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A small blackish muntjac has been discovered in the west Quang Nam province of Vietnam. DNA sequencing of six individuals showed that this is a distinct species. A description is given. The new find is the third new ungulate species to be discovered in Vietnam in five years. The paper predicts that more species remain to be found in the Annamite mountain range and includes DNA analysis from two more undescribed species. The DNA analysis indicates that there has been a radiation of several related muntjac species in the Annamite mountains, also that the new truongson muntjac shows high intraspecific genetic diversity indicative of a large effective population size. The paper discusses the significance of these finds from an evolutionary and conservation viewpoint. The authors urge conservation authorities in the IndoChinese countries to devote special measures to protect this site of dynamic evolution which retains a combination of relict endemics (e.g. Pseudoryx) as well as on-going speciation.

In 1998 we wrote a paper describing DNA-based methods for estimating and monitoring both census ( N c ) and effective ( N e ) population sizes in natural populations (Schwartz, Tallmon & Luikart, 1998). The purposes of this paper were: (i) to review some relatively unexplored, but promising ways that multi-locus genetic data can be used to monitor and detect population declines, (ii) to encourage further development and validation of recent DNA-based N c and N e estimators. These approaches are urgently needed to help detect natural population declines in their early stages, so that management actions can be taken to avoid population endangerment and subsequent extinction. Crandall, Posada & Vasco (1999) have criticized the N e portion of our review claiming that we: (i) excluded phylogenetic estimators of N e that employ coalescent theory, (ii) did not distinguish between historical and current estimates of N e , (iii) did not define which N e we were estimating.

We would like to thank Cunningham and King (2013) for their comment on our paper about evaluating different indices of success (Howe & Milner-Gulland, 2012), particularly for their support for research that aims to develop methods for assessing the outcomes of conservation projects. However, there were a number of misunderstandings that have arisen from their interpretation of the original paper. The response from Cunningham and King (2013) does not appear to recognize the aim of this paper, which was to develop robust indices of success that are consistent between projects and evaluators for conservation programmes and projects in general. This paper does not attempt to evaluate the Darwin Initiative, relative to other conservation programmes, nor is it a criticism of the methods by which the Darwin Initiative measures its own success. For the purposes of this study, the Darwin Initiative is used solely as a database to evaluate the potential of three indices of conservation success. The Darwin Initiative was chosen because of its international reputation, the length and depth of its documentation and because confounding variables are reduced as projects are generally independent of each other, have similar budgets, the same duration and similar broad goals. By failing to recognize the purpose of the paper, Cunningham and King (2013) overlook the findings of this study that demonstrate it is possible to develop robust outcomebased indices of conservation success for comparison of projects within a funder’s portfolio and that output-based indices show similar results. This study does this in two stages: firstly by evaluating the internal consistency of the indices chosen and, secondly through a comparative assessment of the indices’ rankings of project success. On the first aim, we demonstrate that although there were systematic differences between scorers, the relative rankings between them were consistent. Cunningham and King (2013) suggest that we disapprove of the Darwin Initiative’s use of outputs as a measure of success; however, we do not state that this is the Darwin Initiative’s only measure of success nor do we find that it fails as an indicator of success in its own right. In fact, the output index (based on the standard measures) and the subjective ranked outcomes index (based on narratives within the final reports) were fairly consistent between assessors in their rankings of project success. However, the nuanced differences between the indices led us to conclude that different indices pick up different facets of project success, and therefore there is a need for multiple indices. When the data for this study were collected in 2006, with the full support of both the Darwin Initiative and the Edinburgh Centre for Tropical Forests (the organization then in charge of administrating and evaluating the Darwin Initiative), we were provided solely with the final reports and Darwin Standard Outputs as our raw data. The logical framework referred to was introduced in 2001; consequently, there were only 3 years of completed projects (2003–2006) that contained logical frameworks at the time of our study. Therefore it was not possible to use this information. If this study were to be repeated, this information would be an interesting complement to the indices used. As correctly stated by Cunningham and King (2013), a number of the standard measures are inputs rather than outputs; however, for the purposes of this study, only the outputs were used (as discussed in the Methods section) and therefore our use of outputs as a comparative index was valid. Regardless of whether the Darwin Initiative uses the standard measures as a method of evaluating their projects, a large number of conservation and development programmes and projects still use outputs as a method of evaluating their progress or success, and therefore it was a useful comparison to include an output-based index alongside two outcomebased indices for this comparative evaluation study. Finally, Cunningham and King (2013) discuss the Darwin Initiative’s current method of evaluating their projects by reviewing how successful the project has been at achieving its purpose statement (outcome). The method described is similar to, but perhaps less quantitative and repeatable than, the ranked outcomes method developed in this study. Their method allows for effectiveness to be evaluated on a project-by-project basis, whereas the different bs_bs_banner

Can wildlife learn harmful and maladaptive behaviours from each other? If so, insights into social learning among animal populations in response to anthropogenic stimuli are of wide interest and applicability. Donaldson et al. (2012) address social learning in human–wildlife interactions involving food provision. ‘Food-conditioned’ animals are subject to operant conditioning in which learning about anthropogenic food arises from repeated exposure to human stimuli, behavioural responses to those stimuli and reinforcement of behavioural responses because of food reward (Whittaker & Knight, 1998; Samuels & Bejder, 2004). In this paper, Donaldson et al. carefully negotiate the asocial/social learning dichotomy. The former arises from individual responses to the availability of anthropogenic food (e.g. in species that receive limited maternal care and are solitary as juveniles and adults). However, responses to provisioning may also arise through social learning in cases where individual animals are repeatedly exposed to the feeding behaviours of conspecifics that exploit anthropogenic foods. Operant conditioning describes a learning process that may be acquired individually or socially (Sargeant & Mann, 2009). Social learning involving responses to anthropogenic food may then be facilitative or supplementary. The specific focus of this paper is the long-term illegal provisioning of bottlenose dolphins (Tursiops aduncus) by members of the recreational fishing community in Cockburn Sound, south-western Australia. Social learning is available in this case, as individual bottlenose dolphins exist within complex social structures that involve extended juvenile dependence and enduring relationships between animals (Connor et al., 2000; Lusseau, 2003; Sargeant & Mann, 2009). Two variables emerge from the analysis as predictors of individual animals learning to accept food from recreationists engaged in fishing; the use of areas with high densities of recreational boats and association with previously conditioned dolphins. The harmfulness of such learned behaviours may include increased wildlife morbidity/ mortality, interventions to address ‘problem’ wildlife (which may involve culling) and, therefore, compromised sustainability resulting in loss of economic opportunities. These findings highlight the facilitative function of social learning in determining the development of individual animal behaviour responses to anthropogenic stimuli. They should motivate others to explore how these findings may extend beyond food provision (to other anthropogenic stimuli) and into a range of applied fields. One important application of social learning in human–wildlife interactions is sustainable tourism management. Knight (2009: p. 180) observes that we now live ‘. . . in an age when our visual appetite for wildlife has never been greater’. Wildlife viewing has rapidly moved into the mainstream of commercial tourism (Knight, 2009). More social species, including cetaceans, have become the focus of global interest, as tourists seek to observe critical behaviours such as maternal care, feeding and social interactions between conspecifics in the wild. The International Fund for Animal Welfare estimates that the whale watching industry now exceeds $2.1 billion per annum, catering for 13 million whale watchers, and generating 13 000 jobs (O’Connor et al., 2010). Neves (2010) describes this evolution as one form of periodic transformation in the global capitalist economy, and it is a transition that also raises the likelihood of social learning among individual animals that are subject to human stimuli through intensive and prolonged interactions with commercial tourism business and their clients. The provisioning of wild dolphins in a tourism context is unusual but does occur (Orams, 1995). More broadly, the provisioning of wild animals, common in the past (e.g. bear feeding in Yellowstone National Park; Davis, Wellwood & Ciarniello, 2002) continues in some tourism contexts, both directly (e.g. use of food to attract pelagic birds, sharks) and indirectly (e.g. attraction to camp sites because of food storage and disposal, Knight, 2009). This raises questions of diminished behaviours and reduced ‘wildness’ associated with efforts (deliberate or otherwise) to make wildlife viewable (Knight, 2009). However, the relevance of social learning in human–wildlife interactions should extend beyond provisioning, to other forms of tourism (e.g. swimming with dolphins) and specific behaviours (e.g. bow riding – the behaviour of dolphins in which they swim or ‘ride’ the crests of waves formed from the front of moving boats) that may increase the bs_bs_banner

