Management and Control of Populations

of

Foxes, Deer, Hares, and Mink

in England and Wales,

and the Impact of Hunting with Dogs

Macdonald, D.W.¹, Tattersall, F.H.¹, Johnson, P.J.¹, Carbone, C.¹, Reynolds, J. C.², Langbein, J.³, Rushton, S. P.⁴ and Shirley, M.D.F.⁴

¹Wildlife Conservation Research Unit, Dept. of Zoology, South Parks Rd., Oxford, OX1 3PS; ²The Game Conservancy Trust, Fordingbridge, Hampshire, SP6 1EF; ³Wildlife Research Consultant, "Greenleas", Chapel Cleeve, Minehead, Somerset TA24 6HY; ⁴Centre for Land Use and Water Resources Research, University of Newcastle upon Tyne, Porter Building, Newcastle upon Tyne NE1 7RU

1. Why seek to control populations of foxes, deer, hares, and mink in England and Wales?

2. What methods are available to control populations of foxes, deer, hares, and mink in England and Wales?

3. What do simulation models suggest about the effectiveness of methods to control populations of foxes, deer, hares, and mink?

Much more research is required on the basic mortality and dispersal biology of each of the species before we could use modelling in any tactical management of these species. The results of the modelling are therefore ‘general’.

4. How effective are methods to control populations of foxes, deer, hares and mink in England and Wales?

5. How acceptable are methods to control populations of foxes, deer, hares, and mink?

6. What would be the impact on populations of foxes, deer, hares, and mink of a ban on hunting with dogs, and how would this affect different interest groups?

7. In which areas is there agreement or dispute, and what data are lacking?

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Contents

1. Introduction

1.1. The debate

1.2. Background

1.3. How does management and control affect populations of wild mammals?

2. Why seek to control populations of foxes, deer, hares, and mink in England and Wales?

2.1. Introduction

2.2. Why control fox populations?

2.3. Why control deer populations?

2.4. Why control hare populations?

2.5. Why control mink populations?

2.6. Conclusions

3. What methods are used to control populations of foxes, deer, hares, and mink in England and Wales?

3.1. Introduction

3.2. What methods are used to control fox populations?

3.3. What methods are used to control deer populations?

3.4. What methods are used to control hare populations?

3.5. What methods are used to control mink populations?

3.6. What are the alternatives to culling?

3.7. Conclusions

4. What can simulation models tell us about the effectiveness of population control methods?

4.1. Why use simulation models to estimate effectiveness?

4.2. What types of models have we used to estimate effectiveness of population control methods?

4.3. What do population-based models predict about the effectiveness of different population control methods?

4.4. What do individual-based models predict about the effectiveness of different fox and mink population control methods?

4.5. What does Population Viability Analysis tell us about the effectiveness of hunting to control fox populations?

4.6. Conclusions

5. How effective are methods to control foxes, deer, hares, and mink in England and Wales?

5.1. What do we mean by ‘effective’?

5.2. How effective are methods to control fox populations?

5.3. How cost-efficient are methods to control foxes?

5.4. How effective are methods to control deer populations and damage?

5.5. How effective are methods to control brown hare populations?

5.6. How effective are methods to control mink populations?

5.7. Conclusions

6. How acceptable are current legal methods to control populations of foxes, deer, hares, and mink?

6.1. What do we mean by acceptability?

6.2. How humane is hunting with dogs?

6.3. How humane is shooting?

6.4. How humane is trapping?

6.5. How humane is snaring foxes?

6.6. How humane are other methods?

6.7. Conclusions

7. What would be the impact on populations of foxes, deer, hares, and mink of a ban on hunting with dogs, and how would this affect different interest groups?

7.1. Introduction

7.2. What would be the impact of a ban on hunting foxes with dogs?

7.3. What would be the impact of a ban on hunting deer with dogs?

7.4. What would be the impact of a ban on hunting hares with dogs?

7.5. What would be the impact of a ban on hunting mink with dogs?

7.6. Conclusions

8. Acknowledgements

9. References

10. Appendix 1: Species descriptions

10.1. Red fox (Vulpes vulpes)

10.2. Deer

10.3. Hares

10.4. American mink (Mustela vison)

11. Appendix 2: What are the current legislative restrictions on control methods in England and Wales?

11.1. Statutory basis for control

11.2. What additional legislation affects control of foxes?

11.3. What additional legislation affects control of deer?

11.4. What additional legislation affects control of hares?

11.5. What additional legislation affects control of mink?

1. Introduction

1.1 The debate

Whether or not hunting some mammals with dogs should continue to be allowed in Britain has been highly controversial, with staunch protagonists on both sides. The twin issues of how some British mammals should be managed, and whether hunting with dogs makes a useful contribution, have formed an important part of the debate.

Of course, hunting in the broadest sense has its roots in prehistory, but hunting with dogs in Britain is also an ancient occupation (Macdonald, 1984; Macdonald & Johnson, 1996; Strutt, 1883). For example, stag hunting was an obsession with many of the Norman Kings. Of William the Conqueror it is said that it was better to have been a stag than his subject, so rigidly did he enforce the harsh Forest laws, under which the penalty for killing a red deer stag was death (Whitehead, 1964). Fox hunting also has a venerable history and was seen as a useful service. Strutt (1883) reproduces an engraving dated early 14^th century, of a fox being unearthed by digging for a waiting dog to catch, with a bystander blowing a horn. He also records, "in the 43^rd year of Edward III [i.e. 1370], Thomas Engaine held lands in Pitchley in the county of Northampton, by service of finding at his own cost certain dogs for the destruction of wolves, foxes, & c., in the counties of Northampton, Rutland, Oxford, Essex and Buckingham".

During this long history, the balance of motivations, for example between pest control and recreation, has doubtless varied, as it may at any time from place to place or even occasion to occasion. So too, the effectiveness of this pursuit in achieving different goals may vary. To evaluate effectiveness necessitates clearly identifying the goals, and distinguishing the criteria on which performance can be judged. It may also be necessary to consider whether the goals are well founded. For example, evaluating the contribution of hunting to the goal of pest control is only relevant insofar as it is established that the fox is a pest, and that limiting its numbers reduces the nuisance it causes.

Issues of this sort are not uncommon – worldwide they are the grist of the inter-disciplinary science of wildlife management. Invariably, wildlife management and conservation is made challenging not only because people hold contrasting views on the desirability and legitimacy of different goals, but also because the yardsticks whereby effectiveness may be measured are often incommensurable. That is, even when measures have been accurately made, they are often calibrated in such different units (e.g. population size, money, aesthetics, cruelty) that they cannot systematically be compared to give a single ‘right’ answer. In short, science can greatly inform debate, but ultimately judgement will decide it. The case of hunting with dogs illustrates vividly these difficulties: it is perceived by different people to have different goals and diverse consequences, some of which are contradictory, many are technically difficult to evaluate, and most are measured in incommensurable units.

In this report, we address the issues of how populations of foxes, deer, hares, and mink are controlled, and the impact on them of hunting with dogs. In this introductory Chapter, we cover some of the important biological background, including the processes underlying changes in mammal populations. In Chapter 2 we ask why these species are controlled, and in Chapter 3 we ask what methods are used. The effectiveness of hunting with dogs in comparison with other methods forms a key part of our report (Chapters 4 and 5). In Chapter 6 we outline the acceptability of hunting with hounds and other methods, with particular emphasis on the associated welfare implications. Finally (Chapter 7), we assess the potential impact of a ban on hunting with dogs on populations of foxes, deer, hares, and mink, and on wider issues related to their management and control. Throughout, we comment on regional differences and the need for further work. Descriptions of each species are given in Appendix 1.

1.2 Background

This report was written under contract to the Committee of Inquiry into Hunting with Dogs. The Committee’s terms of reference are:

" To inquire into:

To report the findings to the Secretary of State for the Home Department."

Our report relates to two contracts:

Contract 5: MANAGEMENT OF THE POPULATION OF FOXES, DEER, HARES AND MINK, AND THE IMPACT OF HUNTING WITH DOGS.

Contract 6: METHODS OF CONTROLLING FOXES, DEER, HARES AND MINK.

1.2.1. Aims

The aims of this report are based on the research objectives outlined in Contracts 5 and 6, which we have formulated as a series of questions:

In answering these questions, we have focussed on five interest groups: farmers, game managers, foresters, fisheries managers, and conservationists. Of the two species of hare, we have concentrated on the brown hare, and among the six deer species we have concentrated on red and roe, as only these two continue to be hunted with dogs in England and Wales.

1.3. How does management and control affect populations of wild mammals?

1.3.1. What do we mean by management and control of populations?

Man’s treatment of wild mammals has become a controversial subject in Britain. Unfortunately, its many aspects easily become confused by careless use of English. In this report we adopt precise meanings for a few common English words.

The term ‘population’ is a convenient shorthand for ‘the animals living in a defined area’. A population can only rarely be considered in complete isolation, for example on islands. Usually, immigration and emigration from surrounding areas are important, especially in small areas, which have a high ration between their frontier (for immigration and emigration) and area (for births and deaths).

By ‘management’, we mean any deliberate interventions by man to manipulate the number, structure, distribution, and impact of an animal population. An immense variety of techniques and approaches are used to manage mammal populations (see reviews in Caughley & Sinclair, 1994; Bookhout, 1994). We use the term ‘population control’ to mean that numbers are held within desired limits by management. Since animal populations have an intrinsic tendency to increase when resources allow, management for population control usually involves intervention to either increase mortality (e.g. culling) or to reduce productivity (e.g. contraception). Methods specifically aimed at ‘damage control’, such as fencing, chemical repellents or habitat modifications, may sometimes be more appropriate than population control.

An important distinction is between intended and achieved population control. Except in the short-term, killing individuals does not necessarily limit a population (because these individuals may have been part of a seasonal surplus, or because they may rapidly be replaced through births and immigration), or diminish any damage caused. A parallel distinction is important between perceived and proven pest status.

Various people and organisations have an interest in managing mammal populations. These ‘interest groups’ often wish to achieve different, sometimes conflicting, aims, including:

There is often little contact between interest groups, resulting in many separate management strategies within any region, some of which might be diametrically opposite in aim, method and outcome. For example, the brown hare is simultaneously managed as a pest (for eating young crops), a game animal, and a species of conservation concern.

1.3.2. What happens if populations are not managed?

At any one time the number of animals in a population depends on the history of two opposing processes: animals entering the population through births and immigration, and animals leaving the population through deaths and emigration. A great many factors simultaneously influence these processes, among them resources, competition, behaviour, predation, and weather.

Under favourable conditions, populations of animals have the capacity to grow by producing more young than are lost through mortality. This intrinsic growth rate can sometimes be very fast. However, even unmanaged populations are eventually limited by resources such as food, water, or den sites. A population that has increased to the maximum supportable by a limiting resource is said to be at ‘carrying capacity’. Populations can be held below carrying capacity by mortality unrelated to a limiting resource, such as disease.

Often, as populations grow and approach carrying capacity, their growth does not increase at a constant rate, but starts to slow down; processes not operating at low population density come into play to decrease births or increase losses, or both. This change in growth rate with density is known as ‘density-dependence’, and its effect is to regulate or stabilise numbers, as a thermostat regulates temperature. Other processes affecting populations are not related to density and have either a constant or a random effect. There is thus a distinction between processes that limit the population and those that regulate it, the latter being the density-dependent subset of the former.

A useful concept is the ‘equilibrium level’, where the gains and losses of the population are in balance, so that overall the population is neither growing nor declining. The population density at which this balance occurs depends on the prevailing combination of density-dependent and density-independent processes. In modern environments dominated by man, virtually every process acting on even unmanaged wildlife populations will be linked to human activities, and many or most of the possible equilibrium levels will be heavily influenced by man.

Furthermore, populations do not reach an equilibrium level and stick there. In reality, population density will fluctuate around an equilibrium level, as in any system under regulation (e.g. the temperature of a hot-water tank). In addition, if the factors influencing births, deaths, immigration, and emigration are continually changing (an inconstant or continually perturbed environment), the population may never reach an equilibrium, except as a transient state. Thus, stability becomes an important concept in population biology. Again, one can identify processes that tend to increase stability, and others that tend to cause chaotic fluctuations.

1.3.2.a. What is the evidence for density-dependent regulation of fox populations?

There are two main strands of evidence for density-dependence in foxes. First, fox populations that are dense relative to food resources are generally less productive than those that are less dense. Second, there are wide variations in the proportions of vixens that reproduce each year, and in their average litter sizes (Macdonald, 1980; Layne & MacKeon, 1956; Englund, 1970; Lindström, 1982, 1983, 1988; Chirkova, 1955). These two aspects of productivity appear to be related to crowding effects, with lowest productivity tending to occur where fox density is high or food supply poor (Harris & Lloyd, 1991; Lindström, 1992, 1998; Lindström et al., 1989; von Schantz, 1981; Harris, 1977; Pils et al., 1981).

Where food is not limited, crowding itself can suppress reproduction, with only the dominant female breeding. Behavioural mechanisms by which this occurs include harassment of subordinates, infanticide and cannibalism of subordinate vixens’ cubs, and possibly the dog fox courting only the dominant females (Macdonald, 1977, 1980). A hormonal mechanism whereby stress leads to lowered productivity through foetal reabsorption has also been identified (Hartley et al., 1994). Consistent with this mechanism, Heydon & Reynolds (2000b) found that in populations where productivity was low, performance was depressed consistently at all stages of pregnancy, from conception to birth.

Similar evidence of reduced productivity in relation to density and resources is found in a wide variety of other mammal species (e.g. arctic fox, Angerbjörn et al., 1991; racoon dog, Helle & Kauhala, 1995; badger, Cresswell et al., 1992; white-tailed deer, Swihart et al., 1998). This is significant because the hormonal processes governing reproduction and reactions to stress are basically the same in all mammals.

1.3.2.b. What is the evidence for density-dependent regulation of deer populations?

For red, roe, and fallow deer, three types of density-dependent changes which influence population dynamics have been reported: an increase or decrease in age at maturity; variation in the proportion of adult females conceiving or calving each year; and variation in first-winter mortality of juveniles. Each of these clearly has the potential to reduce population growth as density rises or resources become more restricted (Albon et al., 1983; Langbein, 1991; Langbein & Putman, 1992a; Gaillard et al., 1992; Hewison, 1993, 1996; Putman et al., 1996). However, none of these authors studied populations actually at equilibrium, and it seems that while density-dependent factors do reduce recruitment rates when conditions are poor, they rarely result in zero-growth or regulation to equilibrium. On the other hand, effects of density-independent factors (e.g. severe winters, drought) are usually regarded as destabilizing rather than regulatory, due to their stochastic (random) impact on mortality or recruitment. In the case of deer, density-dependent changes may generally function to compensate or to dampen wider, erratic fluctuations in population numbers resulting from density-independent variation (Putman et al., 1996).

1.3.3. What can happen when populations are managed?

Equilibrium levels reached by unmanaged populations may not be at a density that is compatible with human interests (e.g. commercial stalking is incompatible with very low densities of deer, and restoration of water vole populations may be incompatible with high densities of American mink). In these cases, human intervention through management may be necessary to increase or decrease densities, or dampen fluctuations.

Management can increase populations in two ways: by enhancing gains (e.g. through increasing limiting resources or increasing immigration via translocation); or by reducing losses (e.g. by removing causes of mortality such as predators, or by vaccination against disease or parasites). These techniques are commonplace in game management and wildlife conservation (Caughley & Sinclair, 1994; Bookhout, 1994).

Where the aim is to reduce the population density of a wild species, the mechanism is to create a density-dependent force that is sufficient - in combination with all the other processes acting on the population - to reduce density to the level required and hold it there. In principle, the options are to increase mortality or decrease productivity, but the latter is not yet practical (sections 3.6.1.a.iv and 3.6.1.b.v).

In the UK context, the aim of population control is usually to limit a population, rather than to exterminate it. However, there are cases where extermination might be a goal, such as American mink from sea-bird islands. Exterminating a pest is generally difficult. First, very intensive culling is usually required to launch an abundant and productive population into decline, because reducing density relieves density-dependent constraints, thereby facilitating increased growth rates. Clearly, to achieve a sustained decline will require a level of culling that exceeds the maximum rate of growth the population can attain. Second, culling is naturally density-dependent, becoming more difficult as population density decreases. Thus, effort must increase as density declines.

Where a wild animal is harvested, for example, for venison, management may involve a combination of positive management and selective culling. Carefully planned harvesting ensures that animals are killed without reducing the numbers produced in future years. Counter-intuitively, the maximum sustainable yield is produced by populations below carrying capacity, due to density-dependent processes.

1.3.3.a.Can control exert a selective pressure on a population?

In principle, any selective mortality has the potential to influence the genetic makeup of a population if it affects a heritable trait. For example, if population control selectively increased the mortality of animals with a particular heritable appearance or behaviour, not only might there be a short-term reduction in the proportion of such animals in the population, but individuals with these traits might leave fewer offspring, thereby eventually changing the genetic profile of the population. If old animals are selectively culled, this will affect the composition of the population, but not its genetic makeup, as these individuals would have already bred and thereby passed on their genes.

In practice, control methods generally do not exert sufficient selective pressure to have an impact on the genetic makeup of quarry species, and the few exceptions require very heavy selection and continuous maintenance. For example, over the last 300 years black herds of fallow in Epping Forest have been maintained by stringently culling any other colour strains, but recently, increased immigration into the herds by animals from expanding populations in Essex is making this difficult (Langbein, 1996).

The most common example of man’s attempts to exert selective pressure through culling is the management of deer to maximise high ‘quality’ trophy heads. Selective culling of males with antlers which show ‘poor’ characteristics in terms of shape and size is used to promote particular characteristics of antler form. Such approaches to culling are so ingrained in tradition that in some European countries the removal of ‘poor’ quality stags or bucks is a requirement laid down in cull plans agreed by hunting authorities (Gill, 1990). The ability to exert some degree of selection for particular characteristics of antler shape is well proven (e.g. Whitehead, 1993; Ueckermann & Hansen, 1983). However, overall antler size and complexity is probably more influenced by environment than genetics, not least in view of the considerable nutritional demands on male deer in the brief period of antler growth (Fennessay & Suttie, 1985).

Concern is often expressed over loss of genetic variation through inbreeding. For example, deer hunts sometimes choose to kill a stag that has been with a herd of hinds for several years, to prevent inbreeding (MDHA submission to the Inquiry). Suffice it to say here that these concerns are largely erroneous. While problems arising from inbreeding do occur in (already highly inbred) domestic animals, and in very small, enclosed, populations of wild mammals, there is very little evidence for inbreeding in free-living wild mammals, which have sophisticated social systems to avoid it, and even less to suggest that when it does occur it has harmful effects.

1.3.4. What features of the quarry species’ population processes are important to their control?

1.3.4.a. Fox population control in terms of fox biology

The fundamental aims of fox management in the UK are to reduce populations or prevent their increase. For this to happen, losses of foxes through deaths and emigration must equal or exceed gains through births and immigration.

In a large geographical region such as a whole county, immigration and emigration will be minor relative to births and deaths. Rural fox populations produce 2-3 cubs per adult annually. Thus if a population of foxes before breeding was 100, it would increase unless 200-300 foxes died before breeding the following year. Culling could account for most or all such deaths, but accidental (e.g. road traffic) or natural causes (e.g. disease, starvation) are also important, and might be sufficient on their own.

At a more local scale (e.g. within the confines of a single farm or shooting estate), two further aspects of fox biology – territoriality and dispersal – become important. To reduce fox density in such a small area, culling must remove not only any resident territory-holding foxes and their offspring, but also any ‘replacement’ foxes that would normally have been excluded by the territory-holders. These will be either foxes encroaching from neighbouring territories or foxes dispersing from territories farther away. As a result, culls can be locally as high as 25 foxes per square kilometre, even though rural fox densities are typically only 0.5-4.0 per square kilometre in autumn, after cub production.

Because it draws from a pool of potential replacement foxes in the surrounding countryside, intensive local culling does create a ‘sink’ effect, but it is wrong to imagine that local culling creates a vacuum that sucks foxes in from far away. Foxes on distant territories cannot be aware of the vacant space, so the local culling effort increases their risk of dying only if they are already committed to dispersal behaviour and actually arrive in the culling area. In spring and summer, when no dispersal occurs, the impact of localised culling on fox numbers does not extend more than a few kilometres outside the culling area.

How does local culling fit into the regional context? The countryside can be pictured as a mosaic of ‘sinks’ and ‘sources’. In sink areas culling has ensured that mortality exceeded the local fox productivity, while in source areas culling has been insufficient to prevent an increase in fox numbers. Irrespective of whether a local culling effort meets its local aims (e.g. lower predation on game birds), it is inescapably a component of fox mortality in the region as a whole. Indeed, because dispersal allows high ‘bags’ to be attained on quite small areas of land, localised fox culling may contribute substantially to the total cull of foxes in a larger region. If many local culling efforts take place within a region, the impact of these alone could amount to regional limitation of fox numbers.

1.3.4.b. Deer population control in terms of deer biology

Although not all the deer species are territorial, many of the above points for foxes also apply here. For roe, which do exhibit territoriality of both males and females, the vacuum left by culling a mature male is often filled quickly by one or two younger males taking over or splitting the vacant territory. Even for non-territorial species, such as red and fallow, source and sink effects may be significant in achieving population control through culling, as their seasonal or daily home ranges will often cover several different landholdings. This is particularly so in southwest England where average estate sizes are comparatively low (c.100ha) compared to the individual mean range sizes of red deer hinds (c.400ha) and stags (1000ha; Langbein, 1997). Deer culling undertaken only on some small estates but not on surrounding land, may thus often have rather more wide-reaching effects. However, maintenance of close control over the structure of local populations may be difficult within small estates, as for much of the year males of the larger species often live in separate ranges from females.

The potential natural increase of deer populations is somewhat lower than for the other species considered in this report: red and fallow deer rarely produce more than a single young per year, and while roe frequently produce twins, they average nearer 1.5 young per adult female. Most lowland red, fallow and roe populations can sustain an annual cull of 20-25 % of the autumn population (equivalent in number to c. 25-33% of the spring population). In the more extreme conditions of the Scottish Highlands, culling levels of only 10-20% may suffice to prevent increases (Ratcliffe, 1987; Ratcliffe & Mayle, 1992). Based on present nationwide estimates for spring deer numbers of >1,000,000 for all six of our deer species combined (Staines et al., 1998; Harris et al., 1995), this nevertheless suggests an annual cull requirement of around 250,00 deer merely to maintain numbers at current levels. This size of cull is not unusual by comparison to other European Countries: in Germany where deer culls are monitored in some detail, the annual cull of roe deer alone regularly exceeds 1 million.

Britain today lacks significant natural predators of deer, and long-term research indicates that without management deer populations approaching an equilibrium level sustain a heavy impact on their habitat and health, and on human interests (see section 1.3.2.b). In order to obtain a sustainable balance among the varied conservation and economic objectives of deer management in the UK, population control is inevitable. The actual numbers of deer that can be sustained without causing unacceptable impact on vegetation can, nevertheless, be manipulated by careful consideration of deer in long-term plans for the design and restructuring of forests.

1.3.5. Regulation of wildlife management and control in Europe

The regulation of wildlife (including game and pest) management varies considerably among European countries (Myrberget, 1990). Although historically England, Wales, and Scotland were very advanced in enacting statutory protection measures, today the organisation of wildlife management in the UK is much less institutionalised than in many other European countries. Current legislative restrictions on control methods are described in Appendix 2. Below we briefly review key components of European regulations of wildlife management:

2. Why seek to control populations of foxes, deer, hares, and mink in England and Wales?

2.1. Introduction

In this chapter, we consider the reasons why different interest groups seek to exert control over populations of foxes, deer, hares, and mink in England and Wales. Specifically, we consider the reasons for control by five major interest groups that are potentially affected by one or more of these species: farmers, game managers, foresters, fisheries managers and conservationists. In Britain, the methods that are used to control populations (rather than damage) always involve culling, and are dealt with in Chapter 3.

In general, the major reason why people seek to control populations of foxes, deer, hares and mink in England and Wales is that they believe these animals cause damage, for example, to livestock, nesting birds, or crops. Other reasons include prevention of the spread of disease and maintenance of ‘healthy’ populations. The exploitation of animals for sport or animal products is a separate motivation from the desire to control populations to limit damage. Neither motivation will necessarily result in effective population control (section 5.1), although both have the potential to do so, as both involve culling. In exploitation, the desire may be to provide a high yield of sport or animal products. The population level at which this optimum yield is attained will usually be greater than the population level desirable purely for damage limitation (section 1.3.3). Conflicting management aims therefore arise for species such as some deer and hares, which are simultaneously considered pests, game species, quarry, and are of conservation concern.

2.1.1. How well do the reasons for control relate to population control strategies?

Caughley & Sinclair (1994) point out that the original reason for the existence of a control campaign is frequently forgotten, after which the action itself becomes the objective. They discuss the case history of deer control in New Zealand where the official reason given for control operations has changed several times between 1920 and 1992. None of these changes had any effect on the management action in place. The means became the end.

People who deliberately seek to control animal populations generally reason that they do so because they consider a particular species to be a pest because of the damage it causes. As we demonstrate below, however, population control is not always a simple response to recent damage, nor is it necessarily an effective form of damage control, as there is often no simple link between the abundance of a species and the damage it causes.

2.1.1.a. What is the link between farmers’ experiences of stock losses and their fox control?

Farmers’ responses to damage in terms of culling methods and policy aims are related to both farm size and region. Heydon & Reynolds (2000a; section 2.2.1.a.i) found that only on farms of less than 200ha was recent (last 12 months) experience of livestock losses associated with a higher likelihood of (independent) culling effort; this was not the case for any other farm category. Recent losses did not increase or decrease the involvement of communally organised methods like hunting with hounds.

The aim of eliminating losses of livestock, game, or wildlife prey was also more common on small (<200ha) farms, while on larger estates (>200ha) reducing losses to an acceptable level was more often the aim. Small farmers in the east Midlands were more tolerant of losses due to fox predation than were their counterparts in mid-Wales and west Norfolk. Importantly, only a quarter of farmers had purely local aims: three-quarters cited regional control of fox numbers as an aim of their culling regime.

