Do the effect of inbreeding depression disappear after the first generation of outbreeding?

An animal that is the result of generations of inbreeding tends to have lower fitness. (More diseases etc.) This is usually explained by harmful but recessive mutations that exist in all populations, but in inbred animals end up occuring on both copies of a gene in a chromosome-pair.

If the problem of inbreeding is that the two sets of genes in the chromosome pairs are too alike, does that imply that the offspring of an inbred animal will revert to "normal" fitness as long as the inbred parent was mated with an unrelated animal? Seeing as the inbred parent can only transmit one copy. Or is the situation more subtle?

This depends on if you are talking about animals in the wild, such as isolated populations, or humans, or whether you mean research animals. Inbreeding tends to breed homozygosity, so as you said, if you have a recessive disease causing allele within a bloodline, then even if only one of the founders was a heterozygous carrier, then inbreeding can give rise to afflicted homozygotes. However, for lab strains, it is actually desirable to use completely inbred lines where all animals are as near to 100% identical (from a DNA sequence perspective) as you can get. This makes genetic mutational analysis far easier (possible).

It also depends on the alleles in question. Not every mutation is purely dominant or purely recessive. You can have situations, such as with some tumor suppressor genes, that diversifying the population by outbreeding to someone with a functional tumor suppressor, will still cause the offspring to have only a single working allele, and therefore an increased probability that they can sustain a random mutation that will knock the working gene out. You can also have cases where having one functional copy of the gene isn't enough, so the offspring are haploinsufficient. You need two doses of both working copies of an allele to display the non-afflicted phenotype, though the afflicted heterozygote has a better time of it that the afflicted homozygote, who has none of the working allele.

You also have X-Linked conditions. If the maternal lineage is inbred, then all of the males will be afflicted and females will be carriers.

Another possibility is that the outside breeder is a carrier of the mutant allele so the offspring remain homozygous or the outside breeder has an allele of the gene that comes with its own defect. They may not be afflicted, as they could be heterozygous for a working copy, but their offspring with the inbred line could be heterozygous afflicted because neither of the alleles they inherit are functional, it is just that they are not functional in different ways.

Sexual reproductions main advantage is that it is supposed to increase diversity in the population and increase survivability of the species. When you have isolated populations or inbreeding by choice, then this can circumvents the benefits. Not all inbreeding, as seen with laboratory strains and selective breeding of domesticated animals, is deleterious. It is just that the breeder needs to be careful of the choices made and cognizant of how the genetics can work. Animals displaying a diseased phenotype should not be allowed to produce offspring, and you need to evaluate the siblings of that animal as well as the parents to insure that you do not have problems in that lineage.

Pond Boss Forums STOCKING A NEW POND Types of fish to choose Outbreeding depression

Alan not sure I understand the question. I have written a good bit on many subjects here on PB including outbreeding depression and other genetic topics. As best I can recall I have not seen any info connecting outbreeding depression with environment (location). However definitionally there is a connection (see red text in Wiki material at bottom). Outbreeding depression is a very complicated process and often with unknown cause. If you have a more specific question I will give it a shot - but often on PB the answer is "we just don't know".

A couple years back a group of us including Bob and Dr. Anderson , Bill C. and others discussed genetics vs adaptation (phenotypic plasticity) and time. It was used in a article Bob did on giant BG.

In that context genetics are very long term (we thought) and plasticity = adaptive behavior (adaptation) = phenotypic plasticity was short term changes due to environment. I have thought that phenotypic plasticity over the long term leads to genetic change. I have not seen that related to outbreeding depression which is genetic. For example the reduction in fitness in generations 2 onward in HBG .

I would be interested in seeing what I wrote on your question.

I am placing below an old PB article which might help with the concepts.

By Eric West

Everyone who owns a pond or lake with fish has wondered what makes fish behave in certain ways. That question is often debated on the Pond Boss Forum and by anglers everywhere. The answers are often based on observation and experience which are in truth subject to all the human biases. In short there is a lot of disagreement on what makes fish grow and thrive. This applies not only to the casual observer or anglers but also to the scientists who study the what and why of fish behavior and how that affects growth. Is it hunger, fear, reproductive urges, anger , competition or other biological processes that control what fish do ? What part is based on instinct (genetics) and does conditioning play at part? All of this comes into context when we try to establish a bluegill or other forage base for a pond which will have bass or other high end predators. Early growth usually is the primary factor determining recruitment to age-1and thus establishment of a viable forge base.

Well with so much disagreement out there we may not get an iron clad answer but hopefully there will be some though provoking ideas. The authors of a recent study The Effect of Vegetation Density on Juvenile Bluegill Diet and Growth in the Journal of Freshwater Ecology 2012, 1–11 by Daniel E. Shoup , Michael A. Nannini & David H. Wahl discuss a number of their thoughts which are set out below.

At the start the authors noted that the role of vegetation density and its influence on juvenile bluegill diet and growth remains unclear. Even after the many studies that exist. They acknowledge much disagreement in the literature about how vegetation density affects foraging results and thus growth of juvenile bluegill. Several studies have found reduced foraging return for bluegill when they forage in structurally complex habitats, whereas others have found that bluegill growth was unaffected or even increased when foraging in complex environments. However these studies were with predators present. That is not the case when a typical bluegill or other forage base is first started in a new or renovated pond without predators.

In the Shoup et al study eight experimental 0.4-ha ponds (one acre with a mean depth of 1m) were used to evaluate the effects of habitat complexity on growth of small bluegill. Each pond was stocked with 15 kg of young-of-year bluegills (30–50 mm total length, approximately 20,000 fish per pond) to produce a realistic density for small ponds. The ponds contained varying amounts of vegetation (plants) and no predators.

The result was - by the end of the experiment, bluegill from the low vegetation treatment ponds were significantly longer – twenty (20%) percent than bluegill from the high vegetation treatment ponds. These results suggest that bluegill chose to forage in a vegetated habitat even in the absence of predation risk, resulting in reduced growth. The question is why and what caused the bluegill to stay and forage in the plants even when that was not optimum for growth and energy usage (energetics).

The difference in growth rates of bluegill in low and high vegetation density could be caused by several mechanisms. First, the high vegetation treatment may have had
lower prey abundance, and bluegill were unable to find sufficient food. However, this hypothesis does not appear likely as fish in both treatments appeared to find and ingest ample food resources. Mean stomach fullness values (1.4–4.3 mg dry prey/g wet predator) were higher than those reported for bluegill feeding in Lake Opinicon, Ontario. Stomach fullness was also similar to fish fed a 2.2% daily ration (g wet prey/g wet bluegill), the midpoint of the range of daily rations (1–4%) reported.

Also considered was that differences in energetic value or digestibility of the prey types eaten by fish in the two treatments may have differed. Bluegill in the high vegetation
treatment ingested more gastropods (snails) and odonates (dragon- and damselflies) and fewer chironomids (midges) than did the fish in low vegetation biomass treatment. However, all the commonly eaten prey types in this study have similar energy densities. No published information exists on the digestibility of these prey types to bluegill, but rainbow trout are able to assimilate similar amounts of energy from all three prey types. Therefore, it seems unlikely that energetic or digestibility differences among prey types could account for the difference in growth rate of bluegill between treatments.

Another noted possibility was that the potential difference in search efficiency or handling costs could account for the differences in growth rates between treatments. Bluegill that forage in vegetation expend more energy searching for prey. A consequence observed was while fish from both treatments had similar stomach fullness (suggesting similar consumption rates), it is likely that bluegill in the high vegetation treatment expended more energy to ingest their food, leading to reduced growth.

Regardless of the mechanism, the study results demonstrate that increased vegetation densities reduced bluegill growth rates even in the absence of pelagic (open-water) predation risk.

Diets of bluegill indicated that they fed mostly in littoral or benthic habitats.
Percent of pelagic zooplankton in the diet, excluding benthic ostracods (seed shrimp), was 51% by weight in both treatments. Although pelagic (open-water) zooplankton are often considered the optimal prey type for bluegill, several studies have demonstrated that bluegill eat a high proportion of macroinvertebrates (larger invertebrates) , presumably because they use littoral habitat to avoid predation risk associated with pelagic habitat . Similar predator-mediated habitat and diet shifts have been found for other fish species. Because piscivorous fish were not present in the study ponds used by Sloup et al, it was surprising that bluegill foraged so heavily in the vegetated habitat. Either bluegill cannot accurately assess predation risk or some other mechanism causes bluegill to forage in vegetated habitat. The propensity of bluegill to forage in vegetated habitat could be genetically linked or related to phenotype . In both cases, fish would not be expected to alter their habitat use in response to the absence of predators over short time scales. Bluegill may also select habitat due to temperature preference rather than foraging return highlighting the potential for mechanisms other than predation risk. Previous laboratory and enclosure studies have found that Eurasian perch (Perca fluviatilis) and roach (Rutilus rutilus) occasionally chose to feed in vegetation rather than open water even when no predator was present.

The authors’ bottom line - additional research is needed to determine the pervasiveness of these behaviors and the underlying mechanisms. Right back where we started – no iron clad answers but plenty to think about.

