Cervus elaphus (red deer)
- Summary of Invasiveness
- Taxonomic Tree
- Notes on Taxonomy and Nomenclature
- Distribution Table
- History of Introduction and Spread
- Risk of Introduction
- Habitat List
- Biology and Ecology
- Notes on Natural Enemies
- Means of Movement and Dispersal
- Pathway Causes
- Impact Summary
- Economic Impact
- Environmental Impact
- Threatened Species
- Social Impact
- Risk and Impact Factors
- Uses List
- Similarities to Other Species/Conditions
- Prevention and Control
- Gaps in Knowledge/Research Needs
- Links to Websites
- Principal Source
- Distribution Maps
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PicturesTop of page
IdentityTop of page
Preferred Scientific Name
- Cervus elaphus Linnaeus, 1758
Preferred Common Name
- red deer
International Common Names
- English: European red deer; maral deer
- Spanish: Ciervo colorado; ciervo común; venado
- French: cerf commun; cerf elaphe
Local Common Names
- Denmark: kronhjort
- Germany: Edelhirsch; Hirsch; Hirsch, Edel-; Hirsch, Rot-; Rothirsch; Rotwild
- Iran: gawasn
- Italy: cervo
- CERVEL (Cervus elaphus)
Summary of InvasivenessTop of page
Cervus elaphus, the red deer, is a species of large deer that is native to much of Europe and western Asia, as well as parts of North Africa. The IUCN classes it in its native range as ‘least concern’, although some populations are listed as endangered or rare (IUCN, 2017). It has been introduced to several parts of the world for hunting or deer farming; the populations in Australia, New Zealand and South America are listed among the 100 worst invasive alien species by the IUCN (Lowe et al., 2004). They have invaded a wide range of climatic zones from the tropics to temperate areas, and habitats from lowland rainforests, southern beech forests and sclerophyll eucalypt woodlands to savannahs and alpine grasslands. They may affect the structure and dynamics of the native vegetation, and may compete with native herbivores. They are implicated in the introduction of bovine tuberculosis into wildlife maintenance hosts in New Zealand. The mixed status of introduced wild C. elaphus as pests and/or as hunting assets, and farming for venison production or as trophy animals with consequent genetic mixing within the genus, complicates their legal status and management in many jurisdictions.
Taxonomic TreeTop of page
- Domain: Eukaryota
- Kingdom: Metazoa
- Phylum: Chordata
- Subphylum: Vertebrata
- Class: Mammalia
- Order: Artiodactyla
- Suborder: Ruminantia
- Family: Cervidae
- Genus: Cervus
- Species: Cervus elaphus
Notes on Taxonomy and NomenclatureTop of page
Evidence from mitochondrial DNA (Ludt et al., 2004; Skog et al., 2009) suggests that the genus Cervus evolved in central Asia where one or two primordial taxa still exist, C. elaphus bactrianus and C.elaphus yarkandensis. About seven to eight million years ago, as climates cooled, these ancestral deer spread and were separated into (a) an eastern group called wapiti or elk (C. canadensis) with subspecies in eastern Asia and North America and related species in east and south Asia (e.g. sika, Thorold’s, sambar and rusa deer – C. nippon, C. albirostris, C. = Rusa unicolor and C. = Rusa timorensis, respectively), and (b) a western group called red deer (C. elaphus) with several subspecies in western Europe, the Balkans, the Middle East and North Africa. The IUCN assessment of red deer (IUCN, 2017) recognises seven sub-species of western red deer within three main genetic lineages based on geographic separation – one in western/central Europe, one in eastern Europe and the Middle East and the third in North Africa and Sardinia (Lorenzini and Garofalo 2015). However, IUCN (2017) note that the taxonomy of C.elaphus remains unsettled. Its separation into many more subspecies is not supported by the analysis of Ludt et al. (2004) analysis, although (IUCN, 2017) note that further sub-specific taxonomic revision would be beneficial.
Red deer and wapiti have often been considered as different forms of the same species, under the name of C. elaphus, but most recent studies conclude that they are two separate valid species, as summarized in their entries in the IUCN Red List of Threatened Species (IUCN, 2017).
Incidentally, the use of the common name ‘elk’ for C. canadensis in North America can be confusing because in Europe ‘elk’ is the name given to Alces alces – known as moose in North America.
C. elaphus freely hybridise with wapiti (and sika) deer (C. canadensis and C. nippon) in the wild, producing fertile offspring, when made sympatric by human activities, e.g. in New Zealand with both wapiti and sika (Fraser, 2005, Nugent and Fraser, 2005), or in Scotland with sika (Goodman et al., 1996), which supported older views that the wapiti-red deer cline of populations formed a single ring species around the northern hemisphere (King, 2005). C. elaphus can also produce fertile offspring when artificially inseminated by Pere David’s deer (Elaphurus davidianus) (Fennessy and Mackintosh, 1992) and sambar deer (C. unicolor) (Muir et al., 1997).
DescriptionTop of page
C. elaphus across its native range is a highly variable species – hence the taxonomies dividing it into many sub-species. Among the western populations (the ancestors of introduced populations) the largest occur in the Carpathian Mountains (males weighing up to 500 kg) and the smallest on islands in the Mediterranean (males weighing up to 100 kg). The ‘island rule’ where large species on islands tend to be smaller than on the mainland, applies to cervids (e.g. Simard et al., 2008; Sinclair and Parkes, 2008). New Zealand red deer body weights averaged 103 kg for adult males and 75 kg for adult females in one population sampled (Nugent and Fraser, 2005).
Food quality and abundance has a large influence on the body size of C. elaphus. For example, in New Zealand, deer from high density populations that had passed through an irruptive increase and decline and had less food per capita (Caughley, 1970; Forsyth and Caley, 2006) weighed 55 – 65% of the weight of animals in low density populations yet to irrupt (Challies, 1985). The former were smaller, rarely bred as yearlings and had higher mortality rates (Challies, 1985).
