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Can we save large carnivores without losing large
carnivore science?
Running title: Saving large carnivore science
Benjamin L. Allen1*, Lee R. Allen2, Henrik Andrén3, Guy Ballard4, Luigi Boitani5, Richard
M. Engeman6, Peter J.S. Fleming7, Adam T. Ford8, Peter M. Haswell9, Rafał Kowalczyk10,
John D.C. Linnell11, L. David Mech12, Daniel M. Parker13
Author affiliations:
1University of Southern Queensland, Institute for Agriculture and the Environment,
Toowoomba, Queensland 4350, Australia. Email: benjamin.allen@usq.edu.au
2Robert Wicks Pest Animal Research Centre, Biosecurity Queensland, Queensland
Department of Agriculture and Fisheries, Toowoomba, Queensland 4350, Australia. Email:
lee.allen@daf.qld.gov.au
3Grimsö Wildlife Research Station, Department of Ecology, Swedish University of
Agricultural Sciences (SLU), SE–73091 Riddarhyttan, Sweden. Email:
Henrik.Andren@slu.se
4Vertebrate Pest Research Unit, New South Wales Department of Primary Industries, The
University of New England, Armidale, New South Wales 2351, Australia. Email:
guy.ballard@dpi.nsw.gov.au
5Department of Biology and Biotechnologies, University of Rome, Sapienza 00185 Rome,
Italy. Email: luigi.boitani@uniroma1.it
6National Wildlife Research Centre, US Department of Agriculture, Fort Collins, CO 80521-
2154, United States of America. Email: Richard.M.Engeman@aphis.usda.gov
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7Vertebrate Pest Research Unit, New South Wales Department of Primary Industries, Orange,
New South Wales 2800, Australia.Email: peter.fleming@dpi.nsw.gov.au
8Department of Biology, University of British Columbia, Kelowna, Canada V1V 1V7. Email:
adam.ford@ubc.ca
9School of Biological Sciences, Bangor University, Bangor, Gwynedd, LL57 2DG, United
Kingdom. Email: p.m.haswell@bangor.ac.uk
10Mammal Research Institute, Polish Academy of Sciences, 17-230 Białowieża, Poland.
Email: rkowal@ibs.bialowieza.pl
11Norwegian Institute for Nature Research, PO Box 5685, Sluppen, NO-7485 Trondheim,
Norway. Email: John.Linnell@nina.no
12Northern Prairie Wildlife Research Centre, United States Geological Survey, 8711 -37th
Street, SE, Jamestown, North Dakota 58401-7317, USA. Email: david_mech@usgs.gov
13School of Biology and Environmental Sciences, University of Mpumalanga, Nelspruit 1200,
South Africa.Email: Daniel.Parker@ump.ac.za
*Corresponding author.
Abstract
Large carnivores are depicted to shape entire ecosystems through top-down processes.
Studies describing these processes are often used to support interventionist wildlife
management practices, including carnivore reintroduction or lethal control programs.
Unfortunately, there is an increasing tendency to ignore, disregard or devalue fundamental
principles of the scientific method when communicating the reliability of current evidence for
the ecological roles that large carnivores may play, eroding public confidence in large
carnivore science and scientists. Here, we discuss six interrelated issues that currently
undermine the reliability of the available literature on the ecological roles of large carnivores:
(1) the overall paucity of available data, (2) reliability of carnivore population sampling
techniques, (3) general disregard for alternative hypotheses to top-down forcing, (4) lack of
applied science studies, (5) frequent use of logical fallacies, and (6) generalisation of results
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from relatively pristine systems to those substantially altered by humans. We first describe
how widespread these issues are, and given this, show, for example, that evidence for the
roles of wolves (Canis lupus) and dingoes (Canis lupus dingo) in initiating trophic cascades
is not as strong as is often claimed. Managers and policy makers should exercise caution
when relying on this literature to inform wildlife management decisions. We emphasise the
value of manipulative experiments, and discuss the role of scientific knowledge in the
decision-making process. We hope that the issues we raise here prompt deeper consideration
of actual evidence, leading towards an improvement in both the rigour and communication of
large carnivore science.
