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- Materials and methods
Reintroductions from captive breeding programmes for endangered species have been criticized for their low success rates and high resource requirements (Snyder et al. 1996). However, they may be effective when species cannot be recovered by in situ conservation alone (Balmford, Mace & Leader-Williams 1996; Bowkett 2009). Captive programmes can have multiple objectives, including education and research (Converse et al. 2013): however, the main focus is usually the long-term viability of the target species (IUCN 2013). This is usually achieved by establishing captive insurance populations to minimize the short-term extinction risk and by reintroducing individuals into the wild to re-establish self-sustaining populations (Armstrong & Seddon 2008). In the short term, large captive populations can provide greater insurance value; in the long term, they will allow larger releases, which can improve the chances of a successful reintroduction (Griffith et al. 1989; Wolf et al. 1996; Fischer & Lindenmayer 2000).
However, releases deplete the captive population and can reduce its viability, generating a trade-off between the ‘insurance’ and ‘reintroduction’ objectives. Both aspects need to be considered, even though their relative importance can vary among programmes. Indeed, for several species, the decision to retain individuals in captivity after releases had started has ensured that eventual reintroduction failure did not result in their overall extinctions (Odum & Corn 2003; Winnard & Coulson 2008). Even for captive populations with no immediate prospect of reintroduction, releases can help in evaluating management actions and understanding threats (Rodriguez, Barrios & Delibes 1995; Letty et al. 2000). Here, larger releases can provide more information, but at the cost of reducing the viability of the captive population and negatively affecting eventual reintroductions. The biological aspects of the trade-off between reintroduction and insurance have previously been recognized (McCarthy 1994). However, the implications for cost-effectiveness have not been considered: larger captive populations can provide better insurance value, but can be more expensive to maintain. Given limited conservation resources, this aspect of the trade-off must be explicitly addressed.
The cost of the captive population can also vary depending on the length of time that individuals spend in captivity, which in turn may affect reintroduction success. Releasing early life stages may be cheaper because of reduced husbandry requirements per individual: however, younger animals generally have lower survival. Therefore, the survival and fecundity of the respective life stages will influence the trade-offs associated with the choice of actions (which life stage to release) and the importance of fundamental objectives (insurance and reintroduction). Longer periods in captivity may also imply a lack of exposure to natural selection, which can reduce survival following release, as observed for several taxa (hereon, we refer to such reductions simply as ‘post-release effects’; see for example Jule, Leaver & Lea 2008). Later life stages may incur greater post-release effects in species in which selection affects mainly early stages, such as amphibians (Wells 2007). Again, previous analyses of this trade-off have not included cost as an important decision-making criterion (Burgman, Ferson & Lindenmayer 1994; Sarrazin & Legendre 2000).
In this study, we use population models to assess the cost-effectiveness of alternative release strategies for species with complex life histories. For an ongoing programme for a critically endangered amphibian, we identified the optimal release rates for eggs or subadults, in regard to both insurance and reintroduction objectives and management costs. Under a long-term programme, assuming a captive population with a stable age distribution, large releases of subadults were the optimal choice. Conversely, for a short-term release plan, mixed releases of variable proportions of both eggs and subadults provided larger and cheaper wild and captive populations.
- Top of page
- Materials and methods
The first objective of ex situ programmes for critically endangered species is often the establishment of a viable captive population as insurance in the event of extinction in the wild (Conway 2011). The growth rate of the captive population will determine whether this objective can be met, and how it can be balanced with future reintroduction efforts. Populations with high predicted growth are more likely to be able to sustain large release rates and still maintain viability. For P. corroboree, long-term persistence of the captive population could be ensured even when releasing a large proportion (0·98 on average) of either eggs or subadults, due to high survivorship and productivity in captivity.
Releasing different life stages will also change the stable age distribution of the captive population and may affect specific objectives regarding its structure (e.g. representation of genetic diversity). The cost of a stable and constant captive population depends on the maintenance requirements of different stages. In the stable age distribution scenario for P. corroboree, the difference between maintenance costs for eggs and other life stages was not sufficient to change the optimal strategy. However, if differences are significant (e.g. when breeding adults need large individual enclosures), different release strategies may have different costs, influencing the optimal decision when cost is an objective.