Generally, C. elaphus have a plain brown upper body pelage with a light grey or pale throat and belly pelage in females and a darker belly pelage in males. They have a dark muzzle and hooves. In winter the pelage thickens and has brown or greyish guard hairs and a faint dorsal stripe. Adults typically have no spots and only males have antlers, which in the southern hemisphere introduced populations are shed in late winter (August).
DistributionTop of page
The native range of C. elaphus extends from the primordial subgroup (Yarkand and Bactrian deer) in western China and around the Caspian Sea, through maral red deer in the Middle East, to the core red deer populations over most of temperate Europe from Estonia and Bulgaria in the east and Norway and Sweden in the north to Spain and Portugal in the west. The relationships of the insular outliers on Sardinia and Corsica (the latter population being extinct by 1970 but reintroduced form Sardinia in 1985 – Puddu et al., 2009) are unclear. Some claim that they are related to the only North African red deer, found in the Atlas Mountains (Wilson and Mittermeier, 2011), but mitochondrial DNA phylogenies suggest that they are derived from Italian mainland populations taken to the islands in the late Neolithic (Puddu et al. 2009; Randi et al. 2001).
Introduced populations of free-ranging C. elaphus are found over 120 000 km2 (44%) of the three largest islands of New Zealand (Nugent and Fraser, 2005), 78 730 km2 of eastern and southern Australia (West, 2011), and about 50 000 km2 of Argentina and Chile south of 34°S (Flueck et al., 2003). Red deer have not reached the limit of suitable range in these countries, and may not do so if active management to stop their spread is successful (Nugent et al., 2011).
Smaller wild populations occur in Uruguay (Safari Club International, 2016), Peru, South Africa, and California, and members of the species were also released on the island of Bioko (Fernando Pó) in Equatorial Guinea, where they may still persist (Long, 2003).
Red deer are also present as exotic animals on farms and game ranches in several countries.
Distribution TableTop of page
The distribution in this summary table is based on all the information available. When several references are cited, they may give conflicting information on the status. Further details may be available for individual references in the Distribution Table Details section which can be selected by going to Generate Report.
|Continent/Country/Region||Distribution||Last Reported||Origin||First Reported||Invasive||Reference||Notes|
|Afghanistan||Localised||Native||Not invasive||Moheb et al., 2016||Bactrian deer, Tarim group of red deer|
|Armenia||Localised||Not invasive||WWF, 2016||Maral group reintroduced in 2015|
|Azerbaijan||Present||Native||Not invasive||Guliyev, 2014|
|-Xinjiang||Localised||Native||Not invasive||Skog et al., 2009||Yarkand deer|
|Georgia (Republic of)||Present||Native||IUCN, 2017|
|Iran||Localised||Native||Not invasive||Firouz, 2005||Maral group|
|Israel||Absent, formerly present||IUCN, 2017||Extinct|
|Jordan||Absent, formerly present||IUCN, 2017||Extinct|
|Kazakhstan||Present||Native||Not invasive||Moheb et al., 2016||Bactrian deer, Tarim group|
|Kyrgyzstan||Localised||Native||Not invasive||Moheb et al., 2016||Listed as vulnerable, Tarim group, Bactrian deer|
|Lebanon||Absent, formerly present||IUCN, 2017||Extinct|
|Syria||Absent, formerly present||IUCN, 2017||Extinct|
|Tajikistan||Localised||Native||Not invasive||Moheb et al., 2016||Bactrian red deer|
|Turkey||Present||Native||Not invasive||IUCN, 2017|
|Uzbekistan||Present||Not invasive||Moheb et al., 2016||Reintroduced; Bactrian red deer|
|Algeria||Present||Native||Not invasive||IUCN, 2017||Barbary red deer|
|Equatorial Guinea||Localised||Introduced||1954||Not invasive||Long, 2003||Bioko (Fernando Po) island. ‘Still presumably established’.|
|Morocco||Localised||Not invasive||IUCN, 2017||Reintroduced|
|South Africa||Present, few occurrences||Introduced||Not invasive||Long, 2003||A few escapes from captivity|
|Canada||Present only in captivity/cultivation||Introduced||Agriculture and Agri-Food Canada, 2016||Farmed in some provinces, including a few in the Atlantic provinces (422 as of 2014). Numbers have declined from approx. 17500 in 2006 to approx. 7500 in 2016.|
|-Ontario||Present only in captivity/cultivation||Introduced||Not invasive||Agriculture and Agri-Food Canada, 2016||Farmed – 1000 in 2016|
|-Quebec||Present only in captivity/cultivation||Introduced||Not invasive||Agriculture and Agri-Food Canada, 2016||Farmed – 6092 in 2016|
|Mexico||Present, few occurrences||Introduced||Not invasive||Gallina and Escobedo-Morales, 2009||Farmed but with escapes|
|-California||Present, few occurrences||Introduced||Not invasive||Safari Club International, 2016||Farmed with escapes (San Luis Obispo)|
|-Kentucky||Present||Introduced||Not invasive||Safari Club International, 2016||Woodlands Nature Refuge|
|-Maine||Present only in captivity/cultivation||Introduced||Not invasive||Safari Club International, 2016||Game ranches|
|-Texas||Present, few occurrences||Introduced||Safari Club International, 2016||Free-range on some game ranches|
|Argentina||Widespread||Introduced||1906||Invasive||Flueck et al., 2003|
|Chile||Widespread||Introduced||1916||Invasive||Flueck et al., 2003|
|Peru||Localised||Introduced||1948||Not invasive||Long, 2003|
|Uruguay||Localised||Introduced||1930s||Safari Club International, 2016||Introduced from Argentina. Also now on game ranches.|
|Albania||Absent, formerly present||IUCN, 2017||Extinct|
|Austria||Present||Native||Not invasive||IUCN, 2017|
|Belarus||Present||Native||Not invasive||IUCN, 2017|
|Belgium||Present||Native||Not invasive||IUCN, 2017|
|Bosnia-Hercegovina||Present||Native||Not invasive||IUCN, 2017|
|Bulgaria||Present||Native||Not invasive||IUCN, 2017|
|Croatia||Present||Native||Not invasive||IUCN, 2017|
|Czech Republic||Present||Native||Not invasive||IUCN, 2017|
|Denmark||Present||Native||Not invasive||IUCN, 2017|
|Estonia||Present||Native||Not invasive||IUCN, 2017|
|France||Present||Native||Not invasive||IUCN, 2017|
|-Corsica||Localised||Not invasive||Puddu et al., 2009||Reintroduced 1980s|
|Germany||Present||Native||Not invasive||IUCN, 2017|
|Greece||Present||Not invasive||IUCN, 2017||Reintroduced|
|Hungary||Present||Native||Not invasive||IUCN, 2017|
|Ireland||Present||Not invasive||IUCN, 2017|