Keywords: apex predator; behaviourally-mediated trophic cascades; adaptive management;
experimental design; mesopredator release hypothesis; science denial
Introduction
Large carnivores are some of the most charismatic and ecologically-influential organisms on
Earth. Through their interactions with other animals, large carnivores may affect faunal and
floral communities across multiple trophic levels (Darwin 1859; Leopold 1949; Hairston et
al. 1960 ). This process is known as a trophic cascade (Paine 1980), and is a concept now
fully entrenched amongst ecologists, conservation biologists and many land and wildlife
managers.
Seldom have such novel ecological concepts been so rapidly mainstreamed to the extent that
they are identified as one of the 20 most influential topics in biodiversity conservation
(Bradshaw et al. 2011 ). Yet the ‘mesopredator release hypothesis’ (MRH) and its cousins the
‘large-carnivore control-induced trophic cascade hypothesis’ (TCH) and the ‘behaviourally-
mediated trophic cascade hypothesis’ (BMTCH) have done exactly that, so much so that
these concepts are now routinely advanced as scientific and moral justification for what are
essentially highly normative standpoints concerning desired conservation outcomes.
Inherently value-laden, religious terms are now frequently used in academic discourses about
the ecological roles of large carnivores – terms such as hero, doctrine, dogma, demonising,
virtuous, saviour, scapegoat, sanctification, sinners and saints (e.g. Jones 2002; Soulé et al.
2005; Anahita and Mix 2006; Allen et al. 2011a ; Letnic et al. 2011 ; Mech 2012; Chapron and
Lopez-Bao 2014; Middleton 2014; Johnson and Wallach 2016). Unfortunately, but perhaps
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motivated by the dire status of many carnivore populations, a growing number of studies rely
on weak inference to assess the roles of large carnivores in ecosystems (e.g. Allen et al.
2013b; Ford and Goheen 2015). Such practices might stimulate short-term gains in carnivore
conservation and motivate some segments of the public to care about it, but these
communication practices risk undermining long-term confidence in large carnivore science
and scientists (Fleming et al. 2012 ; Sarewitz 2012; Middleton 2014). The actual science of
large carnivore science is getting lost, being replaced by catch phrases, slogans, sound bites,
YouTube clips, fake news and post-truth politics, or the simplification and popularisation of
unsubstantiated or unreliable theories and hypotheses. This tension between scientific rigour
and pursuit of quick conservation gain raises the critical question: can ecologists save large
carnivores without losing large carnivore science?
As described in several studies (summarised, for example, in Crooks and Soulé 1999;
Hayward and Somers 2009; Terborgh and Estes 2010; Eisenberg 2011; Estes et al. 2011 ;
Ritchie et al. 2012 ; Ripple et al. 2014b ; but for a clear definition see Ripple et al. 2016b ), the
core theoretical processes associated with the MRH, TCH and BMTCH are:
1. Mesopredators and herbivores induce declines in smaller fauna and flora,
2. Large carnivores induce declines in mesopredators and herbivores,
3. Lethal control, harvest or hunting of large carnivores by humans induces declines in
large carnivores, increases in mesopredators and herbivores, and ultimately causes
undesirable outcomes for biodiversity and ecosystems,
4. Cessation of large carnivore control, harvest or hunting and/or active large carnivore
encouragement, including reintroduction, induces declines in mesopredators and
herbivores, which ultimately causes desirable outcomes for biodiversity and
ecosystems, and
5. Documentation of the MRH, TCH and BMTCH in some studies has been common
enough that these processes should be considered universal across ecosystem types
and independent of carnivore size or phylogeny.
The way these theories have been mainstreamed are perhaps best encapsulated in the short
online video titled How wolves change rivers (Sustainable Human 2014), which has been
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viewed over 18 million times since early 2014, but which does not consider the contrary (and
often superior) evidence for the processes it claims. Proponents of the MRH, TCH, and
BMTCH argue that these hypotheses should be accepted by scientists and society as
ecological laws by default (not as mere theories or hypotheses) and that the burden of proof
for demonstrating their reality should be placed on those who do not believe them (Estes et
al. 2011 ). These theories also provide the scientific justification for many admirable and
worthwhile efforts to restore large carnivore populations to densities and distributions
reminiscent of former times (Ripple et al. 2014b ; Ripple et al. 2016a ), although historical
ecological benchmarks have not been determined for most systems (e.g. Hayward 2012).