Imagine a construction project whose objective is to build a bridge. The quality of the resulting bridge might be debated and disputed, but that it has been delivered, and that the project is responsible for its delivery, is usually relatively simple to verify. It is much more difficult to evaluate success where a project aims to deliver something whose existence is costly or technically challenging to monitor, and something whose status may be affected (positively or negatively) by a range of influences that have nothing to do with the project. Conservation interventions may seek to influence the conservation status of a range of species, improve the management of an area of habitat, or change the attitudes and behaviour of a population of people. All of these things are challenging to measure in themselves. Importantly, they can also be affected by a myriad of factors external to a project (global commodity prices, national politics or law enforcement, and shifting social norms) making it challenging to tease apart the influence of the project on the sought after outcome. Conservation projects have widely been criticized in the past for poor evaluation (Saterson et al., 2004; Brooks et al., 2006). In their paper, Howe & Milner-Gulland (2012) look at the question of what indices are appropriate for evaluating success in conservation. Howe & Milner-Gulland (2012) follow others by distinguishing outputs (the amount of something delivered by a project, e.g. number of workshops held, papers published or posters distributed) and outcomes (the long-term consequences of the project, e.g. change in population size of a target species). Outcomes are what the project ultimately aims to deliver, but they can be very costly to measure. A recent study of the costs of monitoring presence or absence of a variety of species of conservation concern in the dry forests of Madagascar illustrates the challenge of monitoring outcomes directly. Sommerville, Milner-Gulland & Jones (2011) found that monitoring that could robustly detect change over time would be unrealistically costly for the vast majority of species as would cost more than the budget for the entire intervention. The UK government launched its Darwin Initiative at the Rio summit in 1992. Since then, it has invested £88 million in biodiversity conservation projects in 154 countries (DEFRA, 2012). This fantastic programme provided Howe & Milner-Gulland (2012) with an unrivalled opportunity to investigate how much agreement there was in rankings of project success as evaluated using different indices (one based on reported outputs, and two based on subjective scoring of information about outcomes) and also, which explanatory variables best predicted success as defined by the different indices. Their finding that ranking of projects using the outputs-based indicator was well correlated with the ranking from the subjective outcomes measure is interesting and worthy of further exploration. However, as the authors themselves note, there is no quantitative, independent data on outcomes available against which to measure the success of the various indices. Because outcomes are so difficult to measure directly, and may also not be achieved over the small timescale of a funded project, indices based on outputs will always be needed. Underlying this approach is an assumption that there is a mechanism that links delivery of the outputs with delivery of outcomes. This is often not explicit. If assumptions as to linkages between outputs and outcomes were more explicitly spelt out, both in project proposals and reports, alongside the evidence upon which the assumption is based, output measures would become more valuable for assessing project success. Howe & Milner-Gulland (2012) also investigated the internal consistency (how different assessors would score an individual project using the same index) of two of the possible indices. While they found a high level of agreement between different assessors scoring the same projects with the same index, the existence of an outlier was revealing. The majority of their assessors came from a very similar academic background, while the one from a different discipline (pest-management rather than conservation), scored projects quite differently. It is likely that world view plays an important role in what an individual considers as success in bs_bs_banner

Despite the impression of doom and gloom concerning the fate of the world’s wildlife that one can gather from the media, there are in fact many success stories from recent decades where populations of various species have stabilized or recovered. However, learning to live with these conservation success stories is often harder than bringing about the initial recovery. European large carnivores are a case in point, especially the wolf (Canis lupus). Following the changes in public opinion and policy that began appearing in the 1970s, 1980s and 1990s, wolf populations have begun to recolonize many parts of Europe, such as Scandinavia, the French and Italian Alps, and the German lowlands, from which they had been absent for many decades. The response has been an explosion of conflict, both of a material and economic nature (caused mainly by depredation on livestock; Kaczensky, 1999) and of a social and political nature (focusing mainly on the symbolic aspects of this recovery; Skogen, Mauz & Krange, 2006; Dickman, 2010). The case of wolves also illustrates the diversity of conflicts that can emerge, that go far beyond the simple economic conflicts and into areas where perceptions are just as important as reality. Seals are a clear marine parallel to terrestrial wolves in both their conservation history and the range of conflicts that their conservation has produced, with salmon taking the place of sheep as the main victims, and farmers and hunters being replaced by fishermen of various types as the main protagonists. Addressing such conflicts, be they in Scottish estuaries, European mountains or Scandinavian forests, requires a three-step process. Firstly, there is a need to recognize the genuine existence of a conflict and the equal legitimacy of the diverse dimensions along which a conflict can be expressed. Secondly, there is a need to understand the functional mechanisms behind the conflict, including both the ecological and the sociological/political aspects. Only when these two steps have been completed is it possible to move onto the third step of identifying potential conflict reduction, mitigation or compensation measures. The paper by Graham et al. (2011) is a welcome contribution to the second step of this heuristic framework and is set in the context of the conflict between seals and salmon fishermen. Killing carnivores and seals is controversial with the public and may endanger their conservation status (Treves, 2009). Graham et al. (2011) seek to explore the evidence for the existence of problem individuals (sensu Linnell et al., 1999) among the seal population, which are responsible for a disproportional impact on salmon. The idea is that a more selective removal (i.e. killing) of these individuals, if they exist, would create less of a conflict with the public’s sensibilities and with seal conservation than a more widespread use of lethal control across the whole seal population. This testing of an underlying management assumption is a crucial step toward a robust and knowledge-based management system. In their paper, Graham et al. (2011) draw on multiple lines of evidence to indicate that there may well be some individuals that disproportionally make use of rivers and that seals in rivers feed more on salmon and trout than other seals. Taken in isolation, none of their data would be strong enough to draw conclusions. Their sample size for investigating seal diet is particularly small. However, when taken together, the multiple lines of evidence certainly present a strong case for the existence of problem, or ‘rogue’ seals, although there is clearly a need for more research on the topic. Comparative telemetry data from rogue and nonrogue seals would have been particularly interesting as it has often been used in comparative studies of terrestrial carnivores (e.g. Odden et al., 2002), as would the use of welldesigned adaptive management experiments where the impact of seal removal was monitored. The important question is how this new information is used in the context of seal management. This latest paper goes a long way to testing the assumption of the existence of rogue seals, which is part of the model presented earlier by Butler et al. (2006). However, there is still a little understanding of exactly how much time each rogue seal spends in the rivers, which will be crucial to understanding the potential impact on salmon runs of removing these individuals. This opens up for a fascinating and complex debate about the place of lethal control in conflict management of predatory mammals in general (Treves, 2009). On the one hand, lethal control is often used in the belief that removing certain individuals or a proportion of the population will actually achieve measurable changes (such as reduction in Animal Conservation. Print ISSN 1367-9430

Global biodiversity is increasingly under threat and species loss has prompted the argument that our natural world is experiencing a sixth major extinction event (PimmB Woodroffe, 2000; Treves & Karanth, 2003; Whitman et al., 2004). Parasites are increasingly linked to population impacts in wild felid and canid species (Pedersen et al., 2007). Many predator species exist in small and fragmented populations, vulnerable to stochastic events and other ongoing sources of population loss, which can drive them locally, if not globally, extinct. The Iberian lynx is a key example of an imperiled predator and is currently the most endangered felid in the world (Nowell & Jackson, 1996; Baillie, Hilton-Taylor & Stuart, 2004). This species may lose the survival battle without intense and continued conservation intervention. In their paper, Lopez et al. (2009) provide an excellent example of endangered species management, which demonstrates the importance of long-term population monitoring programs and the effectiveness of careful intervention design. Through their monitoring program, the authors identify the emergence of feline leukemia virus (FeLV) in the core population of Iberian lynx on the Iberian Peninsula. This pathogen is commonly found in domestic cats, which are considered the primary host (Arjona et al., 2007). While previous low-level FeLV exposure had been identified in the Iberian lynx population before the 2007 outbreak (Luaces et al., 2008; Meli et al., 2009), no associated mortality or other impacts had been identified. Then, inexplicably, FeLV emergence in 2007 causes an outbreak with mortality levels that threaten the survival of the species. Viral sequence from the 2007 outbreak is distinct from previous infections in the population but consistent with transmission from locally infected cats (Meli et al., 2009). In response to the outbreak, a FeLV control program is undertaken by the authors and other partners, which utilizes a test, removal and vaccination approach in Iberian lynx coupled with reduction of the sympatric feral cat population. It is successful and FeLVassociated mortality in the Iberian lynx ceases. What can we learn from the approach taken by Lopez and colleagues? Firstly, intervention for the Iberian lynx disease outbreak is designed and implemented through a partnership approach between many national stakeholders, project participants and experts. Often in wildlife management crises, consultation and inclusiveness of stakeholders is avoided in order to streamline processes and expedite action. But this can backfire when partner agencies and stakeholders are required in downstream management activities. This paper serves as a reminder that wildlife management is better implemented inclusively rather than exclusively. Secondly, this study demonstrates the dynamic and unpredictable nature of pathogen invasion outcomes in a population. In this case, FeLV invasion in the population was not identified as an important problem but this changed dramatically in the 2007 outbreak. This highlights the danger of defining the threat level of a pathogen for a species. It is becoming clearer that pathogen behavior can vary widely between species, and within and between populations over space and time. A number of factors influence pathogen invasion outcomes as they operate and interact at the level of the host, pathogen and/or the environment, which can influence host susceptibility and/or pathogen virulence and change the nature of an outbreak in a susceptible host population (Hudson et al., 2002). Thus, as in forensic sciences, we are reminded to be cautious with ‘pathogen profiling’ or reliance on expected behavior of a pathogen, as seen here in the divergent nature of FeLV invasion in Iberian lynx or for example, distemper outbreaks in African wild dogs in Botswana (Alexander et al., 1996; Alexander et al., 2008)