Data from a 1981 survey (Macdonald & Johnson, 1996; section 2.2.1.a.ii), provide evidence of a link between control and significant damage sustained at some time in the past (according to the subjective assessment of the respondent). Of 183 farmers who stated they had suffered significant damage by foxes, 77% carried out control, while only 13% of 395 who had not suffered damage controlled foxes [1]. For comparison, 45% of farmers reporting deer damage in 1981carried out deer control, but only 3% of those not reporting damage did so (see also Table 2-7 in section 2.3.1.b)

Regional patterns of reported fox damage generally mirrored those of control (Table 2-1), except in Warwickshire, where 41% of farmers carried out control, but only 26% suffered damage. A farmer’s tendency to carry out fox control might be linked to the amount of damage he perceives he would incur in the absence of any control measures, as well as to the cost of control (see also section 5.3). One measure of this perceived potential for damage might be whether the farmers believes there are ‘too many’ foxes.

We investigated the link between farmers’ tendencies to report that they controlled foxes, with their tendencies to cite different reasons (Table 2-2), and found evidence that perceived pest status affects a farmer’s tendency to control foxes independent of reported damage. In addition to reported damage, the ‘too many’ reason was a significant predictor of whether a farmer controlled foxes [2]. No other reason had this property. So, regardless of whether a farmer reported damage due to foxes, they were more likely to cull them if they thought there were ‘too many’. This potentially explains the regional variation observed in Table 2-1: for example, in Warwickshire, where the highest proportion of farmers said there were ‘too many’ foxes, almost twice as many farmers controlled foxes as reported suffering damage.

In Wiltshire, in 1995, farmers who considered the fox to be a pest on their own farm were more likely to believe that foxes should be controlled everywhere than farmers who did not consider the fox a pest on their farm (Baker & Macdonald, 2000; section 2.2.1.a.ii). Although most farmers thought the fox ‘too numerous’, farmers who perceived a fox pest problem did not estimate a higher density of foxes on their land, and the belief that there were too many foxes was widespread, even among farmers who were not affected by losses.

It seems, therefore, that farmer’s fox control efforts are largely driven by their fears of what would happen if they stopped culling. At present, we have insufficient data to predict whether farmers’ losses would in fact increase if they stopped culling foxes, and it is likely that only an unfeasibly large and long-term field experiment would be able to provide the answer.

2.1.1.b. Is there a link between a species’ abundance and the damage it causes?

Foxes, deer, hares, and mink cause damage primarily through their feeding behaviour. What, and how much an animal eats is subject to many influences including the quality and quantity of alternative foods, competition with other group members, competition with other species, and individual dietary requirements and preferences. These factors mean that change in a species’ abundance will not necessarily translate into a pro rata change in damage.

In North America, some studies, but not all, have found that sheep predation was a function of coyote density (Knowlton, 1999). There is some evidence that sheep predation by coyotes is restricted to a particular sector of the population. Sacks et al. (1999) found that breeding coyotes whose territories contained sheep were the principal predators of sheep, and further that predation was reduced only when territorial breeders known to kill sheep were removed. Comparisons between the North American coyote literature and UK fox literature must be made with caution because coyotes are capable of killing adult sheep, and of hunting these co-operatively.

Perhaps because of the difficulty of estimating fox abundance and of holding all other variables constant, there has been no formal attempt to relate livestock losses to fox abundance. Rowley (1970) presented evidence that lamb killing by foxes may be habitual in certain individuals, giving rise to serious losses locally. There are no grounds for speculating how fox density and sheep density might influence the risk of such behaviour developing, except that reduction of average fox age might lead to reduced occurrence of any acquired behaviour patterns. Knowlton (1999) showed that besides reducing population density, intensive culling substantially reduced the age profile of coyote populations. A similar effect would be expected from fox culling, although Heydon & Reynolds (2000a) did not detect one in heavily culled and moderately culled regional fox populations, possibly because their samples sizes were too small.

Severe damage by deer to woodland and agricultural crops is widely recognized as being associated with high population densities, but recent evidence suggests that there is no simple linear relationship (Gill, 1992; Kay, 1993; Putman, 1994, 1996b; Reimoser & Gossow, 1996; Nahlik, 1995; Tilghman, 1989). Factors such as proximity of the ‘target’ crop to cover and availability of alternative forage may have a greater influence on damage than density (Putman & Kjellander, in press).

For most deer species, damage levels tend to remain low - and relatively constant - until population density passes a certain threshold (which is itself highly variable between areas). Beyond this threshold, impact suddenly and dramatically increases. The existence of just such a curvilinear relationship between damage and density has been shown for white-tailed deer damage to forest regeneration (Tilghman, 1989), although it is usually difficult to demonstrate clearly in the field because of variation in other site factors. The actual extent of damage sustained would seem to be determined by a complex interplay of density with other factors such as forage diversity and quality, landscape and habitat structure, and climate, as well as the particular type of crop affected, its distance from cover, size of planted area, and distance of the vulnerable crop from alternative preferred forages. In emphasising the significance of those latter factors, Reimoser & Gossow (1996) suggest that levels of deer damage to forestry or agricultural crops relate not simply to deer density per se but to the effective balance between food-independent ‘attraction factors’ for deer (e.g. woodland edge) and the natural food supply. Where habitat structure is very attractive to deer yet the natural food supply is sparse, more damage may be anticipated than where the ‘attractiveness’ of an area is low in relation to the forage availability.

2.1.1.c. How realistic are estimates of damage?

Whether or not a species really causes the damage it is accused of is central to the validity of any control programme. If the species is not, in fact, the cause of the damage, then a programme to control its population will be misplaced at best, and counterproductive at worst, as resources will be directed away from the real source of the problem. Accurate estimates of damage are vital to a key series of questions: ‘What are the relative costs of damage and control?’ (section 5.3); ‘What is the potential cost of damage if control is discontinued in the long term?’; and crucially, ‘What population level is required to maintain effective damage reduction?’. At present data are generally insufficient to answer these questions accurately, but we address some of the issues for foxes in sections 2.2.1.c.i and 2.2.1.f.ii, and for deer in section 2.3.2.b.i.

2.2. Why control fox populations?

Fox population control in rural Britain is attempted by a number of disparate interest groups. These include both communally organized groups (fox hunts and fox control societies or clubs), and individuals (professional pest controllers, gamekeepers, wildlife reserve wardens, farmers and landowners). For many of these people, control by culling is perceived as a means of reducing predation by foxes on domestic livestock and poultry, reared and wild game and on other wildlife. However, foxes are also culled for sport, and sometimes to control the spread of disease (Macdonald & Johnson, 1996; Reynolds, 1998). Culling foxes for their pelts, once important (Macdonald & Carr, 1981), effectively ceased during the 1980s (Harris et al., 1995).

In the UK, anyone may kill or capture a fox by a legal method, but they must have the authority to be on the land to do this, otherwise they commit a trespass (an armed trespass if carrying a firearm). Only the landowner or tenant farmer (for sporting rights this depends on the tenancy agreement) is in the position to grant this basic authority. Hence, whatever the personal motivations of those who actually carry out fox control (e.g. gamekeepers, hunts, pest controllers), the motivation of landowners or tenant farmers to cause or allow fox control is paramount.

The conclusions that fox numbers in some regions are suppressed by deliberate culling (Heydon & Reynolds, 2000a), and in most regions have shown strong changes through time (Reynolds et al., unpublished), severely complicate any appraisal of the reasons that motivate fox culling. They mean that any investigation of the impact of foxes on human interests is specific to the fox density prevailing in that region and at that time.

2.2.1. Why do farmers seek to control fox populations?

Here and elsewhere in this report, our use of the term ‘farmer’ encompasses landowners, owner-farmers, and tenant farmers, and implies livestock and arable farming. While we recognise that control for reasons of game conservation might be an important consideration for many among this group, for simplicity we deal with it as game management in a separate section (section 2.2.1.f.ii).

2.2.1.a. Data and approach

Our data in this section come from three main sources:

We detail the approach used to collect and analyse the GC and WildCRU data, because they are used widely throughout this report.

2.2.1.a.i. The Game Conservancy Trust’s ‘Three-region study’

During 1995 to 1998, The Game Conservancy Trust undertook their ‘Three-region study’ to determine the impact of culling - by all the interest groups involved - on fox numbers across large regions, the size of a whole county. Previous studies of fox culling had considered individual methods only, and on either a very local scale or on a national scale that ignored regional variations.

The three large regions - mid-Wales, the east Midlands, and west Norfolk (Figure 2-1) - were chosen to illustrate a range of landscapes, land-use and fox culling traditions, rather than to be representative of Britain as a whole. In the hills and valleys of mid-Wales, sheep farming is the primary motivation for fox culling. Fox density is low and most culling involves the use of hounds and terriers. The east Midlands is an area of mixed agriculture and land-use. It has a mixed regime of fox culling, but hunting with hounds and mounted followers holds centre place. In west Norfolk, game conservation is the commonest motivation for fox culling, carried out by professional gamekeepers. The flat landscape and low fox density are well suited to culling with rifle and spotlamp.

Figure 2-1 The three large regions used in the Game Conservancy Trust's 'Three regions' study.

The study used three principal sources of data: a questionnaire survey to all farmers, a field survey of fox density, and post mortems of dead foxes. The questionnaire aimed to determine numbers culled, reasons for culling, aims, and methods used. After posting a questionnaire to every farm property, the authors checked for bias by telephoning a random sample of the non-respondents. In all, data were obtained from an unbiased 51% of farm properties, giving excellent representation. (Opinion polls and other studies have typically covered less than 1% of farms in their survey areas.). Cull data were also obtained directly from communally organised culling groups, such as foxhunts, and gun-packs.

2.2.1.a.ii. The Wildlife Conservation Research Unit (WildCRU) surveys of farmers’ attitudes to wildlife

The WildCRU holds data from farmer questionnaire surveys, carried out in 1981, 1992, 1995 and 1998. In this chapter, we use data from three of these (1981, 1992 and 1995).

The 1981 questionnaire comprised 130 questions soliciting information concerning diverse aspects of farming practices relevant to wildlife. Among these questions were a series asking about the damage that farmers attributed to foxes, and their attitudes to the humaneness (section 6.1.2.c) and effectiveness (section 5.2.1) of different control methods. These data are reanalysed here to address the specific issues raised in this report.

The 1981 questionnaire was dispatched to 2,288 farmers, and 859 (37.5%) responded in ‘hunt Countries’ in nine regions of England. The region referred to as ‘West Country’ comprised Devon and Cornwall. The regions were selected after examination of distribution maps in Coppock’s (1976) Agricultural Atlas of England and Wales. Within England, each of the major agrarian regions was represented. The questionnaires were distributed by post, together with a pre-paid reply envelope, and a letter stressing our non-partisan position on countryside controversies. The questionnaire was designed after extensive consultation and pilot studies (along with much discussion) with farmers, and with the assistance of questionnaire experts. The questionnaire provided data for several projects, making it effectively impossible for farmers to anticipate how we would analyse their answers. A full account of these results is given in Macdonald & Johnson (1996) and Macdonald & Johnson (in press).

The second WildCRU survey of farmers used in this chapter was carried out in 1992. This was devised primarily to assess farmers’ perceptions of the mole as a pest, and included questions concerning other common pests. With the aid of the National Farmers Union (NFU), 460 questionnaires were distributed to a representative sample of farmers in England, Scotland, and Wales. Of these, 157 (34%) were returned (Atkinson et al., 1994).

The third WildCRU questionnaire used in this chapter (reported in full in Baker & Macdonald, 2000) was carried out in Wiltshire in the summer of 1995. In a postal questionnaire, 220 Wiltshire farmers were asked about foxes and foxhunting on their farms. These included all 120 tenant farmers of Wiltshire County Council (who farmed >5.0 hectares) and a random sample of 100 other farmers. Questions covered topics such as the perceived pest status of foxes on farms, the number of foxes killed on farms, and the methods used; 101 questionnaires (45.9%) were returned.

2.2.1.b. Do farmers consider the fox a pest?

Questionnaire surveys indicate that most farmers consider the fox to be a pest in general, although far fewer consider the fox a pest on their own land. In Wiltshire in 1995, the majority of farmers (64%) believed foxes should be controlled ‘everywhere’ (Baker & Macdonald, 2000), but only around a third (32%) included the fox on a list of animals that they considered a pest on their own farm. In the WildCRU’s 1992 survey, 57% of 157 farmers said the fox was a ‘pest’, while in the PSL survey (PSL, 1995), 23% of farmers perceived the fox to be a serious problem, 56% a moderate or slight problem, and 21% no problem. In the 1981 WildCRU survey, 74% of a national sample of farmers thought foxes should be controlled in the country and 71% in towns.


Figure 2-2 Perceptions of foxes as a problem by different categories of sheep farmers. Data from Produce Studies Limited (1995). n = 831.

There are clear differences between types of farming enterprise, leading to regional differences in the perception of the fox as a pest. In the PSL survey only 4% of sheep farmers perceived foxes as no problem, compared with 35% of arable farmers, and 18% of dairy and other livestock farmers (Figure 2-2). Similarly, the regional distribution of responses reflected land-use, with 16% of farmers in sheep-dominated Wales seeing foxes as presenting no problem, compared with 35% in the more arable East Anglia/Midlands region.

How do farmers rank the pest status of foxes compared with other species? Table 2-3 summarises the results of a number of different studies, including Packer & Birks’ (1999) survey of gamekeeper’s attitudes. In the WildCRU’s 1981 survey of farmers, rabbits were most frequently cited as a causing damage, and foxes were listed fourth most frequently by 33% (Macdonald, 1984; Atkinson et al., 1994). In the WildCRU’s 1992 survey, fewer farmers considered the fox to be a pest than the rat, magpie, or mole (Atkinson et al., 1994). In this survey 11% of farmers rated the fox their worst pest, 9% the second worst, and 10% third worst. In Wiltshire in 1995, farmers were asked to produce a list of the animals that they considered pests on their farm, in descending order of estimated financial damage caused. Around a third of farmers included the fox on their list, fewer than included rabbits (58%), or badgers (42%). Each species listed was assigned a score, according to the position it occupied in the farmer’s list. Scores were averaged for each animal, and the pests then ranked according to the resultant scores attached to each. Overall, farmers ranked the fox as their third worst pest by this measure, after rabbits and badgers respectively, as did farmers of non-dairy stock. The 1998 Quantocks Deer Management and Conservation Group questionnaire (Langbein, 1998; section 2.3.1.a) also asked those farmers which considered deer to be a pest to rank them alongside other species, including foxes, from 1 (most damaging) to 6 (least damaging).

Bearing in mind that farmers’ perceptions may not accurately reflect reality (section 2.2.1.c.i), what damage do farmers ascribe to foxes? Lamb, poultry and piglet predation are the most frequently cited types of damage caused by foxes.

2.2.1.c. How much loss of lambs is attributed to foxes?

Wherever livestock farming takes place around the world, predation by canids (wolves, coyotes, foxes) is a perennial and controversial complaint, and in the case of foxes, one that has generally defied quantification (Lloyd, 1980; Macdonald, 1987; McDonald et al., 1997; Saunders et al., 1997). Predation on livestock is an extremely complex issue in which breeds, stock management, predator density and individual predator behaviour all appear to influence the outcome. There seems to be no doubt that foxes are capable of taking dead, moribund and healthy lambs, but how important this is to sheep farming enterprises remains unclear.

Studies around the world have estimated the proportion of viable lambs killed by foxes at 1-30% in different circumstances (Saunders et al. 1995). In the UK, Hewson (1984, 1990) and White et al. (in press) estimated that <2% of otherwise viable lambs in various Scottish sites were killed by foxes. MAFF (1996) attributed 5% of lamb losses to ‘predator/misadventure’, which includes predation by dogs, as well as various other causes. However, the logistical and scientific problems associated with studies of lamb losses (section 2.2.1.c.i) mean that for the most part we must rely on the farmer’s judgement as to the extent of damage caused by foxes, even though this is likely to be an overestimate.

A total of 649 respondents to the WildCRU’s 1981 survey kept sheep, in flocks varying in size from only two to several thousand. Of these, 54% said they had lost lambs to foxes at some time. There was regional variation [3]: the proportion of sheep farmers claiming a loss in the previous year was highest in the North and lowest in the Midlands & East (Table 2-4). When only flocks larger than 100 were considered, the proportion of farmers suffering losses in the past year was highest in the West Country (73%), and remained lowest in the Midlands & East (49%).

Sheep farmers in the 1981 survey were asked about the absolute numbers of lambs they thought they had lost to foxes in the three years before the study. On average, a farmer estimated that he had lost about two lambs annually (mean=1.74). This varied regionally: the mean in the Midlands & East was 0.75, compared with 4.45 in the West Country, where flocks were largest. As a proportion of the flock, reported losses were also highest in the West Country, with considerable regional differences (Table 2-4).

In their ‘Three-region study’ Heydon & Reynolds (2000a) reported that 24-61% of sheep farmers, depending on region, had experienced lamb losses during the preceding 12 months that they attributed to foxes (Table 2-4). However, the losses reported amounted to only a small percentage of all lambs, in line with an earlier study in an upland area of western Scotland (Hewson, 1984). Maximum values for any single farmer were 5-15%, depending on region. The pattern of lamb losses among regions did not mirror fox abundance, but more likely reflected the vulnerability of lambs under the regionally diverse lambing practices (see also section 2.2.1.c.i).

In their review of available evidence, McDonald et al. (1997) concluded that losses of lambs to foxes were insignificant compared with losses due to other mortality. Improved husbandry would therefore give rise to greater productivity increases than would fox control. While this is probably true, it overlooks three points that are important for the sheep farmer. First, irrespective of other losses, one can nevertheless ask whether prevention of fox predation on lambs is in itself cost-effective (see section 5.3). Second, measured or perceived levels of loss may already be reduced because of fox culling. Thus MAFF "does not consider foxes to be a significant factor in lamb mortality nationally, but it should be stressed that this is against a background of widespread fox control by farmers" (MAFF, 1993). The only study that has aimed to measure lamb losses in the absence of fox control (Hewson, 1990; section 2.2.1.c.i), provided weak evidence that culling foxes did not influence lamb losses. Third, for many upland and marginal upland areas improved husbandry may not be an option, or may not be a cost-effective option.

2.2.1.c.i. How realistic are estimates of fox predation on lambs?

The WildCRU’s 1981 survey, in which just over half of 649 sheep farmers claimed to have lost lambs, demonstrates the problems of relying on farmers’ perceptions of losses. When asked what kind of evidence had implicated the fox in their reported lamb losses, a high proportion of farmers said that foxes had been seen in the area (83%). Just under a half (46%) said they had seen dead lambs at fox earths, and 39% said that at some stage in their lives they had seen a fox attacking a lamb. This figure also varied between the regions [5], and tended to be highest in the southern counties: for example, 53% of Sussex farmers said they had seen such an attack while only 18% of Yorkshire sheep farmers said they had done so.

For predation on lambs, a key investigatory problem is the difficulty of monitoring events in a widely dispersed flock under low husbandry supervision. Lamb remains following predation or scavenging are usually obvious, but identification of the predator is not always easy or reliable. In assessing fox predation, it is crucial to distinguish lambs that were alive when taken, from those that were dead or likely to die, but the diagnostic field signs are difficult to interpret and easily missed. Indeed, the whole lamb may be missing, its removal by predators or scavengers passing unsuspected by shepherds, a fact not appreciated until recently when ultrasound scanning made it possible to forecast the number of lambs that will be born (Reeves, cited in Saunders et al., 1995). These diagnostic difficulties affect the observations of both scientist and farmer.

The difficulty of correct diagnosis suggests that manipulative experiments, as have been carried out for deer (see below) would be a better approach, allowing comparison of flocks exposed to predation with others where predators are removed or excluded. In such an experiment, comparison of the number of lambs weaned between the two types of flock treatment would give a direct measure of the impact of predation on genuinely viable lambs. Unfortunately, foxes are rarely the only predator, and whether predators are removed by culling or exclusion, it is rarely possible to contrive an experimental treatment that does not affect more than one predator species. Distinguishing the role of individual predators then relies on field signs, just as in non-experimental studies. If, alternatively, only one predator species is removed or excluded, compensatory predation by the remaining species will cause the impact of the missing one to be underestimated. The problems with interpreting the effects of compensation are dealt with further in section 7.1.1.

Hewson (1990) attempted to address the influence of fox culling by estimating lamb losses in a 70km² area where foxes were not culled during the three-year study period. Although this study has never been submitted to a refereed scientific journal, it has been quoted as key evidence that lamb predation is not reduced by fox culling (e.g. McDonald et al., 1997; LACS submission to the Inquiry, p. 26), hence it is important to consider its value as evidence.

There are three major flaws with the study. First, Hewson’s report refers to two estates, only one of which controlled foxes, but quantitative data on lamb losses and fox density were not presented from this site; the work was not, therefore, a controlled experiment in any accepted sense. Second, there was no measure of the fox population before, during, or after the study, on either estate. Four individuals were radio-tracked and two earths located, suggesting that the no-culling estate was big enough to hold 2-3 fox territories. Hence, at best, the study considered lamb predation by only nine foxes each year. Third, lambing on both estates was carried out on enclosed ground close to the farm, where supervision was intensive, and predation would be expected to be lower than on the open hill. Ewes were returned to the hill with their lambs already 3-5 days old, by which time twinned or orphaned lambs would have been fostered onto ewes that had lost their own lambs, avoiding much of the previously identified risk of predation (Saunders et al. 1995). Quantification of lamb losses after their return to the hill depended on the researcher regularly searching 70km² of difficult terrain for evidence, which meant there was no real possibility that losses could be accurately quantified or changes between years measured.

Other problems include doubts over the actual lack of culling on the intended ‘no culling’ estate; the method used to indicate fox numbers were limited by food supply (just one female was judged barren from external examination); and conflicting statements about the extent to which carrion was available and used. Overall, we consider the study to be scientifically weak, and not to allow the strong conclusions drawn by Hewson and by LACS.

Among three large regions of England and Wales, Heydon & Reynolds (2000a) reported a pattern of reported lamb losses that appeared to reflect the vulnerability of lambs under the regionally diverse lambing practices, rather than fox density. Thus, losses were most commonly reported in mid-Wales, where much of the lambing happens on unenclosed hill ground with minimal shepherding. In west Norfolk and the west Midlands, most lambing takes place either indoors, or out of doors under intensive supervision. Nevertheless, for all three regions the effect of having a gamekeeper was to halve reported lamb losses, suggesting that intensive local culling of foxes had a marked impact on perceived or actual lamb predation.

2.2.1.d. How much loss of poultry is attributed to foxes?

Among farmers with free-range poultry (excluding large commercial flocks) surveyed by Heydon & Reynolds (2000a) 49-78% reported losses in the preceding 12 months (depending on region – Table 2-5). For poultry, the regional incidence of losses (% of flocks affected) mirrored fox abundance, so that west Norfolk had the fewest occurrences, the east Midlands had the most, and mid-Wales was intermediate. It is noteworthy that, among the three regions, large-scale commercial free-range poultry units occurred only in Norfolk. Such operations may be feasible there only because of the low regional density of foxes, but each operator also put considerable independent effort into fox control. Again, presence of a gamekeeper significantly reduced reported losses.

In 1995, more Wiltshire farmers reported that they had lost chickens (42%), than lambs (16%), game birds (11%), or other livestock, (16%; Baker & Macdonald, 2000). Chickens were generally kept on a non-commercial scale.

2.2.1.e. How much loss of piglets is attributed to foxes?

In recent years there has been a rise in the number of pigs raised on outdoor units (from 5% of the national breeding sow herd in 1987 to a predicted 40% in 2000), and there is a widespread belief among outdoor pig farmers that foxes take piglets, agitate sows and thus increase mortality from overlays, and transmit disease. There have been no scientific studies of the extent of these problems, and we must therefore rely on farmers’ perceptions and individual experiences.

In April 2000 the National Pig Association requested information on this issue from 51 outdoor pig farmers. All of the six farmers who had replied by mid-May 2000 believed that foxes posed a potential problem, and reported losses at some sites but not others. Losses were variously described as ‘persistent’, ‘occasionally severe’, and ‘small’. One farmer estimated losses exceeding 25% on occasions. A survey of outdoor sows by Cambac JMA Research in 1993/94 (privately contracted 'by various bodies') found that 20% of units reported a fox problem (summary from Dr H.J.Guise, pers.comm).

Fox-proof electric fencing can substantially reduce fox predation, but is a big financial investment, and occasional fox culling is still required as a back-up. Local Wiltshire farmers (R. and K. Shepherd, pers.comm.) handling 1700 litters of piglets annually in 1994, recorded an improvement of c.1 piglet per litter after erecting an electric netting fence (both live and dead piglets were counted). There was no such improvement in a nearby unfenced area. The difference of one piglet per litter amounted to 30% of the profit margin. Another farmer with 830 sows in 1995 reported piglet losses attributed to foxes of <1% of all piglets in late winter, rising to >5% (100 piglets/month) in late summer (Reynolds, unpublished data). Annual loss attributed to foxes totalled 390 piglets out of production of 21,000, amounting to £10,315-29,815 (depending on whether piglets were fattened on or off the farm).

2.2.1.f. What reasons do farmers give for controlling fox populations?

Heydon & Reynolds (2000a), asked farmers to indicate their reasons for fox culling on their land. Not surprisingly, these reflected variation in land-use between regions. Thus, 94% of farmers in mid-Wales cited protection of livestock, but only 28% did so in predominantly arable west Norfolk. Conversely, while only 29% in mid-Wales cited protection of game, game interests motivated 75% of west Norfolk farmers (a more detailed exploration of why foxes are controlled for game management is presented below). Most farmers (62-75%, depending on region) gave two or more reasons for culling foxes (Figure 2-3).

