as a response to the evolution of farming regimes. In Europe, two broad trends in agricultural change can generally be recognised, which started in the early 1950s in northern and central parts, and have progressed to the Mediterranean region in more recent times (Potter 1997). In the one hand, farming practices have intensified strongly in the regions with the most productive soils and under the most favourable climatic conditions, involving increased use of pesticides and chemical fertilisers, increased field sizes, improved drainage, reduced crop diversity associated with the cropping specialization at the farm and regional scales, simplified crop rotations, loss of marginal habitats such as ponds and hedgerows, and decline in favourable practices such as undersown leys, among other transformations. On the other hand, however, agriculture has been abandoned in marginal farming areas, as for instance in the mountains and in little productive soils in the Mediterranean. Afforestation, usually with exotic species, is the most common fate of the abandoned agricultural fields. Indeed the conversion of arable land to forestry in marginal farming areas has been actively supported with European Union financial incentives. As a general consequence of these processes, there has been a large-scale loss of low-intensity farming systems, which are one of the most valuable wildlife habitats in Europe, including for game birds.
The assessment of factors causing the declines on one game bird, the grey partridge (Perdix perdix), provided an early evidence for the strong negative effects of intensification on farmland wildlife. A long-term study in the UK revealed a nationwide decline in partridge populations from the early 1950s onwards, in association with the intensification of grain production; declining populations have also been recorded in North America and other European countries (Potts 1986). The decline of the grey partridge was linked with the increasingly poor levels of chick survival, which were shown to have resulted from the low availability of insects that are essential in the diet of young chicks (Potts 1986). The use of various types of pesticides was the primary cause for the reduction in numbers of these non- target species (Potts 1986.). These results, together with others showing that intensive farmland provide hostile habitats for gamebirds, have provided the reasoning behind most game management approaches for this habitat. The main idea has been to find practically oriented and costed management options whereby farmers can continue to farm profitably and maintain high levels of production but at the same time overcome some of the adverse effects of intensive farming (e.g. Sotherton 1991). Given that agricultural intensification is also responsible for populations declines of many bird species in farmland (Pain & Pienkowski 1997), management options aimed to improve habitat for game have been claimed to favour
39 farmland wildlife in general, and not only the game species (Sotherton, Boatman & Robertson 1992, Robertson 1992).
Game management in intensive farmland habitats has adapted to changing agricultural policies, namely to the incorporation of environmental considerations into the Common Agricultural Policy (CAP). A particularly important development was the introduction of the set-aside scheme in 1988 under the CAP, aimed at reducing surplus crop production in the EU, whereby portions of land that had been in arable production were taken out of cultivation (Floyd 1992). A comparable system had been going on in the US for at least 50 years (Ervin 1992). The set-aside scheme has evolved over the time from a pure production control mechanism of the CAP to become also part of the EU environmental policy (Robson 1997). However, this is a complicated system involving several components, including for instance the possibility to grow industrial and energy crops on set –aside, which may have no value to farmland biodiversity. The rules concerning the extent and management of set-aside land have changed every year since its introduction, and so its value to wildlife. Nevertheless, the set- aside scheme has been considered an opportunity to improve habitat quality for game birds in farmland, and thus a number of management approaches have been put forward (Sotherton, Boatman & Robertson 1992, Anon. 1995a, 1996, Genghini & Avoni 1997, Peeters & Decamps 1998). Likewise, a wide range of management prescriptions has been suggested to increase the value of set-asides to declining farmland wildlife in general (Buckingham et al. 1999, Henderson, Vickery & Fuller 2000).
In extensive farmland, the problems of game management are different from that in intensive farming areas. Hunting in extensive farmland has been viewed as a valuable source of income to farmers, which may help sustaining the farming activity itself (Carvalho et al. 1995). Therefore, the problem in these systems is not as much to restore the minimum habitat requirements for the game species as in intensive farmland, but instead to maximise the bags given the constraints of extensive crop production. In these areas, another source of funding for conservation is the agro-environment scheme of the European Union, whereby subsidies to extensive agricultural practices are granted to support the conservation of threatened species and habitats. Agro-environmental and hunting incomes may add up to help supporting a rural economy based on wildlife resources, but they may also compete with each other if hunting activities are viewed as negative to some desired conservation goal.
In both intensive and extensive farmland, there are many approaches for improving habitat conditions for game birds, depending on many factors, particularly the target species and the agricultural setting. Nevertheless, it may easily be recognised that management options usually fall in some of the following categories: a) maximising habitat conditions of field boundary structures such hedgerows, grass banks and ditches adjacent to the cultivated fields; b) maximising habitat conditions of the crop itself, with a strong focus on the crop edge and the creation of crops specifically targeted at gamebirds (game crops); c) management of fallow land and set-asides to enhance their capacity to provide cover and feeding opportunities; d) managing small patches of woodland within the agricultural matrix, essentially to provide cover for species like the pheasant; e) artificial supplementation of critical resources, such as grain and water; e) managing the habitats to decrease the carrying capacity to gamebird predators, and thus reducing predation rates.
In this review we analyse each of the management options usually adopted for gamebirds in farmed landscapes, trying to frame these in the context of current farming techniques and agricultural policies. Then, we used published studies to assess the impacts of each
40 management prescription on the target game species and on other wildlife. Our main aim is to evaluate whether the evidence available at present supports the claim that management of habitats for gamebird hunting has overall positive effects on farmland biodiversity.
Field boundaries
Hedgerows
Rows of trees or shrubs are common elements in many farmed landscapes, with contrasting roles recognised by different people e.g. physical property limits, shelter, sources of products or windbreaks (Burel 1996, Baudry et al. 2000). Hedgerows comprise one of the most important surviving elements of semi-natural habitat in lowland intensive farming landscapes, providing shelter, breeding and foraging habitats to game species (Rands 1986, 1987, Ricci et al. 1990, Aubineau et al. 1998) and other wildlife (Dover & Sparks 2000, Hinsley & Bellamy 2000). However, these structures may also affect negatively some species, such as open- country birds that avoid edges e.g. lapwing Vanellus vanellus and skylarks Alauda arvensis in lowland farmland in Britain (Hinsley & Bellamy 2000). For some species hedgerows may be movement corridors, mitigating the effect of isolation on populations in fragmented landscapes, but for others they may act as barriers, dividing potentially continuous populations into local populations at the field scale (Dover & Fry 2001, Thomas et al. 2001). Nevertheless, the net effects of hedgerows on biodiversity are usually regarded as positive. Game management encourages the creation of new hedgerows and their management to the benefit of gamebird populations. Seemingly, a wide range of hedgerow management prescriptions has been targeted at other wildlife groups, including for instance plants (Hegarty, McAdam & Cooper 1994, Maudsley et al. 2000), birds (Hinsley & Bellamy 2000), and invertebrates (Dover & Sparks 2000, Maudsley 2000, Moreby & Southway 2001). Therefore, game management will have positive effects on biodiversity if it contributes to the maintenance or creation of hedgerows in the countryside, and if hedgerow management for game is at least as favourable to biodiversity in general than neglect or other commonly adopted practices.