|Italy||Present||Native||Not invasive||IUCN, 2017|
|Latvia||Present||Native||Not invasive||IUCN, 2017|
|Lithuania||Present||Native||Not invasive||IUCN, 2017|
|Luxembourg||Present||Native||Not invasive||IUCN, 2017|
|Macedonia||Present||Native||Not invasive||IUCN, 2017|
|Moldova||Present||Native||Not invasive||IUCN, 2017|
|Montenegro||Present||Native||Not invasive||IUCN, 2017|
|Netherlands||Present||Native||Not invasive||IUCN, 2017|
|Norway||Present||Native||Not invasive||IUCN, 2017|
|Poland||Present||Native||Not invasive||IUCN, 2017|
|Portugal||Present||Native||Not invasive||IUCN, 2017|
|Romania||Present||Native||Not invasive||IUCN, 2017|
|Russian Federation||Localised||Not invasive||IUCN, 2017|
|Serbia||Present||Native||Not invasive||IUCN, 2017|
|Slovakia||Present||Native||Not invasive||IUCN, 2017|
|Slovenia||Present||Native||Not invasive||IUCN, 2017|
|Spain||Present||Native||Not invasive||IUCN, 2017|
|Sweden||Present||Native||Not invasive||IUCN, 2017|
|Switzerland||Present||Native||Not invasive||IUCN, 2017|
|UK||Present||Native||Not invasive||IUCN, 2017|
|Ukraine||Present||Native||Not invasive||IUCN, 2017|
|-New South Wales||Localised||Introduced||Invasive||Moriarty, 2004|
|New Zealand||Widespread||Introduced||1851||Invasive||Donne, 1924; Nugent and Fraser, 2005||North, South and Stewart Islands|
History of Introduction and SpreadTop of page
New Zealand: C. elaphus were first introduced to New Zealand in 1851 with at least 250 animals imported and released at over 50 sites in North, South and Stewart islands before 1919. Some were imported directly from Britain but most were first sent to Victoria in Australia and their offspring to New Zealand. Of those directly imported from Britain some came from ‘pure’ herds in Scotland, while others (and all those that came via Victoria) were bred in English deer parks and probably already of ‘mixed’ European parentage (Donne, 1924, Logan and Harris, 1967).
Each population spread at an average rate of about 1.6 km/year (Caughley, 1963), and red deer now occupy most of the native forests in New Zealand. Active management of deer held on deer farms outside these natural ranges is required to stop escaping deer from establishing wild populations in the few areas they have not reached by themselves (Fraser et al., 2003).
Australia: C. elaphus were first introduced to Australia in about 1860 from English game parks and released in all mainland states. Their populations did not spread until relatively recently and herd sizes remain modest (compared with sambar deer, Cervus unicolor, in Victoria). The largest population occurs in the Grampian Mountains in western Victoria (Moriarty, 2004).
South America: C. elaphus were imported from Germany in 1906 into a game park but by 1922 had escaped. However, the current population appears to have resulted from an introduction of 22 animals released in Neuquén Province of Argentina in 1922. German C. elaphus were imported into Chile in 1916 and, as a result of this and of natural spread from Argentina, now occupy large areas of that country. The southernmost herd is located on Argentina’s Staten Island (Isla de los Estados) near Tierra del Fuego. C. elaphus from Argentina were kept on a farm in Peru in 1948 but have subsequently escaped and formed a small wild herd (Long, 2003). A herd was introduced into Uruguay in the 1930s and persists in a small area along with more recent herds held on game ranches (Safari Club International, 2016).
Africa: C. elaphus were introduced to game farms in southern South Africa in 1895 and subsequent escapees apparently still persist (Long, 2003). In 1954, C. elaphus from Spain were released on the west African island of Fernando Pó (Bioko) in Equatorial Guinea, where they may still persist (Long, 2003).
IntroductionsTop of page
|Introduced to||Introduced from||Year||Reason||Introduced by||Established in wild through||References||Notes|
|Natural reproduction||Continuous restocking|
|Equatorial Guinea||Spain||1954||Intentional release (pathway cause)||No||No||Long (2003)||Bioko (Fernando Po) island.|
|South Africa||1895||Escape from confinement or garden escape (pathway cause)||No||No||Long (2003)|
|Australia||UK||1860||Intentional release (pathway cause)||Yes||No||Moriarty (2004)|
|New Zealand||UK||1851||Intentional release (pathway cause)||Yes||No||Donne (1924)|
|Argentina||Germany||1906||Escape from confinement or garden escape (pathway cause)||Yes||No||Flueck et al. (2003)|
|Argentina||1922||Intentional release (pathway cause)||Yes||No||Long (2003)|
|Chile||Germany||1916||Intentional release (pathway cause)||Yes||No||Flueck et al. (2003)|
Risk of IntroductionTop of page
Many countries now have biosecurity legislation that would make deliberate importation and release of new species a more considered decision than those taken by New Zealand, Australia, Argentina and Chile a century ago. Most jurisdictions have apparently considered that the risks posed by C. elaphus are unacceptable as no red-deer-free country has deliberately introduced them for many decades, except some that used to be within their native range (Greece and Morocco), where reintroductions have occurred (IUCN, 2017). However, deer kept behind fences for production of venison and velvet or as a hunting resource present a risk that new free-ranging herds will establish if animals escape. This has happened in some places. C. elaphus were imported into Texas in the 1930s as a game animal and some live in a free-range state. Ranched herds also occur in Kentucky and California, the latter having established a wild herd when animals escaped (Safari Club International, 2016). C. elaphus are also held on ranches in Mexico where there were 179 areas covering 29 329 ha in 27 states in 2009 (Gallina and Escobedo-Morales, 2009). In Canada, C. elaphus from New Zealand were found to harbour the nematode Elaphostrongylus cervi while still in quarantine, leading Canadian authorities to destroy the deer and ban further imports from countries where the parasite is present (Gajadhar et al., 1994). Nevertheless, there are about 7500 red deer held on farms in Canada (mostly Quebec, Ontario and the Atlantic provinces), although the numbers are declining (Agriculture and Agri-Food Canada, 2016). In New Zealand about 900 000 red deer are held as farmed animals (Stats NZ, 2016); these present a risk when they escape into otherwise deer-free areas.