Local fox culling for the benefit of neighbours was widely cited by farmers in all regions (35-54%), but only 6.5% of this group gave ‘good neighbour policy’ as their sole reason for culling. Sport, too, was usually cited in combination with other reasons. In mid-Wales, not a single farmer cited sport alone. In the east Midlands 57% cited sport, but only 14% cited sport alone. No farmer claimed sale of pelts as a reason for culling.

Figure 2-3 Reasons cited by farmers for killing foxes in Wales, Midlands and East Anglia. The rank order of reasons like game, livestock, and sport reflected land-use in the three regions. Importantly, most culling was done for two or more reasons. The figures on the bottom line refer to the proportion of farmers citing more than one reason for fox culling.

In the 1981 WildCRU questionnaire, farmers who believed that foxes should be controlled were asked which of four options they thought justified the control. These were ‘may spread disease’, ‘kill domestic stock’, ‘kill game birds’, and ‘are too numerous’. Overall, the domestic stock reason was most commonly opted for (70%), with ‘too numerous’ next most frequent with 67%. Disease was opted for by 47%, and ‘game’ by 46%. Only protection of game and disease showed significant regional variation [6] (Table 2-2). The variation in protection of game followed regional game-shooting patterns, while disease was cited markedly less frequently in the West Country compared with other regions.

When asked to select from the same five options, almost three-quarters (73%) of Wiltshire farmers said they were too numerous, and over half said because they kill domestic stock. Killing game birds was cited by 34% of farmers, and the spread of disease by 30%; 7% cited ‘other’ reasons. In comparison with the 1981 survey, relatively more Wiltshire farmers who believed foxes needed to be controlled, selected each of ‘kill domestic stock’, ‘kill game birds’, and ‘spread disease’, while relatively fewer believed foxes ‘too numerous’ (Macdonald, 1984).

2.2.1.f.i. Why do some farmers not cull?

In 1981, two-thirds (67%) of farmers surveyed reported carrying out no fox control (they were not asked about culling carried out by others on their land), but among those who reported fox damage, only 24% carried out no control. There was considerable regional variation in proportions of non-culling farmers [7], as would be expected from patterns of land-use. The lowest proportions were recorded in Yorkshire and Shropshire, and the highest in Sussex, Oxfordshire, and Dorset (Table 2-1). There was no evidence that a farmer’s participation in hunting or approval of the active conservation of foxes had any effect on their tendency to control foxes [8].

In the Midlands and Norfolk in 1996, 12% of farmers did not cull foxes or allow fox culling (Heydon & Reynolds, 2000a). Among these, the commonest reason cited was lack of necessity, followed by a perceived benefit from the presence of foxes. Half of the non-culling farmers stated that they would consider culling in the future if the fox population increased. Only one fifth of non-culling farmers (i.e. 2.5% of all farmers) stated that they did not approve of fox culling.

2.2.1.f.ii. How realistic are farmer’s worries about foxes as carriers of disease?

Of 92 farmers who responded to a questionnaire in Wiltshire in 1995, almost a third (30%) believed that foxes should be controlled because they spread disease (Baker & Macdonald, 2000; section 2.2.1.a.ii). These 19 farmers were asked to list the diseases about which they were concerned. Although they identified fourteen diseases, the majority of those suggested present no risk whatsoever with modern veterinary science (Table 2-6). Despite listing as potential fox-borne hazards various inappropriate pathogens, farmers did not cite Bordetella bronchisceptica (Kennel cough) or Toxocara canis as causes for concern, although foxes might transmit these. Other pathogens known to be carried by foxes include Neospora caninum, para-TB (responsible for Johne’s disease) and parvo-virus. The risks of transmission to livestock and domestic animals are unknown. Although the sample was very small, it nonetheless illustrates that if poorly informed, farmers might overestimate some risks associated with foxes, while other possible threats go unrecognised.

2.2.2. Why do game managers seek to control fox populations?

Predator control has been a feature of game management in both upland and lowland Britain since shooting estates first arose in the early 18^th century (Tapper, 1992). Predator control to protect nesting game birds is a skilled and labour-intensive job reliant on the employment of a gamekeeper, and is regarded as essential to ensure the farmer has a viable shoot.(Tapper, 2000).

An important distinction that must be drawn is between the management of a wild game bird population to allow a harvest, and reliance - wholly or in part - on hand-reared game birds (see section 3.6.1.a.ii). Arguably, the former most closely conforms to the sustainable ‘wise use’ of a natural resource, and for many game managers is preferable if it can be attained (Tapper, 2000). However, the trend in recent decades has been towards the latter, partly because wild game birds on farmland have fared badly under modern intensive agricultural practices (e.g. Potts, 1980). This has implications for the timing of predator control (section 3.2.6).

The commercial aspects of shooting also need some clarification. Shooting is a saleable commodity. Although not all farmers sell or lease their shooting, many do. The provision of game shooting is a popular secondary land-use for many land properties; for some upland estates, grouse shooting is actually the primary land-use. This can and does lead to increased demands to intensify the production of game birds, and in recent years bodies like the British Association for Shooting and Conservation (BASC) and GCT have had to address the question of excess in commercially driven shoots. Tapper (2000) and co-authors argue a view on this that places limits based on environmental and ecological impacts of intensive shoots.

2.2.2.a. Do game managers consider the fox a pest?

In game management, man is directly in competition with the fox, since both are game predators. An important difference is that while man withholds predation during the game breeding season in order to benefit from population increase, predation by foxes is particularly intense during this same period because the prey are more vulnerable and because foxes are themselves breeding at this time. Several studies (reviewed by Lindström, 1994; Reynolds & Tapper, 1995a) have indicated that adult foxes selectively provision their cubs with birds and mammals in the size range 0.3-3.5kg (e.g. rabbits, hares, game birds). Because of this breeding season predation, foxes can have a substantial impact on wild and reared game bird and hare population dynamics.

In a questionnaire survey of 66 Welsh and Midlands gamekeepers, the fox was ranked the most serious predator of game (Packer & Birks, 1999; Table 2-3). Similarly, in a 1994 survey of gamekeepers carried out by BASC, 96 % of 1624 keepers who responded said that foxes were present on their land and needed to be controlled. Control was considered necessary to ‘ensure that damage to game, wildlife and livestock was reduced or kept at acceptable levels’ (BASC submission to the Inquiry).

A survey conducted in 1997 by Bristol University for the National Gamekeeper’s Organisation (NGO submission to the Inquiry) asked gamekeepers to rate how serious a problem foxes posed to them. Of the 203 who replied, 63% stated that the fox was a major pest, and only 6% considered it a minor pest. None said it posed no problem to them.

2.2.2.b. What impact do foxes have on wild game bird populations?

The fox is a key predator in many ecosystems (Reynolds & Tapper, 1996), particularly the heavily altered man-made ecosystems of Western Europe. Usually, evidence of the importance of any single predator species is circumstantial: a study of a prey species – usually investigating poor productivity or population decline – finds high predation levels. Studies of this kind identifying foxes as a major predator exist for all British game birds. Accumulated evidence of this kind can be very persuasive that high predation is associated with population decline. Unfortunately, it remains ambiguous: the predation could be the cause of decline, or it could be merely symptomatic of some other cause.

Unambiguous evidence about the impact of foxes on wild game bird populations was specifically sought through research by the GCT. This evidence is of two kinds. First, an experimental study of wild grey prtridges on Salisbury Plain (Tapper et al., 1996), in which a suite of common predators, including foxes, were intensively culled on a 6km² ‘removal’ site for three years. A similar site nearby had no predator removal and acted as a ‘comparison’ area. After three years, ‘predator removal’ and ‘comparison’ treatments were switched between the two areas. Throughout the six years of the experiment, and for one year before and after, partridge numbers and productivity were monitored on both areas. The results were conclusive: under the predator removal regime, autumn partridge densities increased by 75% year-on-year, finishing 3.5 times greater at the end of three years, compared with the non-removal comparison regime. These improved autumn numbers also carried over to build up spring breeding stocks, which increased 25% annually, to finish 2.6 times greater after three years.

The Salisbury Plain experiment provided decisive evidence of the importance of predators for game, but did not indicate which predator species contributed most to the effect. After predator removal ceased, it was shown by radio-tagging partridges that foxes were by far the most important of the suite of predators removed, accounting for 81% of all annual losses and 91% of all breeding season predation losses (Reynolds et al., 1992; Reynolds, unpubl.). By radio-tagging foxes to establish territory and group size (Reynolds, unpubl.), it was also shown that their annual food requirements (600-1000kg/km²) far exceeded what the partridge population could supply (spring density 1-3kg/km²). In fact, game birds as a whole formed less than 1% of fox diet in this area, although foxes killed 20% of breeding females (Reynolds, unpubl.), accounting for most of the experimental effect (there was a further 10% non-predation loss of breeding females). Thus, partridges were certainly not an important determinant of fox density, but foxes were very important for partridges. On Salisbury Plain, the food resources that allowed foxes to maintain such high numbers relative to partridges were rabbits and hares, which together made up 85% of fox diet.

The second type of evidence is not experimental, but is equally important. It is very rare for any field study to quantify predators, predator diet and prey numbers simultaneously. However, if foxes really are important to game, the number of game they eat must make a significant dent in the game population. Reynolds & Tapper (1995b) undertook this research in a mixed agriculture area in northeast Dorset with unremarkable populations of both game and foxes, and showed that the proportion of game birds taken by the resident foxes was substantial compared with the number of birds (24-100% of wild-breeding game birds), their productivity, and the shootable surplus.

For wild red grouse, a comparison between good and bad moors in Scotland and northern England (Hudson, 1992) demonstrated that louping ill was overwhelmingly the most important factor influencing grouse density and productivity; game keeping (incorporating both predator control and habitat management) and climate were also significant explanatory variables. Where predator control is sufficient to allow high grouse populations (and only there), threadworm parasites cause cyclical year-to-year fluctuations in grouse numbers. Low-density grouse moors appear to suffer severe predation pressure from raptors and foxes that kept numbers low. Of all these influential factors, fox density, habitat management, and threadworm burdens can currently be influenced by practical management measures.

McDonald et al. (1997), state that ‘diseases are the main factor controlling grouse numbers’, but overlook Hudson’s (1992) point that on high density grouse moors the impact of predation has already been minimised by intensive predator culling.

2.2.2.c. What damage to released pheasants is attributed to foxes?

Due to the success of rearing (section 3.6.1.a.ii), the pheasant has become an increasingly important game bird during the 20^th century. At the turn of the century, it probably comprised 15% of all game birds shot in the UK, but by the 1980s had increased to >55%, roughly equivalent to 12 million birds per year (Tapper, 1992). It has been estimated that around 20 million birds are released annually. The wild pheasant population has probably declined during the same period, and now comprises only an estimated 10% of all pheasants shot (Tapper, 1999). If these estimates are reliable (they involve extrapolation from small samples), they imply that 40% of released birds (c. 8 million) die annually. An unknown proportion of these will be killed by foxes.

2.2.2.d. What impact do foxes have on wild brown hare populations?

In north-east Dorset, Reynolds & Tapper (1995b) found that in a population of hares (which was not subject to culling by man), foxes effectively wiped out the annual reproductive gains of the population. They did this by taking a biomass of 47-87 kg/km² annually from a hare population with a pre-breeding biomass of 42kg. (This is obviously only feasible because the hare population reproduces during the year.) This research supplemented many earlier studies of hares that had provided circumstantial evidence of the importance of foxes as hare predators (Jensen et al., 1970; Nyholm, 1971; Spittler, 1974; Angerbjörn, 1977; Frylestam, 1980; Häkkinen & Jokinen, 1981; Lindlöf & Lemnell, 1981; Pegel, 1986; Hearn et al., 1987; Dannell & Hörnfeldt, 1987; Angerbjörn, 1989; Small & Keith, 1992; Lindström, 1994), as well as manipulative experiments (Marcström et al., 1989; Tapper et al., 1993). The foxes in northeast Dorset had a more diverse diet than those on Salisbury Plain, and hares comprised just 11% by weight of their food intake (Figure 2-4). As with partridges on Salisbury Plain, the fox’s influence on the hare population was far greater than the hare’s importance to the fox.

Figure 2-4 Diet of foxes in north-east Dorset

In their submission to the Inquiry, IFAW argue that these studies are atypical, and that ‘there is no evidence’ that foxes are having an impact on prey species in most other situations. In fact, field studies of the type described above have not been conducted elsewhere: there is no evidence because it has not been sought. Furthermore, the grounds for regarding predator and prey densities at the Dorset study site as unremarkable are well supported (Reynolds & Tapper, 1995a,b).

2.2.3. Why do conservationists seek to control fox populations?

Foxes eat species of conservation concern throughout the world, particularly in countries where they are not native (e.g. Kinnear et al., 1988). In Britain, they are controlled by conservationists primarily because of their predation on some ground-nesting wild bird populations (Côté & Sutherland, 1995), particularly those that are already fragmented. Well-documented examples include red grouse Lagopus lagopus in the Stiperstones Reserve (Macdonald et al., 1999), and terns, Sterna spp, at various reserves e.g. Sands of Forvie (Patterson, 1977), North Denes reserve, Great Yarmouth (Paul Lewis, RSPB. pers. comm.).

Predation by foxes has become an increasing problem on coastal bird reserves in Norfolk, and many reserve-owning conservation bodies have carried out or commissioned fox culling to safeguard vulnerable bird populations (Reynolds, 1998a). In most cases, this has been a reluctant and controversial policy. Coastal reserves in north Norfolk may previously have been protected against foxes by a cordon of shooting estates and by the intensity of regional fox control. Foxes were apparently absent in west Norfolk earlier this century, probably the result of the very large workforce of gamekeepers (there were 1202 in Norfolk in 1911). Although today there is only one tenth that number of gamekeepers, the proportion of land with professional gamekeepers remains very high compared with the rest of Britain.

In many cases, foxes pose a threat to wild birds only for a short period, usually the nesting season (e.g. Birkhead & Nettleship, 1995). It is uncertain how significant nest losses are to bird species that are typically much longer-lived than game birds. Predator removal often has a large, positive effect on hatching success and post-breeding densities of the target bird species, but no impact on spring breeding numbers (reviewed by Côté & Sutherland, 1997). Since increasing breeding numbers is the usual conservation goal, Côté & Sutherland conclude that predator removal did not generally fulfil the conservationists’ aims. They add that this could be attributable either to inherent characteristics of the birds’ population dynamics, or to ineffective predator removal. To these comments, we add that reserve-based conservation measures that enhance productivity may be annulled by emigration and events occurring away from the breeding grounds.

The RSPB only consider fox control where it can be achieved legally and humanely, and then only if there is a risk of serious damage to conservation, agriculture or human health. Where fox control is necessary, the policy of the RSPB is to use trained staff to trap or shoot.

2.2.4. Why do foresters seek to control fox populations?

The Forestry Commission has for many years undertaken fox culling through its own wildlife rangers, and financially supported local fox destruction groups. In 1992, this ‘good neighbour’ policy was revised following a review of existing literature, changing the emphasis from extensive and systematic fox culling to providing a quick and effective response to lamb killing by foxes. The policy shift was applauded by conservationists, but criticised by sheep farmers and game managers (Chadwick et al., 1997).

A fox culling policy on the part of forestry bodies has no component of self interest except where shooting is leased out. As in arable farming, foxes may have a benefit to forestry interests as predators of pest mammals (see section 5.3.3). Although the impact of foxes as predators on populations of rabbits and voles is uncertain (Trout & Tittensor, 1989; Dyczkowski & Yalden, 1998), both are significant pests of young forest plantations. Foxes are also predators of roe deer fawns (Lindström, 1994), another wildlife species whose control in forestry enterprises costs large sums of money (see below).

2.3. Why control deer populations?

A wide range of people, with various motivations, have an interest in the control of deer populations and damage. Deer damage has long been a major concern to forestry (Gill, 1992a,b), and with recent expansions in deer ranges and abundances (Staines et al., 1998; section 0), there is increasing concern about damage to agricultural crops and pastures (Scotland: Mitchell et al., 1977; Callander & McKenzie, 1991; England and Wales: Putman & Moore, 1998; Packer et al., 1998; Staines et al., 1998). Furthermore, such concern now extends to natural tree regeneration and ground flora in semi-natural woodlands (Mitchell & Kirby, 1990; Cooke, 1994), with potential impacts on the diversity of birds, small mammals, and invertebrates (Stowe, 1987; Hill, 1985; Petley-Jones, 1995). Culling is also used to stop the spread of exotic deer species. On the other hand, however, deer are widely believed to be valued by the public for their aesthetic appeal (Exmoor NP submission to the Inquiry; Scottish Natural Heritage, 1994), form a valuable natural and renewable resource as venison, and generate stalking revenue.

Many interest groups perceive population control by culling as one of several means available to reduce damage levels (see also section 3.6.1.b). British deer species are also culled for sport (deer stalking and hunting with hounds; Whitehead, 1964; Hamilton, 1907), and stalking can generate significant revenues through venison, stalking fees and trophies. These various (competing) objectives often need to be balanced against one another within a single multipurpose landholding or area covered by a co-operative Deer Management Group, and the primary motivation for seeking to control deer will vary according to local land use.

2.3.1. Why do farmers seek to control deer populations?

2.3.1.a. Data and approach

While quantitative data on losses to farm crops because of deer damage remain limited (Putman, 1986; Putman & Kjellander, in press), a number of recent questionnaire-based surveys have helped to assess the perceived significance of such damage in England and Wales. We draw heavily on two of these.

2.3.1.a.i. ADAS questionnaire survey

The most comprehensive questionnaire survey thus far, conducted by ADAS during 1995, was distributed by post to 3322 landholders representing four main user groups (agriculture, forestry, nature conservation, recreation) to assess their attitudes and practices towards lowland deer (Doney & Packer, 1998; Packer et al., 1998). Sampling extended to four geographic regions (Somerset and Gloucestershire; Essex and Suffolk; Northamptonshire; lowland Yorkshire), chosen to represent a range of agricultural uses and landscapes across the known distribution of lowland deer. The questionnaire asked for details of land area; crop or habitat types present; species of deer present; frequency and number of deer seen; nature and estimated cost of damage caused by deer; and methods and perceived effectiveness of damage prevention.

A total of 1546 (47%) valid responses were received, with high response rates across all sectors (ranging from 82% from conservationists to 36% from forestry holdings). The reliability of responses and perceived damage levels reported was assessed in 1997 by a ground-truthing study encompassing 25 agricultural holdings, focussing on damage to cereals and farm woodlands (Doney & Packer, 1998; Packer et al., 1998).

2.3.1.a.ii. Quantocks Deer Management and Conservation Group questionnaire survey

The Quantocks Deer Management and Conservation Group circulated a similar questionnaire during March 1998 to all known landholders within or close to the boundary of the Quantocks AONB, one of the few remaining areas where deer are subject to hunting to hounds (Langbein, 1998). Its purpose was to obtain better information on the occurrence and current management of deer on local estates, including the landholder’s own estimates of the financial losses associated with deer damage, and their views on how the deer should best be managed in future. The questionnaire was sent to 165 addresses of local farms or known landowners, eliciting 68 responses (43%); the actual percentage of landholders replying was, however, much higher, as many responses covered two or more holdings farmed by the same person. Together, the respondents accounted for 14,649ha, an area slightly larger than the AONB itself. Pasture and grass ley accounted for the largest proportion of the land (29%) followed by cultivated land (25%) and moorland (25%).

2.3.1.b. Do farmers consider deer a pest?

In the WildCRU’s 1981 survey of farmers (section 2.2.1.a.ii), 13% of respondents reported that deer caused damage, though there was considerable regional variation (Table 2-7). Deer were most frequently reported causing damage in Dorset.

In the more recent ADAS survey across lowland England and Wales, deer were present on the holdings of 69% of 1192 agricultural respondents, and 38% believed that deer cause significant damage (Packer et al., 1998). Most felt that agricultural damage from deer had increased (42%) or stayed the same (48%) between 1990-1995, but remains a limited and mostly localised problem (Doney & Packer, 1998.). Where deer were present, 18-51% of respondents reported damage, depending on region, and probably influenced by the species its and abundance. The majority of farmers agreed or strongly agreed that "the damage caused by deer causes significant economic loss" in Somerset (45% of 294 with deer present), and Essex and Suffolk (36% of 141), but disagreed or strongly disagreed in Northamptonshire (51% of 35) and North Yorkshire (49% of 100).

In the Quantocks AONB, where large number of red deer are present, 74% of landholders (most being farmers) considered that deer caused significant damage on their land, with most ranking deer as more damaging than rabbits, badgers, or foxes (Table 2-3). Just over half (52%) the farmers said deer were the most damaging of a list of six mammal species (Langbein, 1998a).

2.3.1.c. What damage to farming is attributed to deer and how significant is it?

Cereals were perceived to be affected by deer by almost half (44% of 822 replies with deer) of the farmers responding to the 1995 ADAS survey (Doney & Packer, 1998), followed by damage to trees (29%), grass (6%), root crops (3%), fruit (3%), vegetables (3%), and oilseed rape (3%). For farms growing mainly cereal crops, 17% of respondents claimed no annual cost of deer damage, and 85% perceived deer damage to be £500 or less per annum.

In the Quantocks AONB, 74% of 68 landowners, holding between them 89% of the 14,649ha covered by the survey, agreed that deer caused significant economic losses (Langbein, 1998a). The median annual losses due to deer damage was also estimated at around £500 per holding (mean holding size 92ha) if including the 30% of respondents suggesting zero or <£100 damage; or £800 if restricting analysis to those reporting at least >£100 losses due to deer. As in the case of results from the wider ADAS study above, most farmers believed that cereals were affected (54%); a high proportion of landowners also stated that they had suffered damage to pasture and sown leys (41%), hedges and banks (34%) and woodland (34%).

2.3.1.c.i. How realistic are estimates of the costs of deer damage to agriculture?

Although it is well established that deer cause damage to crops and forestry, without experience deer damage is often difficult to distinguish from that caused by other mammals such as rabbits and hares. Furthermore, even where landholders may have correctly ascribed damage to deer, their estimates of the costs incurred through such damage often bear no resemblance to actual losses.

This was demonstrated recently by an ADAS survey (Doney & Packer, 1998; Packer et al., 1998; section 2.3.1.a), which found that while farmers were mostly (75-80%) accurate in reporting deer species and approximate abundance, they were generally incorrect about the economic value of damage to cereals. Farmers were as likely to underestimate the costs of damage as to exaggerate it. Actual losses due to grazing of winter wheat were assessed during follow-up ground truthing at up to 0.57 tonnes per hectare on farms which were visited regularly by roe or fallow deer, but at lower levels of grazing, a negligible economic loss, or an actual gain in yield, was recorded (Doney & Packer, 1998; Doney, 1998).

The costs of deer damage to crops is made more difficult to estimate because, over a period of months or years, plants can recover to some extent, or even benefit from grazing or browsing. For example, in Hampshire, roe deer cause substantial levels of apparent damage to cereal fields in spring, but by harvest this may often be negated through tillering and increased growth rates (Putman, 1986b).

A further complication to estimating the costs of deer damage is that deer are themselves a valuable resource. During the late 1980s, when the price of venison was low, the Red Deer Commission reported an increase in the numbers of complaints about deer on farmland (RDC, 1989 [in SNH, 1994]).

2.3.1.d. Which species of deer do farmers seek to control?

In the 1995 ADAS questionnaire survey, the perceived significance of damage levels recorded by farmers indicated very little differences between areas in which fallow, red, roe or muntjac deer were present (Packer et al., 1998), although roe deer were the species most regularly associated with damage overall (partly reflecting their wider distribution across all sample areas).

Fallow, red and roe deer were the three species most frequently associated with damage to agriculture in an earlier review (Putman & Moore, 1998), based on the frequency of unsolicited requests for advice received by ADAS between 1985-1989. Different deer species tended to be associated with different types of damage. Most reports of damage to oilseed rape or to nursery crops, garden shrubs, and top fruit involved roe deer, which were relatively rarely implicated in reports of damage to grass or cereals. Reports of damage by red deer were largely in connection with pasture, silage crops, or field cereals, while 76% of all complaints concerning damage to field cereals cited fallow deer. By comparison, little damage was reported to cereals from sika deer, muntjac or Chinese water deer (Putman, 1995; Putman & Moore, 1999).

Damage by fallow (and red) is often localised to areas where large herds, sometimes of 70-200 animals, aggregate on favoured farmland feeding grounds (Langbein, 1996). Potential damage by red deer, partly by virtue of their larger individual size, tends to be viewed most seriously and can be locally significant (Callander & McKenzie, 1991; Langbein, 1998; Putman & Langbein, 1999).

2.3.2. Why do foresters seek to control deer populations?

The habitats preferred by all six of our deer species are associated with open forest or woodland edge, and at the national scale prevention or limitation of damage to young tree plantations is probably the single most important reason for which deer control tends to be undertaken in Britain. Aside from hindering the establishment of new commercial tree plantations, deer may also cause damage to amenity trees, natural tree regeneration from seed, coppice management, and standing timber (Gill, 1992a,b).

In many parts of Scotland, where already high but still rising deer densities have been experienced throughout much of this century, establishment of commercial tree plantations without either deer population control or fencing or both has long been regarded as non-viable (Staines & Welch, 1989). By comparison, deer distribution and abundance south of the border has been relatively restricted in the past, but the significant increases noted over the last 40 years, especially of roe, fallow and muntjac deer, have made this an equally important issue in England and Wales.

2.3.2.a. Do foresters consider deer a pest?

The recent ADAS questionnaire survey by Packer et al. (1998) indicated that 57% of foresters considered that deer cause significant damage, a greater proportion than any other user group (agriculture, conservation, and recreation).

