Invasive Species Compendium

Detailed coverage of invasive species threatening livelihoods and the environment worldwide


Equus caballus [ISC]



Equus caballus [ISC] (horse)


  • Last modified
  • 27 September 2018
  • Datasheet Type(s)
  • Invasive Species
  • Preferred Scientific Name
  • Equus caballus [ISC]
  • Preferred Common Name
  • horse
  • Taxonomic Tree
  • Domain: Eukaryota
  •   Kingdom: Metazoa
  •     Phylum: Chordata
  •       Subphylum: Vertebrata
  •         Class: Mammalia
  • Summary of Invasiveness
  • E. caballus has been domesticated for approximately six millennia, and exists in domesticated contexts worldwide (Outram ...

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Horse (Equus caballus) feral type; Adults, small herd running, USA.
TitleSmall herd of adults running
CaptionHorse (Equus caballus) feral type; Adults, small herd running, USA.
Copyright©Terry Spivey, USDA Forest Service,
Horse (Equus caballus) feral type; Adults, small herd running, USA.
Small herd of adults runningHorse (Equus caballus) feral type; Adults, small herd running, USA.©Terry Spivey, USDA Forest Service,
Horse (Equus caballus) feral type; close-up of head, USA.
TitleClose-up of head
CaptionHorse (Equus caballus) feral type; close-up of head, USA.
Copyright©Terry Spivey, USDA Forest Service,
Horse (Equus caballus) feral type; close-up of head, USA.
Close-up of headHorse (Equus caballus) feral type; close-up of head, USA.©Terry Spivey, USDA Forest Service,


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Preferred Scientific Name

  • Equus caballus [ISC] Linnaeus 1758

Preferred Common Name

  • horse

Other Scientific Names

  • Equus ferus Gentry et al 1996
  • Equus ferus caballus

International Common Names

  • Spanish: caballo

Local Common Names

  • Australia: brumby
  • USA: mustang; wild horse

Summary of Invasiveness

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E. caballus has been domesticated for approximately six millennia, and exists in domesticated contexts worldwide (Outram et al., 2009). Free-ranging (feral) populations exist in Africa, Asia, Australia, Europe, North America, South America, and on the Hawai’ian, Galápagos, and probably other oceanic islands (Grubb, 2005).  Additionally, “semi-feral” horses that are all privately owned yet live for significant time periods on undeveloped lands (which are often public lands) exist in many parts of the world (e.g. in the UK: Dartmoor Commoners’ Council, 2012).  Because of the species’ long commensal association with human civilization, it is regarded alternatively as a companion, a romantic symbol of unbridled power and wildness, a work aide, a competitor with livestock, a problem for wildlife, or a non-native alien (Ryden 1978).  Although free-ranging horse populations can increase dramatically in many of the countries and regions indicated, indications of horses threatening ecosystems, habitats, and species derive from grey literature, unpublished data and observations from land managers and conservation organizations, rather than from peer-reviewed literature.

Taxonomic Tree

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  • Domain: Eukaryota
  •     Kingdom: Metazoa
  •         Phylum: Chordata
  •             Subphylum: Vertebrata
  •                 Class: Mammalia
  •                     Order: Perissodactyla
  •                         Family: Equidae
  •                             Genus: Equus
  •                                 Species: Equus caballus [ISC]

Notes on Taxonomy and Nomenclature

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First described in 1758 by Linnaeus through the domestic horse, Equus caballus (also known as E. ferus) is one of 6-8 extant species of the genus (Bennett and Hoffmann, 1999). Within the family Equidae, 42 of 43 genera are extinct taxa known only from fossils; Equus is the only exception that has species now living (Palmer, 1999; MacFadden, 2005).  E. caballus contains 7 subspecies (some extinct), according to characteristics of early breeds of domesticated horses, historical descriptions of wild horses in Eurasia, and the morphology and distribution of fossils from the late Pleistocene to early Holocene (Trümler, 1961; Bennett, 1992).  Genetic analyses based on either chromosomal differences (Benirschke et al., 1965) or mitochondrial genes both suggest significant divergence among wild forms of E. caballus about 200,000-300,000 years ago, well before domestication began (George and Ryder, 1986). It has thus been suggested that multiple different types were domesticated independently, such as the warmblood (E. c. mosbachensis) in central Europe and the tarpan (E. c.  ferus) in eastern Europe (Azzaroli, 1990).  All four subspecies recognized by Bennett and Hoffmann (1999) that arose in Europe (E. c. mosbachensis, E. c. ferus, E. c. caballus) and North Africa to the Middle East (E. c. pumpelli) are currently extinct in the wild, and exist only as domesticated animals (Bennett and Hoffmann, 1999).  Of the other three subspecies, E. c. alaskae (from Beringia), E. c. mexicanus (from North America), and E. c. przewalskii (from eastern Asia), the latter is the only surviving wild subspecies (Strelkov, 1977; Bennett and Hoffmann, 1999).  All North American horses had become extinct by 11,400 years ago (FAUNMAP Working Group, 1994; Bennett and Hoffmann, 1999).  The different domestic breeds of horses each originally derived from different wild populations that were distributed from Europe to the Middle East (Bennett and Hoffmann, 1999).  Groves (1986) and Corbet (1978) proposed that E. ferus replace E. caballus for the extinct tarpan, given that Linnaeus’ type was a domestic horse.  However, Azzaroli (1984), Forsten (1988), and Bennett and Hoffmann (1999), among others, used caballus.  The International Commission on Zoological Nomenclature ruled in 2003 that the caballus name for the wild species is not invalid simply because it was first used on the domestic form (Grubb, 2005).  Furthermore, physical evidence that ferus is distinct from wild horses consists only of two osteological specimens, and ferus has not been reliably identified with Pleistocene or Holocene local populations (Forsten, 1988).  The status of ferus as a wild rather than feral form is disputed (Epstein (1971) cited in Wilson and Reeder (2005), among others). 

The truly wild (Przewalski’s) horse, which previously ranged from Germany and the Russian steppes eastward to Kazakhstan, Mongolia, and northern China, now exists only in Mongolia and is considered Endangered on the IUCN Red List (IUCN, 2012).  The latest estimate of its global population size is 306 free-ranging reintroduced and native-born individuals, which are all descended from only 13 or 14 individuals (Bowling and Ryder, 1987; Zimmerman, 2011).  Analyses of mitochondrial DNA suggest that E. c. przewalskii is not the ancestor of modern domestic horses (Vilá et al., 2001), and the former’s possession of 2n=66 chromosomes means that it is more different from domestic breeds (all 2n=64) than any pair of domestic breeds are from each other (Ryder, 1994).  However, Przewalski’s horses can hybridize with domestic horses to produce fertile offspring (Ryder et al., 1978; Trommerhausen-Smith et al., 1979), and Przewalski’s horses only possess four alleles at four different serological marker loci that are specific to E. c. przewalskii (Bowling and Ryder, 1987).

There are local vernacular names, such as ‘mustang’ in the USA and ‘brumby’ in Australia, that are used specifically for free-ranging and feral animals.

In addition to E. caballus, the genus also comprises E. africanus (formerly known as E. asinus, the African wild ass; the domestic donkey is a subspecies of this species), E. hemionus (onager), E. kiang (kiang), E. zebra (mountain zebra), E. quagga (formerly known as E. burchelli, the common or Burchell’s zebra), and E. grevyi (Grevy’s zebra), as well as 15 other extinct species (Grubb, 2005; Hooker, 2005).  The Equidae are joined by two other living families in the order Perissodactyla (odd-toed ungulates): Tapiridae (tapirs, including 4 species of tapirs in the genus Tapirus), and Rhinocerotidae (including 5 extant species of rhinoceroses across 4 genera), as well as by 12 other now-extinct families (McKenna and Bell, 1997; Hooker, 2005).


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The general form of E. caballus is readily recognizable to most humans. Among the species in the genus Equus, on average E. caballus has the heaviest body build, the heaviest limbs, and the widest and deepest head (Bennett and Hoffmann, 1999).  Additionally, horses typically possess rounder (i.e., less oval-shaped) hooves, shorter <20 cm) ears, and a hairier tail than other Equus species (Willoughby, 1974).  In contrast to external measurements (in cm) of Przewalski’s horses (body length 220-280, length of tail including hair 99-111, tail length without hair 38-60, ear length 14-18, height at withers 120-146), measurements of domesticated and free-ranging horses are much more variable and include greater maximum sizes for all measures.  Humans have consistently artificially selected for taller individuals, shorter heads, broader foreheads, finer muzzles, and higher withers (Bennett and Hoffmann, 1999).