Scientists lead a double life as scientist and citizen. The scientist seeks objectivity so as to not compromise interpretation of the results of their studies. The citizen is filled with passions and aversions that necessarily introduce subjectivity when navigating the complexity of the world. Biologists are scientists and citizens of local and national communities; correspondingly, biologists versed in the complexity of nature and the uncertainty of science should voice their informed opinions. The scientist and citizen require distinction, nonetheless. Once the results of the research are subjected to rigorous peer review so that the objectivity of science is assured and the inescapable uncertainty emphasized, then scientists ‘. . . can shift gears and tell us (the public) the meaning they themselves find in those results, as well as the actions that they would like to see adopted in response to their findings’ (Limerick & Puska, 2003, p. 25). Separating science and advocacy is crucial and should be complete. Advocacy sometimes surfaces in the pages of peer-review science journals in a scientific report, a practice that should be discouraged. Plenty of avenues for advocacy exist, including the scientific journals that publish scientific reports, in sections clearly identifiable as advocacy, opinion, point of view, perspective and the like. Rob Ramey is a self-proclaimed advocate for changing the Endangered Species Act (see New York Times, 27 June 2004). Like Governor Bill Owens of Colorado who believes in ‘. . . a natural order, which means that some species will always be sublimated while other species are on the ascendant’, Ramey advocated that we should not waste our conservation resources on the ‘little twigs’ (referring to the tree of life) at the expense of the ‘big branches’. ‘We just have to understand that we may have to lose some of the little twigs out there. That means that some groups will lose their ESA cash cows, but it’s for the long-term good’ (Westword, 1/20/05). In the first paragraph of the Animal Conservation paper where Ramey sets the stage for the research, his advocacy continues: ‘If defensible data are lacking and a protected organism is not distinguishable with a high degree of certainty from neighbouring, nonthreatened relatives, considerable financial and logistical conservation effort may be misallocated at the expense of other endangered organisms.’ It appears that the purpose of the Ramey et al. (2005) paper is to make the task of maintaining biological diversity less expensive. We are led to believe that the Endangered Species listing of meadow jumping mouse subspecies was a costly error demanding correction. This is hardly science; nonetheless, this brand of advocacy masquerading as science has its supporters. The Governor’s Office of the State of Wyoming, the organization Coloradan’s for Water Conservation and Development (not an environmental group), The Colorado Association of Home Builders, the Republican Senator Wayne Allard and others applauded Ramey’s work because it resulted in ‘. . . removing a costly listing that has stymied people, businesses, and government in our state’ (Senator Wayne Allard). Ramey et al. (2005) obliterated the distinction between scientist and citizen. He has made abundantly clear that the ‘little twigs’ (what Governor Bill Owens refers to as ‘obscure creatures’) need pruning to save the ‘big branches’ (e.g. condors and bighorn sheep). To do so, science was corrupted. For example, in the absence of data, Ramey et al. (2005, p. 340) wrote: ‘While the absence of evidence does not necessarily mean there is evidence of absence, there do not appear to be any adaptive differences that prevent the Z. hudsonius subspecies in this study from being ecologically exchangeable. We therefore cannot reject the null hypothesis of historic or recent ecological exchangeability.’ In other words, because there are no data showing adaptive differences between subspecies, the subspecies are ecologically redundant. Such a sweeping and important biological conclusion is not legitimate without doing the difficult, time consuming and costly tests of whether related species are actually ecologically equivalent. Additionally, Ramey et al. (2005, p. 338) estimated gene flow from FST values derived from microsatellite data and concluded that the hypothesis of genetic exchangeability could not be rejected, even though estimation of gene flow using FST has been severely criticized and discredited (Whitlock & McCauley, 1999); moreover, Ramey et al. failed to note that analysis of microsatellites using FST grossly overestimates gene flow, ensuring that the null hypothesis cannot be rejected based on the criterion (Hedrick, 1999). Other examples of unsound analysis and mistaken interpretation pepper the paper Sublimate means to express an instinctual impulse (like sexuality) into a socially more acceptable activity; I suspect he meant subjugate (i.e. to conquer).

While we certainly appreciate the concerns of Martin (2006) about the quality of papers published in our journal as well as the quality and integrity of our peer-review process, we would like to reassure readers that his concerns about our review process are unfounded, and in doing so mention also that concerns about quality of analyses are ubiquitous, regardless of the methods used. Martin’s (2006) first objection to the Ramey et al. (2005) publication is with the perceived advocacy in the absence of peer review. While scientific objectivity in conservation biology, in particular, is difficult (many of us become biologists because of our love of nature), Martin is absolutely correct that the peer-review system should, among other things, guard against blatant advocacy, especially in the face of conflicting data. Yet unlike Martin’s letter, the Ramey et al. (2005) paper was passed through a standard peer-review process. Indeed, the first submission was rejected due to two reviewers who were unconvinced by both the analyses and the data. Both reviewers, however, recommended that the authors be encouraged to resubmit if their concerns could be addressed. As the editor assigned to this paper, I concurred with this assessment and rejected the paper with encouragement to resubmit. The Ramey team then collected the additional data (microsatellite data as a nuclear complement to their mtDNA sequence data) and performed additional analyses as suggested by the reviewers, and resubmitted their paper. This revised draft went out for a second round of full review (a more rigorous standard than in many journals). I sent this paper back to one of the original reviewers (the harshest – as is my own personal policy) as well as a ‘fresh’ reviewer (which is also a standard of mine to make sure that the paper still has good flow and consistency). These reviewers were more favourable to this revised draft, but still had concerns which were subsequently addressed by the authors in a final submission. As the editor, I then examined the final paper and found that the concerns of the reviewers had been addressed. While we obviously cannot disclose the identity of the reviewers, they were, contrary to Martin’s speculation, experts in population genetics and conservation genetics. Thus, the peer-review process of this paper was of at least standard rigour. Martin offers two lines of evidence of ‘advocacy’ by Ramey. The first is citing Ramey as someone who ‘is a selfproclaimed advocate for changing the Endangered Species Act’. I would argue that most conservation biologists would be in this same camp. There are many scientific difficulties with the Act as it stands; however, opening up that for review opens up many doors for political input that could even further reduce the scientific input into such questions. Of course, as a journal, we do not solicit opinions about national conservation policy from authors before we make decisions on manuscripts. The decisions are based on peer review (as described above). Martin’s second line of evidence is the following quote from Ramey et al. (2005): ‘If defensible data are lacking and a protected organism is not distinguishable with a high degree of certainty from neighbouring, non-threatened relatives, considerable financial and logistical conservation effort may be misallocated at the expense of other endangered organisms.’ In my view (as an alternative peer reviewer), this is not advocacy but straightforward common sense. Funds for endangered species are very limited. Why would you want to spend these precious resources on taxa that are originally based on weak data and do not hold up to scientific scrutiny. It seems as thoughMartin is advocating to ignore the science altogether whereas Ramey et al. (2005) clearly state ‘If defensible data are lacking . . .’. Rather than making ‘the task of maintaining biological diversity less expensive’, Ramey seems simply to be arguing for those precious resources to be more carefully allocated to species of truly high need for conservation. Finally, Martin critiques his perception of the ‘peer review process orchestrated by the journal Animal Conservation’ and rates it miserable. I am the one who ‘orchestrated’ the peer review along what I thought (and still think) are standard peer-review practices (as described above). In addition to our own peer review, the Ramey work has also been extensively reviewed by 14 peers via the US Forest Service and these reviews are available online (http://mountain-prairie.fws.gov/preble/). As one can quickly see from Martin’s critique, advocacy is alive and well in the science of conservation biology and advocates fall out on both sides of the issues. In this, as in many other cases, conclusions are debatable, and Animal Conservation, like other scientific journals, is the correct place for those scientific uncertainties to be expounded and debated. In this same issue, Ramey et al. respond to the scientific critique of their article. While I disagree with the recommendation of retraction, certainly standard scientific practice would be to strongly support Martin’s recommendation to acquire independent replication of the study for the strongest conclusion. It is through the preponderance of evidence that these legal

One of the most difficult questions in conservation science is when we as humans should intervene in natural ecological and evolutionary problems. Even when there are presumed successes, such as in bringing back populations from decline or ridding areas of invasive species, there is the feeling that we have not necessarily done the right thing – the natural process has to some extent been tampered with. Perhaps the hottest scientific topic in this regard is whether charismatic species (often carnivores) benefit genetically or ecologically when individuals from outside populations are introduced into declining ones. In this issue is a series of articles evaluating the successes, failures and contentious issues concerning introductions and consequent long-term management of the Florida panther. Because of the strong feelings aroused by this set of articles, we feel it is important to clarify the editorial process that took place. The original manuscript by Pimm, Dollar & Bass (2006) was submitted on 28 September 2004. Accompanying the submission was an extensive cover letter explaining the context and history of the work, including a statement from Pimm acknowledging that the management of the Florida panther was controversial. Pimm stated that if Animal Conservation were to accept the paper, he would request that Dave Maehr, one of the principal researchers long involved with the panther issue, be sent the manuscript for comment. The manuscript was reviewed by two external referees. On 10 January 2005, the paper was rejected with encouragement to resubmit if the authors could respond thoroughly and effectively to the referees’ comments. A revised manuscript was received on 16 April 2005. The paper was re-reviewed and acceptance was recommended following further revision. The editors finally accepted the manuscript on 4 August 2005. We subsequently decided to select the paper as a ‘featured article’ and solicited invited replies from experts in the field (including David Maehr, Gus Mills, Scott Creel and others). Through the course of this process, concerns were raised about the provenance of the data used by Pimm et al. (2006), leading to a protracted series of reviews, commentaries and investigations. In the editorial office, we considered the extent to which Pimm et al. (2006) used only publicly accessible information (i.e. a government report by Land et al., 2004) for their analyses. We satisfied ourselves that the information analyzed was appropriately gathered in a meta-analysis-type fashion; that is, individual animals, their relatedness and population structure could be gleaned from the databases publicly available. Following months of investigation, the editors concluded in January 2006 that the results were reliably based on data in the public domain. On this basis, we went ahead with the publication of the Pimm et al. (2006) paper. Conservation science proceeds by examining new questions and collecting necessary information that will better inform whether panthers and other threatened species are more viable when we as humans do meddle. We hope that the publication of these papers will further this science by allowing others to effectively address these issues, both by verification and by extension.