Hedgerow management can be considered under three broad headings: the management of the shrubby/woody elements of the hedge, the management of the basal vegetation on which the head stands, and the management of the immediately adjacent land (e.g. Dover & Sparks 2000). The management of these components affect their usefulness to wildlife. Hedges that are not managed for long times grow tall, thin at the base and develop gaps, and are considered less valuable for biological diversity. In the UK, the Game Conservancy Trust recommends short (approximately 2 m high) hedges and hedges with few trees (Rands & Sotherton 1987, Anon. 1995a). The British Association for Shooting and Conservation (BASC undated) also suggest that many owners will prefer shorter hedges, to reduce effort for trimming and also for safety during a shoot. To encourage berries, trimming should be carried out every third year, or every other year if cutting can be postponed until the months of January/February (Anon. 1995a). The hedge should not be allowed to grow over the adjacent grassy strip, which is where nesting takes place. Therefore, it is usually advised that the sides of the game hedges should be trimmed vertically, whereas trimming to an “A” shape should be avoided, unless a sufficiently wide grassy strip can be maintained beside the hedge. Wide verges and good vegetation cover are also characteristics identified as highly beneficial to game birds (Rands 1986, 1987, Sotherton & Rands 1987, Aebischer et al. 1994)
41 In a lowland Britain, there is a tendency for game areas to have a greater length of hedgerow than non-game areas in both arable and pastural landscapes, but not to a statistically significant extent (Stark et al. 1999). Non-game areas include a greater degree of variation in hedgerow lengths, with more areas with no hedgerow, but some with several km more hedgerow than comparable game areas (Stark et al. 1999). Hinsley et al. (1999) found that in estates managed for game in lowland Britain, hedgerows were in average shorter and narrower than in non-game areas, but with wider verges and more cover in the hedge bottom. This is consistent with the usual recommendations for the management of game hedges (e.g. Anon. 1995a). These authors found a consistent trend, though not statistically significant, for higher bird densities in game than in non-game hedges. This result is inconsistent with studies suggesting that tall and wide hedges are the most favourable for bird diversity and abundance (Parish et al. 1994, Macdonald & Johson 1995, Sparks, Parish & Hinsley 1996). The discrepancy may be because the benefits of wide verges and good vegetation cover of the game hedges served to offset any detriment arising from the lack of height and width in the hedges surveyed by Hinsley et al. (1999); these hedges were generally tall irrespective of game management, and it is known that once above a critical height of about 2-3 m, further increases in height may not significantly affect birds (Macdonald & Johson 1995). It should be noted, however, that while hedges managed for game should be less than 2 m high (Rands & Sotherton 1987), there is evidence that below hedgerow heights of 2 m, bird diversity seems very sensitive to the loss of height (Macdonald & Johson 1995).
In the same study, Hinsley et al. (1999) found a trend for butterfly species numbers to be slightly higher on game hedges, though again not to a statistically significant level. In contrast to the bird and butterfly results, Hinsley et al. (1999) found a non-significant trend for higher numbers and densities of hedge, hedge bottom and nectar plant species on non-game sites. In relation to non-game hedges, game vegetation was dominated by a smaller number of, individually more abundant species, having fewer uncommon species. The marginally more favourable management of hedges for plants in non-game sites may be related to hedge width (Hinsley et al. 1999), as several studies have associated greater plant species diversity in hedge bottom vegetation with wider hedges, though this relationship may be influenced by many other factors (e.g. Hegarty & Cooper 1994, Hegarty et al. 1994).
In summary, it seems clear that hedgerows are a primary habitat for farmland biodiversity, and thus game management may promote their conservation by contributing to retain hedges in game areas. On the other hand, however, management of game hedges may not be more favourable to wildlife than some alternative options, especially in what regards the recommendation for relatively short and narrow hedges, with few mature trees. In the few studies available, game hedges appeared marginally less favourable for wildlife, though the effects are inconsistent for different taxonomic groups and they are confounded by a number of additional factors. However, all these studies were made in Britain, thus different patterns may be apparent elsewhere.
Herbaceous field margins
Herbaceous strips around arable fields are among the major non-cropped farmland habitats in intensive farmed areas, and it is increasingly recognised that they have potential value for a variety of wildlife (Morris & Webb 1987, Hooper 1987, Moreby & Southway 2001). Besides their significance to biodiversity, herbaceous field margins may have important functions in their interactions with agricultural systems, both by providing reservoirs of pollinating insects (Lagerlöf et al. 1992) and of predatory and parasite arthropods which may enhance the
42 biological control of crop pests (Dennis & Fry 1992). These structures are of particular importance in providing overwintering refuges for many species of polyphagous invertebrate predators in arable field systems (Desender 1982, Sotherton 1984, 1985, Wallin 1985, 1986), which then disperse into the crops in the following spring (Wallin 1985, Coombes & Sotherton 1986, Lee, Menaled & Landis 2001). Increasing natural predation of pests potentially reduces the need for pesticide inputs and decreases their undesirable side-effects on non-target organisms.
The values of herbaceous field margins are strongly dependent upon the type and extent of management (e.g. Harwood et al. 1994). It is apparent that farming operations, such as fertiliser and pesticide application can have adverse ecological effects (Kleijn & Snoeijing, Longley et al. 1998, Tsiouris & Marshall 1998). Close ploughing reduces strip width and leaves the remaining area more vulnerable to agrochemical drift. Mechanical disturbance makes the herbaceous strips favourable for annual weeds at the expense of perennial species; competitive species will take advantage of high soil nutrient levels and suppress the slower growing species (Berendse et al. 1992, Boatman et a. 1994). Weed problems, which result from the accidental application of fertilizer, are frequently managed with broad-spectrum herbicides (Marshall & Smith 1987, Boatman 1989). These impacts on plant assemblages are then transferred across the food chain, affecting for instance various invertebrates that provide crucial feeding resources to both game (Hill 1985, Potts 1986) and non-game birds (Wilson et al. 1999, Moreby & Stoate 2001). Like for hedgerows, game management will have positive effects on biodiversity if it encourages the creation and maintenance of herbaceous field margins, and if game management actions are generally favourable for other wildlife.
Management of herbaceous strips for game birds and other wildlife are targeted at reducing the negative impacts of farming operations, while improving certain qualities that make the habitat particularly favourable for the target organisms. In the case of game birds, the management is targeted at improving the use of this area as nest sites and for over wintering of beneficial insects (Anon. 1995a), which in turn allows invertebrates to be available as chickfood items. Herbaceous game strips should preferably be at least 1m wide and sited on a bank, with vegetation composed of perennial, non-invasive grasses and herbs, preferably including tussock-forming species. Regular annual cutting of the grass strip should be avoided, but rotational trimming every 2-3 years may be necessary to prevent scrub- encroachment (Anon. 1995a). The grass should be managed so that there is always old dead grass from the previous year available for nesting. Non-selective herbicides should never be used in this area, and drift from the adjacent crops should be avoided. Livestock should be kept away from the herbaceous field margin.