One area I wish the study would have addressed in more detail is phenotypic plasticity. That is the ability of an individual or population to change due to environmental influences. Can environmental conditions during early development shape individuals’ phenotypes so they become more adaptive to the conditions they encounter? Were the long bluegill that fed in open water that way because longer fish can swim better in open water and were the shorter bluegill that way because being short allows them to maneuver around the weeds better? Plasticity has been shown to effect sunfish (Lepomis) shape, feeding and behavior in some cases.

Might I suggest you help out your bluegill or other forage base fishes with supplemental feeding as it has been shown to be an efficient way to increase growth rates.

Here is stuff on the topic in general

From Wikipedia, the free encyclopedia

In biology, outbreeding depression is when crosses between two genetically distant groups or populations results in a reduction of fitness.[1] The concept is in contrast to inbreeding depression, although the two effects can occur simultaneously.[2] Outbreeding depression is a risk that sometimes limits the potential for genetic rescue or augmentations. Therefore it is important to consider the potential for outbreeding depression when crossing populations of a fragmented species.[1] It is considered postzygotic response because outbreeding depression is noted usually in the performance of the progeny.[3] Some common cases of outbreeding depression have arisen from crosses between different species or populations that exhibit fixed chromosomal differences.[1]

Outbreeding depression manifests in two ways:

Generating intermediate genotypes that are less fit than either parental form. For example, selection in one population might favor a large body size, whereas in another population small body size might be more advantageous, while individuals with intermediate body sizes are comparatively disadvantaged in both populations. As another example, in the Tatra Mountains, the introduction of ibex from the Middle East resulted in hybrids which produced calves at the coldest time of the year.[4]
Breakdown of biochemical or physiological compatibility. Within isolated breeding populations, alleles are selected in the context of the local genetic background. Because the same alleles may have rather different effects in different genetic backgrounds, this can result in different locally coadapted gene complexes. Outcrossing between individuals with differently adapted gene complexes can result in disruption of this selective advantage, resulting in a loss of fitness.

Mechanisms for generating outbreeding depression
The different mechanisms of outbreeding depression can operate at the same time. However, determining which mechanism is likely to occur in a particular population can be very difficult.

There are three main mechanisms for generating outbreeding depression:

Fixed chromosomal differences resulting in the partial or complete sterility of F1 hybrids.[1]
Adaptive differentiation among populations
Population bottlenecks and genetic drift
Some mechanisms may not appear until two or more generations later (F2 or greater),[5] when recombination has undermined vitality of positive epistasis. Hybrid vigor in the first generation can, in some circumstances, be strong enough to mask the effects of outbreeding depression. An example of this is that plant breeders will make F1 hybrids from purebred strains, which will improve the uniformity and vigor of the offspring, however the F2 generation are not used for further breeding because of unpredictable phenotypes in their offspring. Unless there is strong selective pressure, outbreeding depression can increase in further generations as coadapted gene complexes are broken apart without the forging of new coadapted gene complexes to take their place. If the outcrossing is limited and populations are large enough, selective pressure acting on each generation can restore fitness. Unless the F1 hybrid generation is sterile or very low fitness, selection will act in each generation using the increased diversity to adapt to the environment.[6] This can lead to recovery in fitness to baseline, and sometimes even greater fitness than original parental types in that environment.[7] However, as the hybrid population will likely to go through a decline in fitness for a few generations, they will need to persist long enough to allow selection to act before they can rebound.[8]

The first mechanism has the greatest effects on fitness for polyploids, an intermediate effect on translocations, and a modest effect on centric fusions and inversions.[1] Generally this mechanism will be more prevalent in the first generation (F1) after the initial outcrossing when most individuals are made up of the intermediate phenotype. An extreme case of this type of outbreeding depression is the sterility and other fitness-reducing effects often seen in interspecific hybrids (such as mules), which involves not only different alleles of the same gene but even different orthologous genes.

Examples of the second mechanism include stickleback fish, which developed benthic and lymnetic forms when separated. When crosses occurred between the two forms, there were low spawning rates. However, when the same forms mated with each other and no crossing occurred between lakes, the spawning rates were normal. This pattern has also been studied in Drosophilia and leaf beetles, where the F1 progeny and later progeny resulted in intermediate fitness between the two parents. This circumstance is more likely to happen and occurs more quickly with selection than genetic drift.[1]

For the third mechanism, examples include poison dart frogs, anole lizards, and ciclid fish. Selection over genetic drift seems to be the dominant mechanism for outbreeding depression.[1]

1. Introduction

Small and isolated populations often have genetic problems that lead to reduced fitness (Fischer et al. 2000 Reed & Frankham 2003 Spielman et al. 2004 Reed 2005). These problems include inbreeding depression, resulting from reproduction among close relatives, and increased load from fixed mutations owing to genetic drift (Willi et al. 2005). Drift may also reduce mate availability by causing unequal sex ratio or small effective number of S alleles in plants with self-incompatibility systems (e.g. Wagenius et al. 2007). Genetic problems combine with disruption of ecological interactions and demographic and environmental stochasticity to increase the extinction risk of small and isolated populations (Newman & Pilson 1997 Saccheri et al. 1998). Genetic threats are especially severe because they do not remain constant for a given population size rather, they increase with time owing to continuous accumulation of genetic load and inbreeding, leading to a so-called mutational meltdown (Lynch & Gabriel 1990).

An obvious strategy to counteract such genetic erosion is to stock small populations with individuals from surrounding populations. Experimental studies have documented the rapid spread of immigrant genomes within inbred populations due to heterosis (Richards 2000 Ebert et al. 2002 Saccheri & Brakefield 2002) and even small amounts of artificial gene flow into natural populations can quickly reduce inbreeding depression and fitness reductions from fixed genetic load (Westemeier et al. 1998 Madsen et al. 1999 Stokstad 2005). Nevertheless, conservation biologists hold deep reservations about artificial gene flow, mostly because they fear breakdown of coadapted gene complexes and of local adaptation (Tallmon et al. 2004 Edmands 2007). These potentially negative consequences of gene flow and interpopulation outbreeding have been termed outbreeding depression (Templeton 1986). Hence, the main question that still needs to be answered is whether the net effect of heterosis and outbreeding depression is positive or negative.

Before genetic restoration can become accepted in conservation management, several fundamental issues must be resolved. First, how does population size affect the fitness consequences of gene flow in the long term? So far, long-term studies focusing on the effect of interpopulation outbreeding have not considered the size of focal populations (e.g. Edmands 1999 Fenster & Galloway 2000). These studies report outbreeding depression in some interpopulation hybrids. However, in small populations, heterosis can be so strong that outbreeding depression does not lead to a negative net effect (e.g. Heschel & Paige 1995 Willi & Fischer 2005). However, studies which account for population size have not included assessments of later generations.

Second, little is known about the impact of genetic distance between the donor and the recipient populations for genetic restoration in the long term. Increased genetic distance at neutral markers between populations may reflect differentiation in coadapted gene complexes arising from positive epistasis. According to theory, interpopulation outcrossing destroys those epistatic interactions. Although the breakdown may be masked by heterosis (dominance or overdominance) in the first (F1) generation, it becomes more obvious in the second (F2) and later generations when homozygosity increases (Lynch 1991). Therefore, it is important to test whether genetic distance to the source population influences the outcome of artificial gene flow over the long term. The best predictor of the success of interpopulation outcrossing may be the extent of genetic differentiation in potentially adaptive traits between target and partner populations, rather than the geographical or the genetic distances in neutral markers.

Finally and most importantly, for those managing small populations, one needs to know whether the net effect of artificial gene flow depends on environmental conditions. This seems possible because both heterosis and outbreeding depression might differ among environments, as occurs with inbreeding depression (Dudash 1990).

We tackled these questions by crossing plants of 14 populations of the locally endangered, tetraploid and self-incompatible Ranunculus reptans. This plant occurs in distinct populations of varying size at Lake Constance (Central Europe). In a previous study, we found that long-term small populations with low gene diversity suffered from reduced mean fitness, caused by three types of genetic problems. Pairs of plants from small populations were more often cross-incompatible than pairs of plants from large populations, indicating reduced S allele numbers in those populations. Furthermore, plants from small populations experienced higher fitness reductions due to inbreeding depression and increased fixed drift load, reducing female fertility (Willi et al. 2005). Finally, small populations benefit more than large populations in the first generation after interpopulation outbreeding (Willi & Fischer 2005).

In the present study, we crossed plants from populations of differing allozyme genetic distance and genetic distance in quantitative traits over two generations. We reared the offspring under near-natural benign and stressful conditions in an outdoor common garden experiment. Our focus was on heterosis and gene coadaptation, independent of possible adaptation to local conditions. We addressed the following questions: (i) do F1 and F2 offspring of interpopulation crosses show outbreeding depression or outbreeding vigour? (ii) do small and inbred populations enjoy a higher fitness benefit from interpopulation outbreeding? (iii) does F1 and F2 interpopulation hybrid performance depend on genetic distance between target and partner population? (iv) is genetic distance in quantitative traits a better predictor of interpopulation hybrid performance than marker distance? and (v) how does competition modulate these relationships?