Anthropogenic activities are one of the main threats to species living in human‐dominated landscapes and can promote behavioral changes in birds. This paper presents a novel approach to test how a migratory species responds to habitat alterations occurring in nesting territories during winter, when the species is absent. From 2000 to 2016, we collated territory and nest monitoring data for the endangered Egyptian vulture Neophron percnopterus in the north of Spain, to test the effects of overwinter habitat alterations (OHA) around breeding territories on the species reproductive output. We monitored 70 different nest sites and observed OHA around the nesting area in 39 cases. Probability of switching to another nesting site almost tripled after OHA. Pairs that switched experienced substantially decreased breeding success and avoided reusing the nest for 4.8 ± 4.64 years. The presence of determinate landscape elements that provide screening, such as forest patches, increased nest reoccupancy probability after OHA by more than 0.3, to 0. 55 (compared to 0.24 when no screening was present). We also found that the distance and the situation of the OHA were critical factors explaining reproduction probabilities at nest sites. Our results demonstrate how OHA can strongly impact the breeding behavior of long‐lived species. This highlights the need to examine the long‐term impact of OHA rather than focusing only on disturbances during sensitive periods, as is often the case with habitual mitigation measures.

The trade and consumption of bushmeat are a major threat to biodiversity across the tropics. Conservationists have traditionally advocated for stricter regulation and enforcement as a way to control these practices, with less attention given to consumers and the management of the demand. Yet, it is clear that without adequately tackling demand, it is impossible to effectively curb the bushmeat trade. In this paper, we describe an intervention to reduce demand for bushmeat in northern Tanzania. The intervention was centered around the 1‐h radio show My Wildlife – My Community which included 15‐min episodes of the radio drama Temboni. Each episode of the radio drama was accompanied by a 45‐min interactive call‐in show featuring interviews with experts and local information about available community resources. We evaluated this intervention using a Before‐After‐Control‐Impact framework based on longitudinal data from 168 respondents. To account for the fact that respondents volunteered to be exposed to the intervention, in this case the radio show, we used a matching algorithm together with regression to ensure that we could build a credible counterfactual group. Our analysis did not uncover any differences in outcomes between the treatment and control groups, and thus no evidence of the intervention achieving its initial goals. One potential causal mechanism that could have led to this outcome is the low audience penetration rate. Fewer than 40% of respondents listened to the show and among those who did, only about 20% listened to five of more episodes. This research highlights the challenges of implementing and evaluating interventions delivered through mass media in developing countries, and the importance of reporting on interventions even when there is no evidence that they achieved their initial goals. Only through robust evaluation of behavior change interventions and the sharing of lessons learned can conservationists successfully tackle complex issues such as the bushmeat trade.

Chytridiomycosis is an emerging disease of amphibians with global impact (Skerratt et al., 2007). Innate immune defenses may contribute to the ability of some species to coexist with endemic Batrachochytrium dendrobatidis (Bd) (Retallick, McCallum & Speare, 2004; Woodhams et al., 2007b). We appreciate the thoughtful commentaries by Fisher (2007), Garner (2007) and Kurtz & Scharsack (2007) to the paper ‘Resistance to chytridiomycosis varies among amphibian species and is correlated with skin peptide defenses’ (Woodhams et al., 2007a). The commentators endorsed the ecological immunology approach of the work and its utility for informing conservation. They included excellent suggestions for future studies including the following: (1) Examining population, family and individual level variation in amphibian immune defenses. (2) Testing for mucosal antibodies, immune memory and induction of innate defenses in addition to constitutive innate defenses. (3) Determining how behavioral differences among species affect the rate of encountering the pathogen and response to encounter. (4) Understanding how environmental conditions affect innate immunity. The commentators pointed out that immunity can vary among lineages of frogs (Pearman & Garner, 2005) and stickleback fish (Rauch, Kalbe & Reusch, 2006), suggesting that some of the among-species differences we detected might be due to differences among lineages within species. This is true, however, as the work cited by the commentators illustrates, fully exploring variation in disease susceptibility within even a single species requires very large, complex experiments. For our initial exploration of variation among species, we chose to use single sibships to minimize withinspecies variation, maximizing the power of our comparisons among species. This also minimized the probability of chance differences in peptide expression between infected and control frogs. This was advantageous because we could collect peptides only from uninfected control frogs, because depletion caused by norepinephrine induction could have affected the survival of infected frogs. We agree that much remains to be investigated in terms of variation in innate immunity among sibships and populations; however, it seems likely that variation of skin peptide defenses among species is greater than within-species variation. For example, Apponyi et al. (2004) found substantial variation on a large geographic scale in the composition of skin peptides profiled within two species of Australian treefrogs, and also showed that most species produced profiles containing distinctive families of peptides, and that on smaller geographical scales profiles did not differ among individuals within species. Our own studies of peptide variation among individuals of two populations of Rana muscosa in the Sierra Nevada Mountains of California suggest very limited variation in skin peptides (Woodhams et al., 2007b). This suggests that sibship-level differences in skin peptide defenses are not likely to have affected the results of this study in which four species from the same region were compared. Some antimicrobial peptides are constitutive and others may be induced in response to infection (Boman, Nilsonn & Rasmuson, 1972; Zasloff, 2002; Cunliffe & Mahida, 2004; Izadpanah & Gallo, 2005). Here, peptide induction has two meanings: transcription and storage in glands or secretion from glands onto the skin. We did not test for induction of skin peptides in response to Bd exposure in this study. However, the constitutive peptide defenses may be a good measure of immune condition before Bd exposure. Any induction of new skin peptides that may occur in response to Bd infection may not be effective because chytridiomycosis leads to mortality of susceptible species between 18 and 48 days post-infection (Berger et al., 2004; Woodhams et al., 2007a), and production of peptides in granular glands after discharge can take as long as 6 to 421 days (Flucher et al., 1986; Giovannini et al., 1987; Rollins-Smith & Conlon, 2005). Of course the quantity of peptides released from the granular glands after administration of norepinephrine is

[Extract] The paucity of senior female scientists in science, including ecology and conservation, is a growing concern in the western world. 'Where are the women in ecology?' asked a paper in Frontiers in Ecology and the Environment recently (Martin, 2012). 'Women are driven out of research', conclude O'Brien & Hapgood (2012). Despite the increasing popularity of biology, including ecology, among female undergraduates and graduates, the proportion of female scientists in top positions remains low (European Commission, 2009; Martin, 2012; O'Brien & Hapgood, 2012; Adamo, 2013). An increasing number of individuals, institutions and governmental organizations are starting to ask why so many female scientists do not end up being employed in the type of occupations in which they were trained (Rosser, 2008; Hill, Corbett & St. Rose, 2010; Royal Society of Edinburgh, 2012). There are several reasons to address this issue, especially at a time of economic austerity in many countries. Losing trained scientists can represent a sunk cost: conservative estimates put the economic cost of a PhD in the US at c. $500 000 (Rosser, 2008), while each PhD student in the UK receives c. £100 000 from the government to cover stipends and research and training expenses (UKRC, 2012). Moreover, gender diversity is associated with indirect benefits; for example, commercial busi- nesses with gender-balanced staff and management tend to perform better financially (UKRC, 2010a).

The endemic pink pigeon has recovered from less than 20 birds in the mid-1970s to 355 free-living individuals in 2003. A major concern for the species' recovery has been the potential genetic problem of inbreeding. Captive pink pigeons bred for reintroduction were managed to maximise founder representation and minimise inbreeding. In this paper, we quantify the effect of inbreeding on survival and reproductive parameters in captive and wild populations and quantify DNA sequence variation in the mitochondrial d-loop region for pink pigeon founders. Inbreeding affected egg fertility, squab, juvenile and adult survival, but effects were strongest in highly inbred birds (F greater than or equal to 0.25). Inbreeding depression was more apparent in free-living birds where even moderate levels of inbreeding affected survival, although highly inbred birds were equally compromised in both captive and wild populations. Mitochondrial DNA haplotypic diversity in pink pigeon founders is low, suggesting that background inbreeding is contributing to low fertility and depressed productivity in this species, as well as comparable survival of some groups of non-inbred and nominally inbred birds. Management of wild populations has boosted population growth and may be required long-term to offset the negative effects of inbreeding depression and enhance the species' survival.

