Genetic threats, among other factors such as human-induced habitat loss, pollution, alien invasion and overexploitation are threatening persistence of natural populations. However the relative importance of genetic threats in conservation is a subject of current debate. Loss of genetic variability, accumulation of mildly deleterious mutations and inbreeding depression has been identified as the major threats to natural populations (Gaggiotti, 2003); their effects are more pronounced in small, isolated and rapidly declining populations (Caro and Laurenson, 1994). They reduce survivorship, fecundity and mean population fitness which all tend to drive populations to extinction. Genetic threats can be short-term on individual viability and fecundity (have a direct effect on population demographics, potentially making it more susceptible to extinction) or long term (effects on the population’s response to environmental changes) that might lead to extinction as well (Amos and Balmford, 2001). In this essay, the level of threats posed by genetic factors (genetic variability, inbreeding and accumulation of mildly deleterious mutations) to conservation will be discussed under the context of isolated populations, local populations with immigrations and metapopulations. 

1 Loss of genetic variability
Genetic variation allows adaptation by populations to environmental changes, re-establishment after local extinction and range expansion but such abilities might be lost in the absents of migrations, mutations or selection favoring heterozygotes (Gaggiotti, 2003). Stochastic changes in gene frequencies (due to mendelian segregation and family sizes variations) may lead to random genetic drift in small populations resulting in reduced population` s ability to respond to future environmental changes and evolution (Caro and Laurenson, 1994). Genetic drift eliminates low frequency alleles very rapidly when populations are small or declining. This is detrimental for survival of populations in the long-term but there are few cases that clearly demonstrate that genetic drift is an imminent extinction threat (Gaggiotti, 2003). A well documented case of genetic drift is for the Mauritius kestrel (Falco punctatus), where the population was reduced to a single pair in the 1950s and it recovered slowly to over 500 birds. Compared to museum specimens, the kestrel lost about 50% of heterozygosity on microsatellites during the bottleneck (Groombridge et al., 2000), however, Amos and Balmford (2001) comments that this was an exceptional case where a high percentage of heterozygosity was lost.  Furthermore, data from the Antarctic fur seal (Arctocephalus gazelle) indicate that extreme but short-term depletion may have little impact on heterozygosity as this species shows little evidence of genetic erosion after severe depletion by sealers (Gemmell et al., 2001). Genetic drift is a long term threat, conclusions from short-term studies might be biased because of lack of long-term data and projections from models might over-estimate the risk of extinction.

The more important the genetic variability (in terms of both absolute selective value and the proportion of time any advantage is manifest), the more likely it would be to be retained in a species` population (Amos and Balmford, 2001). Most of the genetic variance responsible for an evolutionary response to natural selection (the additive genetic variation) is found in high-frequency alleles, those least likely to be lost in small populations (Poon and Otto, 2000). Hence drift tends preferentially to remove the variability which is currently least important to the organism; implying that the population will under immediate threat of extinction, although consequences in the long-term will be uncertain.

Genetic variation lost during a bottleneck is a function of growth rate (Nei et al., 1975). Little variation will be lost in species that quickly recover and have long overlapping generations (reproduce many times therefore variability present in a cohort is likely to be transferred to future generations) (Gaggiotti and Vetter, 1999).  Species that have gone through bottlenecks (e.g. California sea otter and Indian rhinoceros) do not necessarily show reduced genetic variation, but in those that do, the number of deleterious recessives depends on the rate of bottleneck occurrence; if rate was slow they would have been purged but not fixed (Caro and Laurenson, 1994). This means if rapid population declines are quickly corrected then little genetic variation especially in long-lived species.

The loss of allelic diversity can have serious consequences in the short term if it occurs at loci associated with disease resistance such as the major histocompatibility complex (MHC) in vertebrates (Allendorf and Ryman, 2002). Low variation at the MHC complex after bottlenecks has extinction risks for example in the Arabian oryx (Oryx leucoryx). A study of this species reported only 3 alleles at the MHC class II DRB from 57 individuals (Hedrick and Kim, 2000). It went extinct in the wild in 1972 (because of hunting) and captive populations were vulnerable to tuberculosis and foot and mouth diseases. However, cross-species comparisons are ambiguous as the Northern elephant seal (Mirounga angustirostris) show low MHC diversity but is not prone to a range of diseases compared to  Californian sea lions (Zalophus californianus) that use the same habitats (Hoelzel et al., 1993). Low MHC variability therefore is a serious risk but cannot be set as a rule for all species as some only show negative effects when there are other confounding factors such as reduced effective population size (N_e), limited resources and predation.