Do the effect of inbreeding depression disappear after the first generation of outbreeding? - Biology

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1 Department of Biology, University of California, Riverside, California 92521 [email protected]

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The two principal theories of the causal mechanism for inbreeding depression are the partial dominance hypothesis and the overdominance hypothesis. According to the first hypothesis, inbreeding increases the frequency of homozygous combinations of deleterious recessive alleles thereby decreasing fitness, whereas the overdominance hypothesis posits that inbreeding increases homozygosity and thus reduces the frequency of the superior heterozygotes. These two hypotheses make different predictions on the effect of crossing inbred lines: the overdominance hypothesis predicts that trait means will be restored to the outbred means, whereas the partial dominance hypothesis predicts that trait means will exceed those of the outbred population. I tested these predictions using seven inbred lines of the sand cricket, Gryllus firmus . Fourteen generations of brother-sister mating resulted in an inbreeding depression of 20–34% in four traits: nymphal weights at ages 14 days, 21 days, 28 days, and early fecundity. An incomplete diallel cross of these lines showed genetic variation among lines and an increase in all trait means above the outbred means, with three being significantly higher. These results provide support for the partial dominance hypothesis and are inconsistent with the overdominance hypothesis.

Corresponding Editor: P. Jarne


Received: 6 February 2001 Accepted: 1 November 2001 Published: 1 April 2002

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Local adaptation and loss of local adaptation via outbreeding depression

Only one of the three studied populations (ECO) exhibited survival rates supportive of the hypotheses of local adaptation and the loss of local adaptation in outbred cross types. Although ECO did not meet the ‘home versus away’ criterion of local adaptation ( Kawecki and Ebert 2004 ), it is more important to conservation research that ECO performed better than GRV and STW within ECO River (‘local versus foreign’ criterion) than how ECO performed in GRV and STW Rivers (‘home versus away’ criterion). In addition, a recent review of local adaptation in salmonids ( Fraser et al. 2011 ) concluded that lack of detection of a fitness trade-off (‘home versus away’ criterion) may not be surprising at this small spatial scale (34–50 km). Alternatively, the positive slopes observed between the percentage of ‘local’ genes and survival (‘local versus foreign’ criterion) could be explained by parental effects or other genetic differences among populations, such as fixed beneficial mutations (see Kawecki and Ebert 2004 ). However, upon inclusion of parental effects in mixed-effects models, these relationships remained, albeit with reduced significance (Appendix S4). Also, similar analyses that included parental effects on other fitness-related traits, i.e., juvenile size, condition, and growth, also revealed few relationships ( Houde 2009 ). The inability to detect local adaptation in all three rivers was notable for Atlantic salmon. It suggests that the geographic scale of local adaptation in our study region (the iBoF) may be larger than the scale of ‘river’ (as concluded by Fraser et al. 2007 ), the scale at which traditional conservation strategies have most often been applied in this species ( Garcia de Leaniz et al. 2007 Fraser 2008 ).

Heterosis and outbreeding depression

Outbreeding effects depended on the populations that were mixed and did not correspond with neutral genetic or gene expression differences between populations or the inbreeding histories of the three study populations ( Fraser et al. 2007 Tymchuk et al. 2010 Appendix S1). Heterosis is predicted to occur in outbred cross types between closely related populations, such as ECO and GRV populations, that have a history of inbreeding ( Vergeer et al. 2004 : Pertoldi et al. 2007 ). Yet, we were unable to detect heterosis in outbred cross types having ECO or GRV population ancestry with the exception of F1 E.S juveniles in the STW River (but not in the ECO River). Furthermore, outbreeding depression is not predicted to occur in outbred cross types between closely related populations ( Edmands and Timmerman 2003 Gilk et al. 2004 but see McClelland and Naish 2007 ). However, we may not have been able to detect small changes resulting from heterosis or outbreeding depression because of our limited statistical power in having three sites that allowed only three pure-outbred cross type comparisons per river.

We were unable to detect outbreeding depression by the breakdown of coadapted gene complexes in the second-generation outbred cross types. However, the potential for local adaptation and its loss via outbreeding depression in one of our study’s three closely related populations is of concern from a conservation perspective. Furthermore, simple additive inheritance models on outbred recapture rates fit only ECO sites pooled in ECO River with most outbred recapture rates not being different from the mean of the sites or rivers, although there were some deviations from the parental midpoints (i.e., F1 E.S heterosis and F1 G.S outbreeding depression). F1 heterosis and outbreeding depression may be attributed to dominance and epistatic interactions which are less predictable than additive effects ( Kawecki and Ebert 2004 Edmands 2007 ).

Collectively, our results are consistent with the hypothesis that outbreeding outcomes may be highly variable at small genetic distances ( Edmands and Timmerman 2003 ) and that the genetic interaction between population pairs may be difficult to predict because of random mutation and fixation processes ( Lynch 2000 see Bougas et al. 2010 ). Hence, the reality for endangered species conservation is that outbreeding effects may have to be evaluated on a case-by-case basis.

Inbreeding depression

The unpredictability of outbreeding effects was mirrored by a similar lack of consistency in the ability to detect inbreeding depression that had been predicted to exist in ECO and GRV because of presumed large inbreeding coefficients attributable to population bottlenecks ( Wang et al. 2002 Tymchuk et al. 2010 Appendix S1). Indeed, similar inbreeding coefficients have been associated with severe inbreeding depression in salmon in both captivity ( Kincaid 1983 ) and in the wild ( Ryman 1970 Thrower and Hard 2009 ). Yet, ECO and GRV did not exhibit inbreeding depression. Possible explanations for these results are that these populations may have naturally mixed with relatives at a sufficiently slow rate such that deleterious alleles may have been purged by selection ( Templeton and Read 1984 Allendorf and Luikart 2007 ). Alternatively, deleterious alleles may have become fixed by genetic drift, resulting in few differences between pure and inbred cross types ( Keller and Waller 2002 Hedrick and Fredrickson 2010 ). The former explanation seems unlikely because there was an instance of heterosis when ECO was mixed with STW. In addition, the pooling of individuals with different inbreeding coefficients (F = 1/4 and 1/8) may not have generated a high enough level of inbreeding to detect inbreeding depression, at least for a salmonid fish ( Gjerde et al. 1983 Pante et al. 2001 ). Furthermore, similar to outbreeding depression, we may not have been able to detect small changes owing to inbreeding depression because of the limited power in having three sites per river.

Relative risks of outbreeding and inbreeding

According to one analysis using Student’s t-tests, our results suggest that the risks from one generation of inbreeding do not differ significantly from the risks posed by one or two generations of outbreeding within endangered Atlantic salmon populations. Yet another test, using linear regressions, which had greater mean statistical power (linear regression power = 0.300 ± 0.309 versus Student’s t-tests power = 0.247 ± 0.249), indicated outbreeding depression via the loss of potential local adaptation in ECO. Furthermore, while limited in statistical power because of having just three comparisons, there was a trend in Student’s t-tests (P = 0.136, power = 0.638) for inbred ECO to consistently perform more poorly than pure ECO at all three study sites within the ECO river. Our precautionary interpretation of these trends is that both outbreeding and inbreeding might be detrimental to survival for ECO Atlantic salmon.

More generally, over more generations than studied here, either process might affect the persistence of small populations (see Frankham 2005 ). For example, the decreased recapture rate of the second-generation (F2) outbred cross type (F2 ES.ES) relative to first-generation (F1) cross type (F1 E.S) juveniles could suggest a negative recapture-rate trend for successive outbred generations (see Dobzhansky 1950 ). Even reduced F1 fitness coupled with fitness improvements in successive outbred generations (e.g., F1 G.S versus F2 GS.GS juveniles) may also be concerning. It may take several generations of outbreeding for fitness to recover to the same level as pure cross types because of natural selection for beneficial gene combinations ( Edmands 2007 ), and small fitness declines in the earlier outbred generations could lead to population extirpation before there is time to recover in later outbred generations ( Hutchings 1991 ).

One caveat of our work is that we could not assess the entire salmon life cycle because of the current very high mortality at sea experienced by iBoF salmon and other logistic issues ( Fraser et al. 2007 DFO 2008 ). On the other hand, it is likely that these populations share similar adult life histories ( COSEWIC 2006 ). Consequently, phenotypic changes detrimental to survival because of a loss of local adaptation in outbred cross types may be more likely to occur at juvenile than adult life stages ( Taylor 1991 ).

Conservation implications

Our study is insightful given the conundrum of either continuing to accrue inbreeding or to risk outbreeding depression in the management of small fragmented populations and endangered species. In the case of endangered salmon, our study revealed that different management recommendations may be necessary even for closely related populations owing to varying inbreeding and outbreeding risks. For one population (ECO), both inbreeding and outbreeding may be detrimental to survival and it is recommended that pure noninbred ECO broodstock be maintained for conservation purposes. For the other populations, GRV and STW, inbreeding for one generation may not be detrimental to survival, at least during the freshwater phase of their life cycle, and interbreeding GRV and STW may be acceptable in their long-term management because the survival decline in the first outbred generation did not continue into the second generation (both backcrosses and F2 cross types). Such specific recommendations would not have been possible without experimentation conducted in the wild. The relative costs of inbreeding and outbreeding in the conservation and management of endangered species may, therefore, have to be tested on a case-by-case basis and interpreted very judiciously.