Removal of herbaceous field margins has strong negative effects on a wide range of farmland species, including plants, invertebrates and vertebrates (Wilson et al. 1999). Therefore, management for gamebird hunting may have potential benefits for biodiversity by promoting the retention or creation of these boundary structures in intensive farmland. However, to show that game management do indeed have realised benefits, it would be necessary to evaluate whether there are indeed more grassy margins in game than in non-game areas. Recognising that a number of agro-environment schemes across Europe now provide incentives to the management of field margin habitats, to assess the realised benefits of game management would also require the comparison between the value to biodiversity of grassy margins managed for gamebirds and for other wildlife. To the best of our knowledge, there seems to be no published study specifically addressing these questions. Nevertheless, it may be argued that some of the general prescriptions recommended for gamebirds, such as the exclusion of
43 agrochemicals, may have general benefits to other wildlife. Rotational cutting also benefits several invertebrate groups, by preventing scrub encroachment and thus favouring grassland species (Baines et al. 1998). Tussock grass, which is recommended for gamebirds, seems also to be more favourable to general arthropod diversity than non-tussock grass (Dennis & Fry 1992). Most likely, the needs of game may easily be made compatible with that of other wildlife, providing grassy field margins that are favourable to biodiversity in general.
In summary, it seems clear that herbaceous field margins are valuable habitats for wildlife, particularly in intensive farmland, and that management for gamebird hunting has the potential to promote the conservation of such boundary structures. Therefore, it may have potential benefits for biodiversity. Nevertheless, evidence is lacking to support the view that a larger extent of grassy field margins occur in game than in non-game areas, and that management of these structures for gamebird hunting are more beneficial for biodiversity than competing management regimes. Nevertheless, the information currently available makes it likely that some management prescriptions for game and for other wildlife may be made compatible, with a general positive impact on farmland biodiversity.
Sterile strips
Sterile strips about 1m wide may be maintained between the crop and the adjacent field boundaries or other vegetation, so as to prevent aggressive arable weeds spreading into the crop. This strip is used together with other management actions, such as “Wildlife Strips” and “Conservation headlands”. There is no evidence that sterile strips produce any direct benefit for gamebirds or other wildlife (Anon. 1995a). However, sterile strips may provide a drying out/dusting area for partridge broods.
Crops
Crop edges
Management of crop edges for game birds has been based largely on the concept of “conservation headlands” created by the Game Conservancy Trust in the UK (Sotherton et al. 1989). This management strategy was developed following recognition that the use of various types of pesticides was one of the most important factors with detrimental effects on game birds in farmland habitats (Potts 1986). To address this problem and thus allowing wild gamebird production in intensively farmed arable land, the Game Conservancy developed and tested extensively the concept of selectively sprayed cereal crop margins or “conservation headlands” (Boatman 1990, Sotherton 1991, Sotherton et al. 1993). In this system, the outermost section of the spray boom (in most cases the outermost 6 metres) is either switched off when spraying around these headlands to avoid certain chemicals at crucial times of the year or the headlands are sprayed separately with more selective compounds. Subsequent weed growth in the cereals is tolerated to encourage phytophagous insects (leef beetles, weevils, plant bugs, sawfly larvae and caterpillars) that provide food for gamebird chicks during the spring. Conservation headlands remain the most widely used system to control pesticide application for the sake of gamebird populations, and it has been used in several countries, at least at the experimental level (Van der Meulen, Snoo & Wossink 1996, Kendall et al. 1998, Chiverton 1999). Recently, there have been attempts to expand the ‘conservation headland’ concept to grassland systems (Haysom et al. 1999, 2000).
44 Field-scale experiments have demonstrated the value of Conservation Headlands to gamebirds, by recording up to threefold increase in densities of chick-food insects with consequent increases in the gamebird chick survival (Rands 1985, Sotherton, Rands & Moreby 1985, Sotherton, Boatman & Rands 1989). In other studies, the use of Conservation Headlands has been shown to increase the abundance and/or diversity of spiders (Hassall et al. 1992), butterflies (Dover, Sothertonn & Gobbett 1990), hover-flies (Cowgil 1990), heteroptera (Hassall et al. 1992), carabids (Sotherton et al. 1985, Hassall et al. 1992), small mammals (Tew 1988) and rare arable weeds (Sotherton et al. 1989, Wilson 1994). These insecticide-free areas may also be valuable refugia for arthropod predatory and parasitoid species (Chiverton & Sotherton 1991). The mechanisms behind these effects have also been analysed in some cases. For instance, several studies (Rands & Sotherton 1986, Dover 1989, 1997, Dover, Sotherton & Gobett 1990) have shown that butterflies exploit both the nectar resources and host plants present in Conservation Headlands, and that they were particularly useful in providing nectar resources for butterflies emerging in the spring when nectar plants in the adjacent hedgerows were in low density. These changes in spray regimes were considered to result in increased butterfly populations (Dover, Sotherton & Gobett 1990). However, Dover (1997) noted that Conservation Headlands slowed down butterfly movement along hedges and might reduce the effectiveness of hedgerows as corridors by providing distracting resources.
The benefits of reduced application of pesticides in conservation headlands may extend well beyond the crop edge, by reducing the drift of herbicides and insecticides applied to the crop. Cuthbertson (1988) and Cuthberson & Jepson (1988) demonstrated that moving spray booms 6 m away from the crop edge, as is the case in Conservation Headlands, reduced deposition into hedge bases by over 50% in the autumn and, with the additional shelter of a mature crop, by 75% in the summer. Kleijn & Snoeijing (1997) documented declines in species richness and shits in community structure of plants in field boundaries affected by levels of herbicide, similar to that in drift from adjacent crops. Longley et al. (1997) found that mortality of caterpillars was higher when field boundaries were not buffered from spray drift by a strip of unsprayed crop. Therefore, Conservation Headlands should be beneficial to biodiversity associated with hedgerows and grassy field margins, by preventing the highly negative effects of agrochemical drift from the crops.
Despite their seemingly strong benefits for gamebirds and other wildlife, we could find little published information on the actual use of this management practice in the wider countryside. Furthermore, it is poorly known whether the use of conservation headlands is actually targeted at gamebirds or at other wildlife. In questionnaires answered by land managers in lowland Britain, unsprayed crop edges to the benefit of game birds were recorded in only one game site out of twelve; conversely, the use of conservation headlands was reported by land managers in five out of seven non-game sites, either to the general benefit of wildlife or other reasons (Hinsley 1999, Table 4.2). This surprising result, considering that conservation headlands were initially designed to help gamebird populations, may result from the progressive incorporation of this management technique in general EU agri-environment schemes or other general programmes for biodiversity conservation (e.g. UK Biodiversity Action Plan for Cereal Field Margins – Anon. 1995b; UK pilot Arable Stewardship scheme – MAFF 1998; reduction of pesticides campaign by the Swedish agricultural authorities – Chiverton 1999). Furthermore, this may also be explained by shooting practices in lowland Britain increasingly associated with pheasant releases, and less with the management of breeding gamebird stocks.