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E. caballus exists in domesticated contexts worldwide (Outram et al., 2009), and in many countries there are ‘semi-feral’ populations that are free-ranging but privately owned. The Distribution table lists only fully feral populations.

Within the United States, free-ranging horse populations exist in portions of ten western states, and semi-free-ranging ones in a further two, although in in New Mexico, Arizona, and Montana there are less than about 500 horses per state in only one or two Herd Management Areas (HMAs) (Bureau of Land Management, 2012).  Across the western USA, HMAs are areas that are actively managed for horses or burros (feral donkeys).  Areas historically occupied by free-roaming horses and burros, during at least some period during 1971-2007, total 366,690 km2 across the western USA (Beever and Aldridge, 2011).  As of 2011, areas occupied by free-roaming horses and burros totalled 150,408 km2 (Beever and Aldridge, 2011).  Small populations are also present in several eastern states, but these tend to be more intensively monitored and managed. Free-ranging horses also exist in four Canadian provinces (Alberta, British Columbia, on Sable Island in Nova Scotia, and in the Bronson Forest of Saskatchewan), and in numerous states of México (Findlay and Halley, 2005; Plante et al., 2007; WHOAS, 2012; Saskatchewan Party Caucus, 2009).

The U.S. National Park Service, which has a dual mandate of 1) natural- and cultural-resource conservation “unimpaired for the enjoyment of future generations” and 2) visitor experience and recreation, manages free-roaming horse herds in order to help protect natural resources.  Decisions about how to manage the species are primarily made at the park level; in a number of cases, horses are managed and explicitly considered as part of the cultural landscape.  Of the 23 park units that list horses on their mammal species lists, their occurrence status is “Present in Park” at 17 units, “Historic” at 4 other units (some of which represent recent active-removal management), and “Unconfirmed” or unspecified at the remaining 2 units.  The U.S. Fish and Wildlife Service, which has as its mission statement “… working with others to conserve, protect, and enhance fish, wildlife, plants, and their habitats for the continuing benefit of the American people”, is at various stages in the process of reducing or removing free-ranging horses and burros from several of its units, consistent with its organic legislation and policies.  Free-ranging horses have recently occurred on 6 USFWS refuges, and herd-management areas for free-ranging horses or burros occur within 16 km of an additional 13 refuges.

Primary responsibility for managing free-ranging horses in the USA, however, resides with the Bureau of Land Management, as governed by the 1971 Wild Free-Roaming Horses and Burros Act.  This Act also governs management of the thousands of free-ranging horses and burros that occur on U.S. National Forests, which are managed by the U.S. Forest Service.

Distribution Table

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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/RegionDistributionLast ReportedOriginFirst ReportedInvasiveReferenceNotes


IndonesiaPresentPresent based on regional distribution.
-Nusa TenggaraLocalisedIntroducedGrubb, 2005
IranLocalisedIntroducedGrubb, 2005
MongoliaLocalisedNative Not invasive Bennett and Hoffmann, 1999Reintroduced after extinction in the wild
Sri LankaLocalisedIntroducedGrubb, 2005


NamibiaPresent, few occurrencesIntroduced Not invasive Greyling, 2005

North America

CanadaPresentPresent based on regional distribution.
-AlbertaLocalisedIntroducedWebb et al., 2009
-British ColumbiaLocalisedIntroducedFindlay and Halley, 2005
-Nova ScotiaPresent, few occurrencesIntroducedPlante et al., 2007
-SaskatchewanPresent, few occurrencesIntroducedSaskatchewan Party Caucus, 2009
MexicoPresentIntroduced Invasive E. Beever, U.S. Geological Survey, Bozeman, Montana, USA, personal observation, 2009Probably invasive
USAPresentPresent based on regional distribution.
-AlabamaPresent, few occurrencesIntroducedNational Park Service, NPS
-ArizonaPresent, few occurrencesIntroduced Invasive Bureau of Land Management, 2012
-CaliforniaLocalisedIntroducedBureau of Land Management, 2012
-ColoradoLocalisedIntroduced Invasive Bureau of Land Management, 2012
-GeorgiaPresent, few occurrencesIntroduced Invasive Taggart, 2008
-HawaiiLocalisedIntroduced Invasive Grubb, 2005
-IdahoLocalisedIntroduced Invasive Bureau of Land Management, 2012
-MarylandPresent, few occurrencesIntroduced Invasive Taggart, 2008
-MontanaPresent, few occurrencesIntroducedBureau of Land Management, 2012
-NevadaLocalisedIntroduced Invasive Bureau of Land Management, 2012Invasiveness illustrated in Beever et al. (2003, 2004, 2006, 2008)
-New JerseyPresent, few occurrencesIntroducedNational Park Service, NPS
-New MexicoPresent, few occurrencesIntroducedBureau of Land Management, 2012
-New YorkPresent, few occurrencesIntroducedNational Park Service, NPS
-North CarolinaPresent, few occurrencesIntroduced Invasive Levin et al., 2002; Taggart, 2008
-North DakotaPresent, few occurrencesIntroducedNational Park Service, NPS
-OregonLocalisedIntroduced Invasive Bureau of Land Management, 2012
-South DakotaPresent, few occurrencesIntroducedNational Park Service, NPS
-TennesseePresent, few occurrencesIntroducedNational Park Service, NPS
-TexasEradicatedIntroduced Invasive National Park Service, NPS; Dobie, 1952
-UtahLocalisedIntroduced Invasive Bureau of Land Management, 2012
-VirginiaPresent, few occurrencesIntroduced Invasive Taggart, 2008
-WyomingLocalisedIntroduced Invasive Bureau of Land Management, 2012

South America

ArgentinaLocalisedIntroducedGrubb, 2005
ColombiaLocalisedIntroducedGrubb, 2005
EcuadorLocalisedIntroducedGrubb, 2005
VenezuelaLocalisedIntroducedPacheco and Herrera, 1997


FranceLocalisedIntroducedGrubb, 2005
GreeceLocalisedIntroducedGrubb, 2005
PortugalLocalisedIntroducedGrubb, 2005
SpainLocalisedIntroducedGrubb, 2005


AustraliaPresentGrubb, 2005
-Australian Northern TerritoryWidespreadIntroduced Invasive Nimmo and Miller, 2007
-New South WalesPresentIntroduced Invasive Nimmo and Miller, 2007
-QueenslandWidespreadIntroduced Invasive Nimmo and Miller, 2007
-South AustraliaPresentIntroduced Invasive Nimmo and Miller, 2007
-VictoriaPresentIntroduced Invasive Nimmo and Miller, 2007
-Western AustraliaPresentIntroduced Invasive Nimmo and Miller, 2007
New ZealandLocalisedIntroducedGrubb, 2005

History of Introduction and Spread

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As mentioned in the Identity section, the different domestic breeds of horses each originally derived from different wild populations that were distributed from Europe to the Middle East (Bennett and Hoffmann, 1999).  Genetic evidence suggests that domestication happened independently on numerous occasions.  Given the geographical distance and human-cultural differences separating populations of free-ranging horses across Europe and Asia, it is likely that domesticated horses became feral or semi-feral across those continents independently, rather than in any coordinated or simultaneous fashion.

Horses were introduced to North America via what is now the southwestern U.S. by Spanish conquistadores, who brought the animals with them on sailing ships at the end of the 16th century; some of their steeds and work horses subsequently escaped.  Free-ranging horses spread rapidly across the Intermountain West, facilitated partly by their capture, use, and subsequent release by Native Americans and settlers in the following three centuries.  Dobie (1952) stated that “My own guess is that at no time were there more than a million mustangs in Texas and no more than a million others scattered over the remainder of the West."  Eventually, however, competing land uses and interests led to the capture of horses in increasing numbers; this process was accelerated when the Taylor Grazing Act was passed in 1934.  These land uses and interests included using horses for riding, breeding stock, or both; selling horses for profit; and reducing suspected competition with livestock grazing.  Repeated genetic analyses (e.g.: Bowling, 1994; Goodloe et al., 1991) have shown that most free-ranging herds of horses in the USA analyzed possess a pedigree reflecting mixing of various combinations and proportions of quarter horses, draft horses, and breeds used by Native Americans and early settlers.