Few species populations have been reported to have gone extinct or decline primarily due to loss of heterozygosity (Caro and Laurenson, 1994). The cheetah (Acinonyx jubatus) is commonly cited as lacking genetic variation and this was previously blamed for its poor survival prospects in the wild. In the mid 1980s, a study of 55 cheetahs from southern Africa demonstrated complete lack of variation at each of 47 allozyme loci (O'Brien et al., 1985), skin grafts were not rejected among animals implying that their immune systems were genetically identical. Further studies in 1990s questioned whether wild cheetah’s survival was being compromised by their lack of genetic variation, and Caro and Laurenson (1994) pointed out that disease susceptibility and breeding problems were an issue more for captive cheetahs while predation was more important in the wild population. The claims that cheetah lost 90-99% of its variability in one or more bottlenecks (O'Brien, 1994) are questionable as it would require 16 generations of sib-sib mating (N_e=2) to lose 99% of genetic variation (Amos and Balmford, 2001). A recent study in Namibia (Castro-Prieto et al., 2011) revealed more variation in Class I MHC, but not Class II MHC and also the amount of DNA sequence variation among the alleles was fairly high thus they could recognize a wide array of foreign proteins. However, they noted that although wild cheetahs appear to have enough MHC variation for defense against current common diseases, they may still be at risk from other new emerging diseases. The importance of low genetic variation in cheetah might have been over-estimated as Castro-Prieto et al (2011) concluded by noting that the long term survival of free-ranging cheetahs in Namibia seemed more likely to depend on human-induced rather than genetic factors.

Genetic variability is lost slowly (over hundreds to thousand generations) and the rate of loss is inversely proportional N_e. This implies that variability loss mainly depends on the size of the lowest population to date and for how many generations it has been held at that level; it is slow compared to the timescales over which conservation biology operates (Amos and Balmford, 2001). Thus, although some studies claimed to show a link between known bottlenecks to low levels of genetic variability like in cheetah (O'Brien, 1994) and in Elephant seals (Hoelzel et al., 1993), closer examination reveals that the expected loss of heterozygosity might be far less than thought  (Amos and Balmford, 2001). The level of threat posed by genetic variability considering this aspect is therefore currently not very serious but there is no guarantee that situation will remain the same given the ongoing challenges in threats like habitat degradation, fragmentation, pollution and climate change. 

2 Inbreeding depression
Inbreeding depression is reduced fitness in a population caused by mating between related individuals. Genetic theory predicts that inbreeding will reveal deleterious recessive alleles, that might be shown in reduced growth rates, lower fecundity and high infant mortality that could lead to population extinction (Caro and Laurenson, 1994). In a large pop with random mating these detrimental alleles are present in heterozygous state and therefore their effects are partially masked.  In small populations however, mating among relatives becomes common and the proportion of individuals that are homologous at many loci increases and results in inbreeding depressions. Following this theory, inbreeding depression should be a real cause for concern in conservation for many species that are in decline or that survive in small isolated populations.

Researchers have questioned the importance of inbreeding for the persistence of wild populations since most inbreeding evidence comes from domesticated/captive populations. There are also theories suggesting that most of inbreeding depression can be purged considering the many ways in which species avoid inbreeding (Keller and Waller, 2002). However, there have been accumulation of evidence for inbreeding in wild animal populations, for example the Soay sheep (Ovis aries) in the UK, homologous sheep suffered high parasitism and lower overwinter survival in comparison to heterozygous sheep (Coltman et al., 1999). In south-eastern USA, inbreeding reduced hatching rates, fledging survival and recruitment to the breeding population of red-cockaded woodpeckers (Picoides borealis) (Daniels and Walters, 2000). Also, evidence for inbreeding in natural populations of plants exist and purging does not appear to act consistently as a major force in these populations (Byers and Waller, 1999). In an experiment using Clarkia pulchella that differed in founders` relatedness, persistence time was found to be lower in individuals whose founders were related (Newman and Pilson, 1997). This then suggest that inbreeding depression is important in the conservation of these species.