45 In summary, conservation headlands are probably the best example of a management prescription for gamebirds that has widely beneficial implications for the overall biodiversity. Indeed, there is strong evidence that the reduction of pesticide applications in the crop edge needed to increase the food supply for gamebird chicks also increases diversity or abundance of many other organisms. Despite their value, however, the scanty evidence available suggests that Conservation Headlands are little used in gamebird management in the countryside; conversely, this technique is increasingly promoted and used in the realm of general wildlife conservation programmes.
Beetle banks
The concept of “Beetle Banks” was designed primarily to enhance populations of polyphagous invertebrate predators in arable field systems, to help control aphid pests in the adjacent crop (Anon 1995a). Beneficial insects over-winter in beetle banks as they would do in hedge banks, and then invade the crop in the following spring to eat pest species such as aphids. A beetle bank is a raised ridge across the middle of an arable field and planted with tussock-forming grasses like cock's-foot (Dactylis glomerata L.), perennial rye-grass (Lolium perenne L.), creeping bent (Agrostis stolonifera L.) and Yorkshire fog (Holcus lanatus). It plays a role complementary to that of the field boundary structures (see section Herbaceous field margins), which is essential in the large fields characteristic of modern farming systems. With the small boundary:field ratio of large fields, the final density in the crop of predators originating from the non-cropped boundary areas may be insufficient to influence pest numbers (Thomas, Wratten & Sotherton 1991). Therefore, by providing a within-field over wintering refuge, beetle banks may allow predators with low rates of dispersal to reach the field centre earlier in the spring than they would do otherwise (Wratten et al. 1984).
Beetle banks are also valuable nesting and brooding habitats for gamebirds, and thus their creation has been recommended as a useful management practice (Anon. 1995a, Thomas, Goulson & Holland 2001). The management prescriptions for beetle banks, however, are not specifically targeted at gamebirds, though they may be valuable components within a range of game management techniques on the farm, as a “spin-off” to their primary role as a overwintering habitat for polyphagous predators (Thomas, Goulson & Holland 2001). Nevertheless, beetle banks can contribute useful, albeit lower, densities of gamebird chick- food than conventional margins; these resources are more abunndant later in the season, which may have implications for early hatched chicks (Thomas, Goulson & Holland 2001). Beetle banks also provide considerable quantities of nesting cover for adults, although sheltering conditions may never be as satisfactory as in well managed hedgerows.
There is abundant evidence on the value of beetle banks for predatory insects and spiders, as they provide a nucleus predator population at the field centres, which would otherwise be absent, from which emigration could take place (Chiverton 1989, Thomas, Wratten & Sotherton 1991, 1992, Sotherton 1995, Collins et al. 1996, 1997). Lee, Menalled & Landis (2001) demonstrated that the presence of within-field refuges might buffer the negative consequences of pesticide application on carabids in adjacent fields.
Despite their potential value for gamebirds, there is no evidence that beetle banks are used more extensively in game than in non-game areas, and indeed the opposite may occur (Hinsley 1999, Table 4.4).
46 In summary, beetle banks seem to be a valuable management prescription for gamebirds, though they have not been designed originally with this purpose. Beetle banks are apparently valuable for the overall farmland biodiversity, not only because they provide suitable habitats for a range of species, but also because they may allow reductions in the use of pesticides in crop protection. Nevertheless, there is no evidence that beetle banks are used more extensively in areas managed for gamebirds than elsewhere, and thus the benefits of this technique cannot be ascribed to management for gamebird hunting.
Undersown leys
In traditional mixed farming, spring-sown cereals may be used as a nurse for grasses and clovers so that after the cereal harvest a ley is established without further cultivation. This system was once used extensively in Europe, but is now restricted to a few traditional mixed farms (Potts 1997). Undersowing has been recommended to help partridges as an alternative to re-seeded grassland (Anon 1979). This management prescription seems to have large potential benefits to biodiversity, but it has not been applied widely in the game management context, probably because it causes reductions in cereal yields (Potts 1997). However, in Denmark a scheme was introduced under the agri-environment measures to help restore undersowing in the country (Potts 1997).
In summary, the practice of undersown leys seems to be beneficial for both gamebirds and the overall biodiversity, but it has been used little in management for gamebird hunting. Some efforts have been made, however, to restore this practice under the agri-environment schemes.
Game crops
Game crops have traditionally been grown to provide cover and food for gamebird populations during critical seasons of the year, particularly in winter and in the brood-rearing season (CTGREF 1975, Anon. 1986, Reino, Borralho & Bugalho 2000). Herein we consider game crops as any culture established primarily to provide cover for gamebirds or to their nests (protection against predators or unfavourable climatic conditions), to provide seeds or green vegetation to be consumed by the birds, or to increase invertebrate populations, which are crucial chick-food. In addition, crops may be established to help in the management of shooting days (Anon. 1986). Whatever the objectives to be fulfilled, a wide range of game crop management techniques have been developed over the year (CTGREF 1975, Anon. 1986, Reino, Borralho & Bugalho 2000).
Game crops are widely used in areas managed for gamebird hunting. For instance, Hinsley (1999) reported that game crops were used regularly by the vast majority of land managers in game areas in lowland UK. Game crops are usually established in small blocks or narrow strips (0.2 – 1.0 ha), which are sown with mixtures of seeds that are considered to be the most attractive to game. For instance, mixtures recommended in the UK by the Game Conservancy Council include e.g.: maize (Zea mays) /millet (Panicum effusum), Spring barley (Hordeum vulgare) / Spring tic beans / Kale (Brassica oleraceae acephala), Kale / Maize mixture, Kale / Millet / Maize (Anon. 1986). In the US, mixtures of maize and sorghum (Sorghum sp.) are frequently used to enhance overwinter survival of pheasants (Bogenschutz, Hubbard & Leif 1995)
A number of studies have been carried out to assess the actual effects of game crops on gamebirds. Most these studies have concluded that winter game crops have measurable effects
47 on gamebirds, namely contributing to improved body condition (Bogenschutz, Hubbard & Leif 1995, Robel & Kemp 1997) and to reduced mortality (Robel & Kemp 1997). The limited evidence available suggests that the creation of game crops may have overall positive value for farmland biodiversity. For instance, Dover (1988) and Dover et al. (1992) found game cover crops (Jerusalem artichoke, Helianthus tuberosus L. and Canary grass, Phalaris tuberosa L.) to be particularly good for butterflies. Hinsley et al. (1999) also recorded concentrations of butterfly activity on game crops.