It is sometimes argued that E. caballus should be considered as a native species in North America, despite its absence between the Pleistocene and the arrival of European people a few hundred years ago, on the grounds of studies such as that by Weinstock et al. (2005). This study used analyses of a segment of 349-716 base pairs in mitochondrial DNA of fossil equid remains from 53,000 years ago to historical times to produce a phylogenetic tree that the authors interpret as suggesting that domestic E. caballus are conspecific with most or all of the equids in North America from the late Pleistocene. While there remains mixed evidence of whether contemporary E. caballus should be considered native to North America or not, a majority of mammalogists and phylogeneticists in practice consider them non-native, given that the preponderance of existing evidence is consistent with such a designation.

As reviewed by Nimmo and Miller (2007), horses were introduced to Australia in 1778 by the First Fleet and domesticated animals were released by 1804 (Rolls, 1969; Dobbie et al., 1993).  By the 1830s, feral horses had become relatively common across parts of southern Australia, and they appear to have reached every mainland state by the 1840s (Sidney, 1854 cited in McKnight, 1976; Letts, 1962; McKnight, 1976; Appleton, 1983; Dobbie et al., 1993).  Feral horses were considered pests by some Australians as early as the 1860s, and their ecological impacts were reported as early as the 1870s (Rolls, 1969; Low, 1999).  The geographic-range expansion of feral horses in Australia was facilitated in part by the expansion and spread of human pastoralism and settlement (Dobbie et al., 1993).  For example, horses were often left behind when pastoral landholdings were abandoned in instances of drought or of failure due to farming unfamiliar land.  Similarly, other horses were released once their market value (spurred by their need as remounts for the British Army) declined precipitously after World War I (McKnight, 1976; Burke, 1996).  Increasing mechanization of agriculture nationwide has also led to animal labour becoming obsolete in many locations (Burke, 1996; Walter, 2002).

Risk of Introduction

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Risk of future introductions probably depends more centrally on socio-political than on ecological factors.  The species already exists in domesticated contexts worldwide, and has been domesticated for about six millennia.  Whether horses become feral in other locations may depend, for example, on the existence of land on which individuals could range freely, the costs of gathering up individuals that have escaped relative to horses’ value as mounts or as a food item for humans or their pets, and the social perceptions of and attitudes toward horses in a free-ranging condition.  Both historically and in contemporary contexts, there is anecdotal evidence that the number of domestic horses being abandoned into feral-horse populations increases during periods of economic hardship; however, neither the scope nor the magnitude of this phenomenon is known, range-wide.


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Domesticated horses have traversed every major biome, ecoregion, and vegetation alliance on earth, and with few exceptions, free-ranging horses are probably limited far more by socio-political constraints than by natural selection in which habitats that they can and cannot use.  Grassland and steppe habitats are more likely to respond favourably over ecological time scales to moderate to heavy grazing by free-ranging horses, given their longer coevolutionary history with large-bodied grazers than habitats such as warm or cold deserts (Mack and Thompson, 1982; Berger, 1986).  Given free-roaming horses’ plasticity relative to climate, topography, and diverse flora and fauna, their ability to persist relates to basic environmental requirements regarding thermoregulation, predation pressure, water stress, and nutrition (Bennett and Hoffmann, 1999).

Habitat List

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Coastal areas Present, no further details Harmful (pest or invasive)
Coastal areas Present, no further details Natural
Coastal dunes Present, no further details Harmful (pest or invasive)
Coastal dunes Present, no further details Natural
Intertidal zone Present, no further details Harmful (pest or invasive)
Intertidal zone Present, no further details Natural
Salt marshes Present, no further details Harmful (pest or invasive)
Salt marshes Present, no further details Natural
Cultivated / agricultural land Present, no further details Harmful (pest or invasive)
Cultivated / agricultural land Present, no further details Natural
Disturbed areas Present, no further details Natural
Managed forests, plantations and orchards Present, no further details Harmful (pest or invasive)
Managed forests, plantations and orchards Present, no further details Natural
Managed grasslands (grazing systems) Present, no further details Harmful (pest or invasive)
Managed grasslands (grazing systems) Present, no further details Natural
Rail / roadsides Present, no further details Harmful (pest or invasive)
Rail / roadsides Present, no further details Natural
Arid regions Present, no further details Harmful (pest or invasive)
Arid regions Present, no further details Natural
Cold lands / tundra Present, no further details Harmful (pest or invasive)
Cold lands / tundra Present, no further details Natural
Deserts Present, no further details Harmful (pest or invasive)
Deserts Present, no further details Natural
Natural forests Present, no further details Harmful (pest or invasive)
Natural forests Present, no further details Natural
Natural grasslands Present, no further details Harmful (pest or invasive)
Natural grasslands Present, no further details Natural
Riverbanks Present, no further details Harmful (pest or invasive)
Riverbanks Present, no further details Natural
Rocky areas / lava flows Present, no further details Natural
Scrub / shrublands Present, no further details Harmful (pest or invasive)
Scrub / shrublands Present, no further details Natural
Wetlands Present, no further details Harmful (pest or invasive)
Wetlands Present, no further details Natural

Biology and Ecology

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All domestic and free-roaming horses investigated exhibit a chromosome number of 2n=64, including 16 pairs of metacentric to submetacentric autosomes and 15 pairs of acrocentric autosomes (Bennett and Hoffmann, 1999). The fundamental number of autosomes is FNa=92 (Bennett and Hoffmann, 1999).  Przewalski, domestic, and feral horses are completely interfertile (Groves, 1974). George and Ryder (1986) reviewed biochemical evolution in Equus, and Ryder (1994) reviewed genetics of the Przewalski horse. Investigations of the relationship between coat colours and patterns that have been investigated as genetic manifestations have spanned many decades, producing a large and conflicting literature (Sponenberg, 1996).

Genetics of free-ranging and feral horse populations across the western USA vary slightly from herd to herd, but generally reflect a mix of draft-horse, thoroughbred, quarterhorse, saddle, and cavalry animals of numerous breeds, often with little or no evidence of heritage of Spanish barb horses (Bowling and Touchberry, 1988; NRC, 1991). Genetic data suggest that there has been gene flow between herds whose ranges are separated by tens of km, over shorter and longer time scales (Warmuth et al. 2013; E. G. Cothran, College of Veterinary Medicine and Biomedical Sciences, Texas A & M University, College Station, Texas, USA, unpublished data).

Reproductive Biology

Horses are seasonally polyoestrous; the mean duration of dioestrus is 21 days, but ranges from 7-124 days (Frandson, 1974; Ginther, 1979). Puberty in females starts at 10-24 (mean = 18) months of age (Bennett and Hoffmann, 1999). Duration of oestrus in mares averages 6.5 days, but ranges from 4.5-8.9 days (Ginther, 1979). The recorded range of gestation duration in wild, feral, and domestic horses is 287-419 days, averaging 335 days (Willoughby, 1974). Parturition in the northern hemisphere often occurs in spring or early summer, and twinning is rare (Bennett and Hoffmann, 1999). In both wild Przewalski’s and domesticated horses, the oldest recorded age of breeding is 24 years among females and 36 years among males (Groves, 1974; Montfort et al., 1994). Neonates are precocious, can stand within 1 hour of birth, and begin consuming solid food at as early as 1 week of age (Bennett and Hoffmann, 1999). Male horses exhibit a behavior known as flehmen, which researchers continue to investigate in terms of function and context.

Asynchrony between the timing of birth pulses and food availability can profoundly affect demography of large-mammal species (e.g., Post and Forchhammer, 2008). Mechanisms for avoiding such asynchrony can include the increasing temperature and sunlight that accompany the change from winter to spring triggering the pineal gland to initiate reproductive receptivity (Grubaugh et al. 1982, Goldman 2001). Extended to an ideal optimum, one would hypothesize that feral horses in the northern hemisphere (which typically enter oestrus in early spring and return to anoestrus in early winter) should conceive most frequently near the summer solstice, and give birth most frequently 335-342 days later due to the typical period of equine gestation (Ransom, 2012). Ransom (2012) found that survival probability of a neonate born to a dam in mean body condition at spring peak Normalized Difference Vegetation Index (NDVI) averaged 79.9%, and declined 1.4% for every 10 days that parturition occurred after the peak in NDVI. Administration of immunocontraception shifted the peak of parturition 31.5 days later in the year in a herd in southern Montana and a herd in central Colorado, and each year of consecutive treatment increased the period of mare infertility by an average of 411.3 days (Ransom, 2012).