It has been noted that  the importance of inbreeding depression in the wild does not necessarily imply population decline (Caro and Laurenson, 1994).  Environmental stress such as food and water shortages can reveal effects not seen under less demanding conditions. Inbreeding depression and its interaction with environmental stress was demonstrated in an experiment with mice (Jimenez et al., 1994). Wild caught mice subjected to full-sib mating in captivity showed greater mass loss and lower survivorship when they were returned to their habitat compared to the progeny of outbred matings. When equivalent mice were retained under laboratory conditions, inbreeding depression was less severe. This suggests that the important effect of inbreeding depression lies with its tendency to aggravate the consequences of environmental downturns. However other studies showed evidence for direct connections of inbreeding with population decline, for example in Glanville fritillary butterfly large metapopulation, extinction risk was positively related to inbreeding even after accounting for ecological factors (Saccheri et al., 1998).

The above discussed evidence shows that inbreeding depression is common even in wild populations and can be a short term extinction threat especially if populations are subjected to environmental stress or rapid population decline. The severity of inbreeding and genetic variability loss can be reduced by immigration of a few individuals into a population - what has been termed the rescue effect. In the absence of such migrants, inbreeding depression can contribute to the driving of populations to an extinction vortex.

3 Accumulation of mildly deleterious mutations
In stable environments, mutations with phenotypic effects are usually deleterious since populations tend to be well adapted to their environment (Gaggiotti, 2003). Random mutations are likely to disrupt such environmental adaptations. Selection is efficient in eliminating detrimental mutations (with large effects on fitness) in when N_e is large or moderate. Mild deleterious mutations with selection coefficient s≤1/2N_e behave as neutral mutations and are therefore difficult to remove (Wright, 1931). When Ne drops to a new value, very small deleterious mutations begin to accumulate after approximately 4/N_e generations and can rapidly drive populations to extinction after N_e<100-1000 (Kondrashov, 1995). In small populations, selection is hampered and this increases the role of genetic drift thereby increasing the chance fixation of some of the deleterious alleles from mutations. This result in reduced fitness for the population, which might eventually lead to extinctions (Muller, 1964). Previously, this was thought to be a problem only in obligately asexual species as there is no recombination (offspring will have parent mutations as well as newly arisen mutations) (Muller, 1964), but sexual species are also at risk of extinction due to mutation accumulation (Lande, 1994). If this process repeats, mutations will accumulate and there will be further declines in fitness and population size forming a positive feedback mechanism-a process called mutational meltdown (Lynch and Gabriel, 1990). Recombination in sexual species can slow down mutational meltdown to some extent, but they are not entirely immune from it (Gaggiotti, 2003).

Empirical evidence for mutational meltdown is scarce for wild populations, and this threat might have been overestimated as an artifact of how the mutation effects on mean fitness has been modeled (Poon and Otto, 2000). In some experiment with yeast (Saccharomycetes cerevisiae), Zeyl et al (2001) used 12 replicates of 2 isogenic strains of yeast with genomewide mutation rates that differed by 2 orders of magnitude to demonstrate mutational meltdown. They used an effective population size of about 250 and after more than 100 daily bottlenecks; the yeast with higher mutation rates declined in size and had two extinctions while the wild type remained constant. These results support the mutational meltdown model (Zeyl et al., 2001), but it has been criticized because of controversies in measures of per-genome mutation rates and mean fitness cost per mutation. These measures are thought to be small (Garcia-Dorado et al., 1999) which makes mutational meltdown less likely or less important for most species.