In summary, it is likely that game crops contribute positively to the overall farmland biodiversity, though strong evidence to support this view is still limited. This management technique is used extensively, at least in some countries, suggesting that these positive effects may be widespread.
Fallows and set-asides
In areas of extensive agriculture, the crops are frequently grown on a rotational basis, with portions of land left fallow for one or more years to recover soil fertility after a period of cultivation. The agricultural mosaic that is created under this regime seems to be important to maintain high levels of farmland biodiversity (Suárez, Naveso & De Juana 1997, Delgado & Moreira 2000). In modern farming systems, nitrogen fertilizers are used to maintain soil fertility, and thus there is little reason to keep the rotational cropping system. For this reason, rotations have become simpler or even disappeared to be replaced, for example, by continuous cereals. In recent years, however, changes to the European agricultural policy have caused the emergence of new areas of fallow land, established under the set-aside scheme of the European Union. In this review we have considered set-asides and fallow land in extensive farming regimes under the same general heading, as they share similarities in the kind of management actions that may be taken to the benefit of gamebirds and other wildlife.
Set-asides were introduced in 1988, initially as a voluntary scheme, to reduce surpluses of arable crops within the European Community (EC) (Floyd 1992). Reform of the Common Agricultural Policy in 1992 made pricing support conditional on arable farmers setting aside a given proportion of their land that had been in arable production. Although the rules concerning management of set-aside land have evolved over the time, there are two basic types: non-rotational set-aside (NRSA), equivalent to long-term fallows, whereby land is taken out of production on a long-term basis; rotational set-aside (RSA), equivalent to short- term fallows, whereby land is taken out for one year only. From the beginning, it became apparent that both these types of set-aside could offer opportunities for the conservation of gamebirds and other wildlife, provided properly managed and targeted prescriptions were made available (Sears 1992, Sotherton, Boatman & Robertson 1992, Wilson & Fuller 1992). However, at the outset set-asides were managed by neglect, resulting in criticisms by farmers and conservationists alike (Sotherton, Boatman & Robertson 1992). In the mid-1990s, changes were introduced to this scheme that created the potential to enhance its value to farmland biodiversity (e.g. the “Wild Bird Cover” option in the UK or the “Jachères Environment et Faune Sauvage” in France): (1) the ability to control vegetation in non- rotational set-asides by using non-residual herbicides; (2) the possibility in non-rotational set- asides to plant specifically designed seed mixtures to create valuable cover for birds and the possibility to plant such mixtures in relatively small strips and plots distributed strategically around the farm, thus avoiding large blocks of set-aside (Havet & Granval 1996, Sotherton 1998). Profiting from the various possibilities of the set-aside scheme, numerous management prescriptions for gamebirds and other wildlife have been devised in the past decade (e.g.
48 Anon. 1996, Genghini & Avoni 1997, Peeters & Decamps 1998). The set-aside scheme will continue in Europe at least until 2006, as part of the “Agenda 2000” reform of the CAP.
In the case of gamebirds, set-asides can provide winter cover for holding and showing birds, nesting cover for hens in the spring and brooding cover for foraging chicks when they have left the nest (Sotherton, Boatman & Robertson 1992). Therefore, a mixture of set-aside management options is likely to fulfil the requirements for these types of cover. For instance, winter and nesting cover may be provided by NRSA, whereas either RSA or NRSA may provide brood rearing areas for young chicks (Sotherton, Boatman & Robertson 1992). The value of these habitats for game may depend on the type of vegetation structure that is maintained on set-aside land, and how it changes across the annual cycle. Management neglect or inadequate prescriptions may have negative impacts on gamebirds (Sotherton 1998). For instance, cutting of RSA at the inadequate time of the year may result in wide- scale destruction of partridge and pheasant nests (Poulsen & Sotherton 1992). Management of set-aside fields should be viewed at the farm scale, with small blocks of set-aside scattered across the farm providing habitat for more individuals than allocating the entire set-aside in two or three fields (Sotherton 1998). The kind of management of each set-aside block should then be designed according to the landscape context and the target species.
In RSA, natural regeneration with stubbles left overwinter is recommended to provide winter feeding and holding cover, a chance for chick-food insects to emerge in spring and brood- rearing habitats for chicks (Anon. 1996). There are additional prescriptions to provide winter food, such as sowing quick-growing cover crops (e.g. mustard, Brassica alba) including seed- rich mixtures (Anon. 1996, Buckingham et al. 1999). The available evidence suggests that gamebirds do find suitable winter feeding conditions in RSA. For instance, Moreby & Aebischer (1992) and Moreby & Sotherton (1995) examined the abundance of chick-food insects on RSA compared to conventional cereals, finding three times more insects sampled on the former than on the later habitats. Probably associated with the increased food resources in winter, RSA seems to contribute to higher gamebird densities, survival and breeding success (Sotherton et al. 1994, 1998). The stubbles left in RSA seem also very favourable for a range of bird species, by providing critical food resources to seed-eating passerines (Wilson, Taylor & Muirhead 1996, Buckingham et al. 1999). The value of RSA as nesting and brood- rearing habitats is more variable. Vegetation cover in RSA is frequently destroyed for cultivation in spring, precluding its use as nesting habitat. (Sotherton et al. 1994). Early cutting also turns RSA of limited value for a number of invertebrates and rare arable wildflower (Sotherton 1998). Frequently, vegetation in RSA is too dense at the base to allow gamebird chicks to forage, especially in wet weather, making it unsuitable for young foraging broods (Sotherton 1998). Despite these potential problems, RSA have been shown to support higher densities and more species of birds during the breeding season than in adjacent arable crops, probably reflecting greater food abundance in the former (Henderson, Vickery & Fuller 2000). Berg & Pärt (1994) have found that several bird species were more abundant in forest edges adjacent to young set-asides than in edges adjacent to cereal fields. A study recorded lower densities and diversity of small mammals on one-year set-aside than in adjoining hedgerow and cereal crops, demonstrating that set-aside is not necessarily a suitable habitat for small mammals (Tattersall et al. 1997).
The later into the year the cover in RSA can be left undisturbed, the more benefit it has for game (Anon. 1996). However, this may disrupt farming operation, by giving problem weeds a chance to set seed. To work out this problem, there is a range of possibilities involving the use of selective, partially selective or broad-spectrum contact herbicides to control the invasion by
49 problem weeds (Anon. 1996). The latter option, however, will impair the value of RSA as brood-rearing cover and wildlflower habitat, though nests, sitting hens, clutches and small chicks will not be harmed, and some beneficial insects can still emerge successfully (Sotherton 1998). Cutting the sward early before the nesting season begins is another recommended method of weed control (Anon. 1996). In this case, a short sward should be maintained, and cultivation should be made as late as possible (Anon. 1996).