In general we found that the life time number of mature offspring under natural conditions depends on the inbredness of the mother, whereas the size, egg load and symmetry of a female are independent of her own homozygosity. However, female size and the tree in or time at which females oviposit have strong effects on their number of eggs. It is thus not surprising that homozygosity could only account for a small fraction of the variation.

It can be difficult to separate fitness effects of mother and offspring. In this system the inbreeding coefficient of offspring will be very weakly correlated with that of their mothers' (if at all) and homozygosity does not affect the number of eggs females can lay. Therefore, the fact that mothers' homozygosity affects the number of offspring that reach maturity, suggest that maternal effects and not offspring genotype or number of eggs, plays a role in determining if an egg will develop into a mature wasp. Maternal factors could be the quality of the egg, or the gall from which the developing wasp larvae feed. The potential role of gall quality was recently illustrated by two studies on alternative male morphs in non-pollinating fig wasps that showed that gall size (Moore et al., 2004) and quality (Pereira et al., 2007), determine males' size and morphology.

We could not confirm the correlation between F and the proportion of homozygous loci by using pedigrees. However, we showed that P. awekei's mixed mating system should result in a high variance in F, which means that the multi-locus homozygosity is a good proxy for F (Pemberton, 2004).

We found evidence that outbreeding depression occurs in this fig wasp, but possibly also inbreeding depression. Despite purging of partially recessive deleterious alleles in haploid males, it is not unusual to observe some inbreeding depression in haplodiploids (Antolin, 1999 Henter, 2003). Inbreeding depression could be the result of overdominance or female limited expression (Antolin, 1999). On the other hand, if model 2 is correct and there is no inbreeding depression, this could be explained by the fact that the wasps regularly sibmate (78% of the time) and should be adapted to high levels of inbreeding. Several authors have suggested that this may be an important factor leading to no inbreeding depression (Biémont and Bouletreau, 1980 Peer and Taborsky, 2005 Dolgin et al., 2007). However, more data will be required to distinguish between these two statistical models.

In line with Clarke et al. (1992) who studied another haplodiploid, the honeybee, we found no increase in fluctuating asymmetry as the inbreeding coefficient increased. This contrasts with a diploid species where fluctuating asymmetry was found (Neff, 2004), giving further support to the suggestion of Clarke et al. (1992) that the level of inbreeding may not affect developmental stability in haplodiploids.

The break up of co-adapted gene complexes is the most common reason given for outbreeding depression (Lynch and Walsh, 1998). For co-adapted gene complexes to build up, lineages need to be isolated from one another. Geographical population subdivision is a common cause of isolation and outbreeding depression is frequently found when individuals from geographically distinct populations are crossed (Pinto et al., 1991 de Meester, 1993 Sorati et al., 1996 Aspi, 2000 Velando et al., 2006 Dolgin et al., 2007). Inbreeding can also result in the isolation of lineages from one another and this may allow for the build up of different gene complexes in one population (Luna and Hawkins, 2004 Peer and Taborsky, 2005). The within population outbreeding depression recorded in this study, as well as that of Peer and Taborsky (2005), supports this view, whereas Luna and Hawkins (2004) did not find within population outbreeding depression in the wasp Nasonia vitripennis. It is also important to remember that co-adaptation of gene complexes can occur in one generation from extant variation and does not necessarily need many generations of new mutations and co-evolution (Templeton, 1979). The fact that our population's outbred component should contain many females that have been outbred for only one generation, supports the growing number of papers that show outbreeding depression in the F1 generation (Peer and Taborsky, 2005 Dolgin et al., 2007).

An alternative explanation for outbreeding depression could be if wasps harbour one or multiple strains of Wolbachia that may induce cytoplasm incompatibility (CI Cook and Butcher, 1999). CI causes a slight reduction in fecundity of matings between infected males and uninfected females (Cook and Butcher, 1999). In the pollinating fig wasp, Pleistodontes imperialis, Haine et al. (2006) found different strains of Wolbachia in three cryptic species. This suggests that Wolbachia may lead to reproductive isolation in pollinating fig wasps (Haine et al., 2006). Platyscapa awekei are infected with Wolbachia (OFC Greyvenstein, CJ Erasmus and JMG, unpublished data) and CI may thus be the cause of outbreeding depression observed in P. awekei in this study. If CI occurs in this wasp, note that the male's dispersal behaviour will allow CI strains to spread faster by spoiling the reproductive success of uninfected females.

Another common explanation of outbreeding depression is a mismatch between environment and the genotypes. In these tiny wasps, however, females move over vast distances so that populations 500 km apart are genetically indistinguishable (Jansen van Vuuren et al., 2006 CJ Erasmus and JMG, unpublished data), even when correcting for the highly expected heterozygosity (Hedrick, 2005). In addition, fig trees regulate the temperature inside figs to reduce the environmental variability the wasps experience (Patiño et al., 1994). This means that local adaptation is very unlikely to explain outbreeding depression in P. awekei.

Our finding of outbreeding depression is surprising—this is a wasp that actively disperses to engage in outbreeding, which will result in the lowest fitness. According to the second statistical model where there is only outbreeding depression, both outbreeding depression, as well as the kin selective advantages, should select for a completely inbred population. This prediction is clearly at odds with the data. If the first statistical model is correct, P. awekei females will attain the highest fitness during the third, fourth and fifth cycles of sibmating. In this scenario a female can either sibmate and have offspring that are more inbred than she is, or she can reset the clock by mating with an unrelated male. However, a certain amount of ‘frustration’ is inevitable in such a mixed mating system as lineages will involuntarily wander into areas where the fitness is below the maximum. Fitness can be improved if females can respond facultatively to their own inbredness, a possibility that is not inconceivable since even borage (Drayner, 1956) and field beans (Crowe, 1971) manage to do so.

When we parameterize an unpublished model (RM Nelson and JMG) designed for a system where fitness depends on the inbredness, it predicts an optimal sibmating rate that is higher than that observed. Thus for both statistical models, these wasps outbreed too much. The suggestion of Moore et al. (2006) that male dispersal serves to reduce competition between related brothers seems to be a more plausible explanation than that male dispersal evolved to achieve an optimal, yet ‘frustrated’, mating system.

Pedigree Analysis and How Breeding Decisions Affect Genes

To some breeders, determining which traits will appear in the offspring of a mating is like rolling the dice – a combination of luck and chance. For others, producing certain traits involves more skill than luck – the result of careful study and planning. As breeders, you must understand how matings manipulate genes within your breeding stock to produce the kinds of offspring you desire.

Photo by Dustin Hartje

When evaluating your breeding program, remember that most traits you’re seeking cannot be changed, fixed or created in a single generation. The more information you can obtain on how certain traits have been transmitted by your animal’s ancestors, the better you can prioritize your breeding goals. Tens of thousands of genes interact to produce a single individual. All individuals inherit pairs of chromosomes one from the mother and one from the father. On the chromosomes are genes so all genes come in pairs. If both genes in a pair are the same gene (for instance, “aa” or “AA”) the gene pair is called homozygous. If the two genes in a gene pair are unlike (for instance, “Aa”) the gene pair is called heterozygous. Fortunately, the gene pairs that make a cat a cat and not a dog are always homozygous. Similarly, the gene pairs that make a certain breed always breed true are also homozygous. Therefore, a large proportion of homozygous non-variable pairs – those that give a breed its specific standard – exist within each breed. It is the variable gene pairs, like those that control color, size and angulation that produce variations within a breed.

There are ways to measure the genetic diversity of a population. One method is to measure the average inbreeding coefficient (or Wright’s coefficient) for a breed. The inbreeding coefficient is a measurement of the genetic relatedness of the sire and dam. If an ancestor appears on both the sire and dam’s side of the pedigree, it increases the inbreeding coefficient. The inbreeding coefficient gives a measurement of the total percentage of variable gene pairs that are expected to be homozygous due to inheritance from ancestors common to the sire and dam. It also gives the chance that any single gene pair can be homozygous due to inheritance from ancestors common to the sire and dam. It also gives the chance that any single gene pair can be homozygous.

The types of matings that you choose for your breeding animals will manipulate their genes in the offspring, affecting their expression. Linebreeding is breeding individuals more closely related (a higher inbreeding coefficient) than the average of the breed. Outbreeding involves breeding individuals less related than the average of the breed. Linebreeding tends to increase homozygosity. Outbreeding tends to increase heterozygosity. Linebreeding and inbreeding can expose deleterious recessive genes through pairing-up, while outbreeding can hide these recessives, while propagating them in the carrier state.