The oldest recorded Przewalski’s horse was 36 years old upon its death, but the oldest domestic horse was 61 years old (Willoughby, 1974). Free-ranging E. caballus have been known to survive to at least 30 years old (e.g., mare 63 from Shackleford Banks, a barrier island off North Carolina, USA).

Activity Patterns

Free-roaming horses exhibit a diverse array of movement patterns, on diel, seasonal, and inter-annual timescales (Pellegrini, 1971; Feist, 1971; National Academy of Sciences, 1982).  Some populations predictably migrate to higher elevations during the hottest, driest months (e.g. at more-southern latitudes), whereas other populations see little change in use patterns of different areas within the year.  Timing of daily behaviours and seasonal life-history events will reflect the combination of selective forces of predation, water stress, food limitation, and thermoregulation that are present at a given time and location (National Academy of Sciences, 1982; Ransom and Cade, 2009). In the western USA horses range further from water than do cattle in the same habitat (Beever, 2003); in Australia they have been recorded as moving more than 20 km in a day (average 15.9), and up to 55 km from watering points over a number of days (Hampson et al., 2010).

Behaviour of feral horses has been described in detail by many sources (e.g. Pellegrini, 1971; Feist, 1971; Feist and McCullough, 1976; McDonnell and Haviland, 1995; Pacheco and Herrera, 1997; Hansen and Mosley, 2000; Linklater et al., 2002; Cameron et al., 2003; Ransom et al., 2010). Feral horses in the western USA are not territorial but typically form long-term social groups (‘bands’) that consist of a dominant polygynous male (stallion) and a harem of females and their offspring, though bands may contain two stallions and accordingly more females (Berger, 1986; Ransom, 2012).  Young and older males occur in populations outside of bands; such males may remain independent or ephemerally form bachelor groups (Berger, 1986; Ransom, 2012).  Pacheco and Herrera (1997) found that in llanos of Venezuela where natural predators were rare, feral horses formed bands of 3-35 individuals, and mean group size varied from 15-21 horses.  Cameron et al. (2003) found that mares in bands of feral horses with >1 stallion (and mares which changed band types, due to their own behaviour or to experimental manipulation) were more protective of their foals against the stallions than were mares in bands with only one stallion.  Hansen and Mosley (2000) found from research near Challis (Idaho, USA) and Lander (Wyoming, USA) no evidence that roundups had deleterious effects on the behaviour or reproduction (percentage of mares with live foals, foaling success rate) of feral horses.  Feist (1971) and Pellegrini (1971) each reported extensively on behaviour of feral horses in the USA, Feist from the Pryor Mountain Wild Horse Range at the border of Montana and Wyoming, and Pellegrini from western Nevada.

Population Size and Density

In contrast to ancestral equids, E. caballus rarely exhibits population regulation by predators (but see Turner et al., 1992 for one exception), and exhibits finite (annual) rates of population change (i.e., lambda) that are far higher than what would be predicted by the allometric, inverse relationship between body mass and lambda.


As caecal digesters, horses are one of the least-selective foragers among ungulates of western North America (Hanley and Hanley, 1982), and thus consume more forage of lower quality. Free-roaming horses consume 70-90% of their food as grasses and graminoids, but this can vary markedly across seasons, for example because herbaceous plants are less accessible during winter at higher latitudes (Hubbard and Hansen, 1976; Hanley and Hanley, 1982; McInnis and Vavra, 1987).


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As - Tropical savanna climate with dry summer Tolerated < 60mm precipitation driest month (in summer) and < (100 - [total annual precipitation{mm}/25])
Aw - Tropical wet and dry savanna climate Tolerated < 60mm precipitation driest month (in winter) and < (100 - [total annual precipitation{mm}/25])
BS - Steppe climate Tolerated > 430mm and < 860mm annual precipitation
BW - Desert climate Tolerated < 430mm annual precipitation
Cf - Warm temperate climate, wet all year Tolerated Warm average temp. > 10°C, Cold average temp. > 0°C, wet all year
Cs - Warm temperate climate with dry summer Tolerated Warm average temp. > 10°C, Cold average temp. > 0°C, dry summers
Cw - Warm temperate climate with dry winter Tolerated Warm temperate climate with dry winter (Warm average temp. > 10°C, Cold average temp. > 0°C, dry winters)
Ds - Continental climate with dry summer Tolerated Continental climate with dry summer (Warm average temp. > 10°C, coldest month < 0°C, dry summers)
Dw - Continental climate with dry winter Tolerated Continental climate with dry winter (Warm average temp. > 10°C, coldest month < 0°C, dry winters)

Latitude/Altitude Ranges

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Latitude North (°N)Latitude South (°S)Altitude Lower (m)Altitude Upper (m)
58 46

Natural enemies

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Natural enemyTypeLife stagesSpecificityReferencesBiological control inBiological control on
Canis lupus Predator All Stages not specific
Felis concolor Predator All Stages not specific

Notes on Natural Enemies

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Other than anthropogenic management or hunting of free-roaming horses, wolves (Canis lupus) and mountain lions (Felis concolor) constitute the non-human predators that most commonly prey upon contemporary horses.  Horses comprise 0-70% of the diet of wolves (Canis lupus) across southern Europe (Meriggi and Lovari, 1996), in northwest Spain (Lagos and Bárcena, 2012), and in Alberta, Canada (Webb et al., 2009).  The most-common result of this predation is reduced foal survival, but recent research tracking movements of collared mountain lions in western Nevada (USA) with GPS collars suggests that some individuals in that region are also preying upon adult free-ranging horses (A. Andreasen, Program in Ecology, Evolution, and Conservation Biology, University of Nevada, Reno, Nevada, USA, personal communication, 2012).  Foals of wild Przewalski’s horses were reportedly killed by wolves in Hustai National Park in Mongolia, and it is suggested that this predation could impair translocation efforts (Duyne et al., 2009).  The lack of evidence for contemporary wolf predation on free-ranging horses in the USA may reflect a very limited overlap of geographic ranges of the two species until recent efforts to foster wolf conservation, horses’ natural tendency to prefer landscapes with wider views for escape by running, or both.

Predation on free-ranging horses in North America has most commonly been attributed to mountain lions (Felis concolor) (Robinette et al., 1959; Ashman et al., 1983; Berger, 1983; Knopff and Boyce, 2009; Knopff et al., 2010), although overall, mountain lion predation on free-ranging horses has long been considered uncommon (Berger, 1986), with few exceptions.  Knopff et al. (2010) found that feral horses that were not free-ranging constituted 10-13% of diets of male F. concolor individuals, but close to 0% of diets of females.  Whereas in all of these investigations predation events by mountain lions were sufficiently rare to not limit population growth of horses, mountain lion predation appeared to regulate population size of free-ranging horses on the south-central California-Nevada border during 1986-1991, a period characterized by low to no human hunting of mountain lions in the region (Turner et al., 1992).   At Pryor Mountain Wild Horse Range (at the border of Montana and Wyoming), lambda (the rate of population increase per individual per unit time) declined precipitously during 1993-2004 until it fell below 1.00 (i.e. the population decreased), but humans killed 3 mountain lions during the winter of 2004 and lambda increased rapidly up to and including 2011 (Ransom, 2012).  Turner et al. (1992) hypothesized that, in addition to the relatively high density of mountain lions, the herd-management area of horses probably conferred relatively high hunting success, given that the area is mostly forested with broken microtopography -- key elements for an ambush predator such as F. concolor (Ashman et al., 1983; Turner et al., 1992).  Lack of population regulation by predation also reflects the fact that survival of foals typically varies more strongly with other variables (e.g., weather severity) than with predation, and is not as limiting to population growth in horses as is adult survival (Eberhardt et al., 1982).

Means of Movement and Dispersal

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Horses have been deliberately moved all round the world by humans. Once horses are present and have become feral, spread of populations in suitable habitat can be rapid (Nimmo and Miller, 2007), as population growth rates can be high in suitable conditions (Eberhardt et al., 1982; Wolfe, 1980; Wolfe, 1986), and individual horses can range over significant distances (Hampson et al., 2010).