HabitatTop of page
In its native range, C. elaphus inhabits open deciduous woodland, mixed deciduous-coniferous and coniferous woodland, upland moors and open mountainous areas (sometimes above the treeline), Mediterranean maquis scrub, natural grasslands, pastures and meadows (IUCN, 2017). Introduced C. elaphus are highly adaptable, living in habitats ranging from tropical and subtropical rainforest, to semi-arid eucalypt woodlands to temperate eucalypt forests in Australia; from temperate lowland forests (including southern beech, Nothofagus, forests) to alpine grasslands in New Zealand; and from Valdivian rainforests (including Nothofagus forests) to Patagonian steppes in South America.
Habitat ListTop of page
|Terrestrial – Managed||Cultivated / agricultural land||Secondary/tolerated habitat||Natural|
|Managed forests, plantations and orchards||Secondary/tolerated habitat||Harmful (pest or invasive)|
|Managed grasslands (grazing systems)||Secondary/tolerated habitat||Harmful (pest or invasive)|
|Managed grasslands (grazing systems)||Secondary/tolerated habitat||Productive/non-natural|
|Terrestrial ‑ Natural / Semi-natural||Natural forests||Principal habitat||Harmful (pest or invasive)|
|Natural forests||Principal habitat||Natural|
|Natural grasslands||Principal habitat||Harmful (pest or invasive)|
|Natural grasslands||Principal habitat||Natural|
|Scrub / shrublands||Secondary/tolerated habitat||Natural|
Biology and EcologyTop of page
The red deer introduced to the New Zealand and Australia in the 19th century came from Britain, while those introduced to South America came from western Europe; they were therefore all part of the western group of C. elaphus. These source populations often came from managed herds on game estates and were themselves often of mixed provenance.
In New Zealand, the wild C. elaphus herds dispersed and by the middle of the 20th century had met wild herds of wapiti (C. canadensis = C. elaphus nelsoni) introduced into the south of South Island from Wyoming and sika (C. nippon) introduced into the central North Island from a game estate in England and of mixed heritage. C. elaphus interbreeds with both C. canadensis and C. nippon, producing fertile offspring. C. canadensis and C. nippon are present on farms in Australia but have not established wild populations.
Red deer farming began in New Zealand in the 1980s (Nugent and Fraser, 2005) and active hybridisation with wapiti to increase body size was encouraged (Sika deer are not farmed in New Zealand). The modern ability to import semen or embryos has also facilitated the mixing of these original British red deer populations with larger phenotypes from eastern Europe and with wapiti (Pearse and Goosen, 1999). Escape of these farmed deer of both ‘pure’ and mixed genetic parentage into wild populations will result in mixing of the genomes in these introduced populations (and indeed could do the same in native populations). Despite legal standards for deer fencing in New Zealand, escapes are common (Fraser et al., 2003). This genetic mixing is also happening in South America -- a recent photograph of a wild deer in Argentina with clear C. canadensis characters despite no legal importation of wapiti suggests that Argentinian deer farmers have been importing genetic material from New Zealand (where the export of such material is not illegal).
When liberated in the southern hemisphere, seasonally breeding species such as C. elaphus adjusted to the change in seasonal displacement within two years (Logan and Harris, 1967). In the native range calving peaks in May-June (IUCN, 2017); in New Zealand and South Africa, C. elaphus have a median birth date of 9 December (Caughley, 1971) when a single fawn is produced. The amount and quality of food per capita determines maternal body condition which determines age at first breeding. When animals are in good condition a proportion of females may first breed as yearlings. Males attain sexual maturity as yearlings, but they continue growing until at least 6 years of age and cannot compete for females with other males until then (IUCN, 2017).
The natural lifespan of C. elaphus is about 15 years, but an age of nearly 27 years has been recorded in captivity (IUCN, 2017).
Population Size and Density
In Europe excluding Russia, the species numbered 1.25 million individuals in 1985 and 2.4 million in 2005. Densities are typically 1-5 individuals per km², sometimes up to 15 individuals per km² (IUCN, 2017).
In New Zealand, the density of C. elaphus in native forests was estimated at up to 30 animals/km2 or more when the animals had access to grassland areas. Commercial hunting of wild deer using helicopters began in the late 1960s and reduced the population by about 90% to about 200 000 in the 1980s, mostly living if forested habitats where they were less vulnerable to aerial hunters (Nugent and Fraser, 2005). The intensity of this commercial exploitation has declined in the 21st century, so populations are generally increasing (Forsyth et al., 2013).
There are no estimates of the population size or density of C. elaphus in Australia, There are no estimates of the population size or densities of C. elaphus in Argentina or Chile.
C. elaphus are highly adaptable and both graze grasses and browse herbs and woody plants. For example, in New Zealand the diet of deer living around forest-grassland margins consists mostly of grasses and herbs (70 – 80% of the diet), while the diet of deer living in forested habitats is mostly shrubs and trees (70 – 80% of the diet) (Nugent and Fraser, 2005).