Predicting the threat of extinction aids efficient distribution of conservation resources. This paper utilises a comparative macroecological approach to investigate the threat of extinction in Neotropical birds. Data on ecological variables for 1708 species are analysed using stepwise regression to produce minimum adequate models, first using raw species values and then using independent contrasts (to control for phylogenetic effects). The models differ, suggesting phylogeny has significant effects. The raw species analysis reveals that number of zoogeographical regions occupied, elevational range and utilisation of specialised microhabitats were negatively associated with threat, while minimum elevation and body mass were positively associated, whereas the independent contrasts analysis only identifies zoogeographical regions as important. Confining the analysis to the 582 species restricted to a single zoogeographical region reveals elevational range and number of habitats occupied to be negatively correlated with threat whether the analysis is based on the raw data or on independent contrasts. Analysis of four contrasting zoogeographical regions highlights regional variation in the models. In two Andean regions the threat of extinction declines as the elevation range across which the species occurs increases. In the presence of substantial human populations on high Andean plateaus, a species with a greater elevational range may be more likely to persist at some (relatively) unsettled altitudes. In Central South America, the strongest predictor of threat is minimum elevation of occurrence: species with a lower minimum are less threatened. The minimum elevation result suggests that lowland species experiencing an ecological limit to their minimum elevation (min. elevation >0 m) may be more at risk than those not experiencing such a limit (min. elevation = 0 m). Finally, in southern Amazonia, where there is little altitudinal variation, the only weak predictors of threat are body size, larger species being more threatened, and number of habitats, species occupying more habitats being less threatened. These contrasting results emphasise the importance of undertaking extinction risk analyses at an appropriate geographical scale. Since the models explained only a low percentage of total variance in the data, the effects of human-mediated habitat disturbance across a wide range of habitats may be important.

Many species have strong habitat preferences that directly influence population viability. For successful reintroduction of threatened populations that rely on habitat structures, the correct placement of artificial structures is also important to population persistence. In this paper, we present a hierarchical approach to the problem of translocating animals that rely on permanent habitat structures, in which we first use population dynamics data to identify areas of suitable habitat, and then identify optimal configuration for habitat structures. We use data collected from a non-endangered, conspecific population of the endangered riparian woodrat (Neotoma fuscipes riparia) to examine the degree to which the distribution of dens in translocation sites might influence the likelihood that animals persist in their new environment. We characterize the habitats in which dens occur, analyze their spatial clustering, and compare them to temporal changes in population status for sex and age classes. We compare the potential efficacy of translocation efforts using spatial analysis versus solely habitat-based approaches and identify the optimal spatial configuration of dens that should be considered in this translocation effort. We found that patterns of habitat use were positively correlated with overstorey cover, and animal weight was positively correlated with understorey cover. Woodrats appear to select den locations on the basis of understorey cover, but also benefit from dense overstorey cover and distance to nearest tree. Our results suggest that in translocation efforts, artificial dens should be placed in clusters within a radius of 15 m, as values above and below this value showed negative correlations with body mass. Translocations should occur after reproductive events, which occur in April and August for woodrats in southern California. Our analyses provide practical guidelines in determining appropriate timing and spacing for translocation events in the context of animal condition, minimizing disease transmission, and reproduction.

Although the conservation of endangered species often implies the definition of priority areas for conservation, detailed information on their distribution patterns is seldom available over a large geographic range. The present paper explores the performance of an alternative data analysis approach, artificial neural networks, for assessing distribution patterns of endangered mammals when data are scarce and noisy. This approach was applied to identify wolf occupancy in Portugal based on information on wolf depredation and inquiries. Artificial neural networks were able to discriminate successfully between areas sporadically used by vagrant individuals and areas occupied by resident wolves, with a low estimated prediction error (around 1%). Only 33% of the wolf range is regularly occupied, being fragmented in five nuclei. These nuclei are surrounded by more disturbed marginal areas, with a source-sink dynamic between nucleus and marginal areas. Selection of priority areas for wolf conservation in the context of the Natura 2000 Network increased twofold the proportion of wolf regular range within protected areas. The good performance of artificial neural networks in assessing wolf distribution patterns suggests that this approach may be applied to other species where detailed records of distribution are limited.

A companion paper in this issue describes the mapping of Cape Sable seaside sparrow (Ammodramus maritimus mirabilis) habitat using satellite imagery. In general, those maps are correct. However, testing against an independent survey of the sparrow population does identify errors. Those errors fall into two categories, model errors and bird errors. The model errors result from genuine problems with the model. More interesting for conservation are the bird errors, which are more numerous. They are of two types. (1) Commission errors: some suitable habitat does not contain birds. These are areas where prior events have depleted sparrow numbers, but because the sparrows do not disperse long distances, they cannot occupy it quickly. (2) Omission errors: some birds remain in unsuitable habitat that was formerly suitable. This results from the sparrow's very high site fidelity. Often, they will not leave an area even when it is no longer suitable. The consequences of these bird errors are that some habitat and some birds are not contributing to the species' survival. Thus, an estimate of only the amount of habitat or only the sparrow population may present an overly optimistic view of the sparrow's plight.

Estimating the risk of extinction or decline requires estimates of the variability in vital rates, such as survival and fecundity. This paper describes a method for estimating variance of survivals and fecundities from census data. The method involves calculating an estimate of the variance in survival and fecundity due to demographic stochasticity and subtracting this estimate from an estimate of total variance. The method is demonstrated by applying it to the data set on acorn woodpeckers by Stacey & Taper (1992), and tested in a series of simulations. The results show that the method can estimate the environmental variance with no bias. The main advantages of this method are its simplicity, its independence of the underlying distribution, and its applicability to estimating the variance of fecundity as well as of survival.

To ensure the effective conservation of biodiversity the distribution of species needs to be accurately characterized and areas of high species richness located. For many taxa this can be achieved only by experienced taxonomists. Taxonomic research has a large input from non-professional or amateur researchers, in addition to professionals working at museums or universities. The decline of taxonomy and the number of taxonomists within the professional community has been widely publicized, but the trends in the activities of amateur taxonomists are unclear. Because amateurs contribute many valuable records of species occurrence this may have a disproportionate impact upon the information available for conservation planning and represents an underappreciated threat to conservation planning. We use taxonomic research by UK entomologists in order to evaluate the changing role of both amateur and professional taxonomists. We reviewed contributions by British-based authors to Entomologist's Monthly Magazine over the past century. Our results show that both amateur and professional taxonomy have undergone a long and persistent decline since the 1950s, in terms of both the number of contributors and the number of papers contributed. It is argued that the conservation community needs to help try and reverse the decline of taxonomy.

In a paper in this issue (Beresford et al., 2011), we set out to quantify the extent to which site-based conservation initiatives overlapped with the ranges of globally threatened bird species in Africa, using a GIS approach. We considered Protected Areas and Important Bird Areas (IBAs), both of which cover c. 7% of the land surface of Africa. The former, as noted by both Brooks & Matiku (2011) and Rodrigues (2011), carry significant governmental weight despite not necessarily being identified for their conservation value, while the latter (which are of conservation value to other taxa in addition to birds – Pain et al., 2005) have no legal standing, but are identified using objective criteria that include their value for globally threatened species (Fishpool & Evans, 2001). We find that Protected Areas perform poorly at protecting the ranges of Africa’s most threatened birds, covering just 13.9% on average, and even where a Protected Area overlaps part of an IBA, the part excluded from the Protected Area has higher value for these species than the protected part. As Rodrigues (2011) notes, the limitations of GIS studies that utilize Extent of Occurrence (EOO) maps are well documented. We therefore attempted to reduce the commission errors inherent in the use of EOO maps by analysing extent of suitable habitat (ESH: the area of potentially suitable vegetation types within the altitudinal preferences of the species (Rondinini, Stuart & Boitani, 2005; Buchanan et al., 2008)). While it is desirable to move beyond interpolation approaches based on EOO or ESH and utilize only presence/absence data derived from recent complete field surveys, these are not available for the majority of species. Even in the most well-watched countries, such as the United Kingdom, the most reliable information on the distribution of birds comes from systematic atlas surveys in which ranges are interpolated from sampled field counts which themselves record only a tiny proportion of the individuals present. At least for the foreseeable future therefore, we will have to continue to assess geographical distributions using interpolative approaches, whilst acknowledging the limitations of these methods. ESH estimates were considerably smaller than EOOs for most species, with a concomitant reduction in commission errors and increase in omission errors. Reducing commission errors reduces the likelihood that a species is assumed to be conserved elsewhere (Rodrigues & Gaston, 2001), and hence in our study, the danger of overestimating the coverage of species’ ranges by Protected Areas or IBAs. Omission errors, on the other hand, reduce the estimated overlap between sites and species’ ranges. Considering the information we presented, Rodrigues (2011) suggests that the ESH approach was not well able to distinguish between occupied and unoccupied areas of the EOO, because both approaches have similar ratios of true presences to total predicted presences. We agree that by focusing on other aspects of the analysis and by presenting only a broad overview of the accuracy of the ESH approach compared with EOOs (table 1; Beresford et al., 2011), we did not show the full advantage of the ESH method. A species-level analysis indicates that although the ESH comprises, on average, 28% of the EOO – hence reducing the chance of the species being assigned to areas where it did not occur – this did not make any difference to the accuracy of the estimated distribution for 74 species [47%, assessed by comparing the total numbers of errors (commission plus omission) using ESH vs. EOO]. Importantly, however, it did improve accuracy for 58 species (37%), decreasing commission errors by, on average, 11.8 IBAs per species. This was at the expense of a reduction in the accuracy of maps for 25 species, due to an increase in omission errors, although these averaged just 1.8 sites per species (Fig. 1). While considerable inaccuracies remain, especially in commission errors (table 1; Beresford et al., 2011), we suggest that the increase in accuracy for over a third of

Large carnivores are key visitor attractions in protected areas, but are difficult to see. Thus, supplementary feeding is sometimes used to attract them to viewing sites. Such intervention is contentious but its effects have rarely been examined. This paper analyses a case study of supplementary feeding in Komodo National Park, Indonesia. Using data from daily and annual Komodo dragon censuses, feeding records and financial accounts, the effects of feeding and its cessation on dragon numbers, tourist viewing opportunities and local community benefits were examined. Regular feeding caused dragon numbers to increase at the feeding site, but not year-round. Cessation of feeding caused numbers to decline again to natural levels. However, tourists were less likely to see dragons at the feeding site after cessation, and local community revenues declined with the loss of a market for goats. Solutions lie in finding less intrusive means for tourists to view dragons, and enabling local people to become involved in tourism through training, recruitment and the development of alternative markets.