Pathway Causes

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CauseNotesLong DistanceLocalReferences
Animal production Yes Yes
Escape from confinement or garden escape Yes
Military movements Yes Yes
Self-propelled Yes

Impact Summary

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Cultural/amenity Positive and negative
Economic/livelihood Positive and negative
Environment (generally) Negative

Economic Impact

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Economic impacts of free-ranging horses have received relatively little quantitative research attention. Bastian et al. (1999), in a case study in a livestock-grazing allotment in Wyoming, USA, sought to estimate opportunity costs associated with foregone domestic livestock and wildlife due to wild horses.  They found that marginal opportunity costs associated with horse population sizes beyond the median Appropriate Management Level target exceeded $1900 per horse.  They acknowledged, however, that it is not possible to say that lower population sizes of wild horses represent a more economically efficient allocation of rangeland natural resources unless the total economic benefits associated with wild horses are estimated (Bastian et al., 1999).  Types of economic impact associated with free-ranging horses, which vary in their extensiveness and their ability to be quantified, include: a) loss of crop value due to horse consumption or trampling; b) reductions in consumptive and non-consumptive uses of wildlife due to direct and indirect competition with native ungulates; c) loss of livestock production due to competition with free-ranging horses; d) costs of federal, state, and provincial programmes to generally maintain and administer horse-management infrastructure; e) costs of programmes in d) associated with gathers, roundups, immunocontraception, and other lethal and non-lethal population-control measures in the field; f) costs associated with feeding, veterinary care, and holding of horses removed from the range; and g) costs associated with rangeland degradation when horse population sizes become very high (e.g., costs for recovery of threatened or endangered species; sensuLosos et al., 1995).

Environmental Impact

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Impact on Habitats


In a study comparing numerous ecosystem components of areas used by horses and areas from which they had been removed, across 19 sites in nine mountain ranges spanning 3.03 million ha of the western Great Basin in Nevada, USA, soils were the ecosystem component that differed most markedly between horse-removed and horse-occupied sites (Beever and Herrick, 2006). Soil hardness (i.e., penetration resistance of soil surfaces) at sites used by horses averaged 2.9 times greater than that at horse-removed high-elevation sites and 17.4 times greater at low-elevation sites (Beever et al., 2003; Beever and Herrick 2006). Strong within-year correlation (r = 0.69, P = 0.001) in this research between average soil hardness and the number of horse defecations at each site suggested the likelihood of a direct effect of horses on penetration resistance (Beever and Herrick, 2006). Compaction was most likely at sites where soil texture was finest and least likely at sites with very sandy soils, and also most likely when soils were moistened (Beever and Herrick, 2006). More generally, however, Butler (1993) suggested anecdotally that among some species of hooved mammals, “chipping” of the soils may aerate the soils and thereby increase local productivity at microsites used by ungulates.  In Australia, areas used by free-ranging horses have been found to exhibit soil compaction, erosion, and soil loss (Dale and Weaver, 1974; Dyring, 1990; Whinam et al., 1994; Nimmo and Miller, 2007).  Dyring (1990) suggested that the mechanisms by which these edaphic alterations occurred probably included reduction of micropore space, water content of soils, and rates of aeration and infiltration.


Vegetation constitutes the ecosystem component that has been most commonly investigated, in terms of responses to (or relationships with) grazing by free-roaming horses.  Given that horses are caecal-digesting grazers, they have relatively higher energetic requirements than ruminants of equivalent body mass (Hanley and Hanley, 1982).  Horses, like all large-bodied herbivores, can affect vegetation directly through selective consumption, and rubbing or trampling, and indirectly through nutrient redistribution, trampling of soils, and creation of wallows and other openings that facilitate patch dynamics (Hobbs, 1996).


Free-roaming horses can affect trees by: 1) inducing mortality of native trees through bark chewing (Schott, 2002; Ashton, 2005); 2) clearing out space under tree canopies through rubbing of body parts against tree limbs and trampling soils under trees (Pellegrini, 1971) while spending time in the shade; and 3) reducing tree recruitment, when equid densities remain high for extended time periods.

Non-native plants

Studies from three continents have found that free-roaming equids can facilitate expansion of non-native plants.  They can disperse weeds directly via one of two mechanisms: a) movement of seeds on the exterior of equid bodies to new locations (epizoochory); and b) seed re-distribution via ingestion and subsequent elimination via faeces (endozoochory; Campbell and Gibson, 2001).  A single feral horse in Guy Fawkes National Park, Australia, excreted up to 19,412 seeds per day, and 6.7% of those seeds were viable (Taylor, 1995).  Feral horses in New Zealand (Rogers, 1991) and across Australia (Weaver and Adams, 1996) were also found to facilitate weed expansion.  In Ernesto Tornquist Provincial Park in the pampas region of Argentina, feral-horse dung covered an average of 2.5% of the study area, and cover of non-native species averaged higher in dung piles than in control plots (Loydi and Zalba, 2009). Dung piles (n = 15) had nine non-native plant species growing on them, compared to only one non-native species on any of the 20 control plots (Loydi and Zalba, 2009).  In contrast, however, as dung piles increased in age (assessed via average dung height), species richness, diversity, and evenness increased (Loydi and Zalba, 2009).  In the western Great Basin of North America, frequency of cheatgrass (Bromus tectorum) averaged 1.61-2.63 times higher at horse-occupied compared to horse-removed sites at high elevations (Beever et al., 2008). These observations show that free-roaming equids can facilitate dispersal of non-native plant species.  In at least some instances (e.g. the Australian Alps), the trails created by free-roaming horses can also become invasion windows (Dyring, 1990; Campbell and Gibson, 2001).


Although some shrub species are quite palatable to free-roaming equids (e.g. winterfat, Krascheninnikovia lanata), dietary data from investigations across the western USA suggest that shrub consumption is relatively infrequent, in most seasons and places (e.g., Hubbard and Hansen, 1976; Hansen et al., 1977; Hanley and Hanley, 1982; McInnis and Vavra, 1987. Thus, equids more commonly affect shrubs via trampling and rubbing (Beever et al., 2008).  At 19 sites across the western Great Basin, horse-removed sites averaged 1.1-1.9 times greater shrub cover than did sites used by free-roaming horses, although stem count (i.e., shrub abundance) did not differ significantly between the two types of sites (Beever et al., 2008).  Maximum length of shrubs was significantly shorter at horse-used sites than at horse-removed sites, and a-priori planned stepwise regressions identified presence of horses as a strong determinant of shrub length in both years of the study (Beever, 1999).  The authors concluded that this suggests that lower shrub cover probably reflects, at least in part, trampling of shrub edges and thus dissection of the shrub canopy (Beever et al., 2008).  Trampling of vegetation by free-roaming horses has also been reported from New Zealand (Rogers, 1991) and Australia (Dyring, 1990).  In research around 5 exclosures in Sheldon National Wildlife Refuge in northwest Nevada, USA, percentage cover of sagebrush increased significantly more inside exclosures over a 3-year period than outside them (Davies et al., 2012).

Grasses and graminoids

Given that equids are grazers, one would naturally expect that areas used by equids might possess less cover of grasses and graminoids than unused areas.  Across nine mountain ranges of the western Great Basin, USA, horse-removed sites possessed 1.9-2.9 times greater cover and 1.1-2.4 times greater frequency of native grasses than did horse-used sites (Beever et al., 2008).  Stepwise regressions and information-theoretic analyses of alternative hypotheses suggested that presence of horses most strongly predicted frequency and cover of grasses, across both years of the study (Beever, 1999; Beever et al., 2008).  In the Pryor Mountain Wild Horse Range, Fahnestock and Detling (1999) found that although there was not a consistent effect of herbivory by free-ranging horses on plant cover across sites, its effects were “conspicuous” at some sites.  Whereas plant cover of most functional groups in arid lowlands and more-mesic communities most strongly reflected interannual differences in precipitation, cover of the dominant grass (Elymus spicatus, also known as Pseudoroegneria spicata) averaged 3-12% in long-term horse-grazed sites and 9-28% in grazing exclosures.  In contrast, Greyling (2007) found in the Namib Desert in southern Africa that “(s)ince the horse density is generally low, they do not utilize excessive large percentages of available grazing in the study area”.