ClimateTop of page
|Af - Tropical rainforest climate||Tolerated||> 60mm precipitation per month|
|B - Dry (arid and semi-arid)||Tolerated||< 860mm precipitation annually|
|C - Temperate/Mesothermal climate||Preferred||Average temp. of coldest month > 0°C and < 18°C, mean warmest month > 10°C|
|D - Continental/Microthermal climate||Preferred||Continental/Microthermal climate (Average temp. of coldest month < 0°C, mean warmest month > 10°C)|
Notes on Natural EnemiesTop of page
C. elaphus in their native range are predated by large carnivores (where these are still present) such as wolves (Canis lupus), bears (Ursidae) and lynx (Lynx spp.) (e.g. Jedrzejewska et al. 1997).
In their introduced range, C. elaphus may be preyed upon by dingoes (Canis familiaris dingo), feral dogs, wedge-tailed eagles (Aquila audax) and possibly by red foxes (Vulpes vulpes) in Australia, and probably by pumas (Puma concolor) and condors (Vultur gryphus) in South America – although hard evidence is absent. In New Zealand they have no non-human predators.
C. elaphus may be infected with bovine tuberculosis but most infected animals survive for many years (Nugent et al., 2015). At natural densities they are spillover hosts of tuberculosis and only maintain the disease when held at artificially high densities, e.g. on farms. They are susceptible to several mycobacterial diseases apart from bovine tuberculosis; paratuberculosis (Johne’s disease) and avian tuberculosis have been found in both farmed and wild deer (Mackintosh et al., 2004).
Deer are also capable of being infected with foot and mouth virus, but as with tuberculosis they appear not to maintain the disease in the absence of infected livestock (Davies, 2002).
In general wild C. elaphus are healthy. In New Zealand, of 80 500 deer harvested from the wild between 1994 and 1997 and inspected at processing works, only 2.7% had tissue worm infections, 0.5% had pleurisy, 0.4% had arthritis and 0.3% had bovine tuberculosis (Nugent and Fraser, 2005).
Means of Movement and DispersalTop of page
Introductions of C. elaphus to new countries have all been intentional, typically for hunting or farming, although their escape into the wild is often accidental. Natural dispersal rates of 1.6 km per year were estimated for C. elaphus introduced to New Zealand (Caughley, 1963), but most dispersal in the main countries where the species has been introduced is confounded by secondary releases by people and in later years by escaping farmed deer.
Pathway CausesTop of page
|Acclimatization societies||New Zealand, Australia and Chile||Yes||Yes||Long, 2003|
|Animal production||Farmed deer in many countries are a risk when they escape||Yes|
|Breeding and propagation||Embryo and semen import/export||Yes|
|Escape from confinement or garden escape||Farmed deer are a risk||Yes|
|Hunting, angling, sport or racing||Yes||Yes|
Impact SummaryTop of page
Economic ImpactTop of page
C. elaphus may be ‘overabundant’ in parts of their native range (Nugent et al., 2011). In such case they may cause damage to crops (e.g. Bleier et al., 2012) or present disease risks to livestock (Gortázar et al., 2006).
Introduced C. elaphus (and other cervids) have been blamed for competition with domestic stock and damage to crops in Australia (Davis et al., 2016). However, as in other countries the actual costs of such impacts have not been assessed, perhaps because they are relatively small.
It is possible that the route of infection of bovine tuberculosis from cattle to the wild maintenance host in New Zealand (the brushtail possum, Trichosurus vulpecula) was via carcasses of infected deer, which were prevalent in many areas during the commercial harvesting boom of wild deer during the 1970s and 1980s, and scavenged by possums (Nugent et al., 2015).
Environmental ImpactTop of page
In the native range of C. elaphus, browsing by red deer, especially when at high densities, can alter the structure and succession of plant communities, with consequences for other animals (e.g. Gill and Fuller, 2007).
The major adverse impact of introduced red deer is on native vegetation; in turn, changes in the structure and composition of vegetation affect native animals.
Most of the research on impacts on vegetation has been conducted in New Zealand where the native vegetation evolved in the absence of mammalian herbivores. C. elaphus living in alpine grasslands have a serious impact on the plant communities, particularly on the long-lived snow tussocks (Chionochloa spp.) and herbs. However, the vulnerability of deer to helicopter hunting in these grassland areas enabled their virtual elimination deer by the 1980s, allowing significant recovery of the vegetation (Rose and Platt, 1987). Deer grazing in these alpine grasslands is thought to be a significant threat via competition for food to the endangered, flightless gallinule, the takahe (Porphyrio hochstetteri) (Mills and Mark, 1977)
The impact of deer in forest ecosystems in New Zealand is complex and may have long-term consequences which may not always be reversible even if deer numbers are reduced (Coomes et al., 2003). For example, Nugent et al. (2001) have shown that the most palatable plant species growing in the browse tier of forests are quickly removed by C. elaphus in sites where they are accessible. The deer then rely to a large extent on leaves of palatable species that fall from the canopy. This buffers the expected feedback of declining food availability in the browse tier on deer density and so inhibits the recovery of these understorey palatable plants. Of course the new balance between deer and their food can last only as long as the canopy trees live – and in New Zealand this is unnaturally shortened for palatable species because of the impacts of another introduced herbivore, the arboreal brushtail possum (Trichosurus vulpecula). A recent study modelled the long-term consequences of deer (and introduced mice as granivores) in New Zealand podocarp forests, suggesting that the changes in structure and composition will take centuries to resolve (Forsyth et al., 2015).
In Australia, C. elaphus live mostly in eucalypt-dominated forests and woodlands. Determining the impact of deer on these communities is difficult as they are sympatric with native macropod herbivores, and fire and rainfall impose significant extrinsic influences on such plant communities. The only study of C. elaphus impacts in Australian (in the Grampian Mountains population in Victoria) showed that Acacia spp. and grasses formed he bulk of the diet in winter, although the consequences of this for the plant communities are unknown (Roberts et al., 2015).