Most outbreeding tends to produce more variation within a litter. An exception would be if the parents are so dissimilar that they create a uniformity of heterozygosity. This is what usually occurs in a mismating between two breeds, or a hybrid, like a Cockapoo. The resultant litter tends to be uniform, but demonstrates “half-way points” between dissimilar traits of the parents. Such litters may be phenotypically uniform, but will rarely breed true due to a mix of dissimilar genes.

One reason to outbreed would be to bring in new traits that your breeding stock does not possess. While the parents may be genetically dissimilar, you should choose a mate that corrects your breeding animal’s faults but complements its good traits. It is not unusual to produce an excellent quality individual from an outbred litter. The abundance of genetic variability can place all the right pieces in one individual. Many top-winning show animals are outbred. Consequently, however, they may have low inbreeding coefficients and may lack the ability to uniformly pass on their good traits to their offspring. After outbreeding, breeders may want to breed back to individuals related to their original stock, to attempt to solidify newly acquired traits.

Linebreeding attempts to concentrate the genes of specific ancestors through their appearance multiple times in a pedigree. It is better for linebred ancestors to appear on both the sire’s and dam’s sides of the pedigree. That way their genes have a better chance of pairing back up in the resultant offspring. Genes from common ancestors have a greater chance of expression with paired with each other than when paired with genes from other individuals, which may mask or alter their effects.

Linebreeding on an individual may not reproduce a outbred ancestor. If an ancestor is outbred and generally heterozygous (Aa), increasing homozygosity will produce more AA and aa. The way to reproduce ab outbred ancestor is to mate two individuals that mimic the appearance and pedigree of the ancestor’s parents.

Inbreeding significantly increases homozygosity, and increases the expression of both desirable and deleterious recessive genes through pairing up. If a recessive gene (a) is rare in the population, it will almost always be masked by a dominant gene (A). Through inbreeding, a rare recessive gene (a) can be passed from a heterozygous (Aa) common ancestor through both the sire and dam, creating a homozygous recessive (aa) offspring.

The total inbreeding coefficient is the sum of the inbreeding from the close relatives (first cousin mating), and the background inbreeding from common ancestors deep in the pedigree. Such founding ancestors established the pedigree base for the breed.
The total inbreeding coefficient is the sum of the inbreeding
from the close relatives (first cousin mating), and the
background inbreeding from common ancestors deep in the
pedigree. Such founding ancestors established the pedigree
base for the breed.

Knowledge of the degree of inbreeding in a pedigree does not necessarily help you unless you know whose genes are being concentrated. The relationship coefficient, which can also be approximated by what is called the percent blood coefficient, represents the probable genetic likeness between the individual whose pedigree is being studied, and a particular ancestor.

We know that a parent passes on an average of 50% of its genes, while a grandparent passes on 25%, a great-grandparent 12.5%, and so on. For every time the ancestor appears in the pedigree, its percentage of passed on genes can be added up and its “percentage of blood” estimated. In many breeds, an influential individual may not appear until later generations, but then will appear so many times that it necessarily contributes a large proportion of genes to the pedigree.

The average inbreeding coefficient of a breed is a measurement of its genetic diversity. When computing inbreeding coefficients, you have to look at a deep pedigree to get accurate numbers. An inbreeding coefficient based on 10 generation pedigrees is standardly used, but requires a computerized pedigree database to compute.

The average inbreeding coefficient for a breed will be based on the age and genetic background of the breed. A mating with an inbreeding coefficient of 14 percent based on a ten generation pedigree, would be considered moderate inbreeding for a Labrador Retriever (a popular breed with a low average inbreeding coefficient), but would be considered outbred for an Irish Water Spaniel (a rare breed with a higher average inbreeding coefficient).

Most breeds start from a small founding population, and consequently have a high average inbreeding coefficient. If a breed is healthy and prolific, the breadth of the gene pool increases, and the average inbreeding coefficient can go down over time. Some dog breeds were established on a working phenotype, and not on appearance. These breeds usually start with low inbreeding coefficients due to the dissimilar backgrounds of the founders. As certain individuals are linebred on to create a uniform physical phenotype, the average inbreeding coefficient can increase.

There is no specific level or percentage of inbreeding that causes impaired health or vigor. If there is no diversity (non-variable gene pairs for a breed) but the homozygote is not detrimental, there is no effect on breed health. The characteristics that make a breed reproduce true to its standard are base on non-variable gene pairs. There are pure-bred populations where smaller litter sizes, shorter life expectancies, increased immune-mediated disease, and breed-related genetic disease are plaguing the population. In these instances, prolific ancestors have passed on detrimental recessive genes that have increased in frequency and homozygosity. With this type of documented inbreeding depression, it is possible that an outbreeding scheme could stabilize the population. However, it is also probable that the breed will not thrive without an influx of new genes either from a distantly related (imported) population, or crossbreeding.

Fortunately, most breeds do not find themselves in the position of this amount of limited diversity and inbreeding depression. However, the perceived problem of a limited gene pool has caused some breeders to advocate outbreeding of all individuals. Studies in genetic conservation and rear breeds have shown that his practice contributes to the loss of genetic diversity. By uniformly crossing all “lines” in a breed, you eliminate the differences between them, and therefore the diversity between individuals. Eventually, there will not be any “unrelated line” to be found. Everyone will have a mixture of everyone else’s genes. The practice in livestock breeding has significantly reduced diversity, and caused the reduced diversity, loss of unique rare breeds.

A basic tenet of population genetics is that gene frequencies do not change from generation to generation. This will occur regardless of the homozygosity or heterozygosity of the parents, or whether the mating is an outbreeding, linebreeding, or inbreeding. This is the nature of genetic recombination. Selection, and not the types of matings used affect gene frequencies and breed genetic diversity.

If two parents are both heterozygous (both Aa) for a gene pair, on the average, they would produce 25% AA, 50% Aa, and 25% aa. (These are the averages when many litters are combined. In reality, any variety of pairing up can occur in a single litter.) If a prolific male comes out of this litter, and he is homozygous aa, then the frequency of the “a” gene will increase in the population, and the frequency of the “A” gene will decrease. This is known as the popular sire syndrome. Of course, each individual has thousands of genes that vary in the breed, and everyone carries some deleterious recessive genes. The overuse of individual breeding animals contributes the most to decreased diversity (population bottlenecks), and the increased spread of deleterious recessive genes (the founders effect). Again, it is selection (use of this stud to the exception of others), and not the types of matings he is involved in that alters gene frequencies. Breeders should select the best individuals from all lines, so as to not create new genetic bottlenecks.

Decisions to linebreed, inbreed or outbreed should be made based on the knowledge of an individuals traits and those of its ancestors. Inbreeding will quickly identify the good and bad recessive genes the parents share, based on their expression in the offspring. Unless you have prior knowledge of what the offspring of milder linebreedings on the common ancestors were like, you may be exposing your litters (and buyers) to extraordinary risk of genetic defects. In your matings, the inbreeding coefficient should only increase because you are specifically linebreeding (increasing the percentage of blood) to selected ancestors.

Don’t set too many goals in each generation, or your selective pressure for each goal will necessarily become weaker. Genetically complex or dominant traits should be addressed early in a long-range breeding plan, as they may take several generations to fix. Traits with major dominant genes become fixed more slowly, as the heterozygous (Aa) individuals in a breed will not be readily differentiated from the homozygous-dominate (AA) individuals. Desirable recessive traits can be fixed in one generation because individuals that show such characteristics are homozygous for the recessive genes. Individuals that pass on desirable traits for numerous matings and generations should be preferentially selected for breeding stock. This prepotency is due to homozygosity of dominate (AA) and recessive (aa) genes. However, these individuals should not be overused, to avoid the popular sire syndrome.

Breeders should plan their matings based on selecting toward a breed standard, based on the ideal temperament, performance, and conformation, and should select against the significant breed related health issues. Using progeny and sib-based information to select for desirable traits and against detrimental traits will allow greater control.


Most fitness effects are thought to be due to increased homozygosity for recessive deleterious alleles (Charesworth Charesworth, 1987). Based on studies of Drosophila using chromosome balancer techniques (Sved Ayala, 1970), approximately half cases of inbreeding depression are due to rare recessive lethal or sub-lethal mutations while the rest can be attributed to a large number of mildly deleterious mutations (Charlesworth Charlesworth, 1987). Other researchers have suggested that inbreeding depression, to some extent, may also result from synergistic or epistatic interactions (Charlesworth, 1998). In this study, inbreeding depression due to recessive mutations, not due to synergistic or epistatic interactions, was focused on.