Accurate long-term series of demographic data are available for most populations of southern elephant seals. However, research on the elephant seals of the Falkland Islands began only recently, and information for an accurate forecasting of the future of this population is lacking. In this paper we present data on the current status of the population and its trend in size during the last 11 years. We built an age-structured model of the population and analyzed the effect of variation in demographic parameters on population growth. Elasticity analysis demonstrated that variation in mortality has a more pronounced effect on instantaneous growth rate than an equivalent variation in fecundity. We examined the effects of environmental variability, inbreeding and catastrophes on population viability by a computer simulation using the VORTEX PVA programme. In the stochastic model, the most important factor affecting extinction risk was variability in mortality rates, in particular of the adult classes. We concluded that, although the population does not appear to be at immediate risk of extinction, its small size and isolation make it necessary to accurately monitor the population trends and to acquire additional information on life history and feeding strategies.

The decline of amphibian populations worldwide is a recognized phenomenon. A variety of important causes have been linked to the declines but perhaps the most renowned cause, in terms of popular press exposure, funding dispersal and papers published, is disease. In particular, chytridiomycosis, a disease caused by the amphibian chytrid fungus Batrachochytrium dendrobatidis (Bd), is a lethal emerging infectious disease that has had profound effects in some amphibian species, but occurs without catastrophic effects in others (Fisher, Garner & Walker, 2009b). Bd affects species differentially. In many cases it is the major threat to animals already stressed by other critical challenges, such as habitat destruction, pollution, invasive species and climate change. In addition to its impact on amphibian populations, the spread and presence of Bd is likely to impact human activities via effects on the pet, bait and food trade (Schloegel et al., 2009, 2010).

In their recent paper, Sato H Gharrett et al., 1999; McGinnity et al., 2003; Tallmon et al., 2004), but claimed the close geographical proximity, similar habitat and small genetic differences among remnant populations might considerably reduce this risk. In a recent review, Edmands (2007) stated that concerns over outbreeding should be taken seriously, as the effects of outbreeding can be, in some cases, as damaging as severe inbreeding. Here, we raise concerns about mixing of populations as a strategy to increase genetic variation, and we bring evidence of likely outbreeding depression observed in the rehabilitation project of the endangered freshwater salmonid marble trout Salmo marmoratus, started in Slovenia in 1993 (Crivelli et al., 2000). Since the beginning of the rehabilitation project, four marble trout populations went extinct due to landlsides (Predelica) and severe flood events (Gorska, Sventarska and Zakojska). The seven remnant populations of marble trout living in Slovenian streams are genetically distant (FST=0.66, Fumagalli et al., 2002), present a very low effective population size (for the Huda Grapa population a total population size ranging from 29 to 60 individuals was estimated between 2000 and 2007) and very low genetic variability estimated at 14 microsatellite loci (for Huda Grapa Ho=0.046, Fumagalli et al., 2002). In order to expand the genetic variability of marble trout living in the watershed, a new population (Gatsnik) was created in 1998 in a previously fishless stream by mixing the progeny of marble trout from two remnant populations from isolated streams with similar habitat (Trebuscica and Lipovesck). After two generations, the population of Gatsnik presented the typical features associated with maladaptation and outbreeding depression, that is lower annual survival rate (0.39 0.02 for trout born in the stream, higher than 0.55 for other marble trout populations living in the study area, Vincenzi et al., 2008), reproductive underperformance at the F2 generation and unusual morphological abnormalities (Fig. 1). Moreover, preliminary results from a egg-to-fry experiment performed in the fish farm for the F1 generation of Gatsnik showed 80.4% of mortality from fertilization to hatching (35.6 29.2% for the remnant populations) (D. Jesensek, pers. comm.). The translocation experiment in Gatsnik is still ongoing and may provide unequivocal evidence of outbreeding effects in 3–4 years. Nevertheless, despite the striking contrast of salmonid species in lifehistory traits may hinder generalizations, what we observed in the marble trout population of Gatsnik urges now further considerations on the opportunity of mixing populations of salmonids to increase their genetic variability. As recently reviewed by Willi et al. (2006), small populations are predicted to have low potential to adapt to environmental changes, as genetic variation and potential response to selection should be positively related to population size. Moreover, individuals in small populations may have lower fitness and be inbred, which may decrease an adaptive response to unpredictable environmental changes. Apart from the obvious cases of overexploitation, pollution and human-induced fragmentation, the small population size and low heterozygosity frequently observed in endemic

The current extinction crisis is caused primarily by human impacts upon wild populations. Large carnivores are especially sensitive to human activity; because their requirements often conflict with those of local people, predators have been actively persecuted in most regions of the world. In this paper, the impact of people upon predators is analysed by relating local carnivore extinctions to past and projected human population densities. There are strong associations between high human density and the loss of carnivore populations from a region. Interspecific variation in ability to survive at high human densities probably reflects species' ability to adapt to human-modified habitats. However, regional and temporal variation in individual species' sensitivity to human density is more likely to reflect the activities of local people than the phenotypes of local carnivores. Local culture, government policy and international trade all influence human attitudes to predators and, therefore, the impact of people upon carnivore populations. The importance of these factors may mean that extinction risks for carnivores will continue to increase, even though human population growth is projected to deccelerate during the new millennium. This points to an urgent need for techniques to resolve conflicts between people and predators at either the local or landscape level.

Carnivore translocations are usually risky and expensive, and a number of biological and non-biological factors can influence success. Biological considerations include knowledge of genetics, demographics, behavior, disease, and habitat requirements. This information is critical for determining if the translocation should be attempted, if it could be successful, and how it could be implemented in an efficient and effective manner. We stress that individual species will vary in their responses, and ideas should be tested scientifically. The technical considerations of translocation are closely related to the biological questions. They include legal framework, fiscal and intellectual resources, monitoring capacity, goals of the translocation, logistic challenges, and organizational structure of decision-making. We do not discuss socio-economic aspects of translocation because those challenges require detailed discussion in a separate paper. We suggest that because large carnivores often play key roles in regulating ecological interactions between trophic levels, restoring them is more than a single-species activity. By restoring carnivores in viable numbers, we can take a large step toward recovering ecological integrity of geographically extensive landscapes.

With increasing awareness of disease as an endangering process, an assessment of which pathogens might pose a problem and their patterns of infection in natural hosts is necessary. This paper examines the exposure of sympatric Ethiopian wolves (Canis simensis) and domestic dogs to canine distemper virus (CDV), canine adenovirus (CAV) and canine parvovirus (CPV) in the Bale region, Ethiopia and then relates these data to population trends of wolves. Wolves (n= 30) sampled between 1989 and 1992 had been exposed to CDV, CAV and CPV, but only CAV might be able to persist in this wolf population. Anecdotal and serological evidence suggested that an epidemic of CDV occurred in the dog population of the Bale Mountains National Park (BMNP) in 1992–93. All park dogs born since this time were seronegative to CDV, although some young dogs in the nearby urban population were seropositive. Despite evidence of CAV infection in wolves, none of the dogs sampled in the park were CAV seropositive, although this virus appeared highly seroprevalent and endemic in urban dogs. All dogs tested for CPV antibodies were seropositive. The BMNP wolf population declined in the late 1980s and early 1990s, with rabies responsible for a dramatic population reduction between 1990 and 1992. Although the population declined further up until 1995, it is not possible to assess whether the concurrent canine distemper epidemic in park dogs also affected wolves. Nevertheless, with evidence of rabies, CDV, CAV and CPV infections in sympatric domestic dogs and Ethiopian wolves, canid diseases clearly pose a significant threat to the future persistence of this Ethiopian wolf population.

Sequences from complete mitochondrial cytochrome b genes of 34 tigers support the hypothesis that Sumatran tigers are diagnostically distinct from mainland populations. None of the latter, including Bengals, Siberians, or Indochinese tigers, were found to have fixed diagnostic characters. Phylogenetic analysis of these sequences confirms these results. Within the framework of a phylogenetic species concept, current evidence thus supports the recognition of two distinct taxa, and within the context of this definition they could be ranked at the species level. This paper also documents a previously unrecognized nuclear insert of mitochondrial DNA that includes, minimally, mitochondrial homologues of a control region that lacks the feline mitochondrial repeat sequences, a complete cytochrome b gene, and complete tRNAThr and tRNAsurPro genes. In a phylogenetic analysis of the nuclear cytochrome b-like sequences and various feline mitochondrial sequences, the nuclear insert clusters with lion mitochondrial cytochrome b sequences, which suggests the insert is at least as old as the split between lions and tigers. The results of this study emphasize the importance of doing more to conserve Sumatran tigers. Because they are underrepresented in zoos relative to Bengals and Siberians, an effort should be made to increase captive breeding stocks of Sumatrans. That Sumatrans are a distinct taxonomic entity relative to mainland populations can be used in educational programs to increase conservation efforts within Indonesia.