Management of NRSA for gamebirds may involve natural regeneration in the first year, and then sowing of unharvestable seed mixtures to develop a suitable vegetation cover (Anon. 1996). These include, for instance, mixtures of two crop groups other than legumes (the Wild Bird Cover option in the UK), which can be used to create both winter and brood-rearing cover. Mixtures based on kale seem to be particularly adequate to develop winter cover in set- aside fields. Seed-bearing species such as quinoa (Chenopodium quinoa), millet (Panicum effusum), buckwheat (Fagopyrum esculentum) and sunflower (Helianthus annuus) can also be used, providing valuable food sources for passerines. There are several options available to create a suitable brood-rearing cover, depending on the agronomic characteristics of the site (Anon. 1996, Peeters & Decamps 1998). For instance, mixtures based primarily on cereals can be used to create insect-rich brood cover. Triticale (Triticum ´ Secale) and oats (Avena sativa) may be particularly appropriate to this purpose, because of limitations on fertiliser and herbicide use on set-asides. Mustard and red clover (Trifolium repens) may be worthwhile components of the mixture. As grass/legume mixtures are frequently too thick, movements of feeding chicks may be made easier by the creation of corridors that meander the plot, using a range of available techniques (Peeters & Decamps 1998). Set-asides sown with brood-rearing cover may provide refuge for rare arable wildflowers, especially if specific areas such as the crop edge are designated for their conservation and managed accordingly (Sotherton 1998). Sown cover may also be preferable where suppression of agricultural weeds is a priority (Critchley & Fowbert 2000). However, natural regeneration seems to provide better opportunities for the conservation of annual arable plants, but only in the first 2-3 years and with site management specifically tailored to their requirements (Firbank & Wilson 1995, Critchley & Fowbert 2000). On the contrary, botanical diversity on NRSA begins to increase after five years, and habitats resembling permanent semi-natural grassland may develop in the long-term (Critchley & Fowbert 2000). Henderson, Vickery & Fuller (2000) have shown that NRSA attracted higher densities of birds and a greater number of species than adjacent arable crops.
Another possibility to NRSA is the creation of field margin set-asides next to existing boundaries, especially hedges and woodland edges (Anon. 1996). Up to 20 m next to the boundary may be left uncut to allow broadening of the hedgerow or natural regeneration of woodland edge, thus creating cover for partridges, pheasants and other ground-nesting birds (Sotherton 1998). Set-aside strips can also be placed across the centres of large fields. In both cases cover can be established by natural regeneration or (preferably) by sowing grass mixtures. Like in field margins (see section Herbaceous field margins), gamebird nests are more likely to be successful if a bank is created first to aid drainage (Rands 1986). The value of these set-aside strips is probably similar to that discussed for herbaceous field margins and beetle banks (see corresponding sections above).
In summary, there is abundant evidence that set-asides tend to have overall positive effects on farmland biodiversity, probably because they contribute to mimic aspects of low intensity farming. However, it is not possible to ascribe these benefits to management for gamebird hunting, as there seems to be no clear evidence that set-aside management is more
50 sympathetic to wildlife in game areas than elsewhere. Furthermore, the value for biodiversity of different management techniques may depend on local ecological context and conservation objectives.
Farmland woods
Small woodlands are important components of some farmed landscapes, providing adequate habitat for gamebird species like the pheasant (Woodburn & Robertson 1990a, Genovesi et al. 1999). Gamebird management in farmland may thus include the creation of new woodlands, and the maintenance and management of old ones to the benefit of game (Woodburn & Robertson 1990b).
Pheasant shooting provides the primary incentive to manage farmland woods for gamebirds. During the winter woodlands provide cover, food and shelter, while in spring they comprise a vital component of the breeding territory. Management is thus targeted at increasing both winter densities with benefits for shooting, and the number of breeding birds in the spring. Management of woodlands for pheasants favour relatively small (3-5 ha), mixed woods of different ages and species of trees, including mast-bearing species, with high edge to area ratio, and abundant shrubby cover (woody vegetation 0.2 – 2 m in height), particularly along the woodland edge (Robertson 1992). Woodland edges should preferably be bordering either winter or spring cereals (Robertson 1992). Wide rides (> 30 m in width) cut through existing woods provide alternative edges for birds in large blocks of woodland, favouring winter, though not spring, densities (Robertson 1992). Management prescriptions to achieve these goals are varied, ranging from the creation of new woods with suitable sizes, shapes and tree species composition, to improving the shrubby cover around and within extant woods (Robertson 1992). A great deal of management effort is generally devoted to maintain adequate shrubby conditions, as neglect causes the woods to become dark, cold and inhospitable to pheasants. The establishment of a rotational coppicing regime is a practice recommended for creating open, shrubby conditions for pheasants; the early years of coppice regrowth can provide among the best shrubby conditions for pheasants while the mixture of age classes creates numerous edges for this bird (Bealey & Robertson 1992, Robertson 1992). Creation of glades or skylights by felling small parts of the woodland may also be used to create patches of scrubs (Robertson 1992).
Robertson (1992) suggested that managing woodlands for gamebird hunting contributes for increasing their overall value for biodiversity, primarily by providing the incentive to create or retain a number of features of recognised benefit to conservation, including large rides, shrubby cover along the woodland edge and in its interior, coppicing, glades and clearings in existing woodlands, among other actions (e.g. Warren & Fuller 1990, Fuller & Warren 1990). Indeed, the available evidence suggests that management for gamebird hunting seems to have a large impact on the quantity and quality of farmland woodlands. This is particularly evident in the UK, where the pheasant is the mainstay of lowland game shooting (Robertson 1992). Piddington (1980) estimated that 33% of agricultural properties had planted or retained coverts, belts or spinneys for game and 18% had let their choice of trees for recent planting be influenced by shooting considerations. Similarly, a questionnaire survey to members of the British Country Landowners Association (Cobham Resource Consultants 1983), found that 67% of respondents claimed game interests as a reason for retaining existing woodland, while 56% claimed game was also a reason for planting new woods. These results are in line with observations by Stark et al. (1999), who compared landscape structure in 1 km squares with or without game management. They found that game areas tended to have more, larger,
51 woodlands than non-game areas; game woodlands tended to be older and more likely to be broadleaved, than non-game woodlands (Stark et al. 1999). However, in a study conducted at the farm scale, Hinsley et al. (1999) reported that woodland structure was similar on game and non-game sites. Overall, however, game interests seem to help maintain woodland presence in farmland.
The hypothesis that woodland management for gamebird hunting is more sympathetic to overall biodiversity than alternative regimes is more difficult to sustain, due to the shortage of available information, and contradictory results from the extant studies. For instance, Robertson (1992) sampled butterflies in each of four habitat types within a large wood, two of which were managed specifically for pheasants (woodland rides and hazel coppice), finding the highest butterfly densities and species richness in managed habitat types. However, Hinsley et al. (1999) did not find statistical significant differences in the number of butterfly species recorded in game and non-game woods in a study in lowland Britain. The same study did not record differences between game and non-game woods in the number of bird species and the abundance of most individual species, expect for the Song Thrush, which was most abundant in game woods, and the chiffchaff, which was most abundant in non-game woods. The Song Thrush belongs to a group of songbirds that share similarities with the pheasant in the woodland types they prefer, and that were suggested by Robertson (1992) to be favoured by woodland management for the pheasant. Robertson (1992), however, did not show that this kind of management actually benefits these species, but only that it would have the potential to do so.