Nevertheless, the worldwide influence of the MRH, TCH and BMTCH have been enormous
(Bradshaw et al. 2011 ). In spite of the perceived universality of top-down control of
ecosystems however, there is a large and growing number of large carnivore studies
indicating that such effects are highly context specific and that many of the most rigorous
studies failed to document evidence of trophic cascades (Tables 1–3).
INSERT TABLE 1
INSERT TABLE 2
INSERT TABLE 3
In this brief overview, we summarise six key issues weakening the strength of the available
literature and undermining scientific advancement on understanding large carnivores’
ecological roles. We focus our discussion on grey wolves (Canis lupus) and Australian
dingoes (Canis lupus dingo), which have been claimed to be the only two terrestrial
carnivores for which both the MRH and TCH have been demonstrated (Figure S2 in Ripple
et al. 2014b ). Our aim is not to denigrate these or other large carnivores, decrease interest in
them, diminish the motivation to conserve them, or hinder the pursuit of scientific knowledge
in this field. On the contrary, our aim is to outline the primary issues weakening the reliability
of research on MRH, TCH and BMTCH, to show why wildlife managers and policy makers
should exercise caution when making decisions based on the currently available literature
describing these processes. We agree with many authors that top-down forcing can occur and
that large carnivores can have important ecological roles. However, there are enormous gaps
in our understanding of when and where such effects will occur in most systems. Articulating
the truth about the reliability (or lack thereof) of large carnivore science is, in and of itself, a
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strong conservation message: it is far better to err on the side of caution and preserve large
carnivores in the first place than to falsely believe ecosystems can be quickly and easily
fixed, restored or rewilded by simply bringing some carnivores back (Glen et al. 2007 ;
Marshall et al. 2016 ). We further offer suggestions for overcoming these issues with the hope
that future large carnivore studies will avoid them and better contribute to the evidence-base
needed for the management and conservation of large carnivores and sympatric species.
Issues that weaken the available literature supporting the MRH, TCH and BMTCH
1. There is not enough evidence of any kind, reliable or otherwise
A general understanding of large carnivores’ roles is only beginning to emerge, and much
more work is needed before we can confidently claim what those roles are or the ecological
contexts that shape these roles. Large carnivores unquestionably have ecological effects or
impacts of some description. In principle, every individual animal eaten by a carnivore
represents an impact – the prey animal is dead or scared, the prey’s population growth or
foraging is slowed, scavengers scavenge, decomposers decompose, nutrients enter the soil,
life for the prey’s competitor is now a little easier, the vegetation that would have been
consumed by the prey survives a little longer, and the carnivore lives to kill another day.
Whether the death of that prey animal is a good or bad thing (or not) depends on the
perspective of which animal is favoured over another (Allen et al. 2011b ; Mech 2012) – there
are winners and losers to every interaction (Flagel et al. 2016 ). These interactions all have a
value, contributing to the building blocks of wider ecological and demographic processes,
and evolutionary selection pressures (Darwin 1859; Hairston et al. 1960 ; Kershaw 1969;
Barbosa and Castellanos 2005; Krebs 2008; Molles 2012). But do these individual-level
impacts of a relatively small magnitude combine and accumulate to produce detectable
cascading impacts of a large magnitude on populations and whole ecosystems? Are these
carnivore effects stronger or more important at shaping systems than bottom-up processes?
Can individual carnivores regulate entire food webs? Do carnivore effects always produce net
benefits to biodiversity? Are positive carnivore effects universal across ecosystems and
apparent across all trophic levels?