In this study, there was no significant difference in all parameters except for the body length of adults between outbreeding and inbreeding populations. Adults of the outbreeding population were longer than those of the inbreeding population in G1. Bijlsma et al. (2000) has mentioned that inbreeding can affect most fitness components and lead to reduced viability, lower fecundity, increased sterility, decreased mating success, slower development, and increased susceptibility to environmental stress, consequently resulting in a significant decrease in individual fitness. In this study, a significant difference was only seen in body lengths of adults. There were slight differences in other three parameters. However, such differences were not statistically significant. Keller and Waller (2002) have suggested that if populations remain small and isolated for many generations, they face two genetic threats. First, as alleles are randomly fixed or lost from the population by genetic drift, levels of quantitative genetic variation necessary for adaptive evolution will erode (Lande, 1995). Second, deleterious mutations tend to accumulate because selection is less effective in small populations (Lynch et al., 1995). This could eventually lead to a mutational meltdown for populations with an effective size (Ne) 100. Both processes tend to be gradual. Thus, they do not threaten populations in the shortterm. In this study, inbreeding effects were assessed for only two generations. This might not be sufficient to see effects of inbreeding depression. For extensive data analysis, it is necessary to use more replicates or samples. Five or six replicates were used for the statistical analysis of inbreeding effects in the present study. Other researchers have observed more generations using more replicates or samples to evaluate differences between outbreeding and inbreeding populations. For example, Bijlsma et al. (1999) have used nine generations of Drosophila to see inbreeding effects. Furthermore, they have observed the maintenance of these effects for approximately 50 generations using 50 replicates for each condition.

It has been suggested that inbreeding problems for small populations are far less than expected because they can reduce or eliminate inbreeding depression by evolutionary changes, termed purging at the contributing loci (Ballou, 1997 Battett Charlesworth, 1991 Hedrick, 1994 Templeton Read, 1984). However, several studies have suggested that inbreeding depression strongly depends on environmental conditions and that the inbreeding load becomes more visible under more stressful conditions (Bijlsma et al., 1997 Dahlgaard et al., 1995 Miller, 1994). For example, Miller (1994) has demonstrated that D. melanogaster homozygous for the second chromosome shows a significant increase in inbreeding depression under lead stress, whereas the same homozygote in an unleaded environment does not. In this study, water supply was restricted after D. melanogaster laid eggs to provide a stressful environmental condition. However, this stressful condition probably had no significant effects on inbreeding except on the mean body length of adults.

This study showed a statistically significant difference in the body length of D. melanogaster adults as a proof of negative inbreeding effect. However, the other parameters did not show a significant difference between outbred and inbred populations due to genetic purging. This study demonstrated one additional experimental case related to inbreeding depression in artificial bottleneck populations. For future studies, it might be necessary to use more replicates (or samples) and more generations to see effects of inbreeding depression in bottlenecked populations.

Do the effect of inbreeding depression disappear after the first generation of outbreeding? - Biology

The following is the established format for referencing this article:
Brook, B. W., D. W. Tonkyn, J. J. O'Grady, and R. Frankham. 2002. Contribution of inbreeding to extinction risk in threatened species. Conservation Ecology 6(1): 16. [online] URL:

Contribution of Inbreeding to Extinction Risk in Threatened Species

Barry W. Brook 1 , David W. Tonkyn 2 , Julian J. O'Grady 3 , and Richard Frankham 3
1 Northern Territory University 2 Clemson University 3 Macquarie University

Wild populations face threats both from deterministic factors, e.g., habitat loss, overexploitation, pollution, and introduced species, and from stochastic events of a demographic, genetic, and environmental nature, including catastrophes. Inbreeding reduces reproductive fitness in naturally outbreeding species, but its role in extinctions of wild populations is controversial. To evaluate critically the role of inbreeding in extinction, we conducted realistic population viability analyses of 20 threatened species, with and without inbreeding depression, using initial population sizes of 50, 250, and 1000. Inbreeding markedly decreased median times to extinction by 28.5, 30.5, and 25% for initial populations of 50, 250, and 1000, respectively, and the impacts were similar across major taxa. The major variable explaining differences among species was initial population growth rate, whereas the impact of inbreeding was least in species with negative growth rates. These results demonstrate that the prospects for survival of threatened species will usually be seriously overestimated if genetic factors are disregarded, and that inappropriate recovery plans may be instituted if inbreeding depression is ignored.

KEY WORDS: endangered species, inbreeding depression, life histories, median time to extinction, population viability analysis, purging.

Published: June 27, 2002

Species in natural habitats face threats both from deterministic factors such as habitat loss, overexploitation, pollution, and introduced species, and from stochastic events associated with small population size such events may be of a demographic, genetic, or environmental nature, including catastrophes (World Conservation Monitoring Centre 1992).

Genetic stochasticity encompasses inbreeding depression, loss of genetic diversity, and mutational accumulation (Frankham et al. 2002). Inbreeding is the most immediate and potentially damaging of these (Frankham 1995a). Essentially, all well-studied naturally outbreeding species show depressed reproductive fitness in inbred individuals this phenomenon is known as inbreeding depression (Falconer and Mackay 1996, Lynch and Walsh 1998, Hedrick and Kalinowski 2000). This has been demonstrated in the laboratory (see Frankham 1995b), in zoos (Ralls et al. 1988), and in the wild (see Crnokrak and Roff 1999). Although some scientists have been skeptical about the occurrence of inbreeding depression in wild populations, compelling evidence for it now exists. Of 157 valid data sets across 34 taxa reviewed by Crnokrak and Roff (1999), 90% showed differences indicating that inbreeding was deleterious to reproductive fitness (Frankham 2000).

There is controversy about the contribution of inbreeding depression to the extinction risk for populations in nature. Whereas it is generally acknowledged that any depressive effect on survival, such as inbreeding, will tend to reduce population growth rates, it is not generally accepted that inbreeding itself translates into elevated extinction risks. For instance, Lande (1988) and others (e.g., Caro and Laurenson 1994, Caughley 1994, Dobson 1999) have argued that inbreeding plays an extremely minor role in extinctions, because demographic and environmental stochasticity, as well as catastrophes, will drive small populations to extinction before genetic factors become important. Although Lande (1995) now believes that genetic factors do contribute to extinction, he is referring to accumulations of new deleterious mutations rather than to inbreeding depression. However, inbreeding depression has been linked to population declines and extinctions in both captivity (Frankham 1995b) and the wild (Vrijenhoek 1994, Newman and Pilson 1997, Saccheri et al. 1998, Westemeier et al. 1998, Madsen et al. 1999). All these studies discussed individual cases, but none provided comprehensive evidence covering a wide range of threatened species or gave a clear indication of when inbreeding is important and when it is not.

Levels of inbreeding (F ) are inversely related to effective population size (N e ) and increase with generations (t), as follows (Falconer and Mackay 1996):

Reductions in fecundity and survival are related to F (Falconer and Mackay 1996, Lynch and Walsh 1998). Consequently, inbreeding is expected to have its greatest impact when populations are small and the number of generations is large. The effects of other stochastic factors are also expected to show similar patterns.

The magnitude of inbreeding depression may be reduced by selective purging of recessive deleterious alleles by natural selection, although the relative importance of purging is also controversial (see Byers and Waller 1999, Miller and Hedrick 2001). Furthermore, there is still some disagreement with regard to the differential effects of purging in very small vs. large populations (Frankham et al. 2001). Purging has little impact in very small populations, e.g., with regular selfing or full-sib mating, but has clear effects in moderate to large populations (D. H. Reed, D. A. Briscoe, and R. Frankham, unpublished data).

Interactions are expected between the impacts of inbreeding and both deterministic factors and "nongenetic" stochastic factors. Human-associated threats such as habitat loss, overexploitation, pollution, and introduced species (World Conservation Monitoring Centre 1992) reduce population sizes and increase inbreeding, which in turn reduces individual survival and fecundity and therefore population sizes, creating the potential for an extinction "vortex" (Gilpin and Soulé 1986). Fluctuations in population size resulting from demographic and environmental stochasticity and catastrophes reduce N e , increase F , and therefore increase the risk of extinction (van Noordwijk 1994, Tanaka 2000).

Studies of the effects of inbreeding on extinction risk in natural populations are hampered by difficulties in separating the genetic and nongenetic components. In addition, constraints on time and resources have forced past studies to concentrate on only a few high-profile species. As a result, stochastic computer projections offer the only means of comprehensively investigating the role of inbreeding in extinction. They make it possible to investigate many species, can be performed relatively quickly, and allow for the inclusion or exclusion of inbreeding in concert with demographic and environmental stochasticity and catastrophes this is impossible in field experiments.

Population viability analysis (PVA) is widely used to predict the fate of threatened populations by projecting life histories forward using stochastic computer simulations (see Akçakaya and Sjögren-Gulve 2000, Menges 2000, Beissinger and McCullough 2002). Critically, PVA has been shown to produce unbiased predictions, making it an ideal research tool for this purpose (Brook et al. 2000). Four studies have used PVA to investigate the effects of inbreeding depression on population growth and/or extinction risk (Burgman and Lamont 1992, Dobson et al. 1992, Mills and Smouse 1994, Oostermeijer 2000). However, these studies focused on specific or hypothetical cases, were often projected for only a few generations, and failed to consider the impact of purging. As a result, their overall message was unclear. For example, Burgman and Lamont (1992) found that inbreeding depression had very little impact on the viability of the plant Banksia cuneata, whereas Oostermeijer (2000) found that it had a strong impact on Gentiana pneumonanthe. Dobson et al. (1992) predicted that inbreeding depression would increase the extinction risk of rhinoceros populations and that its impact depended on population size. Mills and Smouse (1994) showed that inbreeding would have an impact on generalized animal life histories, especially those with slow initial population growth.