Percentage cover of forbs was higher at horse-removed sites of the western Great Basin, but stepwise regressions and information-theoretic analyses suggested that factors other than presence of horses (e.g., PRISM-estimated annual precipitation) were most strongly responsible for variability in forb cover across sites (Beever, 1999; Beever et al., 2008).  Much of this forb cover consisted of species that were grazing-resistant or unpalatable to ungulates, such as Wyethia or Lupinus (Beever et al., 2008).  Earlier research on arid ecosystems suggests that high dicotyledon diversity or forb cover can often be associated with higher levels of community degradation (Fogden, 1978; Archer and Smeins, 1991, Kerley, 1992).

Aspects of the plant community

In addition to these differences in particular life-forms of plants between areas used and unused by free-roaming equids, aspects of the plant community also differ.  Across the western Great Basin, 1.82-hectare sites used by free-ranging horses possessed 2-12 fewer species per site than did horse-removed sites, although mean species richness of plants on 50-m transects did not differ between the two types of sites (Beever et al., 2008).  This suggests that free-ranging horses may homogenize plant composition at resolutions of tens to hundreds of meters.  Diversity indices (Simpson’s and log-series; Magurran, 1988) were generally higher at horse-removed sites than at horse-used sites, though usually not different in a statistically significant manner.  Total percentage cover of vegetation was significantly higher at horse-removed sites, and information-theoretic analyses suggested that presence of horses most strongly predicted total cover, across both years of the study (Beever et al., 2008).  In fact, across 11 vegetation response variables, information-theoretic analyses revealed the presence of horses to have the highest variable weight among all fixed-effects explanatory variables, except for frequency and percentage cover of forbs, and maximum length of shrubs (i.e., degree of dissection of the shrub canopy) (Beever et al., 2008).  Total percentage cover of plants was strongly correlated with penetration resistance (i.e. soil hardness), suggesting that compaction may indirectly contribute to changes in vegetation caused by horses, in addition to the direct effects of consumption and trampling.  In a 1999 Range Condition Survey at 40 sample locations on the Nevada Wild Horse Range, biological consultants found 4 locations (10%) in poor condition, 24 (60%) in fair condition, 12 (30%) in good condition, and zero (0%) in excellent condition (Science Applications International Corporation, 1999).  They estimated that percentages of the Wild Horse Use Area (WHUA) in those four categories were 6.4, 59.9, 13.0, and 0 percent, respectively (Science Applications International Corporation 1999); the remainder were in early- to mid-seral (0.9%), early- to late-seral (1.0%), mid- to late-seral (15.0%), and unclassified (3.8%). Trend was reported as "declining" at 12 of the 40 sample locations, "not apparent" at 23 locations, and "improving" at 5 locations. Extrapolation from the study locations to the WHUA reflected the mapping and consequent estimated percentage of the WHUA of the WHUA's 14 range sites. 

To try to address predictions of the Intermediate Disturbance Hypothesis (Grime, 1973; Connell, 1978), Beever et al. (2008) compared fit of linear and quadratic regressions to three alternative measures of species richness of plants within a site to the density of defecations by (horses + cattle) and by horses alone.  In all six cases, the linear model provided the best fit, and richness declined monotonically with increasing intensities of grazing ‘disturbance’ (Beever et al., 2008).  They concluded that either the theory does not apply to grazing in Great Basin aridlands, or that “the long-term past grazing history of these sites has resulted in generally all of them being located on the more disturbed side of the bell-shaped curve.” (Beever et al., 2008).

In addition to performing these landscape-scale analyses of vegetative characteristics at horse-used versus horse-removed sites, Beever et al. (2008) also used horse-trail transects to compare vegetation in areas that were frequently travelled (i.e. along edges of horse tracks) to randomly selected transects within horse-occupied sites.  They found that the magnitude and direction of trends observed at the landscape scale were mirrored by results from horse-trail transects (e.g. lower shrub cover along horse tracks than in randomly located transects, within the same site), though small sample sizes limited the robustness of these conclusions (Beever et al., 2008).  Sample sizes were naturally limited, as only a limited number of horse-trail transects could be established within a 1.82-ha site without committing pseudoreplication (Hurlbert, 1984).  In Australia, plant species richness declines on trails used by horses (Dyring, 1990).

Beever et al. (2003), using two types of multivariate analyses, investigated whether horse-used and horse-removed sites differ across an ecoregion in a holistic or whole-community sense.  They assessed how well the two types of sites, each of which were sampled in two elevational strata (high and low sagebrush-dominated sites), could be discriminated by five data sets collected at 19 sites.  Horse-used and horse-removed sites were extremely well mixed in ordination space using abiotic properties of the sites (e.g., NRCS-estimated productivity, available water capacity, frost-free days).  Furthermore, horse-used and horse-removed sites could not be clearly discriminated in ordination space using cover of key plant species consumed by horses (measured by management-agency horse specialists in horse-effects monitoring), nor by using cover or frequency of all plant species (Beever et al., 2003).  This would indicate that free-roaming horses were exerting little effect on the communities that they occupied.  However, horse-occupied and horse–removed sites were clearly discriminated when using a diverse suite of variables that prior research suggested were sensitive to grazing disturbance.  These variables included density of ant mounds, soil-surface hardness, species richness of plants, grass cover, forb cover, shrub cover, and other variables (Beever et al., 2003).

Water bodies

Analyses of habitat use by free-roaming horses in sagebrush communities of western North America found that horses preferentially favor riparian habitats (Crane et al., 1997). This mirrors the conclusion by Dyring (1990) that horses in Australia preferentially graze heathlands and grasslands, and the findings of Menard et al. (2002) that feral horses preferred marsh–bog–meadow habitats during warm seasons in European pasturelands shared with cattle.  Use of areas near streams can increase run-off (Dyring, 1990; Rogers, 1994), and accordingly reduce water quality (Nimmo and Miller, 2007).  Flush zones, perhaps due to their high nutrient content, are disproportionately used by free-roaming horses in New Zealand, thus resulting in vegetation trampling, alterations in stream flow, and downstream siltation (Rogers, 1991).  In central Australia, free-roaming horses have been asserted to “foul water holes with carcasses and cause accelerated gully erosion” (Berman and Jarman 1988).  Analogously, in the Australian Alps, stream banks frequented by free-roaming horses have been characterized as “extensively churned up and broken down”, due to the action of horse hooves sinking into peaty soils (Dyring, 1991).  Bog habitats in Australia and New Zealand can also be strongly altered by free-roaming horses (Dyring, 1990; Rogers, 1991; Clemann, 2002).  Greyling (2007) found in Namibia that a “100 m radius around the water troughs at Garub is visibly sacrificed.  Trampling of large herbivores inhibits the growth of vegetation and to some extent changes the species composition of mainly annuals.”  However, he found that the “expected degradation gradient radiating out from the water troughs due to over-utilization by the horses was not found.  Neither vegetation species composition, density, nor standing biomass measured at various distances from the troughs confirmed a degradation gradient” (Greyling 2007).  Using existing exclosures for research in two mountain ranges of the western Great Basin, USA, Beever and Brussard (2000b) found that plots near springs in exclosures had markedly greater plant species richness, percentage cover, and abundance of grasses and shrubs, in addition to more small-mammal burrow entrances.  At higher elevations, ungrazed meadows had higher maximum plant heights, higher species richness of plants per quadrat, and (in 2 of 3 pairs of plots) greater species richness of small mammals than grazed meadows (Beever and Brussard, 2000b).

Impact on Biodiversity


See above under ‘Impact on Habitats’.

Small mammals

In areas with relatively high densities of free-roaming horses in Australia, sites with horses exhibited alterations to numerous aspects of mammals.  For example, areas heavily grazed by feral horses and cattle exhibited fewer signs of macropods (e.g. kangaroos) (Berman and Jarman, 1988).  Nano et al. (2003) cited removal of feral horses in the Northern Territory as a possible reason for the recovery of central rock rats (Zyzomys pedunculatus).  Analogously, signs of black-footed wallabies (Petrogale lateralis) were more numerous in a series of surveys conducted after the removal of feral horses from Finke George National Park in the Northern Territory, compared to before removal (Coventry and Robertson, 1980; Mansergh, 1982; Nimmo and Miller, 2007).