In the stable age distribution scenario, the insurance objective for the captive P. corroboree population was to maintain λc ≥ 1, as a simple approximation of viability. Managers may initially seek a higher growth rate, to increase the size of the captive population and release greater absolute numbers in the future: however, resource constraints are likely to impose an upper limit to the captive population size. Once this carrying capacity is reached, the ‘insurance’ objective may simply shift to λc = 1. Similarly, initial releases in excess of the maximum sustainable rate will result in a population reduction (λc < 1) until a lower bound is reached. In our case, we assumed that this bound was equal to the initial population size, although the exact relationship may differ among projects. In a theoretical study, Tenhumberg et al. (2004) suggested that it is generally optimal to increase the size of the captive population as rapidly as possible, and to start releases once this approaches its carrying capacity. In real programmes, the practical challenge for managers lies in estimating the optimal duration of this ‘build-up’ phase and the subsequent proportion of individuals to release. Framing population models in a clear decision-analytic framework can help in assessing the optimal decision.
In regard to the reintroduction objective, the trajectory of the wild population depends on its intrinsic growth rate. If λw > 1, then the population will grow accordingly after the initial releases, and assuming exponential growth, constant releases from a stable and constant captive population will become progressively less important in the long term. On the other hand, if λw < 1, as it was for P. corroboree in this study, continuous releases are needed to prevent the wild population from declining to extinction. Whether such an approach is justified depends on the objectives of the specific programme. In the case of P. corroboree, where the wild population depended on continuous releases, egg releases were less effective than subadult releases, yielding a smaller number of individuals in the wild for every individual maintained in captivity. If cost is an objective, it is therefore necessary to consider that when releasing eggs a greater population will need to be maintained to provide the same absolute numbers of individuals in the wild.
The effectiveness of releasing different life stages will depend on their expected vital rates. In general, individuals that are released later in life will have better survival than those released early and thus provide a greater wild-to-captive ratio. Within the matrix population model framework, the effects of this increase in survival can be summarized, for example, using reproductive values (the expected number of offspring an individual will produce over its lifetime: see for example Sarrazin & Legendre 2000). However, newly released individuals can suffer abnormally high mortality or low fecundity, reflecting a lack of natural selection during the captive phase or adaptation to captivity (McCarthy, Armstrong & Runge 2012). Reintroduced adults can also exhibit abnormally high dispersal, a behavioural aspect observed in several taxa (Le Gouar, Mihoub & Sarrazin 2012). The use of environmental cues for dispersal has been demonstrated for amphibians, particularly for juveniles learning dispersal routes post-metamorphosis (Ferguson 1971). In this case, the effects of post-release dispersal on the establishment of a reintroduced population may be higher for late-age-class release strategies, in which individuals have had no opportunity to learn dispersal routes. Although no information is available in this sense for P. corroboree, post-release dispersal could be considered as additional mortality.
Although such post-release effects can reduce the relative benefit of releasing later life stages, these may still be advantageous where the better survival of older individuals compensates the incidence of post-release effects. For example, Sarrazin & Legendre (2000) modelled releases of Griffon vultures Gyps fulvus in Europe, suggesting that where post-release effects remain small, releases of adults should indeed prove more effective for long-lived species. In this sense, the average longevity of P. corroboree (6–10 years in the wild: Hunter 2007) makes it a comparable case study for several amphibians, mammals and birds targeted by captive breeding efforts (see for example the species listed in Short et al. 1992; Griffiths & Pavajeau 2008; Graham et al. 2013). Additionally, for amphibians and other r-selected taxa, mortality in early life stages can naturally be an order of magnitude higher than for adults, potentially offsetting post-release mortality of older individuals. Amphibians are also less reliant on learnt behaviour than mammals or birds, further reducing the potential for adaptation to captivity (Griffiths & Pavajeau 2008). Finally, for many amphibian species, it might be possible to compensate mortality by releasing thousands of individuals, especially for juvenile stages, as shown, for example, by our results for P. corroboree. Such numbers may not be practical for other taxa, reducing the scope for a solution of the trade-offs in vital rates.
In our case study, results were also generally insensitive to the short-term fecundity of released subadults. This was a result of the longevity of adults and our assumption of no long-term variation in fecundity (from the second year after release all individuals would have the same reproductive output). Greater sensitivity might be expected in the case of long-term variations in fecundity that differed between life stages (e.g. if early-age releases achieved full reproductive potential, and late-age releases never did). In this sense, our results are consistent with those of Sarrazin & Legendre (2000), who found greater sensitivity of reintroduction success to post-release survival than to fecundity for Gyps fulvus in Europe (for reductions both in the short and in the long term).