The attitude among foresters that deer cause damage, and that their numbers therefore need to be rigorously reduced, is beginning to change in favour of more positive management. National and European policies for sustainable forest management are increasingly aimed at delivery of a wide range of benefits, including biodiversity, conservation and recreation, even from those forests still managed primarily for timber production (HMSO, 1994). Persistently high numbers of deer can be detrimental to these goals, but there is increasing consensus that in many types of woodland retention of some grazing by deer or domestic stock is preferable to their total exclusion (Kirby, 1993; Hester & Miller, 1995; Putman, 1986, 1996; Kuiters et al., 1996). Deer play an important role in creating a diverse structure which benefits other species (Ratcliffe, 1998). Current Forestry Authority and Deer Initiative advice therefore recommends that: "Management should aim to maintain healthy deer populations in balance with their environment, rather than to eliminate deer from an area altogether" (Forestry Practice Advice Note 2, 1995)

2.3.2.b. What damage to forestry is attributable to deer, and how significant is it?

Potential damage by deer to trees may take various forms. Browsing (biting off buds, foliage or shoots) and bark stripping (peeling of bark to eat) tend to be most common in winter when other food sources are scarce, and during shoot elongation in spring. Fraying (using antlers to abrade and partially remove the bark from stems and branches) may occur at different times of the year depending on the deer species involved, as male deer mark their territories and clean their newly grown antlers of velvet by rubbing them on young trees (Gill, 1992a; Langbein, 1993; Forestry Authority, 1995). Browsing and bark stripping in particular can result in serious losses of young trees, whether naturally regenerated, or in plantations (Mitchell et al., 1977; Staines & Ratcliffe, 1987). The susceptibility of trees to damage is very variable between tree species, age, deer species, and frequency and type of damage.

Numerous published studies (Welch et al., 1992; Van Hees, 1996; Putman et al., 1989; Ammer, 1996; Langbein, 1997; Cooke, 1994) have experimentally quantified the effects of deer browsing on tree survival and growth form, natural regeneration, relative abundances of canopy tree species, and ground flora. In commercial sitka spruce plantations in Scotland, Welch et al. (1992) concluded that the overall effect of red deer browsing was equivalent to a check of about one year in the time taken to reach a height of 80cm. In semi-natural woodland, Putman et al. (1989) found that an area protected from fallow deer browsing had a density of 6440 saplings/ha after 14 years, in contrast to only 20/ha in an adjacent heavily browsed area. Deer (especially fallow) can significantly hamper establishment of Farm Woodlands (incentive schemes aimed at converting agricultural land to woodland). Of 74 (mainly broadleaved) farm woodland plantations in east Suffolk, 21% suffered substantial damage from fallow deer, with over 20% of the leader shoots damaged in one year; cherry and rowan were the most frequently browsed tree species (Key et al., 1998).

Deer damage (by browsing and bark-stripping) can amount to as much as 40% of the crop of coniferous plantations in the Scottish uplands (Maxwell, 1967), and average losses close to 50% of the leader shoots have been recorded even in Sitka spruce crops (Staines & Welch, 1989). Allison (1990) estimates that deer damage to forestry in Galloway, Scotland, costs £2 million per year. Aside from actual costs of damage or failure of plantings, further costs are incurred through preventative measures aimed at reducing damage, such as fencing, or employment of stalkers to reduce deer numbers. Gill (in SNH, 1994) has estimated the total cost of red and roe deer damage and control in Forestry Commission plantations in Britain to be in the region of £5 million (net of revenues from venison and stalking – c. £1.26m), equivalent to c. 7% of the total revenue generated by Forestry Commission timber sales in 1989/90.

Comprehensive reviews of the damage caused by deer to forestry, and its likely economic significance in England and Wales, are provided by Gill (1992a,b), and Putman & Moore (1998).

2.3.2.b.i. How realistic are estimates of the costs of deer damage to forestry?

Damage to forestry by deer is not regularly surveyed in Britain, there are surprisingly few published data with which to estimate the true financial costs of deer damage, and how that relates to the expense of deer population control and protective fencing. The variability in the few available estimates serves to underline the difficulties in accurately assessing such costs (Gill, 1992b; see below). In Europe, estimates of the annual cost per hectare of woodland range from <£1 for browsing by moose in Sweden (Jantz, 1982) to £85 for red and roe deer browsing in Germany (Spiedel, 1980).

As with cereals, one of the difficulties of assessing the long-term costs of damage to trees is that unless they have been killed outright, they will often recover completely. Sitka spruce affected by terminal bud damage often respond by ‘flagging’, where one of the lateral branches migrates round to take over apical dominance, sometimes leading to a net increase in height compared to undamaged individuals (Staines & Welch, 1984).

A further complication is that different tree species may have different susceptibilities to, and recovery rates from, damage. Extensive exclosure studies in the Netherlands (Van Hees et al., 1996) showed that browsing by red and roe deer had clear effects on mortality and growth rates of silver birch and oak, but less significant effects on beech. Interestingly, however, they note that there may also be an indirect ‘benefit’ of deer browsing in that grazing and browsing reduce the competition suffered from non-crop species.

2.3.2.c. Which species of deer do foresters seek to control?

Where present, all the deer species (with the possible exception of Chinese water deer which do not yet occur in significant numbers) tend to be implicated in damage to forestry (Gill, 1992a,b). In their ADAS questionnaire survey, Packer et al. (1998) found that the species of deer present (red, fallow, roe, muntjac) on each forestry holding had no effect on whether respondents perceived that damage to be of economic significance.

Fallow, sika, and red deer, as preferential grazers, tend to include rather lower proportions of woody browse in their diet than do the smaller roe and muntjac deer. Whereas this might be expected to lead to major differences in levels of tree damage, such differences are often largely negated by the greater total intake rates by the larger three species, as well as their greater tendency to form herds. The three larger deer species also pose a greater potential threat to forestry (and costs in protection) in view of the greater height to which they will browse, and hence the longer period of vulnerability until growing trees are out of reach to the deer. For establishment of Farm Woodlands, fallow and roe tend to be the most frequent species associated with damage, not least due to their wide distribution in England (and increasingly also Wales), and the tendency of fallow to build up high numbers even in agricultural areas offering only small patches of woodland for cover.

2.3.3. Why do conservationists seek to control deer populations?

Grazing and browsing by wild herbivores have always played a role in determining the structure and dynamics of natural ecological systems. It is considered ‘damage’ when the consequences are extreme and/or conflict with human interests or management objectives, but grazing and browsing by deer within conservation communities also has many positive, facilitative effects, such as preservation of ancient wood pasture and associated wildlife species (e.g. Bakker et al., 1983, 1984; Putman, 1986a, 1996b). Persistently high numbers of deer can be detrimental to tree regeneration or lead to losses of ground flora sensitive to grazing, but there is increasing consensus among conservationists that some grazing is beneficial (Kirby, 1993; Hester & Miller, 1995; Putman, 1986, 1996; Kuiters et al., 1996).

2.3.3.a. Do conservationists consider deer to be a pest, and why?

Deer were considered to cause significant damage by 34% of conservationists replying to an ADAS survey questionnaire (Packer et al., 1998; see below); this is comparable with the farmers’ response to the same question.

In a recent survey of National Nature Reserves conducted on behalf of English Nature (Putman, 1996b), questionnaires were sent to managers of 162 sites designated as National Nature Reserves throughout England. Not all sites had deer present; of the 112 site managers recording deer visiting or resident within their reserve, 45% recorded a measurable impact at some level (browsing damage to coppice, lack of regeneration, impact on ground flora). Sites that reported damage from deer were without exception woodland reserves - managers of ‘open sites’ (grasslands, meadows, heath land or fenland sites) generally regarded the presence of deer as neutral or positively advantageous in suppressing encroachment by scrub. Within woodlands, the vast majority of complaints concerned browsing damage to coppice regrowth.

Overall, only 18% of managers of reserves with deer present considered that damage sustained was sufficient to cause difficulty in meeting management objectives for the site; all of these considered that current management measures (culling, fencing of vulnerable areas) were adequate at present to reduce damage to tolerable levels. However, Putman himself points out that such figures should be interpreted with some caution, in view of differences between sites, not only in levels of damage and numbers of deer, but also in methods of damage prevention already in place. As with foxes (section 2.1.1.a), management history thus confounds the responses, and the proportion of managers reporting lack of conflict in meeting management objectives should not simply be equated to those who would find no conflict in the absence of any management (Putman, 1996, 1998).

2.3.3.b. What damage to conservation is attributed to deer?

Grasslands and lowland heaths rely on the maintenance of grazing to maintain their characteristic structure and diversity, and such communities are more likely to be at risk from reduction of grazing than from increasing deer populations. In certain situations, however, there may be a case for controlling the level of grazing, for example, where impact has risen to such a level that it conflicts with other management objectives determined for a particular site. High densities of red deer are well known to have significant impacts on the growth and regeneration of heather in many parts of the Scottish uplands (e.g. Staines et al., 1995; Clarke et al., 1995; Stewart & Hester, 1998). However, in most parts of Exmoor, the contribution to heather off-take by red deer is of fairly minor importance, and much lower than off-take from sheep grazing (Langbein, 1997).

Woodland communities are most prone to damage from over-grazing by deer and other ungulates, with particular concerns about damage to natural tree regeneration and ground flora in semi-natural woodlands, and damage from deer to regeneration of coppice coupes (Mitchell & Kirby, 1990; Putman, 1994a, 1996b; Cooke, 1994).

Within woodlands, heavy grazing pressure may have a number of distinct effects on regeneration of woodland trees, on structure and composition of field and shrub layers, and on species of the woodland floor. Where losses of mature trees through browsing damage or simply old age are complemented by virtual lack of regeneration due to depletion of the seed source or heavy browsing pressure on new seedlings, browsing mammals start to exert a substantial impact on the entire woodland structure (e.g. Peterken & Tubbs, 1965). Even light but selective browsing, taking a preponderance of preferred species such as ash or oak by comparison to less preferred birch, alder or sycamore, can lead to quite pronounced shifts in species composition of canopy trees (Gill, 1992; Putman, 1996b; Van Hees et al., 1996).

Within the Exmoor National Park, Langbein (1997) recorded significantly higher mortality and reduced growth of seedlings of oak, rowan, and beech in a number of old oak coppice woods grazed by red deer and sheep, compared to fenced plots in the same stands. Very few saplings reached heights significantly above the shrub layer in most old coppice stands studied, except those where sheep were excluded and red deer numbers were comparatively low (<5 per 100ha).

Browsing may also lead to elimination of the shrub layer, and the reduced field layer resulting from heavy grazing can reduce the abundance and diversity of invertebrates and small mammals (e.g. Hill, 1985; Putman, 1986a); this may in turn reduce diversity of raptors or mammalian predators (e.g. Tubbs, 1974, 1982; Putman, 1986; Putman et al., 1989; Petty & Avery, 1990). Other species, however, may derive positive advantage from such heavy grazing. Wood warblers, pied flycatchers, and redstarts all depend on the park-like conditions of traditional wood-pastures (Stowe, 1987; Mitchell & Kirby, 1990).

Finally, heavy grazing pressure can result in dramatic changes in the composition and relative abundance of species of the woodland floor, which may be of serious consequence if that flora itself contains rare or threatened species. Recent declines in oxslip populations in many conservation woodlands of East Anglia (e.g. Hayley Wood, Cambridgeshire, Hales Wood, Essex) have been blamed - rightly or wrongly - on the coincidental rapid increases in range and number of fallow deer throughout that region (Rackham, 1975; Tabor, 1993, 1999). Cooke (1994a,b) has reported declines in bluebell and dog’s mercury in Monk’s Wood NNR and other Cambridgeshire woodlands, associated with heavy grazing pressure from muntjac. Deer browsing more generally results in increases in some species: for example, bracken, grasses, mosses, foxglove, and ragwort; and decreases in others: for example, bramble, honeysuckle, ivy, wild rose, and holly (Gill, 1997).

2.3.3.c. Which species of deer do conservationists seek to control?

All deer species pose some threat to conservation of habitats and other wildlife if allowed to reach very high densities. In Scotland, overgrazing by red deer is of particular concern in relation to conservation and restoration of native Scots pine woods (Holloway, 1967; Beaumont et al., 1994; Palmer et al., 1998). In England and Wales, fallow, and in some regions red deer, tend to be the main species implicated in preventing natural tree regeneration, due to their tendency to build up very high densities locally (Putman, 1996; Langbein, 1997). Muntjac are implicated especially in damage to ground flora (e.g. Cooke, 1994).

In the questionnaire survey of National Nature Reserves undertaken on behalf of English Nature (Putman, 1996b), there was no statistically significant association between the type of deer species present and responses regarding ability to meet conservation objectives for the site. However, of the 25% of responses that suggested deer damage is of some concern in meeting management objectives, fallow and muntjac were the most frequently reported problem species (Putman, 1996).

2.3.4. What is the importance of deer population control for public amenity and as source of revenue for landholders?

Aside from the various needs to control of deer to prevent the damage outlined above, the presence of deer must be also be recognised as an aesthetic benefit (their very presence giving pleasure to people), and an exploitable resource, generating income for the landholder through recreational stalking, venison, and hunting.

2.3.4.a. How important are deer for public amenity?

Deer, as the largest terrestrial wild mammals remaining in Britain today, are highly regarded among the public, with people taking pleasure from seeing them in the wild or even merely knowing that they continue to thrive in our increasingly industrialised landscapes. In England, this is true in particular for many of the large traditional ‘deer forests’ such as Epping Forest, the New Forest, and Exmoor. Continued provision of deer viewing opportunities is an important aspect of published management plans for each of these three areas (Langbein, 1996; Putman & Langbein, 1999; Forestry Commission, 2000; Exmoor NP submission to the Inquiry).

The popularity of the red deer within Exmoor National Park (one of the remaining areas where hunting deer to hounds occurs) was highlighted by the results of a survey undertaken by the Park Authority (Park Life, July 1999). This showed that local residents placed red deer above any of the many other ‘special features’ that they valued most about Exmoor. The preservation of reasonable numbers of red deer on Exmoor and the Quantocks is thus of some importance to local tourism; however, for this, the total size of the deer population is likely to be of lesser importance than ensuring that good numbers remain near to those areas most frequented by the public.

2.3.4.b. How important are wild deer as a source of income?

As discussed above, each of the ‘quarry’ species considered in this report tends to be controlled partly (often primarily) because they are perceived to cause significant (if highly variable) damage to the interests of farmers, foresters, gamekeepers or conservation. The quarry species itself may also be of some economic value to the landholder, for example as saleable meat, pelts and for commercial shooting. Unlike the three other quarry species (hares, foxes, mink), the direct revenue potential from deer is relatively high.

Direct income from deer may be derived from a ‘harvest’ of venison, stalking fees or lets, and trophy fees. In view of the considerable size of some deer species, even the annual income through venison sales (c. £1.50-2.50/kg carcass weight) arising from any ‘necessary’ culling may contribute significantly towards covering any direct costs of deer control. Although variable, total revenues from venison in Scotland may exceed £3-6 million (SNH, 1994); no overall data are available for England and Wales, but here incomes are also likely to exceed £1 million per year.

Over and above the carcass value of culled deer, some landowners are also able to sell stalking opportunities on their land. This may be based on fees per outing (often accompanied by a professional deer ranger) or numbers of deer shot, usually with an additional premium charged for any stag or buck, depending on the quality of the (antler) trophy. Charges for accompanied stalking start at around £75 per day but vary widely between areas and between deer species, with a still wider range of additional charges for shooting a stag carrying ‘medal’ quality antler trophies (usually based on criteria laid down by the Conseil International de la Chasses). Alternatively, some landowners may simply let the deer stalking rights to a third party for an annual rent.

The sustainable exploitation of deer as a renewable natural resource can thus be an important motivation for some landholders to maintain deer populations on their land at particular levels. The extent to which each of the differing sources of income from deer can be exploited varies widely between sites, depending on the primary landuse objective of a given estate, the size of the holding or co-operative deer management area, and the habitats and deer species present. For a given estate the optimum size and structure of the deer population will also differ according to whether the main focus of exploitation is on venison sales, stalking lets, or trophies. Optimising return from venison sales usually requires maintenance of a fairly high population density and a high female to male sex ratio (c. >3:1), with the emphasis on ‘harvesting’ the surplus of young males (de Nahlik, 1992). However, maintenance of very high densities and heavily female biased sex-ratios is often inappropriate for management of wild deer populations, due to increasing impact on other landuses, and reductions in health and ‘trophy’ quality if resources become limiting. It may also become increasingly difficult to achieve adequate annual culls to keep population numbers under control.

Commercial exploitation of wild deer, especially through deer stalking, is widespread throughout Europe (Gill, 1990). While roe and fallow deer stalking is widely available throughout much of England, in Scotland the emphasis has long been on red deer stalking in the Highlands. There, it has been a major factor with regard to land management and the development of the large red deer populations. The average number of stags shot per annum has commonly been used as the main yard stick in assessing the capital value of the large Scottish shooting estates or ‘deer forests’ (e.g. Callander & McKenzie, 1991).

2.4 Why control hare populations?

As with deer, issues surrounding the control of brown and mountain hares are complicated by the fact that they are simultaneously considered a pest (by farmers, foresters and, for mountain hares only, gamekeepers), are a game species culled for their meat and for sport, and are of conservation concern.

Once considered abundant, the brown hare has declined substantially since the early 1960s to an estimated British pre-breeding population of 817,500 in the 1990s (Hutchings & Harris, 1996). This pattern of decline has also been seen in much of Europe during the latter half of the twentieth century. In Britain, the brown hare is subject to a Species Action Plan intended to "maintain and expand existing populations, doubling spring numbers in Britain by 2010" (Anon, 1995). However, the species is by no means rare (section 10.3.2): in Britain, brown hares are in fact the most numerous of all the species considered in this report.

In England, the mountain hare is found only in the Peak District and is extinct in Wales. Of the ten species covered in this report, the mountain hare is the only one subject to special protection under international law (Appendix III of the Berne Convention 1979; section 11.4). Control of mountain hares is nonetheless permitted.

2.4.1. Why seek to control populations of brown hares?

Hares eat crops such as oilseed rape and turnip, and particularly eat grasses and cereals. In addition, hares can eat high value market garden crops, and will often browse and kill newly planted young trees and shrubs. Some of this damage can be of economic significance to individual growers.

However, since the second world war, although commonly still regarded as a minor pest (Mellanby, 1981) the general agricultural damage hares inflict has never been serious enough to warrant the MAFF funded research efforts that went into the control of species like the brown rat, rabbit or woodpigeon. In livestock districts today, hares are not numerous and are rarely considered a pest. In arable areas, high numbers of hares on winter corn are considered damaging by most cereal farmers, and regular winter culls by shooting are undertaken where this occurs. Although the economic loss to cereals has not been calculated in Britain, hares spend much of the late winter feeding on winter cereals (Tapper & Barnes, 1986).

In some areas, such as the South Downs, landowners may also seek to reduce the abundance of hares on their land to deter illegal coursing. We have no data on the extent to which this occurs, but certainly it is a common perception, and the AMHB (submission to the Inquiry) state that the police support a policy of hare culling to remove the threat of poachers.

2.4.2. Why seek to control populations of mountain hares?

In some areas mountain hares may cause localised damage to forestry and moorland vegetation. They may also compete with livestock and grouse. Mountain hares are sometimes shot as a quarry. We are not aware of any evidence that the English population causes damage or is controlled, except in the context of game-shooting.

2.5. Why control mink populations?

American mink were the first carnivore to be introduced to Britain since the domestic cat, over a thousand years ago (Yalden, 1999). Mink are popularly regarded as voracious predators that roam the countryside devouring all wildlife, stock or pets in their path (Dunstone, 1993). The reality is, of course, much more prosaic, but as a recently introduced predator, the American mink brings with it some unique problems.

2.5.1. Why do farmers seek to control mink populations?

Mink are controlled by farmers because of their perceived impact on native wildlife, fish, game birds and livestock (particularly poultry, but sometimes lambs), and have a reputation for the surplus killing of confined animals (Dunstone, 1993).

In 1996, researchers at WildCRU distributed a questionnaire regarding mink to 40 farms along the River Thames in west Oxfordshire. Of the 32 farmers who replied, 29 (91%) considered the mink to be a pest (Strachan, unpublished data). The most frequently cited impact of mink was on waterbirds (37.5%), followed by water voles (28%), and fish (22%). Birds and general wildlife were cited by 19% of farmers, and only 12.5% cited damage to poultry.

Harrison & Symes (1989) examined the different types of damage attributed to mink from 195 reports to MAFF in South-west England by farmers during the periods 1961-70 and 1985-86. The killing of poultry and domestic or ornamental waterfowl accounted for 60% of the incidents reported. Although losses are undoubtedly important to individual landowners, where they are taken, poultry form only a small component of the mink’s diet: in one study of 685 scats, chickens and game birds formed less than 1% of the diet (Birks, 1986).

On the Scottish islands of Harris and Lewis, mink are believed to have made the keeping of outdoor poultry almost impossible. Before mink colonised these islands, 90% of the 4,000 registered crofts are estimated to have kept poultry; nowadays fewer than 10% continue to do so. Based on an average flock size of 10 birds, the net annual cost of this constraint to the crofting economy on the two islands is estimated to be £586,000 (M.C. Swan, unpubl. analysis).

2.5.2. Why do game managers seek to control mink populations?

A 1996 survey of gamekeepers in Wales and the West Midlands found that the 66 respondents ranked the mink as the third most serious predator of game, after foxes and feral cats (Packer & Birks, 1999; Table 2-3).

Mink are perceived to have an impact on game birds, ducks, fish, and native wildlife such as water voles. As with poultry rearing, penned pheasants and partridges have been killed after mink have gained access to the rearing enclosures with surplus killing frequently cited. Wild stocks of game birds and recently released game birds are also believed to be at risk, reportedly predated as adult birds or as chicks or eggs during the breeding season. However, we are not aware of any data quantifying the extent to which mink predate on gamebirds.

2.5.3. Why do conservationists seek to control mink populations?

Like farmers, conservationists believe that there is good evidence that mink have an impact on water voles, waterbirds (especially ducks, grebes, coots and moorhens), vulnerable ground-nesting birds such as the corncrake, lapwing or redshank, sea bird colonies (especially terns, gulls, eiders, black guillimot, shearwaters and puffins), and native white-clawed crayfish. There is an extensive and growing scientific literature regarding the impact of mink on wildlife (see reviews by Birks, 1990; Dunstone, 1993; Macdonald & Strachan, 1999; Craik, 1997, 1998). The UK Species Action Plan for the water vole encourages the control of mink as a conservation tool to protect key populations of the vole throughout its range.

2.5.3.a. What is the impact of mink on the water vole?

One of the strongest conservation arguments for controlling mink in Britain is to protect vulnerable water vole populations. The water vole has declined by an estimated 88% of its total population between 1989/90 and 1998 (Strachan et al., 2000). Following pioneering work by Woodroffe et al. (1990) on the Yorkshire Moors, there is increasingly powerful evidence that predation by mink, in association with habitat degradation and fragmentation, is a causal factor in the vole’s decline (Macdonald & Strachan, 1999; Strachan et al., 1998; Lambin et al., 1998; Woodroffe et al., 1990).

A series of studies (Figure 2-5; Macdonald & Strachan, 1999) in the Thames catchment area indicate that the vole decline there has continued for over twenty years, since the mink first arrived in the 1970s. However, as recently as 1990, mink were uncommon in the catchment and water voles were still found in three quarters of the sites they had occupied earlier this century. A 1995 survey revealed a rapid increase in mink numbers and a catastrophic decline in the number of sites at which water voles persisted. No site was found on the Thames at which the two species occurred together. In 1991, Halliwell & Macdonald (1996) live-trapped American mink along 20km stretches of four rivers in the Thames catchment and found a negative correlation between the numbers of mink caught and the numbers of water vole signs.

Studies in the Thames, therefore, lead to the conclusion that, at least when mink numbers increase, their impact on water voles can be very rapid, leading to widespread losses within only five years. Indeed, evidence from both the Windrush and the Soar suggest that within one breeding season, resident mink may greatly reduce numbers of water voles, and within only two years eradicate them locally. Female mink, which hunt close to their nursery dens, may have a particularly rapid impact on local water vole colonies (Strachan et al., 1998).

In a study of mink diet on 11 rivers in Derbyshire, Leicestershire, Staffordshire and Nottinghamshire over 1993-94, Strachan & Jefferies (1996) showed that water vole was the single most important species in the diet of colonizing mink. Indeed, in the May-June sample they comprised up to 32% of the volume of undigested prey remains. Similarly, analysis of 863 scats collected in 1993/94 along the River Soar demonstrated that water voles can be an important component of the diet of mink colonizing a river, but thereafter their importance declines as they are depleted (Strachan et al., 1998).

Figure 2-5 The changing distribution of mink and water voles in the Thames catchment.

2.5.3.b. What is the impact of mink on birds?

Mink have a serious impact on the breeding success of colonial ground nesting sea birds. They are believed to have caused terns colonies along the Scottish west coast between Mallaig and West Loch Tarbert to halve in the 11 years 1987-1998 (Craik, 1998).Craik (1993, 1995, 1997) describes surplus killing by single mink arriving at tern nesting islands before chicks have been fledged.

Less clear is the effect of mink predation on riparian bird species. Early studies led to contradictory conclusions (Lever, 1978; Linn & Chanin, 1978). As mink colonized different parts of Britain, considerable concern was expressed about the possible disappearance of moorhens (Smith, 1988). More recently, Halliwell & Macdonald (1996) found no significant correlations between mink and moorhen censuses from eight 10km sections of four lowland British rivers. In the Upper Thames, taking habitat into account, coot but not moorhen abundance was negatively related to the presence of mink, and fewer coot chicks were raised by pairs living in mink-occupied zones (Macdonald & Strachan, 1999; Ferreras & Macdonald, 1999). These results accord with faecal analyses which suggest that the mink’s impact on moorhens during the breeding season was less than on the coots. The reason for this difference may be that coots nest on the water surface among rushes and reeds, while moorhens nest in shrubs and trees.

In the Upper Thames, coots and moorhens together constituted 10% of the biomass ingested by mink, based on the analysis of 115 scats. Other water birds constituted 2% of ingested biomass. Higher proportions of waterfowl in mink diet have been related to high waterfowl densities and to the scarcity of other prey such as fish and crayfish (Chanin & Linn, 1980; Eberhardt & Sargeant, 1977).