The Ethiopian wolf (Canis simensis) is a very rare, endangered, endemic species surviving in isolated mountain pockets in the Ethiopian highlands, with nearly 50% of the global population living outside protected areas. In this paper we compare the ecology and behaviour of an Ethiopian wolf population living in Guassa, a communally managed area in the Central Highlands, with that of the Bale Mountains National Park in the Southern Highlands. Ethiopian wolves live at lower density in Guassa (0.2 ± 0.05/km2) than in the Bale Mountains, but giant molerats (Tachyoryctes macrocephalus), the main prey for Ethiopian wolves in Bale Mountains, do not occur in the Central Highlands. Faecal analysis identified nine prey categories across wet and dry seasons common to both populations. In total, rodents accounted for 88% of prey volume in wolf diets. Home-range size was positively related to pack size (r2 = 0.85) and there was no difference in mean home-range sizes in both areas. In Guassa, however, wolves spent less time in the presence than in the absence of humans, but wolves spent similar amounts of time in the presence and absence of cattle. These findings suggest wolves can cope with, or adapt to, the presence of livestock and people in communally managed areas. © 2005 The Zoological Society of London.

Every week, an enthusiastic student emails me, or walks through my open door, keen to do research in conservation. The vast majority of these students want to head into the field. This is wonderful; despite being a modeller, my interest and empathy with ecological systems arises from the direct observation of nature. However, when I explain that the majority of the people in our group analyze existing data, synthesize information across different studies, build models or, worse still, evaluate the outcomes of conservation interventions, the conversation invariably ends abruptly. These are not the sexy activities that our David Attenboroughloving students want – they are desk based and quantitative – what could they have to do with the real world? My proclamation that conservation needs more analysts, not more field data, invariably elicits a hostile reception among field ecologists. However, the past two decades have seen a growing trend towards synthesis and analysis. First, there was the inspirational creation of the National Centre for Ecological Analysis and Synthesis (NCEAS) that has transformed the discipline of ecology – the only funding rule – you can not use the money to collect more data. Second, there has been the rise of the evidence-based conservation movement (Pullin et al., 2004; Sutherland et al., 2004) where people systematically synthesize and disseminate existing knowledge about biodiversity interventions. Why is evaluation and analysis so important? First, if we do not know that we are achieving something through an action, why would we persist in taking that action? Ferraro & Pattanayak (2006) sent shock waves through the conservation community when they questioned the evidence for one of conservation’s foundational interventions – protected areas. Just a year earlier, the Millennium Ecosystem Assessment stated: ‘Few well-designed empirical analyses assess even the most common biodiversity conservation measures’ (Millennium Ecosystem Assessment, 2005). If we do not have assessments of past actions then decisions are being made in the dark (Cook, Hockings & Carter, 2010). Second, even if we know an action is beneficial to biodiversity, quantifying the biodiversity benefit of that action is essential for determining future investments (Murdoch et al., 2007) because costeffectiveness analysis demands quantification of benefits. It is remarkable how much valuable data goes unanalyzed. Two recent papers in Animal Conservation – Howe & Milner-Gulland (2012) and Walsh et al. (2012) provide excellent examples of how we can mine old information to inform future actions. Howe & Milner-Gulland (2012) have evaluated 100 final reports from Darwin Initiative projects carried out between 1997 and 2007. They show that there is consistency in results across three evaluation methods, and they uncover a mix of expected and unexpected effects. Similarly, Walsh et al. (2012) take over 30 years of data on the impact of fox Vulpes vulpes baiting on malleefowl Leipoa ocellata conservation only to discover there is no evidence that this popular intervention works. Surely there is a plethora of opportunities for similar analyses (Bottrill, Hockings & Possingham, 2011). The issue then is not, why were not these analyses done earlier, but why are not there a hundred more similar analyses every year? I return to my initial concern – how can we wean wellmeaning students from the desire to collect more data, which in conservation evaluation is challenging during a 3–5-year PhD project, to the analysis of existing data? Ironically, I have just returned from the NCEAS ‘wake’ in Santa Barbara – March 2012. While the NCEAS model is purportedly replicated in many places around the world, my limited experience so far with the children of NCEAS are that none are quite the same. History may show that NCEAS embodies both one of the US National Science Foundation’s best, and worst, decisions.

Statistical models that include random effects are becoming more common in population ecology all the time. Random effects are two or more effects of some grouping factor that belong together, in the sense that we can imagine they were generated by a common stochastic process. Thus, they are similar, but not identical, and therefore assumed to be exchangeable. Models with random effects contain as a description of this common stochastic process so-called prior distributions, with estimable hyperparameters, for instance the mean and the variance for normally distributed random effects. Models with random effects have intrinsically more than one level; therefore, they are often called hierarchical models (Royle & Dorazio, 2008) or multilevel models (Gelman & Hill, 2007). Interest in hierarchical models often focuses on the hyperparameters, but the realizations from the process described by the prior distributions, that is, the random effects, can also be estimated. The main difference to treating a set of effects as fixed is that in a random effects model, the parameters of a grouping factor are no longer estimated independently. Rather, the assumption that they come from a common prior distribution induces a dependence, which means that the estimate of each is somewhat influenced by the estimate of all other effects comprising that factor. This is often called borrowing strength from the ensemble. One consequence of treating a factor as random is that its effect estimates are pulled in towards the overall mean, which is called shrinkage in the literature. If the assumption of a common stochastic process is reasonable, borrowing strength and shrinkage typically leads to better estimates compared with their fixed effects counterparts (Gelman, 2005; chapter 4 in Kery & Schaub, 2012). Hierarchical models are not at all intrinsically Bayesian; rather, they can be analysed using likelihood or Bayesian methods (Royle & Dorazio, 2008), though Bayesian analysis is often easier, especially for ecologists using MCMC engines like WinBUGS (Lunn et al., 2000) or JAGS (Plummer, 2003). One particularly interesting use of hierarchical model is as a formal way of combining information, for instance, coming from different studies. A good example is the feature paper by Halstead et al. (2012), where estimates of survival probabilities of a snake species from radio-tracking data at multiple study sites in the same region are combined in a hierarchical model that is fitted using the WinBUGS software. Sample size at each site was fairly low. Hence, Halstead and colleagues treated each study site as a replicate and estimated the hyperparameters of the distribution that collects together the sitespecific parameters, which were treated as random effects. Hierarchical modelling thus provides a formal way of combining this information, where study sites with more information (e.g. more snakes or with more precise estimates) get more weight in contributing to the estimates of the overall mean and the among-site variability (the mean and the variance hyperparameters). Site-specific estimates are shrunken towards the mean, where the degree of shrinkage depends on the precision of the site-specific estimates. This is a desirable property, as it avoids possible overinterpretation of patterns that may be due to some idiosyncrasy of a small sample. Borrowing strength typically also helps against boundary estimates that are frequent in survival analysis with small sample size. Thus, not only were Halstead and colleagues able to obtain a formal estimate of the regional mean of the interesting parameters, but arguably they also obtained better site-specific estimates owing to the sharing of information among all sites. In a sense, what they did is a meta-analysis. A further advantage of combining data from several studies is increased precision of the site-specific estimates, because they borrow strength from the entire dataset. Increased precision means to have more power to detect an effect, which is especially advantageous when dealing with rare species, where sample size is typically small to very small. A similar approach was chosen by Papadatou et al. (2012) who estimated survival probabilities of several bat species (i.e. where species was bs_bs_banner