The objective of this study was to determine the contribution of inbreeding to extinction risk for a broad range of threatened taxa. We used realistic PVA models that included the effects of purging to project the population dynamics for 20 actual threatened species covering a range of life history types, both with and without inbreeding depression. We also investigated the impact of different initial population sizes and different population growth rates.

Population viability analyses

Realistic population viability analysis (PVA) models were used to project the future population dynamics of 20 threatened species subject to demographic and environmental stochasticity and to catastrophes, with and without inbreeding depression. The study encompassed a range of taxa (five bird species, six mammals, two reptiles, one amphibian, one fish, three invertebrates, two plants), ecologies (carnivores, herbivores, omnivores, autotrophs), geographical origins (North and South America, Africa, Asia, Europe, Oceania), generation lengths (1󈞄 yr), and population growth rates (r = -0.07 to +0.15, as reported by the demographic analysis routine of PVA models). The 20 species are listed in Table 1, and further details on them are given in Appendix 1. The PVA input files are provided in Appendix 2 these cover the age-specific survival and reproductive rates and all stochastic effects.

The individual-based generic PVA package VORTEX, version 8.4 (Miller and Lacy 1999), was used to model the age-structured populations for the 15 vertebrates and for one of the invertebrates, and the cohort-based RAMAS ® Stage, version 1.4 (Ferson 1994), was used to model the stage-structured populations of the two plants and the remaining two invertebrates. Inbreeding depression for juvenile survival is preprogrammed into VORTEX. It was instituted in RAMAS Stage using procedures devised by Burgman and Lamont (1992), except that purging was allowed for (see Appendix 3 for a full description). Because RAMAS uses a cohort-based modeling system, the cost of inbreeding represents an average across individuals. This assumption ignores some of the potential complexities involved in coupling ecological and genetic dynamics. However, RAMAS and VORTEX gave concordant results when compared on the same species with inbreeding depression included (Brook et al. 2000), which implies that our results were not sensitive to this simplification

Because there are quantitative data on inbreeding depression for only a relatively few species, we applied a conservative value of 3.14 lethal equivalents per diploid genome on juvenile mortality. This resulted in an elevated death rate for inbred individuals before they reached breeding age, which may increase the effectiveness of purging. The value of 3.14 lethal equivalents is the median value from a study of 40 captive vertebrate populations (Ralls et al. 1988). The International Union for Conservation of Nature and Natural Resources (2000) "red lists" more than 50% of mammals as threatened, including 25 of the 40 mammalian species examined by Ralls et al. (1988). There are no clear differences in inbreeding depression between major taxa for diploid species (Ralls et al. 1988, Frankham 1998, Crnokrak and Roff 1999). This estimate is conservative, because inbreeding effects occur not only for juvenile mortality but also for adult mortality, reproductive rates, mating ability, etc. (Lynch and Walsh 1998) and are greater in the wild than in captivity (Crnokrak and Roff 1999). In the two cases for which we had direct data, the American alligator (Alligator mississippiensis) and the golden lion tamarin (Leontopithecus rosalia), we used species-specific estimates of lethal equivalents: 4.07 and 5.0, respectively.

Random mating was applied, because this is an assumption inherent in most of the literature on inbreeding in finite populations (see Falconer and Mackay 1996) it is also valid in our case because we do not apply inbreeding depression to mating. Initial conditions assume that levels of inbreeding (F ) = 0, although F is always defined relative to some arbitrary starting point (Falconer and Mackay 1996). Based on the only reliable data from Drosophila (Simmons and Crow 1977), half of the inbreeding depression was assumed to be caused by recessive lethal alleles and therefore subject to purging. The other half was attributed to sublethal alleles of smaller effect that are not much affected by purging. Purging is achieved in VORTEX through both selection and genetic drift (Miller and Lacy 1999). These values are widely accepted as reasonable, and no credible alternative values for these purging parameters exist in the literature.

Catastrophe regimes defined by studies on the particular species were used when available. When no information was available on catastrophes for a species, a default regime of 5% probability of catastrophes per year was applied, based on Mangel and Tier (1994), and a regime of effects was devised by consolidating the data from Fig. 1 in Young (1994) into five independent catastrophe classes. The individual catastrophes imposed an additional 32, 47, 62, 77, or 93% mortality, with a probability of occurrence for each class of 1% per annum.

Initial population sizes (N ) of 50, 250, and 1000 were used, corresponding approximately to the red-listed categories of "critically endangered," "endangered," and "vulnerable," respectively (International Union for Conservation of Nature and Natural Resources 2000). Insects, small vertebrates, and short-lived plants are widely viewed as having large environmental fluctuations and larger minimum viable population sizes than do large vertebrates and long-lived plants. However, these differences largely disappeared when they were compared on a per generation basis (Sinclair 1996). Because inbreeding operates on a per generation basis (Falconer and Mackay 1996), the effects of inbreeding depression might be expected to be relatively comparable across taxa.

The rate of population growth modeled in the 20 populations mainly reflects historical ecological and human impacts. Although past inbreeding may have had a secondary effect, inbreeding results in an approximately linear decline in fitness with F (see Fig. 14.1 in Falconer and Mackay 1996), so it matters little if populations previously had an F of, say, 0.2. As indicated above, the measure of 3.14 lethal equivalents we use for inbreeding depression derives predominantly from threatened species. Furthermore, any overestimate of the impact of inbreeding depression because of the implicit inclusion of past inbreeding impacts (see Brook 2000) pales into insignificance compared to researchers' underestimates of the impact of inbreeding depression resulting from the fact that they have (1) applied inbreeding depression only to juvenile survival when actual values are up to 3.3 times greater over the full life cycle (see Frankel and Soulé 1981) and (2) taken data from captive populations and applied it to the wild, where inbreeding depression can be up to seven times greater (see Crnokrak and Roff 1999).

All natural populations of threatened species have limited habitat, and it is unrealistic to assume exponential growth. A ceiling carrying capacity (K ) of twice the initial population size was therefore imposed in each case. This is a conservative way of representing habitat limitation, because functional forms of density dependence tend to strongly affect extinction risk (Ginzburg et al. 1990) and generally require the estimation of many additional parameters.

All stochastic simulations were replicated 1000 times and projected forward to estimate median times to extinction. This parameter was used because it has an unbounded scale, in contrast to proportion extinct. In addition, this measure is not biased by occasional run times that are extremely long, which is a problem with mean time to extinction. Most analyses were done using the percentage difference in median time to extinction between the models for a species with inbreeding depression (MTE ID ) and without it (MTE), computed as 100*(MTE - MTE ID )/MTE.

It was not feasible to obtain estimates of MTE in every case, because the VORTEX package has a maximum run length of 2000 yr. In these cases, simulations were projected for 25 generations, and the survivorship curves, i.e., 1 - PE(cumulative) for 1000 populations, were fitted to the lognormal distribution as theoretically predicted by Sæther et al. (2000), with a correction for censored observations, i.e., populations surviving beyond 25 generations, using the parametric regression program in JMP, version 4.04 (SAS Institute 2000). The presence or absence of inbreeding was introduced as a categorical predictor variable, and mean lifespan (ML) was estimated as exp(μ + μ ID + 0.5·ς 2 ), where μ and ς are fitted parameters. This yielded unbiased estimates of the mean time to extinction (lifespan) with and without inbreeding depression. As with the median times to extinction, the percentage difference was calculated as 100*(ML - ML ID )/ML. In a few cases, the lognormal survival analysis algorithm did not converge, so no estimate was obtained. Inclusion vs. exclusion of species lifespan estimates did not alter these conclusions.

The resulting statistics were analyzed by nonparametric methods, because percent differences were not normally distributed. Wilcoxon signed-rank tests were used to test whether the differences for each of the three population sizes and for various groups of taxa were greater than zero. Variation among major taxa was tested using Mood's median test, whereas differences among population sizes and among individual taxa were assessed using Friedman's test (see Sokal and Rohlf 1995). The significance of the relationship between percent differences and population replacement rates was tested using Spearman's rank correlation. All tests were done using MINITAB, version 12, statistical software (Ryan et al. 1994).

Projected population sizes and extinction risk

All 20 species showed a pattern of lower population sizes with inbreeding depression than without it that eventually translated into a higher extinction risk with inbreeding depression. However, the magnitude of the impact of inbreeding depression varied considerably across different species. Population trajectories for four representative species are illustrated in Fig. 1. The median times to extinction for all 20 species are summarized in Table 1. The probabilities of extinction, mean final population sizes, and % heterozygosity remaining after 100 yr, with and without inbreeding depression, are listed in Table 2.