In spite of claims for the universality of trophic cascades and a concomitant shift in the
burden of proof to disprove top-down forcing and prove bottom-up forcing (Terborgh and
Estes 2010; Estes et al. 2011 ), Haswell et al. (2017) shows that at best, detectably large
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cascading effects of top-predators (from a wide range of taxonomic groups) are the exception
and not the rule. Indeed, several studies using strongly-inferential methods demonstrate that
such top-down effects do not always occur, or if they do, they are far weaker than bottom-up
processes (e.g.Gasaway et al. 1983 ; Boertje et al. 1996 ; Hayes et al. 2003 ; Vucetich and
Peterson 2004; Vucetich et al. 2005 ; Brodie and Giordano 2013; Marshall et al. 2013 ; Allen
et al. 2014b ; Ford et al. 2015a ; Sivy 2015; see also Schmitz et al. 2000 ; Bowyer et al. 2005 ;
Sergio et al. 2008 ; McCoy et al. 2012 ; White 2013; McPeek 2014; Kuijper et al. In press ).
Ford and Goheen (2015) showed that of five strongly-inferential experiments investigating
large carnivores’ roles, only two found evidence supporting the TCH. Morgan et al. (2017)
highlight the supremacy of bottom-up processes and articulate the folly of attempting to shoe-
horn or apply outcomes from one ecological context into another. Recent global reviews of
the topic have also reported that ‘little is known’ about 24 of the 31 species of the world’s
largest carnivores, as ecologists are only just beginning to discover their ecological functions
(Ripple et al. 2014b ); or put another way, the MRH, TCH and BMTCH have not yet been
shown for at least 77% of large carnivores. Hence, we do not yet know what the ecological
functions of large carnivores are, and what we do know is from a minority of species in an
even smaller minority of biomes. While these hypotheses might eventually be applied to, and
supported in, a wide number of food webs, evidence supporting these hypotheses are
currently quite restricted.
Ripple et al. (2014b) claim that both the MRH and the TCH have been demonstrated only for
two related species, grey wolves and Australian dingoes, but the evidence-base for these two
species is very limited. In the case of dingoes, the total number of field studies on their
ecological roles is just a few dozen. Of these studies, all but four are observational or
correlative studies conducted in small areas (i.e. a few hundred km2) and over only a few
days (Allen et al. 2013b ; Allen et al. 2015 ). Drawing on this limited pool of empirical data,
the 22 literature reviews of dingoes’ ecological roles produced over the last 10 years have
unavoidably borrowed heavily from each other in what might be called citation inbreeding
(Allen et al. 2014c ). Thus, there is not a growing body of reliable evidence for dingoes’
ecological roles at all, but merely a growing body of largely recycled literature (Table 1; see
also Allen et al. 2011b ). Evidence for the ecological roles of wolves is much stronger than
dingoes, but is still frequently challenged and often found unreliable for similar reasons
(Tables 2 and 3; see also Winnie and Creel 2017). The combination of mixed-outcomes when
testing the MRH, TCH and BMTCH and the absence of studies on most species of large
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carnivore warrants far greater circumspection than is often afforded in syntheses of carnivore
ecology and conservation.
2. Sampling methods for carnivores are often unreliable
Studies measuring the effects of large carnivores roles typically correlate some change or
difference within an ecosystem to some change or difference in carnivore abundance (Ford
and Goheen 2015 ). But such approaches are frequently challenged because of their lack of
rigour (Tables 1–3). These challenges usually fall into three main categories of complaint:
experimental design constraints (e.g. manipulative experiments vs correlations or
observations; alternative hypotheses), predator sampling strategies (e.g. tracking plots,
camera traps, direct observations, movement data etc.), and data analysis approaches (e.g.
indices, occupancy modelling, statistical assumption violations, exclusion/inclusion of
outliers or contradictory data etc.). Counting or indexing carnivore populations can be
difficult and is often associated with large confidence intervals, but analytical methods do
exist to detect broad differences (e.g. Kershaw 1969; Caughley 1980; Underwood 1997; Zar
1999; Quinn and Keough 2002; Krebs 2008; Engeman et al. 2017 ). Unfortunately, many
studies use carnivore sampling methods that are incapable of yielding reliable data on
carnivore abundance, let alone actual rates of predation or perception of risk by prey animals.