While there is great interest in the potential of local knowledge to inform quantitative conservation, there are no formal guidelines for its interpretation or analysis. Our work to quantify retrospective bias in fisher interviews and explore effects on IUCN Red List assessments shows that assumptions made when analyzing local knowledge affect conservation assessments (O’Donnell, Pajaro & Vincent, 2010). Aswani (2010), Daw (2010), and Patterson (2010) suggest several important ways to improve our understanding of local knowledge and its incorporation into conservation and management. We concur with Patterson (2010) who notes that the existence of bias does not render local knowledge useless. As part of ‘re-inventing fisheries management’ (Pitcher, Hart & Pauly, 1998) scientists must confront uncertainty by quantifying and including it in management recommendations (Ludwig, Hilborn & Walters, 1993). Fisher knowledge should be probed with the same analytical rigor as scientifically collected data, and to do so we need methods to quantify and correct bias in local knowledge. Daw suggests research to identify and quantify memory and perception bias that will improve our ability to accurately incorporate local knowledge into quantitative conservation assessments (2010). Psychologists have identified biases that can occur when we recall events from the past (e.g. Bradburn, Rips & Shevell, 1987), but there is currently no approach for quantifying those biases. As Daw (2010) suggests, there is a need for collaboration with psychologists to understand the nature and existence of retrospective bias along with a host of other memory and perception biases (table 1; Daw, 2010). Developing methods of quantifying these biases and/or procedures to correct for them, as we have explored here, would allow researchers to interpret interview responses appropriately, improving conservation assessments. A range of existing techniques should be utilized to help us collect and apply local knowledge as accurately as possible. There is a growing literature about how to design studies, phrase questions, and choose appropriate respondents (e.g. Neis et al., 1999a; Davis & Wagner, 2003; Anadon et al., 2009). Quantitative assessment protocols are being developed that explicitly incorporate local knowledge (e.g. Walmsley, Medley & Howard, 2005). It is necessary to acknowledge memory biases that may be present (Table 1; Daw, 2010) and consider the motivations of respondents to overor underreport (reviewed by Yasue, Kaufman & Vincent, 2010). Comparing multiple sources of local knowledge (‘local’ data-collection methods reviewed by Danielsen, Burgess & Balmford, 2005) with scientific observations can help determine the strengths and weaknesses of both approaches. As well, assumptions about biases should be made explicit in research papers and researchers should experiment with metrics to account for biases and evaluate effects on results. As both Aswani (2010) and Daw (2010) point out, the importance of understanding the perceptions of resource users extends beyond filling data gaps in single-species monitoring. Aswani reminds us that the appropriate application of local knowledge requires a broader understanding of community, beliefs, and economic context of that knowledge (Neis et al., 1999a). For example, our study (O’Donnell et al., 2010) is part of a larger research project to evaluate options for seahorse conservation and management action. We have consulted with a range of stakeholders and experts from fishers to scientists about preferred management options (Martin-Smith et al., 2004). We are currently investigating the spatial and behavioral dynamics of fishing, noted by Aswani (2010) as being necessary for moving beyond managing for single species to a more comprehensive socialecological system. We also agree with Daw’s (2010) suggestion that a better understanding of how humans perceive and recall environmental change can help us understand conflicts in conservation and resource governance. Conflict can arise when there is a gap between fishers’ perceptions and scientific knowledge used to inform management of fishing. From another study conducted in our focal community, Handumon, we

Our study (Sberze et al., 2010) combined three personal contributions: a professor’s (M. C.-H.) experience in neotropical ornithology, a post-doc’s (G. F.) practice with occupancy estimation, and a graduate student’s (M. S.) interest in nocturnal birds. The work was developed in the Biological Dynamics of Forest Fragments Project (BDFFP), a 30-year landscape manipulation experiment focused on biological change in artificially isolated forest patches created by forest clearing in the early 1980s (Laurance et al., 2002). Circumstances beyond control of the BDFFP led to gradual abandonment of cleared land and subsequent growth of c. 30 km of secondary forest (SF) surrounded by old growth (OG) and embedding the experimental patches. Our occupancy study of nocturnal birds was at once an effort to take advantage of unplanned secondary growth (as accurately pointed out in Sekercioglu’s commentary) and a contribution to the ongoing debate about the conservation value of SF (Laurance & Wright, 2009). We are grateful to the editors of Animal Conservation for featuring our paper and to Ken Feeley, Cagan Sekercioglu and Daisy Dent for their kind and insightful commentaries. We aim this reply at clarifying some of our methodological choices and identifying two avenues for future progress. We believe that our study convincingly illustrates the potential conservation relevance of SF thanks to a set of simple methodological choices. Both abundance and occupancy estimates provide biologically meaningful quantifications of population state. Abundance estimates, however, require mark–recapture or counts of individuals at measurable distances from the observer (Williams, Nichols & Conroy, 2002). Our focus on occupancy was in part due to the difficulties of counting nocturnal birds mentioned by Sekercioglu (2010). At the same time, the occupancy approach provided a sound conceptual basis for comparing different types of land cover and allowed us to replicate sampling at a large and reasonably representative spatial extent. While attempting to cover as much ground as possible on foot, MS had to limit the time per sampling point. This temporal constraint reduced the probability of detecting species at each visit, increasing the importance of explicitly accounting for detection failure in the analyses, which we did. Also due to walking constraints, sampling points had to be relatively close to each other. We addressed this problem by quantifying the influence of occupancy in neighboring points upon occupancy at each focal point (spatial autocorrelation). Since we did not find strong evidence of spatial autocorrelation, the procedure gave us confidence that our sampling points are not pseudo-replicates. The spatial analysis was our strongest reason to use a Bayesian framework, although Feeley (2010) correctly suggests that we could have further explored the framework by using informative priors. We chose not to do this because we were interested in finding out whether our data revealed a priori expectations despite the use of priors that did not convey those expectations. Although most of our predictions were supported by the data, we were surprised by the general preference for SF over OG among owl species. We suspect that relatively high abundance of rodents in SF may attract owls, but we have no data to back up this explanation. We agree with Feeley (2010) that birds in SF may be in a demographic sink, that is a place where reproduction fails to keep pace with local mortality (Pulliam, 1988). This should inspire caution in the interpretation of results and motivate further investment in life-history study, a need that all commentaries have identified. Dent (2010) is particularly clear about this point, emphasizing how life-history information will help generalizations about what types of species are more or less vulnerable to landscape change. We believe that multi-taxon occupancy studies backed with detailed life-history information should be a major avenue of future progress (Gardner et al., 2009). The challenge resides in how much rigor will be required of life-history assessments. Life history studies are field-intensive and must be conducted one species at a time: more rigor will mean smaller sets of species and more limited contribution to generalization. Progress on this front requires a combination of smart data-mining, conceptual simplicity, and analytical sophistication. Ideally, this should be done without forgetting phylogenies, because species often share more than the life-history traits that we decide to measure (Harvey, 1996; Owens & Bennett, 2000). Relentless effort in life-history study should be complemented with further development of occupancy studies. This second avenue of progress presents analytical challenges but relatively little additional fieldwork. Recent work by Zipkin, DeWan & Royle (2009) illustrates how multi-taxon models of occupancy developed in a Bayesian framework make it possible to borrow information across species and thus produce inferences about species for which there are relatively few data. Perhaps even more important for the SF–OG comparisons is the infusion of a dynamic perspective in occupancy studies (MacKenzie et al., 2003). Our nocturnal bird study treats OG and SF as static environments for which we estimate average site-occupancy values. It would be desirable to estimate not only the occupancy at a

The subsidization of predators occurs when humans directly or indirectly alter resource availability in such a way as to increase the density of a predator population above levels that would occur without the additional resources. Subsidized predators can drastically impact prey populations because subsidies insulate the predator populations from the effects of declines in prey populations (Sinclair et al., 1998). Such scenarios become problematic for conservation practitioners when the enhanced predator population influences species of conservation concern. Examples of this are widespread: subsidized ravens influence tortoise ecology (Boarman, 2003); subsidized predatory beetles influence insect herbivore populations (Rand & Louda, 2006); subsidized mountain lions influence bighorn sheep demographics (Rominger et al., 2004); and in the example presented by Shapira, Sultan & Shanas (2008), subsidized red foxes Vulpes vulpes influence gerbil ecology. Shapira et al.’s fox-gerbil example is particularly interesting because the influence of foxes is best seen not in altered densities of gerbils per se, but in the behavior of these taxa. In areas near farms, the availability of crops and favorable habitat apparently subsidize foxes. As a result, in areas where foxes are putatively more abundant, gerbils are not only less abundant but are also less active, and show decreased foraging efficiency. This suggests that foxes in this agricultural region have altered gerbil biology not by direct predation alone, but also by creating a landscape of fear (Brown, Laundre & Gurung, 1999; Laundre, Hernandez & Altendorf, 2001), from which the gerbils have somehow rendered a microhabitat-level map of the perceived risk of predation. It is harder, however, to address the effect of actual predation on gerbils as Shapira et al. do not present information on the diet of foxes. Indeed, the differential activity rates and behaviors of gerbils at sites near and far from foxes are potentially confounded by several concerns. Foxes are not uniformly distributed but are found near farms. While this is, of course, a primary point of Shapira et al.’s paper, it nonetheless means that many of the factors making the farmed areas so different from the more natural surrounding areas are also presumably correlated with fox distributions and may also influence gerbil biology. In addition, which of the two gerbil species, or the ratio of the two species, that is being examined at each study site is not clear. This is important, because if the two gerbil species differ in their foraging and behavior, then observed differences in gerbil ecology and behavior that are attributed to foxes may in fact be a function of which gerbil species is visiting the seed tray. Finally, the foraging behavior of the predator itself may differ across study sites. Shapira et al. did not study the ecology of the foxes, but we know that predator behavior is not static. Just as prey alter their behavior to address perceived risks of predation, predator behavior also changes as a function of prey density and behavior (Lima, 2002; Quinn & Cresswell, 2004). Nonetheless, differential intensity of fox use of the landscape does appear to be the most parsimonious explanation for the observed patterns of gerbil ecology, as these patterns match our understanding of how subsidized predators and their prey should interact. In a world in which anthropogenic landscape change is a seemingly ubiquitous concern for conservationists, can we predict a prioriwhere and which predators are likely to become problematic in such situations, and can we make recommendations as we attempt to identify solutions? Although a full examination of these two questions is beyond the scope of this commentary, we know enough about the ecology of predators, of prey and of their intraand interguild interactions to make some generalizations that may carry predictive power and applied relevance. For instance, we know that larger carnivores may suppress populations of other guild members in the same way in which predators influence prey: through predation and by creating landscapes of fear (Linnell & Strand, 2000; Creel, Spong & Creel, 2001). At the Israeli–Jordanian study sites, the generalist and mid-sized red fox apparently outcompetes