Across the western Great Basin, USA, Beever and Brussard (2004) found community completeness of granivorous small mammals to be significantly lower at sites used by free-ranging horses compared to horse-removed sites. In that research, community completeness was defined as the percentage of species detected during a standardized period of trapping effort, compared to the full suite of species predicted to exist at a site given the site’s elevation, geographic location, and microhabitats. Two to four times higher capture rates of the grass-specialist western harvest mouse (Reithrodontomys megalotis) at horse-removed compared to horse-used sites probably reflected the higher grass cover at horse-removed sites (Beever and Brussard, 2004).  Areas used by horses can be more dominated by disturbance-tolerant generalists such as deer mice (Peromyscus maniculatus) and species that specialize in open areas without vegetation (Beever and Brussard, 2004). Although species richness did not differ in a statistically significant manner, completeness differed significantly due to the differential species pools available across the broad study domain, which straddled the Lahontan Trough and areas on both sides of the Trough (Beever and Brussard, 2004).  In the Namib Naukluft Park of Namibia, neither abundance nor species richness of small mammals differed appreciably between horse-occupied and control plots (1-2 replicate sites per treatment) during three short trapping periods in 2004 (Greyling, 2005). The only exception to this trend was that 10 individuals from four species were captured at two control sites dominated by Euphorbia shrubs in November 2004, whereas only one individual was captured at the two comparable horse-occupied sites (Greyling, 2005). However, Greyling (2005) posited that “some of these (not-trapped) species were observed in both areas, and it is believed that all these species do actually occur in both.” During 1993-2005, the population size of horses ranged from 89-149 animals (mean + 1 SD = 122 + 20 individuals), and the area mainly used by the horses encompassed 40,000 hectares or 400 km2 (Greyling, 2005).

Large ungulates

Meeker (1979) and Berger (1985) found horses to be dominant among native Great Basin (USA) ungulates in social interactions (probably due to their larger body size), particularly at spring areas.  As such, horses can induce alterations in the spatial and temporal dimensions of habitat use by sympatric native ungulates (Meeker, 1979; Ganskopp and Vavra, 1987; Coates and Schemnitz, 1994).  For example, horses have been found to displace native ungulates already using water sources, and to delay or prevent native ungulates from approaching sources and drinking (Meeker, 1979; A. Gooch, Brigham Young University, Provo, Utah, USA, personal communication, 2012).  Ostermann-Kelm et al. (2008) used manipulative field experiments to assess whether horses affect the tendency of bighorn sheep to use watering sites at three elevations in Anza-Borrego State Park in southern California.  They observed no evidence of direct interference competition (i.e. aggression) between bighorn sheep and feral horses.  However, the two species were never observed using water simultaneously, nor were bighorn sheep ever observed to water when horses were within sight of the watering source (Ostermann-Kelm et al., 2008). Furthermore, the placing of tethered domestic horses near sites preferred by bighorn sheep reduced the number of sheep groups watering at that site by 76% and increased sheep-group visits to other sites in the region (Ostermann-Kelm et al., 2008).  Horses remained at wash or riparian areas for an average of 0.3, 2.3, and 4.9 hours per day during July-October, at the lower-, middle-, and upper-elevation sites, respectively (Ostermann-Kelm et al., 2008).


Across the western Great Basin, USA, horse-occupied sites exhibited lower species richness of ants and density of ant mounds than did horse-removed sites (Beever and Herrick, 2006). High-elevation horse-removed sites, however, had highly variable densities of ant mounds.  In contrast to other ecosystem components in that research, biologically driven stepwise regressions suggested that factors other than the presence of horses best predicted density of ant mounds (Beever and Herrick, 2006).  In the provincial Namib Naukluft Park in the Namib Desert, Namibia, with low horse densities, Greyling (2007) found that the “general species composition and abundance [of ants] between the Horse and Control plots were similar”.  However, two species were found exclusively in Control plots, and one species exclusively in Horse plots (Greyling, 2007).  Multivariate analyses and classifications produced equivocal results.  Greyling (2007) also found using Student’s t-tests that tenebrionid beetles did not differ significantly in either species richness or abundance between plots in Horse and Control areas.  Sampling at different distances from a water trough suggested no pattern in tenebrionid abundance or species richness as distance from the trough increased (Greyling, 2007).  Whereas four tenebrionid species were recorded only in Control sites, five species were recorded only in Horse sites; in all nine cases, these species were represented by only one or a few individuals (Greyling, 2007).


In the Pampas grasslands of Argentina, Zalba and Cozzani (2004) studied the effects of different grazing regimes of free-roaming horses on bird communities in a nature reserve.  Areas dominated by tall grass (i.e. exclosures and moderately grazed areas) exhibited the greatest species richness and total abundance of birds, though some species were associated with the presence of horses while others were more common in less-grazed areas (Zalba and Cozzani, 2004).  The presence of free-roaming horses was associated with an increased rate of egg predation; this varied from 12.5% in the exclosures to 70% in grazed areas (Zalba and Cozzani, 2004).  However, bird-species richness and diversity were higher in areas grazed moderately than in horse-excluded areas (Zalba and Cozzani, 2004), a result consistent with predictions of the Intermediate Disturbance Hypothesis (Connell, 1978).  Possible interactions of free-roaming equids with the declining Greater Sage-Grouse in the USA were thoroughly outlined by Beever and Aldridge (2011).

Other ecosystem components

Levin et al. (2002) studied numerous ecosystem attributes on six marsh islands of ~10 km2 each that were adjacent to Shackleford Banks, North Carolina, USA.  Three were accessible to horses, whereas the other three were not visited by horses.  Earlier long-term research in marshes suggested that free-roaming horses significantly decrease abundance of Spartina at sites to which they have access (Wood et al., 1987, Hay and Wells, 1991).  Horse-grazed marshes exhibited less vegetation, a higher diversity of foraging birds, higher densities of crabs, and lower species richness and density of fishes than did marshes not grazed by horses.  Fish density was also lower in sub-tidal habitats adjacent to grazed marshes (Levin et al. 2002).

Social Impact

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In a thorough investigation of attitudes of various segments of U.S. society, Kellert and Berry (1980) found the horse to be the second-most-liked animal, trailing only the domestic dog.  Likely in part due to their long-standing commensal relationship with humans, horses have been suggested to embody wildness, power, endurance, loyalty, and beauty, among other characteristics (Beever and Brussard 2000a).  Feral horses have been deemed one of the most contentious management issues during the last 100 years by numerous authors (Thomas, 1979; Wagner, 1983; Symanski, 1994; Bama, 1998; Nimmo and Miller, 2007), galvanizing diverse interest groups into tightly held positions.  When delinquent young men shot and killed 25 horses (that were descended from privately owned horses that had gone feral) near Reno, Nevada, USA, at close range with high-powered weapons and deployed a fire extinguisher on a foal’s face in late December 1998, numerous interest groups were able to raise over $100,000 within 24 hours in response to the atrocity.  The U.S.’s Wild Free-Roaming Horses and Burros Act of 1971 was instituted in response to outrage expressed by animal-rights activists, horse advocates, and schoolchildren over inhumane treatment of horses during control and transport management actions.  The Act deems horses “a national treasure”, “living symbols of the historic and pioneer spirit of the West,” and “an integral part of the natural system of the public lands.”  Implementation of the 1978 Public Rangelands Improvement Act, which allowed “the removal and disposal” of feral horses by virtue of their perceived threat to ecosystem integrity of rangelands, sparked public outcry and fierce battles among numerous stakeholders (Nimmo and Miller, 2007).  A unique twist to human social dynamics is that the free-ranging horse controversy has on some occasions created alliances that occur relatively infrequently (e.g., the alliance of conservation activists, livestock operators, and wildlife-hunting proponents in opposition to unlimited numbers of free-ranging horses) (Beever and Brussard, 2000a).  Both research and management of free-ranging horses in the western U.S. have been strongly shaped by public input, arguably since the early 1970s (Beever and Brussard, 2000a).  Adaptive management (Holling, 1978; Williams, 2011), structured decision-making (Martin et al., 2009), use of citizen science (Miller-Rushing et al., 2012), and analytic deliberation (Stern and Fineberg, 1996) are processes that may assist in objectively addressing the controversy surrounding this topic.