The absence of these data undermines evidence for the proposed link between variation in
carnivore abundance and other reported changes and/or differences in the ecosystem.
Studies concluding that dingoes trigger trophic cascades are derived from non-validated and
often confounded comparisons of population indices between habitats, season, and/or species
(Allen et al. 2011a ; Allen 2012b). Ways to validate some common sampling methods have
been developed (Allen and Engeman 2014). When their methods are scrutinised, the results
of the most oft-cited works are unreliable (Allen et al. 2014c ). Even the results of the best
available manipulative experiments are sometimes contested on grounds that the predator
sampling methods are unreliable (e.g. Table 1).
There is unlikely to ever be any one perfect predator sampling method that suits all
applications, so the use of different sampling techniques and analytical methods across
studies is not particularly concerning. It does not matter if carnivores are sampled using sand
plots, camera traps, snow tracking, GPS collaring, direct observations, or remote sensing (for
example) provided the data are subsequently handled and analysed appropriately. We argue
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instead, that it is important to ensure that that whatever the implicit assumptions of the
methods are, that they are justifiable for the context under which the study was conducted
(Engeman 2005; Allen and Engeman 2014). The weaknesses and limitations of these survey
methods need to be openly acknowledged and discussed – not only in the peer-reviewed
manuscript, but also in the subsequent public discourse. This is where many previous studies
have erred (Table 1), and where improvements must be made if science is to acquire less
ambiguous evidence to support the MRH, TCH or BMTCH (Hayward et al. 2015 ).
3. Alternative hypotheses are seldom tested
Carnivores are just one of many potential causal agents operating in ecosystems (Vucetich et
al. 2005 ; Middleton 2014; Peterson et al. 2014 ; Ford and Goheen 2015). Yet for many studies
claiming support for the MRH, TCH and BMTCH, the study framework is designed to create
evidence for these hypotheses rather than being designed so that evidence for plausible
alternative hypotheses is both tested and compared at the same time (Winnie 2014). Studies
investigating these hypotheses commonly focus on competition, predation/removal or risk of
predation (Tables 1–3). But there are many more interaction types besides these within food
webs, which interaction types can also be strong and often do not conform to simple
expectations (Muhly et al. 2013 ; Saggiomo et al. 2017 ). Invertebrate (Meadows et al. 2017 )
and theoretical (e.g. Finke and Denno 2004; Holt and Huxel 2007; McCoy et al. 2012 ;
McPeek 2014; Kendall 2015) studies highlight many different outcomes of predator removal
or addition, most of which have received little attention in the wider large carnivore literature
(Fleming et al. 2012 ; Mech 2012; Ford and Goheen 2015; Haswell et al. 2017 ). The
consequence of not investigating plausible alternative explanations is that management
actions may completely overlook key processes contributing to declines of fauna (e.g. Allen
2011; Middleton et al. 2013b ; Cooke and Soriguer 2017), and they cannot discover these
processes because the study framework simply corroborates a narrow set of a priori
hypotheses without looking for others.
A clear example of the systemic failure to evaluate alternative hypotheses and ignore contrary
data comes from a series of studies conducted in the Greater Yellowstone Ecosystem, USA
(Winnie 2014). Environmental changes following the restoration of wolves to Yellowstone
National Park are often given as a clear example of the beneficial effects of restoring large
carnivores to ecosystems (Table 2), but there are alternative hypotheses to explain many of
the observed changes (Vucetich et al. 2005 ; Marshall et al. 2013 ; Middleton et al. 2013b ).
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There is strong evidence that wolves alone are not responsible for all the changes attributed to
them (Mech 2012; Winnie and Creel 2017). Many other important changes to the
Yellowstone system occurred around the same time as wolf restoration, and ‘when we tell the
wolf story, we get the Yellowstone story wrong’ (Middleton 2014). Using data from 1961 to
2004, Vucetich et al. (2005) investigated the TCH and showed that changes in climate and
harvest rate are justified explanations for most of the observed decline in Yellowstone elk,
rather than heightened predation by wolves. Indeed, wolf predation was determined to be
compensatory to existing rates of mortality (e.g. from starvation or mortality from other
predators). In addition, early studies on the BMTCH reported that wolves scared herbivores
away from riparian areas, which reduced herbivory on trees and ultimately caused increased
tree growth (Ripple and Beschta 2004; Beschta and Ripple 2007). Not only did these earlier
studies incorrectly identify areas of high predation risk (Creel et al. 2005 ; Kauffman et al.