South America: Veblen et al. (1989) compared the understorey composition in two forests in Argentina dominated by Nothofagus dombeyi, one with abundant C. elaphus and one with no deer. They showed that deer had nearly eliminated one common subcanopy tree (Aristotelia chilensis), had reduced the abundance of many other woody and herbaceous species, and impeded the regeneration of the dominant canopy trees after canopy gaps are created. It has been suggested that C. elaphus compete with native herbivores, such as huemul deer (Hippocamelus bisulcus), and that their presence as prey for pumas (Puma concolor) increases predation on native prey (e.g. Dolman and Wäber, 2008) but hard evidence is lacking (Flueck, 2010). C. elaphus also facilitate plant invasions in temperate forest in Patagonia, Argentina (Relva et al., 2010; Nuñez et al., 2013).
Threatened SpeciesTop of page
Social ImpactTop of page
Unlike some cervids (e.g. rusa deer (Cervus/Rusa spp.) in Australia, or white-tailed deer (Odocoileus virginianus) in the USA), C. elaphus in their introduced ranges rarely live in suburban areas and cause problems.
Risk and Impact FactorsTop of page Invasiveness
- Proved invasive outside its native range
- Has a broad native range
- Abundant in its native range
- Highly adaptable to different environments
- Is a habitat generalist
- Capable of securing and ingesting a wide range of food
- Highly mobile locally
- Ecosystem change/ habitat alteration
- Modification of successional patterns
- Negatively impacts agriculture
- Negatively impacts forestry
- Negatively impacts animal health
- Reduced native biodiversity
- Threat to/ loss of endangered species
- Threat to/ loss of native species
- Competition - monopolizing resources
- Pest and disease transmission
- Interaction with other invasive species
- Highly likely to be transported internationally deliberately
- Difficult/costly to control
UsesTop of page
C. elaphus have long been managed within their native range as a hunting resource, and this was the main motivation for their release in the southern hemisphere (e.g. Donne, 1924), where such hunting is the main contributor to the annual harvests of deer.
Introduced red deer are exploited for productive purposes in three ways. Commercial hunting of wild red deer for venison began in New Zealand in the 1970s and was sufficiently extensive and intensive to reduce the national population by about 90% during the 1970s and 1980s. The harvest size is determined largely by the price of venison in the European market and the costs of harvesting. The annual harvest is currently about 17 000 (Warburton et al., in press). Red deer are also held in semi-natural states on large private ranches where trophy hunters pay for access to large-antlered stags. Genetic manipulation of some of these herds to increase body and antler size is common in New Zealand, Argentina and Mexico, including the import of stags from larger subspecies or wapiti, or imported semen or embryos of such animals (Pearse and Goosen, 1999).
The third productive use of red deer is as farmed animals for venison or velvet production (Gilbey and Perezgonzalez, 2012). Farming was developed in New Zealand in 1969, based initially on the capture of wild deer, and at its height in 2004/2005 about 1.8 million red deer (and hybrids with wapiti) were farmed. Currently about 900 000 deer are held with an annual export of about 14 500 tonnes of venison (Stats NZ, 2016). Red deer are also farmed in other countries outside their native range (Australia, Canada and the USA), as well as inside it.
Uses ListTop of page
- Game animal
- Sport (hunting, shooting, fishing, racing)
Human food and beverage
- Fresh meat
- Meat/fat/offal/blood/bone (whole, cut, fresh, frozen, canned, cured, processed or smoked)
Similarities to Other Species/ConditionsTop of page
The genus Cervus forms a ring of sub-species and species around the Northern Hemisphere. Deer at the western end of this ring (red deer, C. elaphus) have a wide phenotypic variation in body size and antler conformation, but are generally distinguished from the eastern and North American species (wapiti, C. canadensis) and southern species (sika, C. nippon) of the ring by differences in pelage and antler form. Generally, there is an increase in body size and lighter pelage across the genus from west to east. Wapiti are typically the largest species in the genus. Both sexes have a lighter coloured body and larger cream-coloured rump patch (but with dark head, neck and legs) than red and sika deer. Wapiti calves are brown with a dark dorsal strip and may have a pattern of indistinct light spots for the first few months – red deer fawns also have white spots with a distinct double row along their back that are more distinct than those in wapiti. There are differences in antler conformation between the three species, with wapiti sur-royals (the top tines in adults) sweeping backward, while those in red deer stay in the same plane as the lower tines, as do those for sika deer. Sika deer (Cervus nippon) retain the juveniles’ distinct white spots as adults, especially in their summer pelage. See King (2005) for more detailed descriptions of the three species.
Prevention and ControlTop of page
The biosecurity risk chain consists of action to prevent introduction of an exotic species into a new country or area, early detection and rapid response when ‘borders’ are breached, eradication where feasible, sustained control or acceptance of the status quo with no management possible. Introduced C. elaphus are managed under all of these strategies in various places. Their management in New Zealand, Australia, Argentina and Chile is confounded by their mixed status as farmed assets, game animals and biodiversity pests (e.g. Nugent and Fraser, 1993).
It is possible that introduced C. elaphus already present in one country can disperse naturally into neighbouring countries across land borders in South America and South Africa, but the main potential route of establishment is via human liberations. Most countries either prohibit all legal introductions of exotic species unless they pass a risk analysis, or maintain a black list of banned species. C. elaphus are banned in some State jurisdictions in Federal countries, e.g. in Western Australia or in the western provinces of Canada.
Established populations of C. elaphus have been or are being eradicated (permanently removed) or extirpated (removed but at risk from reinvasion) on a few islands (Parkes and Forsyth, 2011; Island Conservation, 2016). They can swim, so the current ongoing attempt to remove them from Resolution Island (20 860 ha) in Fiordland National Park in New Zealand, and the completed removal programme on Secretary Island (8140 ha) (Crouchley et al., 2011) have to limit or manage the expected reinvasion from the mainland herds. See also the case study below.