2.5.3.c. What is the impact of mink on other species?

American mink arrived in Britain at a time when two other members of the guild of semi-aquatic predators were seriously reduced (Birks, 1989; Jefferies, 1989). These two, the amphibious otter and the more terrestrial polecat, are currently increasing and spreading into areas now occupied by American mink. The implications of this recolonization are likely to affect directly population processes of all three mustelids, and thereby indirectly to affect their prey.

2.5.4. Why do fisheries managers seek to control mink populations?

Mink are perceived to have an impact on salmonid fisheries, trout farms, salmon farms, and commercial carp ponds (especially exotic koi carp). Concern is frequently expressed over the effect of mink on angling interests and particularly salmonid stocks (e.g. Lever, 1985).

In order to obtain a dispassionate opinion on the impact of mink on fish stocks, Birks (1990) carried out a questionnaire survey of the fisheries and conservation staff of the then newly formed National Rivers Authority. All ten NRA Regions responded, and nine of them stated that mink were not currently a threat to fish stocks (except perhaps in the artificial situation created by fish farms). Yorkshire Region felt that mink were a problem throughout the river catchments, and Southern Region expressed concern about the possible impact of mink on sea trout during periods of low flow, when this species lies in small pools in rivers and streams.

In its native range, fish are the second most important prey in the mink’s diet, after mammals. The Cyprinidae is the most diverse family of fish inhabiting fresh waters in North America, and are most frequently consumed by mink. In Europe, too, cyprinids are one of the mink’s main prey in fresh water ecosystems, particularly roach and chub. Salmonids, eels, stickleback and bullheads have also been reported as important fish prey for mink (Akande, 1972; Chanin & Linn, 1980; Gerell, 1967a; Skirnisson, 1980; Wise et al., 1981). The representation of various fish families in the diet of British mink (as gleaned from a synthesis of 22 papers on mink diet; Macdonald & Strachan, 1998) is summarized in Figure 2-6. Mink generally take small fish: 97% of salmonids taken were < 25cm (Cuthbert, 1979).

Figure 2-6 Proportion of fish families in the diet of mink in Britain.

2.6. Conclusions

Section 2 Footnotes:

1. c²_[1] = 225.26, P < 0.0001

2. logistic regression, c²_[1]= 13.2, P = 0.003

3. c²_[8] = 37.8, P <0.001

4. A single farmers reported losing 8/8 lambs

5. c²_[8] = 18.5, P = 0.02

6. c²_[8] = 23.7, P = 0.003 and c2[8]= 58.1, P < 0.001 respectively

7. c²_[8] = 15.7, P = 0.04

8. logistic regressions, all P > 0.05

3. What methods are used to control populations of foxes, deer, hares, and mink in England and Wales?

3.1. Introduction

Although a wide range of methods is potentially suited to controlling populations of foxes, deer, hares and mink, those used in the UK are restricted by legislation (see Appendix 2). Effectively, there are four legal methods in operation in 2000: shooting, hunting with dogs, trapping, and snaring. These are all lethal methods – that is, they aim to control the population by killing animals. In recent decades there has been an increasing drive to develop and implement non-lethal methods to achieve population control (Baker & Macdonald, 1999).

While most attempts to control populations involve deliberate culling, not all culling is done with the intention of controlling populations. Other aims include the harvest of natural products like meat and fur, and sport. Furthermore, different interest groups may have different aims for the same culling effort. Thus, culling may be instigated by farmers to control a population and the damage it causes; but culling may actually be carried out by a group of people who suffer no losses due to damage, but who find sufficient interest in the culling itself to spend their own time and money carrying it out.

Yet another layer of complexity is added by the fact that any one interest group will use a range of methods, but the emphasis varies between groups, even within the same area. There is also a degree of fluidity between methods, making strict definitions unrealistic and analysis difficult. For example, a single Welsh hunt might operate as a mounted hunt in lowland areas, a foot pack in open uplands, and as a gun pack in plantations, and in some or all of these situations might also use terriers, rifles or shotguns. Welsh packs, with their flexible modus operandi illustrate the difficulties in distinguishing clearly between methods that use dogs to kill, chase, locate or flush foxes (or indeed, other quarry).

In this chapter, our focus is population control through deliberate culling by man, for any reason. We include, therefore, sports that make no claim to control, such as hare coursing, as well as sports that do make a claim to control, such as mink and foxhunting. Our goal here is to examine which methods are used to cull each species, which interest groups use these methods, when and where. The numbers of animals killed by each method, and whether it effectively achieves management aims, is explored in Chapter 5.

There is an important distinction to be made between methods used to control a population, and methods used to control the impact of a population. In Chapter 2 we covered some of the reasons why people seek to control populations of foxes, deer, hares and mink. One major reason common to all these species is that they can cause damage to human interests, for example by eating crops or killing livestock. In these situations, it must be remembered that population control is not the only effective method, nor necessarily the most effective method, for reducing the impact of damage. Some of the alternatives to culling are explored at the end of this chapter (section 3.6).

3.2. What methods are used to control fox populations?

The methods used successfully to take adult foxes fall into two categories. Nighttime methods require either a lamp or image intensifier to make the fox visible to the operator, or else traps and snares that work in his absence. Daytime methods require the fox to be flushed out of cover. Additionally, the foxes’ need for an underground earth to shelter cubs while they are very young creates a vulnerable period during which it is easier to locate both cubs and adults. Because adults have to provide solid food to their cubs, their activity may also extend further into twilight hours during spring and summer, creating more opportunities for shooting without spotlight or image intensifier.

Most fox culling is done without conscious selectivity, as it is difficult to distinguish males from females, or young from old, at any distance. From April to August cubs can be distinguished from adults, but this too becomes increasingly difficult as autumn approaches. Apart from obvious generalisations (e.g. a preponderance of females among adults shot at cubbing earths; a preponderance of young foxes among those caught in cage traps or shot after attracting with a squeak), few data are available to support the commonly held belief that different culling methods address different sectors of the fox population. Systematic exploration of this problem is difficult, because biologists too must use the same capture methods to study the population.

3.2.1. Data and approach

For fox culling in the context of game management, we used four sets of data held by The Game Conservancy Trust (GCT): the National Game-Bag Census; the Gamekeeper Fox Culling Methods Survey; the Fox Monitoring Scheme; and the Joint BASC/GCT Snares Trial. Our data on hunting with dogs come primarily from questionnaire surveys held by the Wildlife Conservation Research Unit (WildCRU), and from a questionnaire survey to all hunts organised by the Masters of Fox-Hounds Association (MFHA) and the Campaign for Hunting and analysed by the GCT. The most recent data available come from a January 2000 survey by Produce Studies Ltd (submission to the Inquiry). These data sets, and their limitations in terms of reliability and representativeness, are detailed below. Further data on the use of various methods were obtained by Reynolds & Heydon (2000a) from landowners and farmers in three large regions of Britain (mid-Wales, east Midlands, west Norfolk; see section 2.2.1.a.i), and by Baker & Macdonald (2000) in Wiltshire (see section 2.2.1.a.ii). Additional data were used are detailed in the text.

3.2.1.a. National Game-Bag Census (NGC), 1961-ongoing

This is a historical data base detailing ‘bag’ records for game species and predators for an annual sample of c. 500 shooting estates in all parts of the UK (Tapper, 1992). In general, predators (including foxes) were recorded only since 1961, although for a few estates records are available from much further back. Participation is voluntary, hence the sample of estates is self-selecting, and its geographical distribution varies with time as individual estates enter the scheme or drop out. No measure of culling effort is recorded. As a result of these peculiarities, great care is required to interpret the data correctly.

3.2.1.b. Gamekeeper Fox Culling Methods Survey, 1992-93

This survey was intended to fill a gap in our knowledge of the extent to which gamekeepers use different methods to achieve the ‘bags’ indicated by the NGC. The aim was to recruit a sample of c. 100 gamekeepers from around the UK to keep a daily record for 12 months of the effort and success of each culling method used. Recruitment effort was crudely stratified by region, in that regional advisors of Game Conservancy Ltd were asked to identify likely participants and estates that were either typical of the region or, conversely, unusual and of particular interest. Based on these recommendations, 83 estates were approached, and keepers on 47 of these agreed to take part. Head-keepers also involved a further 58 beat-keepers. After 12 months of recording, 65 record books were ultimately returned of which four were incorrectly filled out and were therefore unusable. The final sample therefore consisted of 61 gamekeepers from 36 shooting estates. Each participant was interviewed by telephone following return of the completed record books.

3.2.1.c. Fox Monitoring Scheme, 1994-ongoing

Because the above study suggested a way of monitoring relative fox numbers regionally and over a long time period, it was continued on a simplified basis in successive years. An annual sample of 40-60 self-selecting gamekeepers and amateur or professional fox controllers has contributed records for 6 years. As with the NGC, individual participants have come and gone. With such a small sample this means that the geographical origin of records has shifted with time, requiring very careful interpretation.

3.2.1.d. Joint BASC/GCT snares trial

This study was designed as an experimental comparison between a new type of neck snare and existing snaring practice. Sixty-four gamekeepers were recruited to record their snaring effort and success in a pro forma daily diary. Participants were not selected as a random sample, but rather as a sample of gamekeepers who already used snares to an appreciable extent in their work. Each snare location was described in detail, and the period for which the snare was set was recorded. At each daily inspection, any captures (foxes or non-target species) were recorded, also whether captives were alive or dead. The overall conclusion of this experiment was that the new variety of snare under trial performed no differently from those normally used by the participants. Consequently, data from these snare types may be combined and used as a single source of information on snaring.

3.2.1.e. WildCRU Questionnaire Surveys of MFHA Masters

The WildCRU hold data from a number of questionnaire surveys of farmers and MFHA hunt Masters. Data for farmers is detailed in section 2.2.1.a.ii.

In 1980, the Masters of all 206 MFHA hunts were circulated with a questionnaire soliciting information on the size of their hunts in terms of numbers and type of participants (Macondald & Johnson, 1996). They were also asked to supply the numbers of foxes killed and numbers of days hunting for the previous decade. A total of 80 Masters provided responses, and 60 yielded usable data on tallies. In the three years after this, Masters were asked to complete daily diaries recording the number of foxes moved, hunted and killed on each days hunting. Between 10 and 25 Masters provided data for each year. In 1994, these data were updated by circulating a further questionnaire, again via the MFHA, to hunt Masters. Of 60 circulated, 39 were returned (Macondald & Johnson, 1996). Foot-pack data were also obtained in 1994 (Macondald & Johnson, 1996), when the Welsh Farmers Fox Control Association supplied records for 18 of their (then) 28 member packs in 1993/1994 and 1994/1995 (at that time incomplete). They also recorded the number of lambs reported lost and the number of farms where losses were said to occur.

In 1995, Masters were interviewed from each of the thirteen recognised packs of foxhounds operating in Wiltshire (Baker & Macdonald, 2000). Face-to-face interviews lasting around 2 hours were conducted with Masters from each of the eight packs responsible for most of the hunting in Wiltshire, and the remainder were interviewed by telephone. Interviews provided information on hunt activity, and Masters from 7 of the 8 most active packs provided kill data from their diaries.

3.2.1.f. MFHA/Campaign for Hunting

The MFHA/CH data derive from questionnaires completed by 162 hunts, covering the 1990/91 to 1995/96 seasons. Data include the number of meets under early- and main-season rules, the number of foxes caught above and below ground, estimates of the percentage of foxes found that are killed, and the percentage of foxes found (moved) that are run to ground. For most hunts, the area of the hunt country, and patches of ground that were not hunted (‘no try’ areas) and where permission had been sought but denied (‘no go’ areas) were recorded. The percentage of foxes moved that were killed above ground, and the percentage of foxes run to ground that are dug out and killed by each hunt can be calculated from these data. Analysis was by The Game Conservancy Trust (Reynolds, unpublished).

3.2.1.g. Produce Studies Limited

In January 2000, Produce Studies Limited (PSL) sent a standard pre-coded self completion questionnaire to a Master or Chairman of all the 302 registered hunts for all species via their Masters’ Association (PSL submission to the Inquiry). Almost all (94%) responded. 178 of the replies related to foxhunting. The questionnaire covered five regions: South West including Avon, Gloucestershire, Wiltshire and Dorset; South, from Hampshire to Kent and Oxon to North London; Midlands & East Anglia to Norfolk, Lincolnshire, Nottinghamshire, Derby and Cheshire; North to the Scottish Border; Wales.

3.1.2.h. Are the data reliable and representative?

In none of the data sets could data be verified, and each may be prone to reporting and recording inaccuracies similar to those detected by Heydon & Reynolds (2000a). Inevitably, certain parameters are easier to measure than others; for example, the WildCRU survey (Macdonald & Johnson, 1996) asked Masters to note the numbers of foxes killed (distinguishing kills above and below ground), hunted (but not necessarily killed) and moved (but not necessarily hunted) – these measures are progressively more difficult to make accurately.

In the Culling Methods Survey and the Fox Monitoring Scheme (FMS), recruitment of volunteers aimed to minimise the risk of falsification. First, it was made clear that the task of recording data would be quite tedious; hence only those who were enthusiastic to help with research would take part. Second, absolute confidentiality was assured for each gamekeeper; data would not be revealed to his employer, his neighbours, the local hunt, etc., removing any temptation to falsify records to ‘keep up appearances’. Third, because of the involved nature of the pro-forma diaries, falsification would have been difficult to achieve convincingly. Fourth, because data were recorded daily, there was no risk of memory lapses as is the case with questionnaire studies referring to the previous twelve months. Finally, in the case of the FMS, a proportion of participants were willing to save body parts of foxes (lower jaw, uterus) for later analysis, verifying at least the number of foxes killed by these people.

None of the datasets could be gathered in a way that adequately took account of known regional variation in fox abundance, terrain and other environmental circumstances that might influence the effectiveness of different culling methods. Because of the ways in which participants were recruited, none of these surveys is representative of farmers, shooting estates or gamekeepers in the UK as a whole. In particular, despite obvious regional variation in circumstances, samples were not stratified by region. Only the NGC has a sample large enough to be broken down in this way. The WildCRU MFHA surveys were restricted to registered hunts, which may not be representative of other types of hunt, although this is partly remedied by the data from the Welsh Hill Farmers packs. The Fox Monitoring Scheme set out to involve people who would were interested enough to contribute data over a long time, and is biased towards operators who primarily use lamping to cull foxes.

3.2.2. Hunting with dogs

Foxes are primarily hunted with two very different types of dogs: large, fast hounds with enhanced trailing abilities that can pursue and catch a fox above ground; and very small terriers to locate and corner the fox that has gone to ground. Foxhunting mostly combines the use of the two breeds of dog to provide a unique daytime method for culling a nocturnal animal. However, some groups, particularly in hill country, use terriers without hounds; gun-packs use hounds without terriers to flush foxes from cover; and MFHA-registered hunts now do not dig out foxes (using terriers) unless specifically requested by the landowner. A few hunts – mostly fell packs – do no digging out on hunt days. Foxes are also hunted with lurchers and other ‘long-dogs’ on an ad hoc basis.

3.2.2.a. Mounted hunting with hounds

Mounted hunting with hounds is the most widespread form of fox hunting in England, and is most closely associated with the public perception of ‘foxhunting’. There are currently 184 foxhound packs registered with the MFHA, with 14,720 meets annually; five harrier packs also hunt the fox on horseback (MFHA submission to the Inquiry).

The structure and mode of action of each hunt is reasonably constant. The MFHA submission to the Inquiry gives a full account of this. In essence the huntsmen, their hounds and followers gather at the meet, and move off to a location where the huntsman has decided to searc for ('draw') a fox. The hounds then fan out, searching for fox scent. If a hound finds a fox, the hound barks ('speaks') and if disturbed but not instantly caught, the fox flees, and the chase commences.

Each registered hunt has an exclusive ‘country’ allotted by the MFHA, within which it negotiates permission to hunt from individual landowners. Historically, the purpose of the country was to avoid border disputes between neighbouring hunts – it implies nothing about rights of access.

3.2.2.a.i. Where, when and by whom are foxes controlled using mounted hunting with hounds?

Between them, hunt countries occupy a total of 145,000-164,000 km² (Macdonald & Johnson, 1996; PSL submission to the Inquiry) and in the south of England they are effectively contiguous. About 23% of England and Wales is not included within any registered hunt country (MFHA/CH data), although there may be hunting by unregistered packs there. Furthermore, within each hunt country there are ‘no-try’ areas where no attempt is made to hunt or to seek permission, usually because the land is unsuitable for hunting. On average, such no-try areas amount to 21% of the land included in hunt countries, and permission to hunt is sought but denied on a further 1-2% of allotted hunt countries (MFHA/CH data; PSL submission to the Inquiry). Excluding all of these un-hunted areas, the proportion of England and Wales actually hunted over is 61%.

Table 3-1 presents a regional breakdown of the structure of an average hunt country in 1981, in terms of their mean area, and the mean number of farmers and gamekeepers within them.

The hunting season is usually September to March and can be split into two. In September, ‘cub hunting’ (now frequently referred to as ‘early season’ or ‘autumn’ hunting) takes place. This gives experience to young hounds and is said to promote dispersal (MFHA submission to the Inquiry). Between November and March, main season hunting occurs. Hunts meet 2-4 times a week, depending on the size of their country. The WildCRU data from 1981-1993 show that during an average season there are 71 hunting days. About two-thirds of these consist of main season hunting (44 days, on average), and a third consist of cub hunting (26 days, on average). There is wide variation within these averages, however, with full seasons lasting 33-125 days.

Clearly, mounted foxhunting affects about two-thirds of England and Wales, but how widely is it regarded as a population control method by farmers? One poll (PSL, 1995) reported that hunting occurred on 35% of Welsh farms and 26% of farms in the Midlands and East Anglia combined. Another found that the number of farmers allowing hunting was 63%, 61% and 56% for television network areas covering Wales, the Midlands and East Anglia respectively (NOP, 1996). These values compare with Heydon & Reynold’s (2000a) estimates of 41% of farms in Wales, 82% in the Midlands and 50% in East Anglia. In 1993/94 and 1994/95 respectively, 41% and 48% of farmers in Wiltshire reported that hunting had occurred on their land (Baker & Macdonald, 2000).

However, a farmer allowing hunting on his land does not necessarily see it as part of a strategy for fox control (Baker & Macdonald, 2000): only 31% of farmers in Wiltshire encouraged the hunt. Hunting was not allowed by 6% of Wiltshire farmers, leaving a further 63% who ‘tolerated’ or ‘discouraged’ it. The high proportion of tenant farmers, and the retention of sporting rights (Parkes & Thornley, 1994) by the Council may create this complex situation in Wiltshire. In 1995, the sporting rights on 88 (73%) of the 120 Council farms had been retained by Wiltshire County Farms Estate, and foxhunting was automatically permitted regardless of the farmer’s wishes. It is possible therefore that some Council tenants tolerated hunting on their land only because of their landlord’s policy. For the same reasons, however, tenant farmers without sporting rights would have had no motivation to control fox numbers for the purposes of game management. In Wiltshire, farmers reported that the hunt had visited their farm for an average of 1.8 days in each of the two hunting seasons. There was no statistical relationship between the number of kills made on a farm, and the farmer’s attitude to hunting [9].

In a questionnaire survey of National Gamekeepers Organisation members in 1997 just under half of 203 respondents (employed on shooting estates) cited hunting (48%) as one of the methods used on their ground to cull foxes (NGO submission to the Inquiry).

3.2.2.b. Foot packs

In some (predominantly upland) areas, foxhunts operate as foot packs, with an October-March season. These function in a very different way to the mounted hunts, without horses, and making more use of shot-guns and terriers. There is considerable overlap between foot packs and gun packs (see below); in Wales, a foot pack might operate very much like a gun pack when in plantations (Federation of Welsh Packs, submission to the Inquiry). In general, foot packs use larger numbers of hounds than gun-packs. Some hunts also operate as both foot packs and mounted hunts. Some packs are also part of Fox Destruction Societies, which claim payment (currently £25 per day) for fox control from Forest Enterprise Wales (formerly the Forestry Commission).

In 1993/94, each Welsh foot-pack covered an average of 57.4km² (a considerably smaller area than a mounted hunt), and in total covered 2678km².

In Cumbria, there are 6 foot packs registered with the Central Committee of Fell Packs, and another three are affiliated (CCFP submission to the Inquiry). These make only limited use of shooting.

3.2.2.c. Digging with terriers

Digging out with terriers is widely practised by mounted and foot hunts and other communal fox control groups, as well as by small groups, or individuals such as gamekeepers. As with other methods involving dogs, terrier work has an enthusiastic following for its own intrinsic interest. When used by the hunt, terriers are usually entered into earths where hounds have marked a fox to ground. In other circumstances terriers are entered into holes where fieldsigns and local knowledge suggest an active earth, where tracks show that a fox has entered, or where there is evidence of a cubbing earth (section 3.2.3.c). Terriers can be entered speculatively into any earth, pile of straw bales, very thick cover, etc. to locate and either bolt, corner, or kill the fox. Field-craft skills are critical, as it is illegal to enter a terrier or dig in any place in regular use by badgers. This may be a very limiting condition in hill areas where rock piles are commonly used as shelter or cubbing earths, as neither species leaves much surface evidence in these situations.

Where foxes are dug out, a radio-transmitter collar on the terrier aids economical and rapid digging. Once exposed by digging, the fox must be dispatched humanely, for which a .22 rim-fire pistol or rifle firing a free round is recommended (Harris, 1985). In some cases the fox may be killed underground by the terrier – this is particularly likely with fox cubs. Foxes are usually prevented from bolting by lightly blocking tunnel entrances (e.g. with a spade), and those that attempt to bolt can be dispatched there. Sometimes nets are used to entangle a bolting fox.

The National Working Terrier Federation is an umbrella body for the major working terrier clubs, and has a code of conduct for pest control (NWTF submission to the Inquiry).

3.2.2.c.i. Where, when and by whom are foxes controlled using terriers?

The use of terriers lends itself best to communally organised fox culling efforts covering large areas of ground, particularly mounted hunting with hounds, or foot packs, which have been covered above. Foxes are not dug out by members of the MFHA unless this has been requested by the farmer on whose land the fox has run to ground. Known earths in the area to be hunted may be lightly blocked prior to the meet. Where this is practised, it will obviously influence the proportion of the cull taken by digging. There is also considerable variation in the extent to which hunts practice digging out with terriers: the PSL survey found that hunted foxes were most likely to be dug in Wales (where 53% of 2674 foxes were killed in this way) and least likely to be dug in the Midlands & East Anglia (25% of 2,519).

Although terriers are certainly used in the context of individual fox culling efforts (e.g. by professional gamekeepers), within a typical lowland beat of about 8km² (National Game-Bag Census, 1997 data), with fox breeding group densities of 0.1 to 0.4/km² and ongoing culling by other methods, each terrier would be entered to only a few (1-3) earths annually and would not build up much experience.

Nevertheless, 46% of 214 NGO gamekeepers use terriers (NGO submission to the Inquiry; ‘Terriers’ and ‘hunting’ were separate options under the question "How are foxes controlled on your beat?"). Among BASC gamekeeper members (BASC, 1994), 57% used terriers to cull foxes.

In Wiltshire, farmers did not report the use of terriers to control foxes outside the context of hunting with hounds (Baker & Macdonald, 2000).

3.2.2.d. Lurchers

The use of a spotlamp with a running dog (large lurcher or greyhound) rather than a rifle is popular in some areas of the UK. This tends to be practised as an unauthorised (and therefore illegal) sport rather than as legitimate fox control, and as a result no data are available.

3.2.3. Shooting

Foxes are shot in two main ways: at night with a spotlight and rifle (‘lamping’); during the day by groups or individuals. They are also shot at the cubbing earth. Gun packs and shooting at earths may combine shooting with the use of dogs to find, bolt, or flush out foxes. Almost all gamekeepers use shooting to cull foxes (NGO submission to the Inquiry; BASC, 1994), mainly by lamping or by driving foxes to guns. Shooting is probably also widely used by other groups, such as farmers, on an ad hoc basis.

3.2.3.a. Spotlamp and rifle (‘lamping’)

In ‘lamping’ foxes are shot at night with a rifle (usually high powered centre-fire .22, .22/.250, or .243 calibre) with telescopic sight, in conjunction with a powerful spotlamp, usually from a 4WD vehicle. Lamping requires good vehicule access, an absence of cover, and terrain that allows safe shots. Red light reflected from the fox’s retina can be detected from over a kilometre when there is no mist, although the fox must be close enough for its body shape to be distinguishable before a shot can be fired (Anon, 1998). Squeaking sounds often bring foxes running towards the lamp; this trick is most successful with young, naïve foxes, allowing rapid culling by this method during autumn.

A spotlamp and rifle can also be used on foot or from a stationary high-seat, but away from cubbing earths this is extremely inefficient, because the likelihood of a fox passing in range within a reasonable space of time is very low. Most foxes shot from high seats are either at a cubbing earth, or are shot opportunistically during deer culling operations.

3.2.3.b. Gun-packs and standing guns with shotguns

These methods involve the use of a small pack of hounds, or a team of human beaters, to flush foxes out of cover towards a line of standing guns. This approach is most often used in dense woodland, especially commercial softwood plantations. The choice of hounds or human beaters varies regionally depending on availability, but hounds are clearly better in very dense cover and can also trail and catch wounded animals. Some foxes may be caught and killed by the hounds before they reach the guns.

For safety, and because the opportunity to shoot an emerging fox is usually brief, shotguns are almost invariably used. Gun-packs are communally organised, and are especially active in late winter/early spring.

In Wales, there are currently 30 gun-packs; while these operate largely on foot, one or two mounts may be used to keep up with the hounds (The Welsh Farmers Fox Control Society, submission to the Inquiry).