In the study by Hinsley et al. (1999) the only consistent differences between game and non- game sites were found for woodland plants, with more canopy and shrub species in non-game than in game woods; the number of native tree species was highest in non-game woods, as it was the proportion of species that were native. There were no significant differences in the number of field species, nectar species and ancient woodland indicator species (Hinsley et al. 1999). Bealey & Robertson (1992) analysed the effect of coppice management for pheasants on the ground flora, reporting that few species or groups demonstrated clear responses to management, particularly where small-scale coppicing was carried out. However, in another treatment, they detected a slight but non-significant depression in Ancient Woodland Indicator species, whereas weed species significantly increased following management.
The studies carried out so far have not analysed in detail the impacts on biodiversity of woodland management for gamebirds at the landscape scale. One interesting observation by Hinsley et al. (1999) bears some relevance to address this issue. These authors have suggested that variation in woodland characteristics were larger in non-game than in game sites, suggesting that management contributed to a greater similarity among game woodlands. The results of this process may be the reduction in habitat diversity at the landscape scale, making it less adequate to support a wide range of species with different habitat requirements. Another problem is related to the creation of woods within the farmland habitats. Although this is certainly favourable for woodland species, steady declines have been recorded primarily in farmland, species suggesting that the creation of new woods may reduce further the availability of habitats.
In summary, management of farmland woods for hunting is widespread in some countries, and it has been claimed to have overall positive effects on biodiversity. However, the extent to which these potential effects actually occur is unknown. Furthermore, whereas for some groups there is some evidence that this kind of management may indeed be beneficial, for
52 others the information is contradictory or it even suggests that the effects can be negative. However, game interests seem to help maintain woodland presence in farmland, which may be overall positive to biodiversity in farmland. It is nevertheless doubtful whether this management may contribute positively for the conservation of typical farmland species, which have declined steadily in recent years.
Supplementary resources
Provision of grain
To increase the survival of birds through the lean seasons and to ensure that birds enter the nesting season in good condition, it is common practice to provide supplementary feeding. For instance, supplementary food for pheasants is provided during the autumn and winter, taking three typical forms: gamekeepers daily spreading grain along straw covered woodland rides; in feed hoppers placed throughout pheasant woods; or by dumping piles of grain spoil (Robertson 1992). All three techniques lead to higher winter pheasant densities (Robertson 1992).
Provision of supplementary food is frequently considered beneficial to other bird species (Hinsley et al. 1999). However, although it is likely that some seed-eaters may benefit from increased food resources during the winter lean season, hard evidence to support this view is still lacking. Furthermore, there is the possibility that congregating large number of birds at predictable feeding sites may increase the potential for the spread of infectious diseases (Pennycott et al. 1998) and predation risk.
In summary, it seems likely that the provision of grain may be of some value to other seed- eating birds, but evidence to support this view is lacking. Likewise, there may be negative effects of this artificial feeding regime, which are also poorly documented.
Provision of open water
In hot, dry climates, gamebird populations may be limited by the abundance and spatial distribution of water, through effects on survival (Degen 1985), reproduction (Koerth & Guthery 1991), population dynamics (Rice et al. 1993) distribution and habitat use (Brennan et al. 1987, Borralho et al. 1998). One common management action to address this problem is the preparation of summer water points (e.g., small dams and ponds, artificial watering devices) to improve the survival of both adult and young (Coles 1979, Benoklin 1988, Pépin & Blayac 1990, Borralho et al. 1998). We could find no study assessing the effects of this management action on species other than the target gamebirds. It is likely, however, that the provision of open water may benefit species with water requirements similar to the target gamebirds.
In summary, there is little doubt that the provision of open water benefits gamebirds in areas with hot climates. However, no information is available regarding the value of this management technique to other wildlife, though it is likely that a few species may be favoured.
53 Habitat manipulation to limit predation pressure
In the highly artificial farmland environment, predation seems to have a particularly important role in controlling gamebird populations (Tapper et al. 1996). Therefore, predator control tends to be strongly advocated as a key management action in farmland habitats. Although the most common practice associated with predator control is the actual killing of predators, either legal or illegal, there are also habitat management options targeted at reducing predation levels. These include, for instance, the burning of bramble (Rubus spp.) thickets in Portuguese game estates to reduce cover to carnivore mammals (R. Borralho pers. comm.). In the UK, Kenward et al. (2001) suggested a number of habitat management prescriptions to reduce predation levels by buzzards (Buteo buteo) at pheasant-release pens. At present little is known about the effects that these may have on overall farmland biodiversity, apart from reducing probably the abundance of predators. The impact of gamebird predators on target and non-target populations is discussed at length in section 3.1.
Conclusions
From the present review, it is clear that habitat management for gamebird hunting in farmland comprises a very large array of approaches and techniques, which are designed to allow the target gamebirds to meet their ecological requirements, given the constraints of arable crop or livestock production. This is generally difficult in present-day, highly intensive farming systems, where the habitat features needed by the birds to secure food and shelter are generally very restricted or even lacking. Therefore, most management prescriptions are designed to mimic some of the features characteristic of more extensive farmland, including increases in the quantity and quality of marginal (hedgerows, herbaceous field boundaries, etc.) or semi-natural (e.g. farm woodlands) habitats, reduction in agro-chemical inputs (e.g. Conservation Headlands), and recuperation of more favourable farming practices (e.g. maintenance of winter stubbles or undersown leys), among other actions. Another set of management practices involves the direct supply of critical resources, such as food and water, to increase gamebird populations beyond the carrying capacity of the habitats available.
Management actions reducing the intensity of farming practices, seem to have overall positive impacts on farmland biodiversity. Indeed, most farmland wildlife in intensive farming systems seem to be limited by some of the factors that also drive gamebird populations, and thus alleviating these limitations may have large beneficial effects. This view is supported by a number of well-designed, replicated experimental studies, in which the application of management actions have been shown to result in positive effects for both gamebirds and many other organisms. Conservation headlands probably represent the best example of a measure designed for gamebirds, which proved unequivocally to possess wide positive implications for farmland biodiversity.