Differences in median times to extinction due to inbreeding depression

For N = 50, median times to extinction with inbreeding were shorter than those without inbreeding for all 20 species (Table 1). The mean reduction attributable to inbreeding was 36%, and the median 28.5% (Wilcoxon W = 210, P < 0.001). Results for N = 250 and 1000 yielded conclusions similar to those for N = 50 (Table 1). The mean and median percent differences attributable to inbreeding depression were 40 and 30.5%, respectively, for N = 250 (Wilcoxon W = 210, P < 0.001) and 34 and 25%, respectively, for N = 1000 (W = 171, P < 0.001).

For the same 14 species with complete MTE data, the impact of inbreeding differed slightly but significantly with population size, giving medians of 19, 24, and 26% for N = 50, 250, and 1000, respectively (Friedman S = 12.0, df = 2, P = 0.002).

Comparisons among taxa with regard to the impact of inbreeding depression

There were significant differences among species in terms of the impact of inbreeding on extinction risk (Friedman S = 36.9, df = 13, P < 0.001). However, the impact of inbreeding was similar across major taxa, which is to be expected if the effects of inbreeding depression scale to generations. There was no significant variation among mammals, birds, poikilotherm vertebrates, invertebrates, or plants in the magnitude of the difference (Mood's median test M = 0.53, df = 4, P = 0.97). Tests of the impact of inbreeding for N = 50 and N = 250 (for which the data were complete) were significant for all taxa (see above), including vertebrates (Wilcoxon W = 120, P < 0.001), mammals (W = 21, P = 0.02), birds (W = 15, P = 0.03), and invertebrates plus plants (W = 15, P = 0.03). A full exploration of any potential differences arising from different life history strategies or ecologies would require the examination of many more species. However, we do not know of any evidence to date that relates inbreeding depression to life histories.

Relationship between impact of inbreeding and population growth rate

The major variable explaining differences among species with regard to the impact of inbreeding was the initial intrinsic population growth rate (r ). As seen in Fig. 2, the relationship for N = 250 runs was positive and highly significant (Spearman's rank correlation = 0.831, P < 0.001). Similar relationships were evident at all population sizes.

Inbreeding depression markedly reduced the time to extinction for a broad range of threatened taxa the median reduction in median time to extinction (MTE) was 25󈞋%. This was consistent across initial population sizes of 50, 250, and 1000, and there were no obvious differences among major taxa. However, there was a strong effect of initial population growth rate. These impacts of inbreeding will be underestimates (see Methods). Our findings indicate that evidence from a few species (Dobson et al. 1992, Newman and Pilson 1997, Saccheri et al. 1998, Oostermeijer 2000) applies across a wide range of taxa. Claims that purging will eliminate the impact of inbreeding depression on extinction risk are refuted by our results.

Independent evidence that populations are not driven to extinction before genetic factors can affect them comes from comparisons of levels of genetic diversity in endangered and related nonendangered species. Genetic diversity between endangered and related nonendangered species is a widely accepted comparison for general meta-analyses (see Frankham 1995a, Haig and Avise 1996, Frankham 2000) and for innumerable individual species such as the cheetah (May 1995), northern hair-nosed wombat (Taylor et al. 1994), Mariana crow, Ethiopian wolf, Mauritius kestrel, and others (see Frankham et al. 2002, Chapter 3). Most endangered species have less genetic diversity than related nonendangered species (see references above D. Spielman, B.W. Brook, and R. Frankham, unpublished data), although there are a few examples, e.g., the Indian rhinoceros (Dinerstein and McCracken 1990), that do not fit this general pattern. Because the proportionate loss of heterozygosity equals the inbreeding coefficient (Falconer and Mackay 1996), most endangered species are already inbred. If "nongenetic" factors drove species to extinction before inbreeding was a problem, there would be no such difference. Further, loss of genetic diversity is related to reduced fitness (Reed and Frankham 2002). Although there are a number of ecological factors that may also plausibly correlate with standing crops of genetic variation, the evidence for these is weak and inconsistent. The predominant factor explaining differences in levels of genetic variation among species is population size: Soulé (1976) and Frankham (1996) attribute approximately 50 and 72% of this variation to population size, respectively. In addition, a careful examination of the ecological predictors presented in Nevo et al. (1975) reveals that these are likely to be surrogates of population size. Other explanatory factors include phylogeny (vertebrates, invertebrates, plants, etc.) and range extent, both of which can also be explained as an effect of population size (Frankham et al. 2002).

Why did other researchers conclude that inbreeding depression has little impact on extinction risk? One reason is the time factor. The study by Burgman and Lamont (1992) considered only a few generations. In contrast, our study was not affected by duration because we took populations to extinction. Lande (1988) and others have suggested that other factors would be likely to cause extinctions before inbreeding depression became a problem. Our study has shown that the impact of inbreeding is less when the population growth rate is negative, as often happens as a result of human impacts. However, this effect has been severely overestimated, given the overall picture revealed by our results.

In what circumstances is inbreeding depression likely to make important contributions to extinction risk? Our results indicate that it will be important for most naturally outbreeding diploid species. However, inbreeding will have little time to act in populations that are declining rapidly due to deterministic pressures such as habitat loss (see Fig. 2). Inbreeding will have less impact on naturally inbreeding species, because on average they have lower inbreeding depression (see Husband and Schemske 1996). It will probably be less in species with polyploid ancestry, because they seem to suffer less inbreeding depression than do equivalent diploids (Husband and Schemske 1997). Species that exhibit large variations in population size due to demographic and environmental stochasticity and catastrophes are likely to be particularly sensitive to inbreeding. Populations that have had very small effective population sizes (N e < 500) for a long time, or those that have recovered from population bottlenecks, should be less sensitive to inbreeding depression due to the purging of deleterious recessive alleles. However, the effects of purging often seem to be relatively small (see Byers and Waller 1999, Miller and Hedrick 2001) Frankham et al. (2001) found no significant difference in extinction risk between purged populations (formed by crossing very highly inbred populations) and nonpurged, wild, outbred populations when both were deliberately inbred.

Our results have important conservation implications. First, ignoring inbreeding depression will substantially underestimate extinction risk. Many population viabilitiy analyses (PVAs) have been and still are being carried out without considering inbreeding depression. Almost all PVAs done using RAMAS software have omitted inbreeding depression, as have most PVAs done using software written for case-specific analysis. We are aware of only two plant PVAs that included inbreeding depression (Burgman and Lamont 1992, Oostermeijer 2000). Even those PVAs done using VORTEX software, which normally incorporates inbreeding depression, include it only for juvenile survival, despite the fact that inbreeding depression affects all components of the life cycle (Frankel and Soulé 1981).

The second concern is that inappropriate recovery programs may be devised if inbreeding depression is not taken into account. Reproductive fitness is normally improved if inbred populations are outcrossed (see Westemeier et al. 1998, Madsen et al. 1999, Ebert et al. 2002). If this is not done, an inbred population with low fitness may continue to decline, as happened with the Illinois population of the greater prairie chicken, Tympanuchus cupido (Westemeier et al. 1998). Attempts to recover the population by habitat improvement failed to halt its decline, and it recovered only after outcrossing with a population from another state. Third, the relative impact of inbreeding on median time to extinction (MTE) is similar over a range of different population sizes (although the absolute value of MTE increases as population size increases), and even relatively large populations (N = 1000) are susceptible to the deleterious effects of inbreeding. This is because the inbreeding coefficient increases rapidly when N e is temporarily reduced because of fluctuations in population size and is not subsequently mitigated (a "ratchet effect"), except through migration. Further, the relative impact of all stochastic effects on extinction risk decreases with increasing population size (see Menges 1992, Frankham et al. 2002). Fourth, funding priorities for conservation and restoration will be distorted if the impacts of different factors on extinction risk are not correctly understood.

Our results provide strong evidence that inbreeding depression elevates extinction risk in most outbreeding threatened species. They emphasize the importance of avoiding inbreeding and maintaining genetic diversity in species of concern to conservationists.


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We thank K. B. Kulasekera for his valuable advice on the statistical analyses, and Sam Scheiner and two anonymous referees for their comments on the manuscript. This study was supported by grants from Macquarie University, where the first two authors carried out much of the work for this paper, and the Australian Research Council.

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Do the effect of inbreeding depression disappear after the first generation of outbreeding? - Biology

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Joseph H. Rausch, 1,2 Martin T. Morgan 1,3

1 School of Biological Sciences, Washington State University, Pullman, Washington 99164-4236
2 [email protected]
3 [email protected]

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The minority cytotype exclusion principle describes how random mating between diploid and autotetraploid cytotypes hinders establishment of the rare cytotype. We present deterministic and stochastic models to ascertain how selfing, inbreeding depression, unreduced gamete production, and finite population size affect minority cytotype exclusion and the establishment of autotetraploids. Results demonstrate that higher selfing rates and lower inbreeding depression in autotetraploids facilitate establishment of autotetraploid populations. Stochastic effects due to finite population size increase the probability of polyploid establishment and decrease the mean time to tetraploid fixation. Our results extend the minority cytotype exclusion principle to include important features of plant reproduction and demonstrate that variation in mating system parameters significantly influences the conditions necessary for polyploid establishment.


Received: 18 February 2005 Accepted: 29 June 2005 Published: 1 September 2005

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