3.2.3.c. Culling at the cubbing earth

The cubbing earth provides a focal point within the territory where adults as well as cubs may be culled. Foxes culled at the cubbing earth must be either shot (with a rifle or shotgun), dug out or caught in nets (after sending down a terrier to locate and bolt or corner the fox), or trapped using cage traps set into the tunnel entrance (effective only for cubs older than 8 weeks; usually ineffective for adults).

Earths used for cubbing are difficult to recognise early in the spring, but become more obvious as evidence of occupation accumulates around them. Among gamekeepers, the aim will be to destroy resident breeding females as early as possible in the season and about 24% of breeding earths are located before cubs are active above ground. Correspondingly, 25% of vixens killed at the earth are killed before cubs can be culled or even counted, unless by the use of a terrier.

3.2.4. Snaring

Only neck snares are allowed under UK legislation. These are set on any route-way likely to be used by a fox, and will be successful only if they remain undetected. This is reasonably easy to achieve (Reynolds, 1998), making the snare a powerful tool against wary adults. The recommended way to dispatch a fox captured in a snare is to shoot it with a shotgun from close range (10-15 m) (Reynolds, 1997). At this range, choice of shot size is not important, and death is instantaneous.

A high proportion (81-86%) of gamekeepers use snares (NGO submission to the Inquiry; Reynolds unpubl.; BASC, 1994). Snares are the only culling method available where prolonged use cannot result in a fraction of the population being untrappable through selection against the unwary. One indication of this is that catch-per-effort for snares peaks in mid-winter when dispersal is at its height, rather than summer or autumn when the highest proportion of the population is naïve. A further indication is that in ecological research on foxes, where animals captured in snares are tagged and released, recaptures and multiple recaptures are not uncommon (J.C.Reynolds, D.W. Macdonald, pers obs.)

Snares are unpopular in sheep-farming country during the lambing season due to the risk of lambs being caught. In upland regions, snares are most often set around a buried carcass bait, with a surrounding fence that keeps sheep out but allows free passage by foxes – this arrangement is known as a ‘midden’.

3.2.5. Trapping

In rural areas foxes are generally difficult to catch in live-capture traps (Harris, 1985; Macdonald, 1987; Reynolds, pers. obs.). Among professional gamekeepers, live-capture traps account for just 1% of all foxes taken. .

3.2.6. What strategies are adopted to attempt to control fox populations?

Consciously or not, farmers and landowners adopt various strategies to achieve their fox control aims. These strategies determine the combination of culling methods used, the timing of culling, and the amount of effort put into each method (and indeed, whether to cull at all). Because different methods are best suited to particular seasons, these aspects are closely inter-linked.

The choice of strategy will vary at a local level from one estate to another according to the needs and preferences of individual farmers. Independent culling efforts are most likely to involve night shooting, trapping, and snaring, often carried out by professional gamekeepers. In addition, there may be a regional strategy for fox control, involving the use of communally organised hunting with hounds and terriers, particularly in upland areas. These require the consent of several or many farmers, and are often supported by subscriptions from farmers, sometimes with formalised state backing. In Scotland, for example, there is limited financial support from the state for fox control groups; similar support was abandoned in England and Wales in 1979, but is continued in Wales by Forest Enterprise.

Other aspects of culling strategies will also change through time. For example, gamekeepers may have shifted away from snares in favour of night shooting with a rifle and spotlight (Reynolds & Tapper, 1994). As there are no historical data on the use of different methods, these are perceptions only, but it seems certain that during the last 30 years or so there has been a rise in popularity of lamping for fox culling. This has been helped by development of lighter and more powerful spotlamps, but also represents a move from the older tradition of managing wild gamebirds - requiring intensive predator control during spring and summer - towards an increasing reliance on hand-reared gamebirds (which are in rearing fields and therefore protected from foxes during spring and summer). Because lamping becomes less effective and snaring more effective as vegetative cover grows, the newer tradition favours lamping. The change of emphasis to autumn/winter lamping inevitably means that greater numbers of foxes will be culled to achieve the same aims.

3.2.6.a. How do local strategies translate into regional effects?

Although a regional impact on fox numbers is an aim for the majority of culling efforts, few people are in a position to organise a fox culling strategy over large geographical areas. Communally organised hunts and fox destruction clubs are better placed to do this than anyone else, but they do not have exclusive command over fox culling. Heydon & Reynolds (2000a) found that although organised groups operated on 88% of farms on which culling took place in Wales, the Midlands and Norfolk, 33% to 91% of these farms (depending on region and farm size) carried out additional fox culling independently. In some regions, therefore, the net impact on fox numbers is more the incidental result of local actions than the outcome of regional planning.

In their ‘Three-regions study’, Heydon & Reynolds (2000a) found fox control strategies used by farmers varied with a distinctly regional character. Communal methods such as hunting with hounds, gun-packs, and digging with terriers were practised on almost every farm in mid-Wales,

where fewer than 10% of land properties had a professional gamekeeper. Spring/summer culling was uniquely important here, and was reflected in high fox mortality during these two seasons. Gun-packs (involving hounds to drive foxes out of cover to standing guns) were used only in Wales. In Norfolk, the bulk of culling was carried out independently by professional gamekeepers on large estates, hence shooting with a rifle and spotlamp, and snaring, were the methods most commonly used.

3.2.6.b. How prevalent is fox control?

The prevalence of fox culling (which occurred on 88% of farms across mid-Wales, the Midlands and Norfolk) indicated by Heydon & Reynolds (2000a; see section 2.2.1.a.i for details) is much higher than that previously suggested. A 1981 survey reported that nationally only 33% of farmers replied "Yes" to the question "Do you attempt to control foxes?" (Macdonald, 1984; Macdonald & Johnson, 1996; Table 2-1). The difference between these studies may be explained by the form of question asked. Heydon & Reynolds asked, "Is any fox culling carried out on your land?" When they asked, "Are foxes culled by anyone other than a hunt or gun-pack", 31% to 67% (depending on region) indicated that they took undertook such culling independently. In a more recent poll (Produce Studies Limited, 1995), 80% of farmers in Wales and 47% in an area covering the Midlands and East Anglia stated that fox control in some form occurred on their land. These values, however, included non-lethal control via electric fencing and assumed that where farmers did not consider foxes a problem there was no attempt to control them. The fact that some forms of control (particularly hunting) may occur on a farmer’s land without his encouragement (Baker & Macdonald, 2000) creates further difficulties in interpreting levels of active control.

3.3 What methods are used to control deer populations?

In contrast to the other species covered by this report (fox, mink, hare), only two groups of lethal methods (shooting and hunting with hounds) are in general use for controlling deer populations in England and Wales.

3.3.1. Hunting with hounds

The history of hunting deer with hounds stretches back at least as far as the Norman Conquest, when, Royal ‘Forests’ were created to provide and restrict sport and venison for the King. Such Forests, including amongst others The Forest of Exmoor, the New Forest, and Epping Forest, were widespread in England and Wales in Medieval times (Rackham, 1986), and some still exist today, stocked mainly with fallow deer. Although deer hunting itself is so ancient, the present-day style of hunting by riding to a pack of hounds is comparatively modern (Burton, 1969). On Exmoor, mounted hunting of red deer (staghunting) was revived around the middle of the 18^th century, and, with some interruptions, it has now had a 200-year history. Staghunting on the Quantocks dates back to 1917, when red deer were reintroduced there, and the Quantock Staghound pack was reinstated with government support as a source of food supply and to revive this sport.

Today, organized hunting of red deer with hounds is practised in only a small area of southwest England (West Somerset and parts of North Devon), and only three hunts are registered with the Master of Deerhounds Association (MDHA). Some small, unregistered hunts, also mainly in the West Country, hunt roe deer with hounds. Until recently, fallow deer were hunted with hounds in the New Forest, but this ceased in 1997 following the Forestry Commission’s ban on deer hunting. While hunting with dogs is currently limited to red and roe deer in the West Country, there are no legal reasons why other species such as muntjac and sika could not be hunted in the same way.

Hunting red deer is referred to as ‘staghunting’, but both mature males (stags) and mature females (hinds) are hunted. The processes of stag and hind hunting vary slightly (MDHA submission to the Inquiry). Before a hunt, stags are selected by the ‘harbourer’ (an experienced local deer expert), who watches their movements at dawn. The chosen stag is separated from the herd by a group of experienced hounds (‘tufters’) before the remainder of the pack are introduced to the scent. Hinds are not harboured because individuals are almost indistinguishable. Instead, if hounds encounter a herd of hinds and calves, they will chivvy them until the herd fragments, and a single hind (with or without calf at heel) breaks away to be hunted alone. Once the stag or hind has been brought to bay it is killed by a shot at close quarters with a modified shotgun or humane killer.

3.3.1.a. Where, when and by whom are red deer controlled using hunting with hounds?

At present, three registered staghunts operate in Devon and Somerset (see also Error! Reference source not found.). According to a recent survey (PSL submission to the Inquiry), the three staghunt countries cover roughly 3,900km², of which 3% is not hunted over for safety or access reasons.

‘Autumn stags’ (usually 5-6 years or older) are hunted from the middle of August to the end of October. Thereafter hunting concentrates on hinds until the end of February, from which time ‘Spring stags’ (mainly 2-4 years old) are hunted until the end of April.

The Devon & Somerset Stag Hounds (DSSH) normally hunt three days a week (Tuesdays, Thursdays, and Saturdays) throughout the season, although occasional ‘bye-days’ may also be added on other days of the week. The other two hunts rarely hunt more than twice a week. As a result of the recent bans on deer hunting imposed by the National Trust and the Forestry Commission, which together hold very significant areas of land within the Quantock Hills, the Quantock Stag Hounds now commonly only hunt one day a week within their own area, with additional meets by invitation within the DSSH country.

Among landowners who responded to a questionnaire circulated by the Quantock Deer Management And Conservation Group (Langbein, 1998a; section 2.3.1.a), 37.5% used or permitted only hunting with dogs as a control method, and a further 27% used both hunting with dogs and shooting. Thus, 64% of farmers in this area use hunting with hounds to control red deer.

The support for hunting among farmers may be higher still: during 1998 over 90% of local farmers (100) with at least ten or more acres of land within the Quantock Staghunting country (including tenant farmers not necessarily always able to grant access for hunting) agreed to sign a letter in support of hunting by the Quantock Staghounds.

3.3.1.b. Hunting roe deer with hounds

While red and fallow deer have consistently been regarded as noble quarry of the Forest or Chase down the centuries, the status and hunting of roe deer has had a more chequered history (Whitehead, 1964). Following their demotion in Ancient Forest Laws in 1338 from ‘beasts of the Forest’ (preventing them and other animals such as red deer, fallow, and wild boar from being hunted other than by order of the King or his appointees), to ‘beasts of the warren’, roe deer gradually became exterminated throughout much of England and Wales by the seventeenth century. However, helped by some re-introductions and sustained re-afforestation schemes, roe have gradually re-colonised most English and many Welsh counties over the last 200 years. For example, in 1800 Lord Dorchester reintroduced the species in Dorset, and within fifteen years the first pack of roebuck hounds was formed to hunt them. Several packs of roebuck hounds hunted in Dorset during the 19^th century (including the Blackmore Vale Hunt and Charborough Hunt) and seem to have continued in Dorset until the First World War (Whitehead, 1964). With their gradual spread further west, by 1900 roe also started to be hunted again occasionally in Somerset (e.g. by the Seavington Hounds near Chard).

Since then the practice of roe deer hunting has continued intermittently. Few published details are available, as most roebuck hunts are not registered, nor recognised officially by the Master of Deerhounds Association. However, it is known that two buckhound packs (Cheldon Buckhounds and Exe Valley Buckhounds) currently hunt roe deer in parts of the Devon and Somerset hunt countries used by the three packs of staghounds. The hounds used to hunt roe deer are usually either bassett/harrier crosses or beagles. Buckhounds commonly pursue their quarry followed only by a small core of mounted followers (huntsman, Master, whipper-in), with the remainder of followers on foot. The procedures for hunting roe are similar to those described for hind hunting above, with an average hunt resulting in a kill normally lasting around 1 to 1.5 hours (Exe Valley Buckhounds, pers. comm.). At the end of the chase roe deer tend to lie down, rather than being brought to bay, and are then dispatched with a modified shotgun or humane killer.

Hunting of male roe deer usually takes place between the end of August and end of October, and also from April to early May, whereas females are hunted from October to February. Each of the two packs usually hunts once a week during the season, with a total of around 35 meets per pack. The main motivation for roe deer hunting is to provide sport with a total estimate of only 30-40 roe deer killed by hunting per annum by both the above packs combined.

3.3.2. Shooting

Shooting by stalking with a rifle or large bore shotgun is the most common method used to cull deer in England and Wales, as well as in Scotland and Northern Ireland (BASC and BDS submissions to the Inquiry). Shooting culls of deer, particularly as part of organised Deer Management Groups (DMGs – groups of adjoining landholders co-ordinating their deer management) are the method of deer control recommended by government (MAFF submission to the Inquiry).

3.3.2.a. Data and approach

Because there is, as yet, no single organisation in England and Wales which collates all culling data for deer, there are surprisingly few data available on the exact numbers of deer shot. The British Association for Shooting and Conservation’s (BASC) 1996 survey on stalking "Deer, Deer Stalking and the Future" provides the most comprehensive account yet of deer stalking in this country, although results accrue necessarily from a self-selected sample of deer stalkers.

BASC sent the first questionnaire for the survey to a randomly selected 10,000 of their c. 120,000 members; 67% responded. This identified 13.4% of members who were active deer stalkers, and a further 22% interested but not active. A second questionnaire, to active stalkers only, produced 408 replies, providing a small but widely distributed sample (BASC, unpublished).

3.3.2.b. Shooting by rifle (deer stalking)

Shooting deer by rifle, generally known as deer stalking, is by far the most common and widespread of the lethal methods of legal deer population control (of any of the species) undertaken in England and Wales, as indeed it is in Scotland (Callender & McKenzie, 1991), and most other European countries (Deutscher Jagdschutz Verband, 1997). Rifle culls tend to be taken either from a high-seat (an elevated platform usually paced against a tree), or by stalking carefully up to the deer on foot at ground level to a safe shooting position. Many deerstalkers use a trained dog to ‘point’ to deer during stalking, and to locate fatally shot or injured deer.

Deer stalking may be undertaken for various complementary reasons, including population control to assist with crop protection, to obtain venison for own consumption or profit, or to provide income and sport through letting the deer stalking rights or offering accompanied stalking to shooting clients.

In the 1996 BASC deer stalking survey (see above), 24% of the total nationwide cull reported related to red deer, 49% to roe deer, 17% to fallow deer, 7% to muntjac, 3% to sika and <1% to Chinese water deer.

3.3.2.b.i. Where, when and by whom are red deer controlled using stalking?

Red deer are stalked with rifles throughout their range (see Error! Reference source not found.), particularly in Scotland, but also in England and, to a lesser extent, in Wales. The open season runs from August to April for stags, and November to February for hinds (see also Table 11-1).

Stalking is generally the main method of controlling red deer numbers across most private and public landholding types (forestry, moorland, farmland), including by Forestry Commission (Forest Enterprise), and on MOD land. In the Quantocks, shooting was used to control red deer by 50% of farmers who responded to a questionnaire circulated by the Quantock Deer Management And Conservation Group (Langbein, 1998a; section 2.3.1.a).

From their 1996 survey, BASC used a very conservative estimate of 9% (not 13.4%) to calculate that an estimated 10,000 of their members were active deer stalkers. Of these, 87.6% (8700) were ‘recreational’ stalkers and 12.4 % (1300) were ‘professional’ deer stalkers (rangers/ghillies etc). However, the professional stalkers accounted for 40% of the total deer cull. Extrapolated to the nationwide estimate of 180,000 deer shot by their members alone, these percentages suggest that around 72,000 deer are shot by professional rangers, and c. 108,000 by recreational stalkers.

3.3.2.c. Shooting by shotgun (for damage control)

Prior to the Deer Act of 1963, shooting with shotguns was a common method of deer control. In the West Country, any perceived surplus in deer numbers not taken by hunting was generally accounted for by means of organised shotgun drives (HMSO, 1951; Bonham-Carter, 1991). However, the Deer Act 1963, and most recently, the Deer Act 1991, prohibited the use of any smooth bore gun (shotgun) to kill deer for the purpose of general deer management, although they can be used for damage prevention or to kill an injured animal (Appendix 2). No separate data are available for the proportion of the cull currently taken in this manner.

3.3.3. What strategies are adopted to attempt to control deer?

Traditional approaches to deer control involve attempting to reduce population sizes. While this is often an appropriate response, it is associated with various problems. In particular, many deer populations range widely, over several estates; this means neighbouring landowners must coordinate their management efforts to achieve effective control. However, even effective control of deer numbers will not necessarily deliver an equivalent reduction in impact, since there is no simple relationship between damage and deer density (section 2.3.2.b.i).

In practice, effective control will generally only be achieved via integrated management involving both direct management of the deer population and independent control of its impact (Ratcliffe, 1998; Chapman & Harris, 1998; Putman & Langbein, 1999, 2000). Non-lethal methods (section 3.6) form a useful tool for management of deer and control of damage in an integrated approach. Strategies to manage local distributions and to reduce the significance of damage include permanent and temporary wire fencing, tree guards and shelters, chemical repellents and habitat manipulation.

In a recent large-scale questionnaire survey conducted by ADAS in lowland England, the majority of respondents with farming or forestry landholdings perceived organised culling via a Deer Management Group (DMG) as being the most effective way of preventing deer damage (Packer et al., 1998). Encouragement and formation of DMGs is the main aim of the Deer Initiative, a Forestry Commission-led partnership of a wide range of countryside and welfare organizations.

3.4 What methods are used to control hare populations?

Hares are only a minor pest, and are usually culled for sport and sold to game dealers (Harris & McLaren, 1998). Where they do need to be controlled, organised hare shoots are generally used (Stoate & Tapper, 1993)

3.4.1. Hunting with dogs

3.4.1.a. Hunting with hounds

Hunting hares with beagles appears to have been well developed by Tudor times and was popular in the 18^th century (Stuttard, 1981). Today, hunting hares with hounds takes place on foot with packs of beagles or bassets, or on horseback with harriers. In England and Wales 102 packs of hounds (of which 72 are beagles, 10 bassets and 20 harriers) hunt hares and are registered with the Association of Masters of Harriers and Beagles (AMHB) or the Masters of Basset Hounds Association (MBHA) (AMHB and MBHA joint submission to the Inquiry).

The AMBH (submission to the Inquiry) estimate that hunts are normally restricted to 1-2 square miles and last 30-90 minutes. Typically, the hunt consists of a number of short chases and checks, as the hare evades and then is flushed by the hounds. The huntsmen draw a nearby field with a pack of beagles until a hare is put up or its scent is found. This is then pursued until the line is lost and the hare escapes, or the animal is caught and killed by the hounds.

Only one beagle pack (the Holme Valley Beagles) take mountain hares in England.

3.4.1.a.i. Where, when and by whom are brown hares hunted with hounds?

As with foxhunting, each hare hunt has its own country, registered with the AMHB, and also used by the MBHA. These average 1,300km² (PSL submission to the Inquiry), and cover an estimated 90% of rural England and Wales. In 1989 and 1990, the Game Conservancy Trust asked one hunt Master to map where his hounds ran on each of 14 days of hunting. The average of the 14 days was roughly three square kilometres (300ha).

Most beagle and harrier packs hunt twice a week (AMHB and MBHA joint submission to the Inquiry). The season extends from after harvest begins in late August or early September and extends until the end of March. AMBH rules determine that there is no hare hunting after March, though there is no statutory requirement on this.

3.4.1.b. Hare Coursing

Hare coursing as a sport was well established in Roman times in Gaul (Stuttard, 1981). Today, two types of legal hare coursing operate: competitive (‘two-handed’) coursing, and ‘single-handed’ coursing. Coursing competitions, usually with greyhounds, but also with other breeds including deerhounds or salukis, are organised events carried out by coursing clubs, which take place on land with permission of the land-owner.

Under rules laid down by the National Coursing Club (Stable & Stuttard, 1971; NCC submission to the Inquiry), coursing takes place on open ground to allow the hare to escape. Competition coursing is not considered a control method, and the aim is not to kill the hare (NCC submission to the Inquiry), although the hare will be killed if it is caught by the dogs. Hares are driven onto the running ground by beaters (‘driven coursing’), or are put up by participants and their dogs (‘walked-up coursing’). Dogs are released once the hare is sufficiently far away (about 80 yards). Competitions are usually knock-out competitions between pairs of dogs; the winning dog is chosen by a judge on horseback who awards points on the basis of how well they perform in relation to each other and the hare. Coursing competitions using deerhounds follow most NCC rules with additional Deerhound Coursing Club rules (DCC submission to the Inquiry). Competition coursing with lurchers follows NCC rules "as far as is applicable" (Association of Lurcher Clubs [ALC], submission to the Inquiry).

‘Single-handed’ coursing is usually carried out with a single lurcher, and the object is to catch and kill the hare. This form of coursing is not formally recognised as a field sport by the Countryside Alliance (ALC submission to the Inquiry), and does not abide by NCC rules. As with any other form of hunting, this type of coursing is illegal only when carried out without the landowner’s permission.

3.4.1.b.i. Where, when and by whom are hares coursed?

There are 224 greyhound coursing clubs affiliated with the NCC. The ALC estimates the total number of lurcher owners to be 112,422. Of 328 lurcher owners who responded to a questionnaire survey by the ALC, 52% coursed. There are no data on extent to which farmers and other landowners consider coursing to be a useful form of pest control, but we note that "In many cases the courser pays [in cash or in kind] for the right to course on the farm" (ALC submission to the Inquiry).

On average, between the 1990/91 and 1998/99 seasons, there were 1934 competition courses held over 93 days annually (data taken from the NCC submission to the Inquiry). The number of coursing events run on an estate ranges up to 112 (Altcar) and these generally take place on arable farmland with a gamekeeper.

National Coursing Club rules do not allow coursing between March 11^th and September 14^th; the ALC regulations also stipulate a closed season.

3.4.2. Shooting

Hare shooting is the most significant form of hare culling in the countryside and the method most frequently adopted by farmers in arable areas as a means of pest control (Tapper, 1987).

Hare shoots (using shotguns) are normally undertaken over the whole farm area, and smaller farms sometimes combine to take in a wider area of ground. The day is organised as a series of drives, and unlike game bird shoots, there is no separation between beaters and guns. Most participants (usually 20-40 people) carry guns and will be involved in either standing or walking lines during the course of the day. Separate areas of the farm are surrounded and the drives move inwards making a tighter area. Hares either are shot as they flush running forward, or are taken as they break out through one of the lines of guns.

February is the most common month for organised hare shoots, being after the end of the game bird shooting season, and at a time when most hares are feeding on winter cereals (Tapper & Barnes, 1986). Hares shot on hare shoots are treated as game, and provided the shoot is undertaken before March 1^st it is usually sold to a game dealer who markets the hares in Britain or in Europe.

An unknown number of hares are also shot with rifles by or at the instigation of farmers, to reduce numbers and avoid the problems associated with illegal hare coursing. Although there are no data on this, contact with landowners and farmers suggests that this is quite a common strategy (Pye-Smith, 1998; S. Tapper, pers.comm.). It is extremely easy for hare numbers to be substantially reduced by steady attrition over a period of weeks or months, using a .22 rifle.

In 1997 an estimated 200,000-300,000 brown hares and 40,000-100,000 mountain hares were shot in the UK (Cobham Resource Consultants, 1997). The number of mountain hares known to be shot in England and Wales has been typically no more than 50 individuals annually during the 1990s (National Game-Bag Census, The Game Conservancy Trust, unpubl.), all on a handful of shooting estates in the north of England.

3.4.3. Other methods

The smaller mountain hare is a quarry for falconers using the larger raptors such as the golden eagle. Hares have also traditionally been snared and netted. All of these methods remain legal today.

3.5. What methods are used to control mink populations?

Mink are controlled using three methods: hunting with hounds on foot, trapping (either break-neck or live-trapping followed by shooting), and shooting.

During the 1960s, the Ministry of Agriculture, Fisheries and Food trapped over 5,000 mink in England and Wales, with a similar effort in Scotland by the Department of Agriculture and Forestry. By the mid-1970s this effort was emerging as futile and was abandoned (Birks, 1986). Subsequently, highly intensive mink control has been implemented patchily, where fishing interests prevail. Otherwise, attempted mink control has been haphazard or none-existent for the last two decades.

3.5.1. Hunting with hounds

Since the otter was legally protected from hunting in Britain, at least six otterhound packs began hunting mink instead. Several new packs hunting mink have also been formed, and there are currently 20 packs registered with the Master of Minkhounds Association (MMHA submission to the Inquiry). There are also some unregistered packs.

The hunt operate on foot, and drag rivers and streams across several farms where permission has been granted or in response to a request from fish farmers and others (MMHA submission to the Inquiry). Dens are located by dogs as they search the riverbank, and attempt is made to bolt their quarry. The dogs may catch the animal as it is flushed into the open or it may be treed and then shot or dislodged to fall to the hounds. As with foxhunts, minkhunts use terriers to kill mink which have gone to ground.

In January 2000, a Produce Studies Ltd survey (PSL submission to the Inquiry; section 3.2.1.g) estimated that mink countries covered 76,000km², 61% (46,000km²) of which was actually hunted over. The hunting period runs from April to September when water level is low and the mink are raising young.

In 1996, Strachan (unpublished data) sent a letter with simple questions concerning mink to 40 landowners along 24km of the River Thames in west Oxfordshire. Of 32 respondents, 6 (19%) said they would use the mink hounds to control mink.

3.5.2. Shooting

Over the period 1995-98, Strachan et al. (unpublished) carried out interviews with farmers, gamekeepers and lock-keepers in the Thames Valley regarding their mink control practices. Shooting by farmers and gamekeepers was ad-hoc and largely incidental to other control activity (e.g. while shooting rabbits, foxes or pigeons). The majority of mink were killed during the winter months or in the autumn when juveniles were dispersing.