Aerial surveys in central Australia undertaken in the late 1980s by the Australian Conservation Commission of the Northern Territory (CCNT) showed a population of approximately. 200,000 horses, a much-larger total than was previously assumed (Dobbie et al., 1993).  Consequently, the CCNT offered financial assistance to pastoralists who wanted feral horses removed from their land (Symanski, 1994).  However, the culling of horses drew outrage from numerous animal-rights groups internationally, and the International Court of Justice for Animal Rights tried and convicted several members of the Australian government for massacring horses (Symanski, 1994; Nimmo and Miller, 2007).  Due in part to public outreach and education, and the rate of horse-population increase, culling has subsequently received wider use across Australia.

Risk and Impact Factors

Top of page Invasiveness
  • Highly adaptable to different environments
  • Is a habitat generalist
  • Pioneering in disturbed areas
  • Tolerant of shade
  • Capable of securing and ingesting a wide range of food
  • Highly mobile locally
  • Benefits from human association (i.e. it is a human commensal)
  • Long lived
  • Fast growing
  • Has high reproductive potential
  • Gregarious
  • Has high genetic variability
Impact outcomes
  • Damaged ecosystem services
  • Ecosystem change/ habitat alteration
  • Increases vulnerability to invasions
  • Infrastructure damage
  • Modification of hydrology
  • Modification of natural benthic communities
  • Modification of nutrient regime
  • Modification of successional patterns
  • Negatively impacts agriculture
  • Negatively impacts livelihoods
  • Negatively impacts tourism
  • Reduced native biodiversity
  • Threat to/ loss of native species
  • Transportation disruption
Impact mechanisms
  • Competition - monopolizing resources
  • Fouling
  • Herbivory/grazing/browsing
  • Interaction with other invasive species
  • Trampling
Likelihood of entry/control
  • Highly likely to be transported internationally deliberately
  • Difficult/costly to control


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Domestic horses are used worldwide; uses in different parts of the world include agriculture, transport, sport and recreation, or food. Use of domestic horses is not discussed in detail in this datasheet.

Free-ranging horses, given their emotional and romantic appeal to numerous stakeholders, can provide a diversity of values, from economic (via tourism and non-consumptive uses such as viewing horses) to socio-cultural (especially for some Native American peoples) value.  A range of material values (via processing of carcasses) have occurred in the past, but these are not universally acceptable.  If there were no stigma associated with using parts of horses that perished due to age, misstep, drought, or overwintering death, the large body mass of horses suggests that each individual could provide for many other needs.

Uses List

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Animal feed, fodder, forage

  • Fodder/animal feed
  • Meat and bonemeal


  • Draught animal
  • Pet/aquarium trade
  • Ritual uses
  • Sociocultural value
  • Working animals (miscellaneous)

Human food and beverage

  • Meat/fat/offal/blood/bone (whole, cut, fresh, frozen, canned, cured, processed or smoked)


  • Skins/leather/fur

Medicinal, pharmaceutical

  • Traditional/folklore

Similarities to Other Species/Conditions

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Comparisons with other Equus species are listed in the ‘Description’ section, but E. caballus, although tremendously variable in size, shape, and pelage (coat) colour and markings, is likely to be readily distinguishable from all other species by individuals from most human societies.  Willoughby (1974), Bennett (1992), and Bennett and Hoffmann (1999) discuss differences among the living and extinct subspecies of E. caballus and other equids.

Prevention and Control

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To the author’s knowledge, neither biological, chemical, cultural, nor mechanical control measures have been used on free-ranging populations of E. caballus.

Because horses have been domesticated for nearly 6 millennia, control by utilization has been pervasive and common over the past two centuries; in such instances, horses change from being feral to domesticated.  Uses have included transportation, battle mounts, a work ‘implement’, recreational pursuit, and in some cases in Eurasia, consumption as food by humans or dogs (Beever and Brussard, 2000a).

In Australia, a national population size nearing half a million or more individuals has led to more-aggressive measures to reduce the prevalence of environmental degradation and density-dependent die-offs of horses.  In particular, Nimmo and Miller (2007) reported that in Australia “Options for population control of feral horses are quite limited but include ground shooting, mustering and trapping, helicopter shooting, immobilisation, and fertility control.”

In the western USA, control measures have been restricted since 1971 to removals (by roundups and aircraft-assisted gathers), sex-ratio skewing, and various types of reproductive control. In particular, control has largely been effected through periodic roundups and gathers of individuals, concentrating animals by hazing them on horseback or with a helicopter, or by using bait or water inside corrals.  Gathers occur every 3-5 years on larger Herd Management Areas, and sometimes less frequently on areas with smaller population sizes.

Immunocontraception has been deemed the socially most acceptable method of population-size control, and research has occurred for over two decades on various formulations and adjuvants, and frequency, duration, and method of administration (Kirkpatrick et al., 2011; Ransom, 2012). Other methods of fertility control include: adjustment of the sex ratio (via biased gathers and re-releases) to limit reproductive rates and Ne (effective population size); surgical ovariectomy (i.e. removal of ovaries from females); intrauterine devices (which mechanically prevent fertilization); immunocontraceptives that trigger the animal’s immune system to prevent pregnancy (e.g., porcine zona pellucida [PZP] vaccine, or PZP formulations designed to extend the period of contraception); gonadotropin-releasing hormone (GnRH) vaccines; and steroid hormones (e.g. natural and synthetic progestins and oestrogens) (Asa and Porton, 2005). Immunocontraceptives have not yet been utilized widely, but have primarily been administered to horses that have been gathered, while in captivity before either being moved to holding facilities or released back out into a free-ranging state. Research has been initiated to investigate the feasibility, efficacy, and possibility of any injury by delivering the immunocontraceptives via darting (see Roelle et al. 2010).

A plethora of investigations have studied the physical, physiological, genetic, behavioural, phenological (of cycling and parturition), and demographic (related to survival, age structuring, and population processes) side effects of these various approaches to fertility control (Gray and Cameron, 2010; Ransom, 2012). Though observed and inferred growth rates range widely (Roelle et al., 2010), annual population growth rates across the western USA of horse herds have hovered around 18-25%, meaning that without control, population size doubles every 3-4 years.  However, removals ensure that population size remains below the level of maximum sustained yield, thus preventing the enabling of density-dependent population self-regulation.

Monitoring and Surveillance

Across the western USA, numerous techniques are used to enumerate free-roaming horses.  Such techniques, which can all be performed from the ground or from aircraft, include strip or line transects, mark-resight and mark-recapture approaches, simultaneous double-counts, and removals shortly after other enumeration procedures.  Other approaches, which do not have currently have sufficient fine-tuning to be fiscally feasible in diverse contexts, include genetic techniques (e.g., hair snares, faecal DNA), and use of forward-looking infrared (FLIR) remote-sensing technology.

Ecosystem Restoration

Little to no research has been performed on ecosystem response to experimental manipulation of feral-horse densities, perhaps due in part to the logistical challenge and socio-political controversy that may accompany such action.  Beever and Brussard (2000b) described results of comparisons of plant and small-mammal measures inside and outside of grazing exclosures in two mountain ranges, in areas where cattle use was light or actively excluded.  Research at 19 sites from 9 mountain ranges across >3 million ha of the western Great Basin (Nevada, USA) compared penetration resistance, density of ant mounds, measures of vegetation, and abundance of granivorous small mammals and squamate reptiles at sites used by free-ranging horses vs. sites from which horses had been actively removed for 10-14 years (Beever et al., 2003; Beever and Brussard, 2004; Beever and Herrick, 2006; Beever et al., 2008).  The results of this research are described in the Impacts section.

Gaps in Knowledge/Research Needs

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Much remains to be learned of the synecology of free-ranging horses, and how their effects on other ecosystem components vary across ecological contexts and key environmental gradients (e.g., rainfall, elevation, seasonality, temperature) (Beever and Aldridge, 2011).  A number of applied research questions related to the ecological relationships of free-ranging horses occur in Beever and Brussard (2000b) regarding experimental and monitoring designs for grazing effects and using exclosures; Beever and Brussard (2004) regarding more-time-intensive field methods, use of manipulative studies, studies at watering areas, and exploration of other indicator taxa; Beever and Herrick (2006) regarding the relative importance of direct effects, indirect effects, and feedbacks in grazing-induced ecosystem alterations; and Beever et al. (2008) regarding greater replication of horse-trail transects to examine fine-scale grazing effects and how far they extend laterally from the trail centre, and how and in what contexts horses do or do not increase abundance of non-native plants.


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23/11/2013: Original text by:

Erik A. Beever, US Geological Survey, Bozeman, MT 59715, USA.

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