Meltdown models ignore the effect of beneficial and backward mutations. Consideration of these mutations might imply that only very small populations would face the risk of extinction due to genetic stochasticity (Poon and Otto, 2000). Also, new mutations may be compensatory or suppressive, which might restore fitness losses incurred by previous mutations without requiring true reversals (Kimura, 1990). Thus currently it is impossible to give clear evaluation of the importance of meltdown process.

4 Local extinction in the presents of migration
Inbreeding depression is normally reduced by immigrants that are heterozygous for deleterious recessive mutations (Whitlock et al., 2000), and by heterosis mean fitness of populations may be enhanced. However, outcrossing can reduce mean population fitness if hybridisation disrupts coadapted gene complexes or favourable epistatic interactions (outbreeding depression). Few studies have demonstrated outbreeding depression as it requires tracking beyond the F₁ generation. A study of song sparrows (Marr et al., 2002), showed signs of outbreeding depression in the F₂ generation, and measures of fitness was low in the F₂ generation of crosses of the tidepool copepod (Tigriopus californicus) from different populations (Burton, 1990). However, this effect of breaking up coadaptations is only magnified if the genetic distance between the two populations has increased greatly (Edmands, 1999). Thus, the threat of outbreeding might not be very serious in most wild populations since it takes many generations in contrasted environments for genetic distance to be significantly very large.

The reduction or increase of fitness in a population after receiving immigrants also depends upon interactions among several genetic and non-genetic factors (degree of epistasis, demography, behaviour, environmental etc.) (Tallmon et al., 2004). It might therefore be difficult to predict whether any given immigration event will effect genetic rescue especially when conservation managers lack understanding of the interactions between the genetic and non-genetic factors. However, Gaggiotti (2003) reviewed studies on plants such as Lotus scoparius, Ipomopsis aggregata and Silene diclinis and concluded that outbreeding depression may be common in the wild but the potential benefits of outbreeding usually outweigh the threats of outbreeding depression.

5 Extinction in metapopulation context
There has been some theoretical work showing that metapopulation can be subject to extinctions due to genetic factors. Genetic variation can be lost through population turnover. This would be more pronounced when colonizing propagules are formed by individuals from the same deme than from all extant demes (Maruyama and Kimura, 1980). However, if habitat patches differ in quality (the typical case in source-sink metapopulations), then population turnover not have large effects (Gaggiotti, 2003). Moreover, sink populations can maintain a large proportion of variation in the presents of migration.

The mutational meltdown theory was extended to cover metapopulations (Higgins and Lynch, 2001) using individual-based models with stochastic, demographic, environmental and genetic factors. They concluded that mutational meltdown may be a significant threat to large metapopulations and would exacerbate the effects of habitat loss or fragmentation on metapopulation viability. However there is little empirical evidence supporting predictions made in these theories (Gaggiotti, 2003).

Conclusions
For the above discussed genetic threats, inbreeding seems the most likely to exacerbate decline and hasten extinction especially where the reduction in Ne has been very great and under stressful environmental conditions. Most of genetic threats take many generations to be detected; caution must therefore be taken when making conclusions from studies because what may be insignificant for now might be a threat in the future. In most systems we do not know the threshold where fitness will be an imminent threat of extinction. Also, selection intensity on particular measures of fitness (or life history traits) can vary over time and space, thus the cumulative effects of selection on multiple traits will interact to produce overall fitness effects. This implies that short-term studies of a few traits might result in misleading conclusions.

The division between demographic, environmental and genetic is artificial since extinction processes often operate together and their synergy may have a stronger impact, especially for populations of intermediate sizes which were previously thought not to be under extinction risk. For very small populations, extinction risk is more influenced by demographic and ecological stochasticity rather than genetic threats. Neither genetic nor demographic factors per se are responsible for most of the populations decline; they only become important after populations have been driven to very low levels, particularly by human activities. Human disturbances such as poaching, habitat fragmentation, introduction of invasive organisms and pollution present the greatest challenge to populations in the wild than genetic threats. In populations that are less affected by humans (e.g. Checkerspot butterfly), extinction still result from environmental rather than genetic causes (Ehrlich and Murphy, 1987). Conservation efforts should therefore be distributed proportionally to threats posed by any factor, and as shown by most studies, most genetic threats are currently not priorities for many species.

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