3.5.3. Trapping

Mink are either caught and killed in spring (break-neck) traps, or are caught alive in cage traps, and then killed. Spring traps must be approved to catch mink in the Spring Trap Approval Order 1995 (i.e. Fenn Mk 6, Springer No 6, Kania 2000, BMI Magnum 116). They risk killing non-target species (in a riparian habitat these are chiefly weasels, stoats, polecats, water voles, rats, and moorhens, but also possibly young otters).

Cage traps are best made of 14-gauge weldmesh, and should be wrapped in hay and baited (e.g. with sardine or day old chick carcasses) to provide food and shelter, and to increase their attraction to the mink. Traps must be checked daily, ideally first thing in the morning. Live-caught mink must be killed, as it is illegal to release them back into the wild. For humane killing, a 0.22 rimfire rifle, or 9mm and 0.410 shot-pistol or shotgun is recommended, but in practice it is not always easy to shoot a small, moving, target within a cage, and some practitioners simply drown the animals; this is particularly cruel (section 6.4.2.a).

Trapping is the main method of mink control in Britain, and is widely recommended (Advisory Service of the Game Conservancy, 1994; Strachan, 1998; MAFF, 1998). As a specific conservation tool at water vole sites, mink control by cage trapping is recommended in the months of January, February, March, and April (around the time of the rut but before birth). This will allow for the efficient removal of all resident female mink (Macdonald & Strachan, 1999).

Gamekeepers, water bailiffs, lock keepers, angling clubs, farmers, nature reserve managers, and private landowners may all trap mink, but there are few data on the extent to which it is used. In a 1996 survey of farmers along 24km of the River Thames, just over half (56.3% of 32 replies) said they took action against mink. Of these 18 farmers, 6 (19% of all those surveyed) used trapping (Strachan, unpublished data). In 1995-1996, 21 of the 47 Lock keepers on the River Thames upstream of Teddington, trapped for mink (Strachan, unpublished data). Trapping effort was very variable, from none (even where mink were known to be present) to up to 10 traps on one Lock island. Trapping was generally in response to the presence of mink.

3.6. What are the alternatives to culling?

In recent years, ethical and conservation concerns over culling have led to increasing interest in non-lethal methods of population and damage control. Non-lethal methods of control fall into three categories:

Environmental alteration techniques are by far the most widely and reliably used; behavioural alteration and fertility control methods are still largely experimental, although they have been used effectively in some practical applications. The nature of the species, the damage it causes, and management aims, define the type of control needed. An important aspect of behavioural and environmental alteration approaches is to isolate the aim of damage reduction from any other aims, such as sport or meat.

Non-lethal methods of damage and population control could prove more effective than lethal control, because non-lethal methods aim to minimize perturbation, for example by retaining the predator in its original territory (Tuyttens & Macdonald, 2000). This avoids the density dependent population responses and immigration which can result from culling (section 1.3.2.a), whilst allowing the animal to continue with whatever effect it has on limiting other prey numbers or excluding conspecifics.

3.6.1.a. Non-lethal methods of fox control

The most widely used non-lethal method of fox control is exclusion using physical barriers. Hand-rearing of game-birds is a management technique which protects them from foxes at a vulnerable stage, but also creates other problems. Two further non-lethal approaches have been widely discussed: manipulating foxes’ food preferences (conditioned taste aversion or CTA) and fertility control.

Physical barriers such as wire netting are valuable for preventing loss of poultry, game-birds or livestock held in small areas, but are not practicable to protect them on any wider scale. Electric fencing has been used with partial success to protect wild ground-nesting birds on nature reserves, but usually needs to be backed up with lethal methods. It is most useful for colonial nesting species whose nests are concentrated in small areas (e.g. terns), or to a very few individual nests (e.g. stone curlews), and for a limited time (Minsky, 1980; Haddon & Knight, 1983; Patterson, 1977).

3.6.1.a.ii. Hand-rearing of game birds

Hand-rearing of gamebirds is itself a way of avoiding predation (and other causes of mortality) at a vulnerable life-history stage. However, it does create other problems, including:

3.6.1.a.iii. Conditioned Taste Aversion (CTA)

Conditioned taste aversions are a well-researched component of vertebrate behaviour through which an animal’s innate food preferences are modified as a result of experience (Reynolds, 1999). Characteristically, a single bad experience of a food that is mildly damaging (e.g. food poisoned by harmful bacteria or laced with an emetic) leads to a robust and lasting aversion towards any food that is similar in taste. In captive or laboratory animals, the power of CTA to cause lasting avoidance of certain food types, even in the absence of alternatives, is very impressive.

Since the early 1970s wildlife biologists have speculated that CTA might be exploited to curtail unwanted predation. In the UK, it has been envisaged that individual foxes might be made averse to game birds - for example - by feeding them dead game birds that have been laced with a mildly toxic chemical (Reynolds, 1999; Cowan et al., 2000). Once ‘educated’ to avoid the chosen prey, a fox would become a valued resource, because it would keep out ‘uneducated’ foxes through territorial behaviour. A single, loaded, bait would be sufficient to ‘teach’ an individual that the referent food type (e.g. game birds) was bad to eat. Thus a short baiting programme at the appropriate time of year might be all that was required to achieve management aims.

CTA has recently been the subject of recent intensive field research by both The Game Conservancy Trust and Central Science Laboratory (MAFF). The WildCRU at Oxford University have explored a related topic - the creation of food-aversions using a bitter-tasting substance (Baker & Macdonald, 1999). It is now clear that deployment difficulties and non-target hazards make CTA non-viable for fox management in the UK (although the WildCRU and CSL are continuing related research on repellents). The major problems identified are:

3.6.1.a.iv. Fertility Control

The concept of controlling wildlife populations by reducing productivity rather than by increasing mortality has been around since the 1970s. For canids, chemosterilisation has been attempted in the USA using a number of chemicals (Balser, 1964; Asa, 1997). Hormonal control has also been trialled, both in captivity and in the field (Linhardt & Enders, 1964; Linhart et al., 1968; Oleyar & McGinnes, 1974; Allen, 1982). However, because reproductive biology is so similar in all mammals, both kinds of fertility control carry risks for non-target species, including humans. The safest approach to fertility control for wild animals is by exploiting the body’s immune system (immuno-contraception) to create antibodies to sperm-coat or ovum cell wall proteins (Tyndale-Biscoe, 1994).

Immuno-contraception for foxes has been the subject of intensive research over several years by Australian and French government scientists. In Australia, the chief concern is to drastically reduce or eliminate the (introduced) red fox. In Europe, the interest is primarily to ensure that it remains possible to control rabies outbreaks through oral vaccination, despite increasing densities of foxes (Artois, 1997). Compared with Australia, a much more modest impact on fox numbers is sought to ensure efficacy of rabies vaccination.

Immuno-contraception was expected to be long-lasting, cheap, safe, reversible, more species-specific and more humane than traditional methods for reducing fox populations (Kirkpatrick & Turner, 1991). Initial hopes that contraceptives could be administered using modified viruses as vectors have largely been displaced by the less ambitious route of oral administration through baits. Despite enormous expenditure in America, France and Australia, many other practical problems remain to be resolved before a workable methodology is available. One view of this is that the probability of success and the environmental risks linked to species-specificity mean the money required for development would be better used in traditional methods of population control (Artois, 1997; Artois, pers.comm.). Optimism remains high within research teams (Boue, pers.comm.), though the time-scale to development of a vaccine for foxes in captivity is at best 5 years. One can speculate that development and testing for environmental application would take at least 10 years from now (2000).

3.6.1.b. Non-lethal methods of deer control

Non-lethal measures are more widely used against deer than any other species covered in this report. Non-lethal measures to limit the impact of deer on agriculture, forestry or conservation areas include a variety of fencing, individual tree guards or shelters, chemical repellents, species choice and habitat manipulation aimed at altering the behaviour and feeding patterns of the deer. In addition, fertility control is being developed to control (mainly enclosed) populations.

3.6.1.b.i. Physical barriers

Traditional wire mesh fencing is probably the most widely used form of protection against deer damage. In the planting year 1991/92 the Forestry Commission alone spent over £6 million on fencing, primarily to protect against deer damage (Gill, 1997). The Forestry Authority (Pepper, 1992) recommends the minimum fence heights between 1.5m (for roe and muntjac) and 2m (for red deer).

Full height (1.8-2.4m) high tensile fencing is also used extensively along many motorways in Britain and elsewhere, as it is also the most reliable method of reducing numbers of road-traffic accidents related to deer (e.g. Ward, 1982; SGS Environment, 1997). Electric fencing, can provide temporary protection to areas vulnerable only for a short period of time, but is unreliable and often ineffective (Pepper et al., 1992), and requires maintenance.

Where the protection of new plantings of small numbers of trees is required, individual trees may be protected with guards or shelters (Pepper et al., 1995). A wide variety of such shelters are now commercially available. Plastic guards can provide protection from roe deer, but welded wire mesh guards are most effective for both broadleaves and conifers, particularly against larger species. Tree growth shelters (opaque plastic tubes), are widely used to establish broadleaved trees and shrubs and have the added benefit of providing a suitable microclimate for growing trees. As with fencing, the height and specification of the tree guards or shelter is important. The cost of individual tree protection tends to be high (currently c. £1.50-2.50 per tree).

3.6.1.b.ii. Species choice

Where new trees are to be planted the choice of species will depend on the particular amenity, sporting, or commercial objectives for the forest, and will be limited by site factors (soil type, climate, exposure) as well as by the anticipated future economic needs for timber. Some scope may nevertheless remain for the forester to chose those species least susceptible to wildlife damage.

As a general guide, willows, aspen, and silver fir are highly preferred by deer for browsing, and Norway spruce, lodgepole pine and ash are particularly susceptible to bark stripping. Sitka spruce, Scots pine and Corsican pine are less vulnerable (see reviews by Gill, 1992a,b). Damage even to low ranking species may nevertheless be severe where more highly preferred species are not available, and hence provision of some alternative browse alongside the main crop may be useful (e.g. Langbein, 1993). Damage to coppice woodlands across England shows clear differences in vulnerability to damage between coppice species, depending on the deer species present (Putman, 1994). Birch regrowth was browsed heavily by roe and red deer, but left virtually untouched by fallow, the latter causing more damage among chestnut and ash. Alder and maple seem least palatable to deer in general, while willow is universally preferred, again suggesting some scope by focussing attempts at coppicing at the least vulnerable species given local conditions.

3.6.1.b.iii. Chemical repellents

Chemical repellents may be applied to reduce deer damage in two main ways: as barrier repellents, leaving an ‘olfactory fence’ which animals will not willingly cross; or as feeding repellents, applied to individual vulnerable plants which, by scent or taste, repel or inhibit feeding. Only one barrier repellent (Renardine) is currently approved for use under the Pesticides Registration scheme, but it appears to be ineffective against deer. The Forestry Authority has tested the efficacy of over 65 chemicals or proprietary compounds sold for application to individual plants (trees), but results have generally been disappointing (Pepper, 1978; Pepper pers. comm). While various products still under development may have some potential to provide protection against deer damage, a wide variety of other folk-lore treatments (e.g. lion dung or human hair) proved to have no repellent properties. Only one, Aaprotect® has so far been found to give consistently good results, but even this offers protection for conifers for one winter only. A fundamental limitation with most repellents is that new foliage remains unprotected unless the repellent is re-applied after every growing season (Gill, 1997).

3.6.1.b.iv. Behavioural and habitat manipulations

The impact of deer may be manipulated by changing their patterns of habitat use or foraging behaviour, so that they cause less damage in vulnerable areas. Damage sustained by any given agricultural area or woodland block will be a function both of the amount of usage that area sustains (in terms of ‘deer-hours’ spent within it) and the proportion of that time which is spent actually feeding on vulnerable crops.

Both the attractiveness of an area for deer, and the probability that they will feed upon vulnerable species, can be manipulated by appropriate changes in management. Valuable crops may be protected from damage by intercropping them with more palatable ‘sacrifice crops’ (Petley-Jones, 1995; Putman, 1998), or animals may be drawn away from sensitive areas by increasing attractiveness of alternative sites through increasing forage quality or cover (Langbein, 1997). Conversely, habitat manipulations may also be employed to reduce the attractiveness of vulnerable areas for deer (e.g. removal of nearby cover). For example, changes in coupe size may affect the amount of usage of regenerating coppice or plantations, with large open areas often being less attractive to deer (Kay, 1993; Putman, 1994).

While such measures are very likely to have the potential to alleviate damage, especially if used in combination with other forms of damage control, their actual effectiveness has not been studied in detail (Putman, 1998).

3.6.1.b.v. Non-lethal population control

In the United States, fertility control trials have been undertaken on enclosed or island populations of various ungulates, including deer (white-tailed deer: Kirkpartick et al., 1997, Turner et al., 1992; Garrot, 1995; fallow deer: Fraker, pers comm), with some degree of success. However, numerous potential behavioural and physiological side-effects have yet to be fully investigated, and it is unlikely at this stage that such methods would ever become practical and cost-effective for general use on free-ranging deer populations. As with foxes, a major problem is simply the large numbers of animals which must be treated to achieve any significant reduction in recruitment rates (Garrot, 1996; Putman, 1997; Moore, 1998).

Although strictly speaking not a non-lethal method, a ‘natural’ method of population control might in theory be achieved through reintroduction of large predators of deer. Re-introduction of wolves and lynx to Britain is likely to be problematic (Macdonald et al., 2000), particularly in view of the vast population of free-ranging sheep likely to suffer predation (Gill, 1997). However, the gradual re-establishment and spread of wolves in Spain, France, Norway and Sweden illustrates that wolves can co-exist with modern agriculture in parts of Europe (Boitani, 1998), although not without problems.

3.7. Conclusions

Section 3 footnote

9. Fishers Exact, P=1

4. What can simulation models tell us about the effectiveness of population control methods?

4.1. Why use simulation models to estimate effectiveness?

There are many factors that influence the distribution and abundance of animals in landscapes; these are complex and often poorly understood. The abundance of a species within a particular area is dependent upon an interaction of four basic processes: reproduction, mortality, emigration and immigration. These processes are, in themselves, determined by other landscape-dependent factors such as the availability of food and den sites, and the incidence of anthropogenic factors such as culling. The abundance of an individual species can therefore be considered as the resultant of a complex interacting system of different life history processes.

Discovering genuine trends in population sizes of animal species is, therefore, a difficult task. There are so many confounding factors occurring simultaneously, that, compounded with the difficulty of obtaining an accurate census of numbers, it may be impossible to distinguish actual effects from the morass of ‘noise’. In addition, mammal populations tend to act on a time-span of years or even decades, and so population trends caused by the cessation or escalation of control measures may not become evident for some time.

Clearly, in order to assess the impacts of culling on the abundance of culled species it is necessary to assess the relative impacts of all of these different processes as well as the impact of culling on population size. The complexity of these systems means that they are ideal subjects for investigation using modelling approaches. The purpose of modelling is to simplify complex systems so that the processes determining their functioning can be understood. Population dynamics modelling has a very long history. Lotka (1925) and Volterra (1926) developed the first mathematical models to investigate how species competed. These were the first of a series of analytical models that have been developed over the last 70 years. These models were based on numerical analyses of differential equations. These models did not pretend to analyse how populations of animals might behave in real landscapes, and as such their use was purely strategic in so far as they illustrated general ecological principles rather than addressed tactical issues. Strategic models of this type are particularly useful for evaluating the impacts of individual life history parameters and culling strategies on broad populations and have considerable value didactically, because they can be used to assess the broad impact of changes in life history parameters, such as cull rate, on population size.

4.2. What types of models have we used to estimate effectiveness of population control methods?

In this Chapter we use both strategic and tactical modelling approaches to investigate the impacts of variations in life history parameters and culling on the likely dynamics of the controlled species. We first develop a series of matrix population models to investigate the impacts of culling at the large population scale for foxes, hare, red deer and mink. We use this approach in a didactic sense to investigate firstly, under what conditions large scale population change could be influenced by culling and secondly to provide a comparison with other simple population analyses based on net gains or losses of individuals form populations. We then go on to develop individual-based models for the fox and mink to investigate the impacts of different culling strategies on the populations of these two species at the scale at which population control is practised. We focus on the fox and the mink since these species have very specialised space use patterns and behaviours. Finally, we use population viability analysis to examine long term trends in fox population dynamics under different scenarios.

4.2.1. Modelling at the level of the population

Matrix population models are a relatively recent technique used in population ecology. They were independently developed by Bernardelli (1941), Lewis (1942) and Leslie (1945, 1948), but were not widely adopted by ecologists until the 1970s. The work by Leslie (1945) is by far the most influential in the field of ecology. By expressing the basic age-specific projection equations of a population in matrix form, complex calculations can be reduced to matrix algebra. The eventual rate of increase of a population, and the stable age distribution, can be derived from such matrices. Jenson (1995, 1996) described a simple and easily-employed matrix model for density-dependent population growth that requires no new functions, matrices or parameters. It is directly analogous to the models produced by Lotka and Volterra, except that it takes into account population age structure, and so is useful for looking at systems where culling is applied differentially to different stages of the life-cycle. Used strategically, matrix population models can be used to assess the sensitivity of an animal population to culling, and the likely result of escalating or ceasing any management strategies employed.

The purpose of the population modelling approach in this inquiry is didactic. The approach is not intended to simulate actual population changes, but instead to examine the sensitivity of an animal population to culling. This is a complex issue, as culling might be differentially employed on different age classes of the animals involved, and this may have a delayed effect on population dynamics.

For the matrix population models we attempt to model a population sufficiently large that immigration and emigration are made irrelevant (because there is no ‘outside’ for these migrants to come to or from). In our individual-based and population viability analysis, however, we specifically look at the possible influence of migration on population persistence (see below). We reconstruct the population dynamics of the species in question using established values for life histories and then impose different levels of culling mortality on these populations, to see what effects these have on long term trends in the population.

For this study, matrix population models have been constructed to model populations on a regional or national scale. The use of matrix algebra overcomes some of the theoretical objections to differential population models since the matrix operates over discrete periods of time. The age structure of the population is taken into consideration, and so differential mortality may be applied to each age class of the animal modelled.

By modelling at the population level, particularly with matrix algebra, we lose the fine-scale interactions found in individual-based models (see below). However, these effects become less important when looking at a regional or national population. Matrix models are also ‘data-hungry’, in that they require age-specific life history data (yearly production and probability of survival for every age of the animal’s life), and this often requires extrapolation of the available data.

4.2.2. Modelling at the level of the individual

Whilst simple matrix models can be used to assess the impact of culling at the overall population scale, they takes no account of how culling is undertaken in practice. In order to assess how culling measures such as hunting and shooting might impact on real populations it is necessary to consider how the control is practised, and more particularly, where it is practised.

Populations of culled species do not exist as homogeneous individuals spread at random through landscapes: they have very specialised interactions with the landscapes in which they are found. Furthermore the behaviour patterns they adopt in landscapes will have knock-on effects on how culling is practised. The majority of species utilise a permanent or semi-permanent den or nest for breeding. In such a situation animals forage for food away from the nest in a more or less defined area that makes up a home range. A home range may be more or less exclusively occupied by a single animal, as in the mink, or by a small social group with dominant adults, as in the fox. Humans involved in managing these species are well aware of the implications of these behavioural and space use patterns for population control and have adapted their own culling behaviour to ensure that their control utilises these patterns to their own advantage. The complex space use-animal behaviour interaction and human culling behaviours require a different modelling approach for fox and mink. We focus on the fox and the mink since these species have very specialised space use patterns and behaviours

Spatially explicit population dynamics models predict the distribution of animals in the landscape on the basis of interactions between the landscape structure and individual behavioural processes such as home-range behaviour, territoriality and dispersal, and the life-history processes of births and deaths. In these models space provides reference points on which the populations processes occur and the distribution of organisms amongst these points emerges as the model is run. They are inevitably much more complex than models based on statistical methods because they attempt to simulate individual processes, but they offer a potential route for investigating species with complex social interactions and highly dynamic distributions such as the mink and fox.

Spatially-explicit models can be broadly classified as either population based or individual based depending on the level at which the life history processes are modelled. In individual-based models the processes of mortality and reproduction are estimated at the level of the individual and the overall effects on the population are derived from summation of the life-histories of each individual. These approaches are also termed i-space configuration models (Caswell & John, 1992). In population-based models mortality and reproduction are not modelled at the level of the individual; rather they are applied to groups of animals such as individual cohorts or age classes that constitute components of the total population. In these models the various life history data used may be drawn from a statistical distribution that represents that observed in the population.

These modelling approaches require a detailed knowledge of the effects of all biotic and abiotic factors on the dynamics of fecundity, mortality and migration within the individual population. In addition, some understanding of the basic behaviour of the species is needed in order to be able to link the life history processes to the space where they occur and this information is often poor (Lima & Zollner, 1996). Whilst the utility of using these approaches has been debated (Bart, 1995; Wennergren et al., 1995) these models have been used to investigate the distributions of many other species of conservation and pest significance with considerable success (Akcakaya et al., 1995; Liu et al., 1995; Rushton et al., 1997; Rushton et al., 1999; Rushton et al., 2000). Spatially explicit models are inevitably more tactical than the matrix models described above, in the sense that they attempt to simulate processes of individual behaviours such as the social group behaviour of the species and in this case, the human behaviour associated with different types of culling.

We develop these approaches in landscape where the fox and the mink are culled or are considered pests. For fox we utilise the three study areas subjected to hunting and intensively studied by Heydon et al. (2000) and Heydon & Reynolds (2000a, b). For mink we utilise the river catchment of the Thames, which has been extensively studied by Macdonald and associates (Macdonald & Strachan, 1999).

4.2.3. Population Viability Analysis

Population Viability Analysis models (PVA) are a form of modelling widely used to explore the risk posed to a population by various threats. In the context of conservation, PVA models are often used to identify factors which limit the abundance of a population. In terms of pest control, comparable analyses may serve instead to reveal the likelihood that a given source of mortality will limit a population’s numbers or diminish its persistence. Our population viability analyses use the commercially produced program Vortex (version 8.03) (Lacey, 1993). We used Vortex because it is a widely used system, amenable to sensitivity analyses, and for consistency with the earlier explorations of fox population dynamics by Macdonald & Johnson (1996).

4.2.4. Summary of models used

The models used for each species are summarised in Table 4-1. In population based models, each step in the calculations is based on average estimates for a given subset of the population and often fixed estimates of mortality and fecundity are used. The model results tend to be easier to interpret, but they are less realistic. These models are best used to explore broad patterns in population dynamics and to understand the possible interactions between key population parameters. These models make unrealistic predictions for very small populations because the mortality and fecundity estimates are based on fractional estimates of individual numbers.

In individual-based models, the program keeps track of each individual. The life history processes of fecundity, survival and migration are determined by age and sex classes and by predefined probability distributions. These models generally require greater computational power and the results may be more complicated to interpret, but they have the advantage that they are more realistic. We often use these models to understand the inherent variation in population processes. These models produced variable data and statistical analyses are often required to interpret the results.

Spatially explicit models examine population dynamics within a spatial framework and are useful when the distribution of habitat features or resources (food, den sites etc.) have an important influence on the distribution of the animal. This is particularly important where a species has specific habitat requirements and preferred habitats are not unevenly distributed in the landscape. Spatially explicit models can be designed to look at a range of questions and can be tailored to explore issues related specific species and human animal interactions.

The population viability analysis can be used to examine the influence of different population parameters and environmental variability on population size and population persistence. A recent version allows for simple metapopulation structure (Lacey, 1993). They are widely used and so the results of these studies can be readily compared with other studies in the literature. Their generality has its drawbacks, however, as these models are less flexible. For example, in Vortex, only one mortality rate can be used for all age classes above age at first breeding.

4.3. What do population-based models predict about the effectiveness of different population control methods?

4.3.1. General methodology for each species

To model at the population level, Jenson’s (1995) method was employed for foxes, red deer, brown hares and mink. This involves taking the matrix population models developed by Leslie (1945), and making them density-dependent. This latter method requires knowledge of the stable age structure of the population under investigation, and its carrying capacity.

The general methodology for each species was as follows:

First, national estimates for carrying capacity were calculated. These assumed the population could grow unrestricted by any factors other than intrinsic mortality. Population densities for different habitats were obtained (from estimates provided by Macdonald et al., 1998), and multiplied by the national coverage of each habitat type, derived from the Institute of Terrestrial Ecology’s Land Cover Map (LCM) held in the Countryside Information System.

Predictions of actual population sizes and distributions from satellite-derived land cover data such as these has been attempted, but found to be unreliable (Cardillo et al., 1999). The concept of ‘carrying capacity’ was therefore used to represent the maximum possible land area that could be occupied by the mammal species in the UK, assuming that distribution is both determined and limited solely by habitat.

Initial population sizes in the model were assumed to be half of this calculated carrying capacity.

These estimates are almost certainly too high, as not all habitats included in each land cover category will necessarily be suitable habitat for the mammals studied. Mammal distributions also show considerable geographic variation which may be independent of land cover. Caution must therefore be exercised in interpreting these results.

Age-specific life history data for fecundity and survival were obtained from the available literature. These data were then used to simulate population sizes using population models, written in the programming language C.

One of the simplest models of population growth is the discrete time logistic equation (May, 1974):

,

where are the population sizes at time t and t+1 respectively, d is a density dependent function involving carrying capacity and r is the intrinsic rate of increase. Matrix models of the form used by Jenson (1995) are direct analogies of this generalised equation, but the numbers of individuals of each age class are tracked, instead of just the total population size. The most important aspect of a matrix model is the element known as the Leslie Matrix, which appears as follows:

,

where F is the fecundity of age group (the number of young produced per individual of each age during the time period t) and P is the probability of survival from one age group to the next. The Leslie matrix (or a version of it) takes the place of the intrinsic rate of increase in the above logistic equation.

The matrix model was used to derive population trajectories over time for the four species modelled in this fashion – the red fox, the brown hare, the red deer and the American mink. Given sufficient time, the proportion of the total population in each age class will remain constant between generations – this is the stable age distribution, and is a property of the life history variables provided. These population trajectories represented the national trends of the four species over fifteen years.