It should be pointed out, however, that demonstration by the above studies that at least some gamebird management practices have potential benefits to biodiversity, do not imply that gamebird hunting in itself has benefits to biodiversity in the wider countryside. These studies thus show what can be achieved, and not what is typically being achieved in the real world. The reasons for this are at least threefold:
1. Management actions in experimental settings reflect best practice, which may differ strongly from the typical practice adopted by landowners and estate managers; these
54 differences may have far-reaching consequences for the value of each practice on biodiversity. 2. Some of the gamebird management actions may not be widely used, because they are costly or difficult to implement technically; other practices, easier to implement and more cost-effective may be widespread. Indeed, one study in the UK has shown that Conservation Headlands and other valuable management actions seem to be more widely used in non-game areas, for the general benefit of wildlife, than in game areas, where hunting is largely based on released birds. 3. Most studies have compared gamebird management actions with the normal agricultural procedures. However, there is nowadays a range of agri-environment schemes promoting farming practices more sympathetic to wildlife and to which gamebird management prescriptions should be compared. For instance, although both gamebird and wildlife management promote the conservation of hedgerows, they tend to recommend hedges with different structural characteristics. It should be recognised, however, that some of these agri-environment schemes have adopted techniques initially devised for gamebird management.
In summary, it is likely that if a range of gamebird management actions, carefully planned, adapted to the local ecological context and conservation objectives, were implemented in a farm, this would lead to a far richer farmland biodiversity than what could be found in comparable farms under the average agricultural management. This view is supported by a lot of studies analysing the effects of isolate game management actions, as well as whole-farm experiments like that carried out under the Allerton Project in the UK (e.g. Boatman & Brockless 1998). However, it is doubtful whether these results can be extrapolated to real, field conditions, where a number of additional factors seem to play a much stronger role in determining the levels of farmland biodiversity. Given this problem, there is clearly a need for studies analysing the real effects of management for gamebird hunting on biodiversity, which may differ from the potential benefits inferred from experimental settings.
55 4. CONCLUDING REMARKS
Despite differences between countries, management of gamebirds and their habitats is relatively common in all countries considered. Landscape-scale habitat management only occurs in grouse moors in Great Britain, and in some areas in the Iberian Peninsula where private estates are managed for red-legged partridges. Small-scale habitat management for gamebird hunting interests is particularly common in farmland areas, for partridges and pheasants. Predator control is common in all countries, and is practised as part of the management of all species that are commonly hunted. Only for species that are hunted in small quantities (mountain gamebirds in the UK, France or Spain), predator control is not carried out.
Unsurprisingly, most management practices carried out for gamebird shooting interests have their desired positive effects on target gamebirds. However, sometimes these management practices, to be effective, have to be carried out in optimal (best practice) circumstances, which are not always achieved in the real world. This is the case, for example, of predator control, where its efficacy may vary enormously according to population and ecological variables. In many instances, predator control programs have to be intensive and large-scale to achieve an effect on game populations.
Releases and restocking are common, and possibly increasing practices, which tend to cause short-term boost in the numbers available for shooting, though they may have long-term negative effects. In some circumstances releases/restocking contribute for depressing wild gamebird populations, by spreading diseases and parasites, or by competition with local wild populations. Another usual consequence seems to be the decrease in habitat quality if habitat management is neglected in areas where most money is invested in releases. Releases/restocking tend to be associated with intensive forms of hunting, and not to game management based on the sustainable exploitation of breeding populations. The exceptions are few and primarily related to the recovery of populations depressed by overshooting or low habitat quality, when associated with other management actions.
The effect of management actions on species other than wildlife is even less well documented, and results of available studies are inconsistent. Predator control has the potential to be positive for species that are limited by the same set of predators as game, or for maintaining balanced predator guilds, but studies to prove this are needed. In general, releases/restocking may have several negative consequences for wildlife populations and genetic diversity, though very few of the effects suggested have been supported by hard scientific evidence. Future research is needed, and it should probably be focused on ecosystem and community- level processes.
Habitat management seems to be the practice that is overall more positive for both gamebirds and other wildlife, but even in this case effects are mixed. Shooting activities in areas where hunting is the main land use may contribute to maintain habitats that otherwise may disappear, and a change to alternative land uses is likely to be negative for biodiversity, at least in some cases (for example, upland grouse moors in the UK). However, management of these areas for gamebird is beneficial for some species, but detrimental for others, so the overall biodiversity value of management will thus depend on conservation priorities in each area.
56 Habitat management in farmland areas seems to be highly valuable, both for gamebirds and for other species, particularly those practices that mimic the characteristics of extensive farming. That is the case of the creation of herbaceous field margins, the creation of wild bird cover in fallows and set-asides, or the setting up of conservation headlands, which are probably the best example of a management prescription for gamebirds that has widely beneficial implications for the overall biodiversity. This view is supported by a number of well-designed, replicated experimental studies, in which the application of management actions have been shown to result in positive effects for both gamebirds and many other organisms.
The effects of game management practices are obviously conditional upon the extent to which they are actually implemented for hunting purposes in the wider countryside. As stated in this review, quantitative information on the extent of such practices is lacking in many countries, particularly in Spain, Portugal and Finland. With the exception of areas where the primary land use is hunting (such as grouse moors), management of habitats and predators is more common and intense in farmland than on the uplands, whereas the opposite is true for the management of bags. In many cases, however, evidence is lacking to support the view that a larger extent of valuable management prescriptions, such as grassy field margins, set asides or conservation headlands, are implemented in game than in non-game areas, and that management of these structures for gamebird hunting are more beneficial for biodiversity than competing management regimes. The latter aspect is particularly relevant in farmland habitats, given the wide range of agri-environment funding opportunities currently available in EU countries. Furthermore, the value for biodiversity of different management techniques in farmland may depend on local ecological context and conservation objectives, as it happens in areas where the main land use is hunting. These problems underline the potential shortcomings of inferring the effects of game management from the kind of experimental studies carried so far, which may be powerful for elucidating the processes behind the patterns associated with each management technique, thus demonstrating what can be achieved, but which are limited in their capacity to reveal what is typically being achieved in the real world. Therefore, there is clearly a need for studies analysing the real effects of management for gamebird hunting on biodiversity, which may differ from the potential benefits inferred from experimental settings.
Overall, it is likely that a range of gamebird management actions, carefully planned, adapted to the local ecological context and conservation objectives, could have positive effects on biodiversity, particularly in intensive farmland habitats and in areas where hunting may help maintaining habitats that would otherwise be transformed in less ecologically valuable land uses. On the other hand, however, ongoing trends for the intensification of the hunting regimes, with large-scale releases but habitat management neglect, may be detrimental to overall biodiversity. There is clearly a strong need for more research to assess quantitatively the net effect of gamebird management actions on biodiversity, preferably over large spatial scales and long time frames. Also, socio-economic studies would be needed to investigate how best to encourage those game management practices most valuable for biodiversity.
57 5. ACKNOWLEDGEMENTS
Many people contributed with data or papers to the workpackage, particularly Elisabeth Bro (who coordinated data and comments coming from a number of people in the Office National de la Chasse et de la Faune Sauvage) and Rui Borralho, but also Javier Viñuela, Santi Mañosa, Steve Tapper, Arjun Amar, Nicholas Aebisher and Jari Valkama. We are also grateful for critical reading of the manuscript and constructive comments by Elizabeth Bro, Stephen Redpath and Robert Kenward.
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