Preventing spread: Current distributions of free-living C. elaphus in New Zealand, Australia and Argentina/Chile have not reached their potential. Preventing further spread is not a trivial option as effective management depends on the way the deer are likely to spread – by natural dispersal, by deliberate introduction, or by escape from farms. In most cases all three mechanisms are involved.
Natural dispersal is the most difficult mechanism to manage, as seen in the spread of the species across the border between Argentina and Chile (Nugent et al., 2011). Although illegal, deliberate liberations appear to be a major cause of spread in Australia (Moriarty, 2004) and to date has not been effectively managed. Escape from deer farms is inevitable, so management either requires a zonation system where deer farming can be prohibited or an early detection-rapid response approach.
A study of deer farmed in areas of New Zealand where no wild deer were present showed that over seven years 27 escape events were reported for 58 farms. The mean number of animals escaping per escape was 13 (range 1 – 270); in 15% of events the animals were not recaptured and had to be shot to ensure wild populations did not establish. Inadequate fences, human error and storm damage to fences each accounted for a third of escape events, suggesting a reactive approach to inevitable events is required. The project has a telephone hotline for deer farmers to use to call for assistance when they have an escape. Being more proactive e.g. by enforcing fencing standard, only accounts for a third of the risk (Fraser et al., 2003). See also the case study below.
The principles of sustained control are that the target population must be reduced enough to mitigate the impacts on some asset they adversely affect (Choquenot and Parkes, 2001). Most introduced deer populations are harvested by recreational hunters who see the animals as a resource and asset. A question faced by land managers who see the deer as pests is whether this harvest is sufficient by itself or in combination with other tools to act as a control on the deer numbers.
C. elaphus in New Zealand are legally classed as ‘wild animals’ and are the property of the Crown unless legally taken. They may be legally hunted with no restrictions on the numbers, age or sex by anyone with a firearms licence who has the permission of the land owner. The New Zealand Department of Conservation manages about 33% of New Zealand and most of the range of red deer. It generally views the animals as pests and where it does not itself control deer it encourages recreational and commercial hunters to kill deer.
Nugent and Choquenot (2004) explored the relative cost-effectiveness of commercial harvesting, recreational hunting and state-funded culling as red deer control tools in New Zealand, where recreational hunting of deer for trophies or for meat is a popular activity with about 31 000 hunters taking about 41 700 C. elaphus per year – at least the last time these data were estimated in 1988 (Nugent 1992). They concluded that payment of incentives to commercial aerial hunters would usually be more cost-effective than state-funded culling, at least in non-forested habitats, but payment to recreational hunters would not be effective. In general, even when there are no restrictions on how many deer can be taken by recreational hunters (as in New Zealand) it appears that the harvest is not large enough to reduce deer densities by enough to protect more sensitive biodiversity assets (Nugent and Fraser, 2005). In one area of New Zealand occupied by wapiti, C. canadensis, a hunting organisation conducts extensive culls of red deer and hybrids to maintain the wapiti phenotype (Fiordland Wapiti Foundation, 2016).
Commercial harvesting of C. elaphus, for venison can reduce populations by a large amount under certain conditions (Parkes et al., 1996). The important factors are competitive access, the presence or absence of refugia for the deer (e.g. deer in grasslands are much more vulnerable to helicopter shooting than those in forests), and the cost of harvesting across deer densities versus the price of venison (Warburton et al., in press). Commercial exploitation of C. elaphus in New Zealand began in 1930 when official government culling also took and exported deer skins, with up to 103 000 being taken in one year. This ceased when a market for venison was developed in the late 1950s. In the late 1960s this harvest was largely taken by hunters shooting from helicopters. About 133 000 red deer were harvested in 1972 at the peak of the industry, which reduced populations in alpine grasslands by up to 90%, but less in forests. Currently the economics of the industry means about 17 000 deer are shot each year (Warburton et al., in press).
The final strategy to sustain control of deer is via official government culling, again only used to any extent in New Zealand. Large-scale culling of deer was conducted by government agencies first across all herds, and then in priority water catchments (Caughley 1977). Since the advent of the commercial harvest, government culling is limited to deer-free areas to manage escapees (Fraser et al., 2003), a few areas with high-priority biodiversity assets affected by deer (e.g. to protect takahe (Porphyrio hochstetteri) habitat, culling of deer has been conducted since the rediscovery of the endangered birds in 1948 -- Fraser and Nugent, 2003), and attempts to eradicate deer from some islands in Fiordland National Park (Crouchley et al., 2011; MacDonald et al., in press). These projects account for a few hundred deer each year.
In the mid-1980s the growing farmed deer industry in New Zealand led to large numbers of wild C. elaphus being captured and transferred to farms (Parkes et al., 1996).
See also the case study below.
In Australia, the status of C. elaphus depends on the State in which they live. They are classed as pests in South Australia, the Australian Capital Territory and Queensland (as well as the States in which they do not occur), but as game that can be managed as pests if justified in New South Wales, and simultaneously as pests, protected wildlife and game in Victoria (Forsyth et al., 2016). There is some limited government control of red deer in National Parks in some states (Natural Resource Council, 2016). Most States allow recreational hunting under permit (i.e. a hunter must have both a firearm licence and a hunting permit) but this is usually limited in National Parks where firearms are restricted or prohibited.
South America: In Argentina, the policy to manage C. elaphus in national parks is to stop further spread and contain numbers at or below current densities by allowing recreational hunting. This appears to have failed, as the range of deer expanded by about a third between 1985 and 2005. The hunters focus on adult stags as trophies so are unlikely to be an effective control tool by themselves (Nugent et al., 2011). In Chile, recreational hunting is not allowed on public lands and that on private land is mostly paid trophy hunting on ranches (Nugent et al., 2011).
Further Information on Prevention and Control
This section consists of three case studies from New Zealand. A more detailed account of deer control in two regions of New Zealand is provided by Fraser et al. (2003).