In all modelling exercises, it is important to test the sensitivity of the model to the input parameters. To do this, upper and lower bounds for the life history variables were estimated, and sets of input data were generated at random between these bounds using a method called Latin Hypercube Sampling (Vose, 1996). Age-specific fecundity and survival were varied in this way, along with additional mortality caused by the culling methods employed. The model was run for 15 years for each of 1000 parameter sets. The population sizes that result from this analysis can be compared with the input parameters using partial correlation to discover which life history or control parameters have significant effects on the total population size. Binary logistic regressions of the same data set will reveal which life history or control parameters have significant effects on population increase (or presence/absence for less fecund species). This step reveals which culling parameters have a statistically significant effect on population size. We can then use that information to predict what level of control of these species would be required to prevent population increase, or cause population decrease.

The statistically significant culling factors elucidated in the previous modelling stage were used to investigate the sensitivity of fox, hare, red deer and mink populations to culling. Each significant parameter (here called JUVENILE CULL, SUB-ADULT CULL and/or ADULT CULL) was given a value between 0 and 0.9 (i.e. by culling between 0% and 90% of the age class). All combinations of culling mortalities were used to discover the response of the long term population trajectory, and thus levels of culling that would be needed to halt population increase or cause a population decrease.

The elements of the matrix population models are derived from life history characteristics of the species under investigation. Age specific fecundity and survival must be either available or estimated to construct these models.

4.3.1.a. Assumptions made

Models are simplifications of reality, and therefore assumptions must be made in the formulation of any modelling system. As long as the effects that these assumptions make on the output of the model are taken into account when interpreting the results, these assumptions should not reduce the usefulness of modelling. In the context of population modelling, the purpose is not to mimic reality, but to investigate the sensitivity of animal populations to anthropogenic change, and thus these assumptions can be accommodated.

Some general assumptions that apply to matrix population models are:

Any culling mortality (JUVENILE CULL, SUB-ADULT CULL and/or ADULT CULL) is applied to the pre-breeding population. Natural mortality is assumed to occur before culling mortality – the former is density-dependent, the latter is density independent. No pregnant animals are assumed to be culled.

The same density-dependent function was applied to both birth and death, as it is in the logistic equation. Although it would be more realistic to separate the effects of density dependence on these two processes, this would increase the difficulty of parameter estimation. It is simpler, for this investigatory approach, to estimate density-dependence using parameters that can be estimated – the carrying capacity (as described above) and the stable age structure (which is a property of the life history variables used).

4.3.2. How effective are methods to control fox populations?

4.3.2.a. Approach

In the matrix population modelling approach, anthropogenic control was simulated by culling a fixed proportion of the pre-breeding population. This culling was density independent, and different levels of culling could be applied to different age groups, namely the juveniles (defined as individuals yet to reach breeding age), sub-adults (individuals in the first year of breeding) and adults (individuals in a second or later breeding season).

The aim of this modelling approach is to discover what levels of culling are required to result in a long term population decline in Great Britain.

4.3.2.b. Data used

Red foxes are the most catholic species covered in this inquiry with regards to habitat preference, being found in nearly all land cover categories. Densities were estimated to be 1.5 foxes/km² for urban and suburban areas, 0.025 foxes/km² for bare land, bogs, salt marshes and other unfavourable habitats, and 2.5 foxes/km² for all other terrestrial habitats. The final estimation of the maximum number of foxes that Great Britain could sustain, given no anthropogenic restrictions, is approximately 434,400. See Table 4-28 for more details.

Estimates of annual fecundity and survival of foxes were derived from the available scientific literature. For the construction of matrix models, age-specific information for these two parameters is required. In many cases, these can be estimated from the available information.

Fecundity is at a peak in vixens of 2-4 years old (Harris & Smith, 1987), at which time the mean litter size for a vixen is 4.5 cubs (Harris, 1979). This potential fecundity of vixens is reduced by 20% due to late-term reproductive failure (Corbet & Harris, 1991). These data lead to the fecundity schedule in Table 4-2.

Survival data can be estimated from the age structure of a population. Culling occurs in most fox populations studied, and therefore survivorship data is likely to underestimate actual fox numbers in the absence of a cull.

The mean age structure over several populations of foxes was determined from Corbet & Harris (1991, p365). This can then be used to calculate age-specific survival – the probability of surviving to the next age class (Table 4-3).

4.3.2.c. Assumptions made

In addition to the general assumptions of the matrix models, the following initial assumptions are made in the fox models:

To derive age-specific fecundities it was necessary to extrapolate the mean fecundity of vixens in their prime to older animals. However, it is known that fecundity decreases in vixens over the age of four, and the mean yearly fecundity for all vixens of all ages was 3.14, the same as the mean yearly fecundity of the three regions given in Corbet & Harris (1991)

Density dependence is known to limit fox fecundity (see section 1.3.2.a). In these matrix models, as explained above, the effects on mortality and fecundity of density-dependence are combined into a single function.

The survivorship data used in the matrix modelling were taken from fox populations subjected to control. Therefore, levels of natural mortality are estimated to be higher than in reality, so the subsequent models underestimate fox populations and thus estimates of culling levels required to control fox populations are a little high (that is, a lower culling mortality will produce the same change). However, the general principles remain the same because the survival data used was taken from a number of fox populations with different levels of population control.

4.3.2.d. Results

The intrinsic rate of increase of the modelled fox population can be calculated from the Leslie matrix by eigen analysis. This results in a value of 1.13, reflecting the relatively high reproductive rate of fox populations. Given no form of population regulation (either natural or anthropogenic), this rate of increase indicates that the fox population will increase by 12.2% every year. This potential for population growth is almost never realised, because the habitat restricts the maximum size of a population and density-dependent processes mean that population growth rates decrease as the carrying capacity is reached. The stable age structure of the modelled population is shown in Table 4-4., and the results of the partial correlation in Table 4-5.

The sensitivity analysis of the matrix model showed that long term population size of the model was significantly affected by applying ADULT CULL and SUB-ADULT CULL on a yearly basis, but JUVENILE CULL did not significantly affect population trends. This is interesting, because it suggests that modelled fox populations are able to compensate for high juvenile mortality, presumably by lowered mortality of the other age classes and an increased fecundity.

The intrinsic rate of increase indicates the ease with which the population recovers from perturbation. Because this rate of increase is achieved only at low densities, it is important to examine long term population trends under different culling regimes. When animals are removed from a population, compensation may occur, so that natural mortality or emigration is lower in the following year, allowing the population to grow faster than it did before the cull (see section 1.3.2). By using the population models we can discover the sensitivity of the fox population to man-made perturbation by imposing a yearly cull of different age groups. By examining all levels of cull mortality, we can investigate the responses of a population to this control

The sensitivity analysis reported above indicated that fox population sizes are significantly affected by SUB-ADULT CULL and ADULT CULL. These two parameters were therefore varied in tandem, resulting in an estimation of the overall predicted effect on final population size after fifteen years of simulation.

For the comparison of sub-adult fox cull and adult fox cull, the results can be seen in Figure 4-1, and summarised in Table 4-6

The model suggests that a control of more than one age class will produce the most effective control of fox populations in the long term.

Figure 4-1 Illustrates the predicted effect of a fox cull. The red area indicated parameter values that result in either no increase or a decrease in total fox population size after fifteen years, the blue area indicates the parameter values resulting in net increase after 15 years.

Without ADULT CULL, over 30% of sub-adults must be removed from the population each year to see a small decrease in total fox numbers after 15 years. At this level of SUB-ADULT CULL, if over 10% of the adult foxes are also removed each year, then there will be a greater than ten-fold reduction of fox numbers after fifteen years.

A high value for sub-adult cull (greater than 70% per year) will result in extinction of fox populations within 15 years, whatever the level of adult cull.

The values given in Figure 4-1 and Table 4-6 represent the response of the modelled population only in terms of proportion of individuals in each age class affected by the levels of each CULL parameter. To compare this output to actual culling intensity it is necessary to convert the proportion affected by these parameters into densities affected.

To calculate the density on the ground of individuals culled, the stable age class distributions resulting from the matrix model were converted into fox density per age class by multiplying the proportion of the age class by the density per km² of the whole fox population at carrying capacity. Data for real fox densities were available from Heydon et al. (2000). These values are an average fox density of 1/km² and a maximum of 1.2/km².

This calculation allowed the production of population response curves for different levels of CULL parameters. These plots are shown in Figure 4-2.

Figure 4-2 The percentage of the fox population remaining after 15 years, assuming an average carrying capacity of 1.0/km² (diamond symbols) and a maximum carrying capacity of 1.5/km² (red symbols) against different denisities of adults and sub-adults culled/km² for Wales, West Midlands and East Norfolk. These desnities were adjusted to exclude jueveniles culled (proportion juveniles = 0.563) assuming even selection of the age classes, giving 0.40, 0.21 and 0.23 respectively. This corresponds to limiting the population by 50-10%, causing population extinction over 15 years.

4.3.3. How effective are methods to control hare populations?

4.3.3.a. Approach

In the matrix population modelling approach, anthropogenic control was simulated by culling a fixed proportion of the pre-breeding population. This culling was density independent, and different levels of culling could be applied to different age groups, namely the juveniles (defined as individuals yet to reach breeding age), sub-adults (individuals in the first year of breeding) and adults (individuals who have reached peak reproductive performance).

The aim of this modelling approach is to discover what levels of culling are required to result in a long term population decline in Great Britain

4.3.3.b. Data used

Brown hare are found primarily in open grassland and farmland; and occur at a wide range of densities, even in similar habitats, and year-to-year variation is considerable. On arable land, average densities can be 14.4 hares/km², for pastoral and managed grasslands 9.8 hares/km², and for less favourable grasslands, densities average 7.2 hares/km². The final estimation of the maximum number of hares that Great Britain could sustain, given no anthropogenic restrictions, is approximately 1 million.

Brown hares have 1–4 litters per year (average 3) and 1–4 leverets per litter (max 10) (Macdonald & Barrett, 1993). With a mean of 2.5 leverets per litter, and 3 litters per year, the mean fecundity for a female in its reproductive prime (3-4 years old, according to Frylestam, 1980) is 7.5 young per year. Fecundity is known to be lower in hares younger and older than this age group (Frylestam, 1980), so we assumed a slow decrease in yearly reproductive output up to the age of 7 (the average lifespan of a wild hare, Corbet & Harris, 1991). Hares mature at about 8 months old, so the first age group (0-1 year) can produce approximately 4/12 of the young that a hare in her prime can achieve. These data lead to the fecundity schedule in Table 4-7:

Survival data (Table 4-8) can be estimated from the age structure of a population. Culling occurs in most hare populations studied, and therefore survivorship data is likely to underestimate actual hare numbers in the absence of a cull.

The mean age structure over several populations of hares was determined from Tapper (1991); age categories are less than 1 year, 1–2 years, 2–3 years, 3 years or older). This can then be used to calculate age-specific survival – the probability of surviving to the next age class. We have extrapolated the age structure of the population to account for hares above the age of 3. This extrapolation is based on the fact that although first winter mortality is thought to be higher, mortality is otherwise constant with age (Broekhuizen, 1979). The number of hares in the 3+ category remains the same as that given in Tapper (1991), approximately 22.

4.3.3.c. Assumptions made

In addition to the general assumptions of the matrix models, the following initial assumptions are made in the hare models:

Extrapolation of prime fecundity was required to obtain age specific production for older hares. With no data available on this, we assumed a slow decline from an average of 7.5 leverets per year for a hare in her prime, to 4 leverets per year for a seven year old individual. Because of this assumption, the model is more likely to underestimate hare productivity than overestimate.

The matrix models used here cannot accurately take into account the multivoltine pattern of breeding in hare populations (i.e. the fact that they have multiple litters in a year). However, the lowest age class is given a value for fecundity to account for the fact that individuals can achieve breeding success in the same year that they were born, at an age of 8 months. In real populations, fecundity of female hares changes throughout the year, and this is not accounted for either. Optimistic values for hare productivity have therefore been used throughout this modelling section. The effect of the failure of the matrix model to allow for multivoltine breeding is only likely to be important in small populations, as density dependent effects on fecundity will come into play in populations close to the carrying capacity.

The initial age structure of hare populations was derived from bag counts from spring shoots on estates in England. The data are therefore assuming that the numbers of hares shot of each age is representative of the age-structure of the pre-breeding hare population.

Survival data was estimated from a culled population. The matrix model is likely to output hare populations that are under-estimates in the absence of a cull

Extrapolation of the survival data was needed for hares over 3 years old. However, it is known that adult hare mortality stays constant with age (Broekhuizen, 1979), and we were able to use this information to estimate the number of hares from their fourth year onwards.

4.3.3.d. Results

The intrinsic rate of increase of the modelled hare population can be calculated from the Leslie matrix by eigen analysis. This results in a value of 1.65, reflecting the high reproductive rate of hare populations – given no form of population regulation (either natural or anthropogenic), this rate of increase indicates that the hare population will double every year. This potential for population growth is almost never realised, because the habitat restricts the maximum size of a population and population growth rate decreases as the carrying capacity is approached. The stable age structure of the modelled population is given in Table 4-9.

The results of the partial correlation are as follows (Table 4-10):

The sensitivity analysis of the matrix model showed that long term population size of the model was significantly affected by applying juvenile cull, sub-adult cull and adult cull on a yearly basis. Hare are so productive that huge populations can result from unperturbed populations. Removal of breeding individuals can substantially affect the long term population trends, as can removal of leverets. However, even with large levels of CULL parameters, the resultant populations in the long term are still huge (just not as huge as in situations without a cull), so it is necessary to investigate further to find the true effects of culling on the predicted population trends.

The intrinsic rate of increase indicates the facility of the population to recover from perturbation. Because this rate of increase is only achieved at low densities, it is important to examine long term population trends under different culling regimes. When animals are removed from a population, compensation may occur, so that natural mortality or emigration is lower in the following year, allowing the population to grow faster than it did before the cull. By using the population models we can discover the sensitivity of the hare population to man-made perturbation by imposing a yearly cull of different age groups. By examining all levels of cull mortality, we can investigate the responses of a population to this control

The sensitivity analysis reported above indicated that hare population sizes are significantly affected by culling of all age classes, juveniles, sub-adults and adults. These three parameters were therefore varied in tandem, resulting in an estimation of the overall predicted effect on final population size after fifteen years of simulation.

For the comparison of sub-adult hare control and adult hare control, the results can be seen in Figure 4-3 and summarised in Table 4-11. With only sub-adult cull and adult cull parameters, the model output indicates that hare populations are impossible to control unless virtually every individual is removed. This is because of the high reproductive potential of each individual in the population.

For the comparison of juvenile hare control and sub-adult hare control, the results can be seen in Figure 4-3, and summarised in Table 4-12. The model output indicates that control of hare populations is possible through application of both juvenile cull and sub-adult cull. The effect would be to remove individuals before they have a chance to breed, and thus population control is achieved. Control levels need to be quite high to get this effect, with more than 50% JUVENILE CULL each year without sub-adult control, down to more than 10% JUVENILE CULL with more than 70% sub-adult cull.

For the comparison of juvenile hare control and adult hare control, the results can be seen in Figure 4-3, and summarised in Table 4-13.

Figure 4-3 Illustrates the predicted effect of a hare cull. The red area indicated parameter values that result in either no increase or a decrease in total hare population size after 15 years, the blue area indicates the parameter values resulting in a new increase after 15 years. Each graph shows the effects of varying two of the three culling parameters.

The predicted effects of juvenile control combined with adult control is similar to those given above, but ADULT CULL must be more severe than sub-adult cull to give the same effect.

The values given in the above tables and figures only represent the response of the modelled population in terms of proportion of individuals in each age class affected by the levels of each CULL parameter. To compare this output to actual culling rates it is necessary to convert the proportion affected by these parameters into densities affected.

To calculate the density of individuals culled, the stable age class distributions resulting from the matrix model were converted into hare density per age class by multiplying the proportion of the age class by the density per km² of the whole hare population at carrying capacity. Hare population densities were taken from Table 4-28 – average hare density is 10.3 / km² and a maximum hare density is 14.4 / km².

Figure 4-4 The percentage of the hare population remaining after 15 years, assuming an average carrying capacity of 10.3/km² (blue symbols) and a maximum carrying capacity of 14.4/km² (red symbols against different desities of adults, sub-adults and juveniles culled/km².

This calculation allowed the production of population response curves for different levels of the CULL parameters. These plots are shown in Figure 4-4.

4.3.4. How effective are methods to control deer populations?

4.3.4.a. Approach

In the matrix population modelling approach, anthropogenic control was simulated by culling a fixed proportion of the pre-breeding population. This culling was density independent, and different levels of culling could be applied to different age groups, namely the juveniles (defined as individuals yet to reach breeding age), sub-adults (individuals in the first year of breeding) and adults (individuals who have reached peak reproductive performance).

The aim of this modelling approach is to discover what levels of culling are required to result in a long term population decline in Great Britain.

4.3.4.b. Data used

Red deer can utilise a range of habitats, including woodland, grassland, moor and scrub. Density is highest in open woodland habitats (22.5 deer/km²) and lower in grassland and moor habitats (15.5 deer/km²). These estimates are not achieved by English populations of red deer, which approach nearer 1 deer/km² even in open woodland habitats. However, the purpose of this exercise is to estimate the maximum density that the species could achieve, and so the higher values for Scottish populations were used. The final estimation of the maximum number of red deer that Great Britain could sustain, given no anthropogenic restrictions, is approximately 1.4 million individuals.

Scottish populations of red deer in open upland habitats have been extensively studied, and age specific life history data have been published (Lowe, 1969) and used in matrix models (Usher, 1973). However, the Scottish populations have a low fecundity in comparison with data collected from the populations in the south west of England, who dwell in a more productive environment (Ratcliffe, 1987). Data from the south west indicate 73% of yearling hinds and 91.5% of older hinds are pregnant (Langbein, 1997; data from the two regions have been averaged). We have assumed that fecundity is not constant with age, but reaches a peak by the 7^th year, then declines slowly thereafter. This is in accordance with the fecundity schedule of Scottish upland populations (Lowe, 1969).

These assumptions produce the fecundity schedule in Table 4-14 (values for Scottish upland deer have been included for reference). In the absence of data which is more representative of English populations of red deer, the age-specific survival – the probability of surviving to the next age class – of the Scottish populations was used (Lowe, 1969) (Table 4-15).

4.3.4.c. Assumptions made

In addition to the general assumptions of the matrix models, the following initial assumptions arise from the deer models:

The models are designed to model populations in Great Britain, however, it is the English populations of deer that are of specific interest in this Inquiry. As the fecundity of English deer is greater than that of Scottish deer, we have chosen to use the English data, as it will produce a ‘best case scenario’. However, if these results were to be extrapolated to Scottish populations, caution must be used in coming to the same conclusions as in the following section.

Extrapolation was necessary to derive age-specific fecundities for deer over 2 years old. This information was available for Scottish deer, but these populations are less productive than English populations, and so could not be used. However, the pattern of age specific fecundity could be derived from this life history data – in Lowe’s (1969) data, fecundity steadily increases to a maximum at 6-7 years of age, and declines thereafter. The same pattern was used in the matrix models, ensuring that the mean productivity of hinds older than 2 remained the same in Langbein (1997).

In the absence of age-specific survival data for the English populations of red deer, details from the life table of Scottish deer have been used.

4.3.4.d. Results

The intrinsic rate of increase of the modelled red deer population can be calculated from the Leslie matrix by eigen analysis. This results in a value of 1.16, reflecting the relatively high reproductive rate of deer populations – given no form of population regulation (either natural or anthropogenic), this rate of increase indicates that the deer population will increase by 14.8% every year. This potential for population growth is almost never realised, because the habitat restricts the maximum size of a population and population growth rates decrease as the carrying capacity is reached. However, this intrinsic rate of increase indicates the facility of the population to recover from perturbation. The stable age structure of the modelled population is given in Table 4-16, and the results of the partial correlation in Table 4-17.

All of the CULL parameters had a significant effect on the long term population trends of this matrix model. The annual population growth of this matrix model is comparable with that of the fox, however, this is due to different reasons. Red deer produce many less young than foxes, but they live longer and so the population growth rate is similar. Because fewer young are produced in deer populations, the matrix model is more sensitive to removal of yearling animals, each of which has a long and productive life ahead of it.

The intrinsic rate of increase indicates the facility of the population to recover from perturbation. Because this rate of increase is only achieved at low densities, it is important to examine long term population trends under different culling regimes. When animals are removed from a population, compensation may occur, so that natural mortality or emigration is lower in the following year, allowing the population to grow faster than it did before the cull. By using the population models we can discover the sensitivity of the red deer population to man-made perturbation by imposing a yearly cull of different age groups. By examining all levels of cull mortality, we can investigate the responses of a population to this control

The sensitivity analysis reported above indicated that red deer population sizes are significantly affected by culling of all age classes, juveniles, sub-adults and adults. These three parameters were therefore varied in tandem, resulting in an estimation of the overall predicted effect on final population size after fifteen years of simulation.

For the comparison of sub-adult deer control and adult deer control, the results can be seen in Figure 4-5, and summarised in Table 4-18. This simulation suggests that the model is quite sensitive to the sub-adult cull parameter. Fifteen years of removing sub-adults at levels of 10% or greater will result in population decline. The populations are more resistant to adult cull – without removal of sub-adults, adult cull values greater than 10% will halt population decline, but greater than 50% adult cull is predicted to be required to see a significant sustained reduction in numbers.

For the comparison of juvenile deer control and sub-adult deer control, the results can be seen in Figure 4-5, and summarised in Table 4-19. The matrix model combining juvenile cull and sub-adult cull predicts that any level of removal greater than 10% of either of these age classes will result in modelled population decline. Significant culling of the fawns and yearlings will result in population extinction within 15 years – a level of juvenile cull and sub-adult cull of more than 30% will achieve this, although if more juveniles are culled, less sub-adults need to be removed to achieve the same result, and vice versa.

For the comparison of juvenile deer control and adult deer control, the results can be seen in Figure 4-5 and summarised in Table 4-20. This simulation compares well with the sub-adult cull / adult cull scenario described above. It therefore indicates that the modelled population will decline if fawns are culled at an intensity of greater than 10%, but can withstand up to 50% adult cull given no juvenile cull.

The values given in the above tables and figures only represent the response of the modelled population in terms of proportion of individuals in each age class affected by the levels of each CULL parameter. To compare this output to actual culling rates it is necessary to convert the proportion affected by these parameters into densities affected.

Figure 4-6 The percentage of the deer population remaining after 15 years, assuming an average carrying capacity of 6.5/km² (blue symbols) and a maximum carrying capacity of 22.5/km² (red symbols) against different densities of adults, sub-adults and juveniles culled/km².

To calculate the density of individuals culled, the stable age class distributions resulting from the matrix model were converted into red deer density per age class by multiplying the proportion of the age class by the density per km² of the whole deer population at carrying capacity. Average deer population density were not taken from Table 4-28 as these respresent an average for the whole of Great Britain, and do not represent red deer populations in England. Instead, average deer population densities were taken from Langbein et al. (1998). The average red deer population was therefore set at 6.5 / km². Maximum densities of 22.5 deer/km² were taken from national data presented in Table 4-28. This calculation allowed the production of population response curves for different levels of the CULL parameters. These plots are shown in Figure 4-6.

4.3.5. How effective are methods to control mink populations?

In the matrix population modelling approach, anthropogenic control was simulated by culling a fixed proportion of the pre-breeding population. This culling was density independent, and different levels of culling could be applied to different age groups, namely the juveniles (defined as individuals yet to reach breeding age), sub-adults (individuals in the first year of breeding) and adults (individuals who have reached peak reproductive performance).

The aim of this modelling approach is to discover what levels of culling are required to result in a long term population decline in Great Britain

4.3.5.b. Data used

Mink were introduced into Britain in the 1930s, and have been increasing on a national scale ever since. Mink are riparian, and found in association with both still and running water, as well as estuaries and rocky coastlines. Mink density in these regions are dependent largely on the availability of rabbits, their primary food source, but average density is about 0.35 mink/km of river (Macdonald et al., 1998). The final estimation of the maximum number of mink that Great Britain could sustain, given no anthropogenic restrictions, is approximately 378,600. See Table 4-28 for more details.

The details on mink life histories are sparse. Much of the available data is derived from studies on captive mink, and naturally these do not reflect the mortality factors experienced by wild populations. Mink have one litter per year, four to six young per litter (Corbet & Harris, 1991), although larger litters are possible (up to 17 in captivity). The percentage of barren females is smaller and the litter size tends to increase in older females (Dunstone, 1993). These data result in the fecundity schedule in Table 4-21.

An age structure from Hatler (1976, by way of Dustone, 1993) indicates longevity is about 5 years in the wild, although 10 years has been recorded in captive animals (Macdonald, Mace & Rushton, 1998). This age structure allowed the calculation of age-specific survival – the probability of surviving to the next age class (Table 4-22).

4.3.5.c. Assumptions made

In addition to the general assumptions of the matrix models, the following initial assumptions arise from the hare models:

As before, it was necessary to estimate age-specific fecundity from the data provided in the literature. Mink productivity was assumed to increase with age, as reported by Dunstone (1993).

The survival data for the matrix models were taken from a study of mink populations on Vancouver Island. It is not known whether these data are representative of British mink populations, or whether that population was subject to hunting. It is likely that the data used overestimates mink mortality, and therefore populations should grow faster than indicated in the following results.

4.3.5.d. Results

The intrinsic rate of increase of the modelled mink population can be calculated from the Leslie matrix by eigen analysis. This results in a value of 1.49, reflecting the high reproductive rate of mink populations – given no form of population regulation (either natural or anthropogenic), this rate of increase indicates that the mink population will increase by 40% every year. This potential for population growth is almost never realised, because the habitat restricts the maximum size of a population and population growth rates decrease as the carrying capacity is reached. However, this intrinsic rate of increase indicates the facility of the population to recover from perturbation. The stable age structure of the modelled population is given in Table 4-23. The results of the partial correlation are given in Table 4-24