Environmental stochasticity can also have life stage-specific effects that will influence the relative effectiveness of release strategies (Sarrazin & Legendre 2000). In P. corroboree, egg survival can be affected by environmental stochasticity: since eggs are laid in nests on the ground, they need sufficient precipitation to be flushed to a water body that must retain water throughout the period of tadpole development (Hunter et al. 2009). High mortality of eggs and low recruitment have been observed in drought years (Osborne 1989). On the other hand, wet years can facilitate the spread and virulence of chytrid fungus, again with potential age-specific effects (Kriger 2009). In the light of this complexity, currently not entirely understood for P. corroboree, we chose not to explicitly model environmental stochasticity in our study; however, it may affect the efficiency of egg releases in particular, and monitoring is being carried out to evaluate this possibility.
When the size of the captive population is not constant, retaining individuals in captivity for a longer period will increase the overall financial cost of a programme and may generate conflicts where limited resources are available (such as space or human resources at zoo institutions). In this case, strategies that envisage releases of a single life stage releases inevitably bear the consequences of this trade-off. Focusing only on releases of early life stages might be appealing to risk-seeking managers with strict budget constraints. This was clearly reflected in our realistic, short-term example for P. corroboree with multiple objectives (insurance, reintroduction and costs). Across the full range of parametric uncertainty, mixed strategies including joint releases of eggs and subadults provided the most cost-effective solution. Although egg releases allowed the lowest costs, they were also less effective towards the reintroduction objective. The reduction of age-class diversity in the captive population might be another concern. Conversely, releasing subadults was predicted to produce a greater presence in the wild. This potential risk-averse solution, however, led to high and increasing costs, making it impossible to meet both cost and insurance objectives. In this sense, mixed strategies allow managers to combine the advantages of releasing different life stages, for example releasing subadults to improve viability and managing egg releases to control the size and cost of the captive population.
Ultimately, the evaluation of the trade-off between additional costs and the predicted improvements in viability associated with releasing later life stages must be solved on the basis of the importance given to each objective. For example, Martínez-Abraín et al. (2011) used population viability analysis to assess translocation options for crested coots Fulica atra in Spain: they found that releasing adults improved viability, but this remained generally poor. They concluded that a 160% increase in costs ‘outweighed’ the marginal conservation benefits of releasing adults rather than juveniles. However, the definition of the threshold above which benefits are outweighed by costs will differ among programmes. Adopting an explicit decision-analytic approach may help define priorities and consequently inform the optimal decision.
The definition of clear objectives is the key to a rigorous approach to decision-making (Possingham et al. 2001). In our realistic scenario, we chose a short time frame to reflect the current requirement of the release programme for P. corroboree and used the number of breeding adults as a metric of success; longer programmes may focus on growth rates or probability of extinction. Our choices are influenced by the current difficulty in mitigating threats for this species, reflected by the expected negative growth rate of the wild population. Reintroduction programmes aimed at establishing self-sustaining populations, or involving species with longer generation times, may need to consider longer time frames, which can be easily accommodated in our approach. Finally, we recognize that legislative or funding constraints can create additional objectives, for example requiring managers to report some metric of success within a given deadline. If such additional objectives can be stated explicitly, the decision framework can be modified to accommodate them.
Additional aspects would also need to be considered in a more realistic analysis. For example, our assumption of exponential growth may be violated by small-population dysfunctions such as Allee effects. We did not model the genetic viability of the species either in captivity or in the wild, although the management of captive populations to minimize inbreeding is recognized as a key component of recovery efforts for P. corroboree (Lees, McFadden & Hunter 2013). Again, different components of the decision process for this problem could be modified to account for these aspects.
The ex situ conservation programme for P. corroboree shows characteristics common to most similar efforts world-wide: the rapid and seemingly irreversible decline of the target species, the unknown feasibility and time frame of threat abatement, the need to minimize time in captivity and maximize production of release candidates, while containing costs. Population models can provide useful information for managing a captive insurance population, while life stage-specific release plans can help in addressing trade-offs between numbers of releases, the probability of establishing a wild population and management costs. Framing models in an explicit decision-analytic framework can assist in evaluating key objectives, uncertainty and trade-offs.