Case study 1: stopping spread
The development of farming or ranching of C. elaphus behind fences began in New Zealand in 1969 and has spread to many other countries. Where these farms are developed in areas without wild populations there is an obvious risk that new populations can establish if animals escape.
In some areas of New Zealand there are no wild C. elaphus but there are red deer farms. A government-funded programme aims to detect and manage any deer that escape from these farms; the results were assessed between 1993 and 1999 (Fraser et al., 2003; Parkes, 2009).
Table 1. Results of a survey of deer farm escapes from Northland and Taranaki regions, 1993 – 1999 (after Parkes, 2009).
Number of deer farms
Number of deer held
Number of escape events reported
% of farms reporting at least 1, 2 or 3 events
25.9, 15.2 and 5.2%
Mean number of deer escaping per event
13 (range 1 – 270)
% times all deer were recaptured
Causes of escapes of 33 events where this was known
2. Human error, e.g. escaped while handling
3. Acts of God, e.g. storm damage to fences
4. Inadequate fences, e.g. jumped intact fence
This survey allowed managers to better allocate their budget between proactive approaches such as inspection and enforcement of fencing standards or reactive approaches such as surveillance, early detection via a telephone hotline for farmers to report escapes, and rapid response to recapture or shoot the escapees. The results suggest managers should allocate 67% of their effort to reacting to events and 33% to enforcing fencing standards, to match the actual risks.
Case study 2: eradication
C. elaphus have swum from mainland South island in New Zealand to many of the 90 islands of Fiordland National Park. The New Zealand Department of Conservation has been removing these insular populations (Crouchley et al., 2011; Nugent and Arienti-Latham, 2012).
Secretary Island (8140 ha) is a rugged island (reaching 1196 m above sea level and mostly forested), separated from mainland deer populations by Thompson Sound with a minimum gap of 950 m and by Doubtful Sound with a minimum gap of 600 m. C. elaphus swam to the island some time before 1970 (Mark and Baylis, 1975), and a campaign to remove them was started in the mid-1970s using ground and aerial shooting and natural bait poisoning (e.g. Parkes, 1983). This did not succeed, but a more determined project was started in late 2006.
Both shooting from helicopters and hunting from the ground were used (Table 2). Novel tools used during the project were modelling to estimate the minimum number of deer known to be alive and the identification and location of individual deer by DNA analysis of faecal pellets that were then matched (or not) with deer shot (Nugent and Arienti-Latham, 2012; Macdonald et al., in press). The key to success in this project was to change the way that ground-based and aerial hunting were used as deer numbers declined. This entailed a detailed survey of the whole island to identify the individual surviving deer (from faecal DNA) and a change in behaviour of the hunters from individual hunters hunting ‘deer’ to team hunting (with GPS and radio contacts) hunting ‘that deer’ – and integration of the ground and aerial effort (Macdonald et al., in press).
Table 2. Number of deer killed by aerial and ground hunters on Secretary Island from November 2006 to September 2014 (after MacDonald et al., in press). 1 Estimated by modelling (Nugent and Arienti-Latham, 2012). 2 Last known female (pregnant) was shot. 3 Estimated from individual faecal pellets DNA. 4 Camera traps identified 4 deer in 2014: one female which was shot and 3 males of which one was shot.
Estimated no. deer remaining
No. deer shot
No deer shot
The DNA samples from deer on the island had low genetic diversity with none of the rarer alleles found on the mainland source population (Crouchley et al., 2011) suggesting that a single invasion event had occurred in the 1950s and reinvasion risks after eradication might be rare and manageable.
Case study 3: sustained control
The flightless New Zealand gallinule, the takahe (Porphyrio hochstetteri), was thought to be extinct until a small population was rediscovered in Fiordland National Park in 1948 (Orbell, 1949). C. elaphus began to colonise this area in about 1930 and because of their dietary overlap with the takahe were thought to be a serious threat to the birds (Mills and Mark, 1977).
Control of C. elaphus, and of a later incursion by wapiti (C. canadensis), began in 1948 and continues to the present. Over 24 000 deer have been removed by ground hunting and shooting from helicopters, which reduced the population from an estimated 8.7 deer/km2 to <0.8/km2 in 2008 (Tanentzap et al., 2009). This virtually eliminated deer use of the alpine grasslands, the main habitat of the takahe, and allowed some improvement of the palatable plant species within the forest habitats (Tanentzap et al., 2009).
One novel tool used in this project relied on the fact that yearling deer are much more vulnerable to hunters when they start to become independent of their mothers. Older deer that have survived many attempts to kill them are very wary and difficult to find. Managers can take DNA from any young deer shot and identify their mother if she is subsequently shot – or that she is still alive in the area (Fraser and Nugent, 2003).
Gaps in Knowledge/Research NeedsTop of page
Where C. elaphus are thought to have adverse impacts on native biodiversity, any control, especially if it requires taxpayer funds, must be justified and the extent of the control (residual or target densities and so annual harvests) must be known or measured. Some of these deer density-impact relationships are known (e.g. for some habitats in New Zealand -- Nugent et al., 2001) but the relationships are not simple (Choquenot and Parkes, 2001) and are largely not known for C. elaphus in Australia or South America.
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OrganizationsTop of page
New Zealand: Landcare Research, PO Box 69040, Lincoln, www.landcareresearch.co.nz
Principal SourceTop of page
Draft datasheet under review.
ContributorsTop of page
14/12/16: Reviewed for Invasive Species Compendium by:
John Parkes, Landcare Research/Kurahaupo Consulting, New Zealand.
Reviewed by: Dr. Werner T. Flueck, Consejo Nacional de Investigaciones Cientificas y Tecnologica and Centro de Ecologia Aplicada del Neuquen, Argentina.
Principal sources: Dr. Werner T. Flueck, Consejo Nacional de Investigaciones Cientificas y Tecnologica and Centro de Ecologia Aplicada del Neuquen, Argentina
Wednesday, May 26, 2010
Distribution MapsTop of page
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