2007; Kauffman et al. 2010 ; Winnie 2012), but they also failed to consider more
parsimonious explanations for increased tree growth in riparian areas, such as the height of
the local water table (Bilyeu et al. 2008 ; Kauffman et al. 2013 ). MacNulty et al. (2016; pg.
27) summarise the present situation when they state that ‘scientific consensus about the role
of wolves in driving [trophic cascades] has yet to emerge, despite 20 years of research by
numerous federal, state and academic investigators’, and that the ‘overarching reason for the
impasse’ is the experimental design constraints on the Yellowstone wolf reintroduction
program. In other words, the lack of rigour and strong inference in testing the MRH, TCH
and BMTCH has generated the controversy over the role of wolves in restoring this
Yellowstone landscape.
In Australia, snap-shot studies comparing fauna abundances in adjacent areas separated by
predator-proof fences are commonly used to highlight the greater amount of biodiversity
present on the side of the fence with a greater number of dingoes (e.g. Letnic et al. 2009 ;
Fillios et al. 2010 ; Letnic and Koch 2010; Brawata and Neeman 2011; Gordon et al. 2017a ).
However, the relative abundances of dingoes is not the only important difference between the
two sides of these fences (e.g. Newsome et al. 2001 ; Fitzsimmons 2007; Allen 2011). A range
of important geological and biophysical differences are also present, not the least of which
are the markedly different herbivore types, densities, and land-use histories, which are also
well-known to structure fauna communities through grazing-induced habitat changes
independent of dingoes or other predators (Tiver and Andrew 1997; Williams and Price 2010;
Parsons et al. 2012 ; Howland et al. 2014 ; Koerner and Collins 2014). The cross-fence
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differences are obvious, but their causes are not. In spite of the appearance of a grandiose
‘natural experiment’, the cross-fence comparisons are often poorly replicated and
confounded. Nonetheless, studies adopting this design have formed the bulwark of claims
about dingoes’ ecological roles (Letnic et al. 2012 ; Allen et al. 2013b ; Glen and Woodman
2013). Until more rigorous experimental designs are implemented, further studies predicated
on correlative, cross-fence differences does little to increase evidence for the ecological role
of dingoes.
The management consequences of failing to address alternative hypotheses are exemplified
by the relatively simple carnivore system in Australia. Johnson and colleagues (2007) argued
that human control of dingoes in the last 200 years caused the continental collapse of
marsupial communities across Australia, but the role of the continental invasion of European
rabbits (Oryctolagus cuniculus; Cooke and Soriguer 2017) and the historical grazing of
introduced sheep (Ovis aries) coupled with drought (Allen 2011) were not properly assessed
as potential causal factors for marsupial decline. Johnson and colleagues continue to assert
that if only dingo persecution stopped, dingoes would suppress introduced rabbits, red foxes
(Vulpes vulpes) and feral cats (Felis catus), and facilitate the recovery of reintroduced
marsupials and other small mammals across the continent (e.g. Johnson 2006; Wallach et al.
2009; Ritchie et al. 2012 ; Letnic et al. 2013 ). But such reintroductions continue to fail largely
because predators – including dingoes – keep quickly decimating reintroduced mammals
(Christensen and Burrows 1995; Moseby et al. 2011 ; Bannister 2014; Armstrong et al. 2015 ;
Bannister et al. 2016 ). All the dingoes occupying Australia did not prevent the historical
establishment and expansion of rabbits, foxes or cats across the continent in the first place,
nor did the presence of dingoes prevent the collapse of marsupial communities following the
advent of these pests. Extant dingo populations, never managed by modern humans across
roughly one-third of the Australian continent (Allen et al. 2015 ), have not facilitated
extirpation of these pests or their impacts, nor facilitated the recovery of marsupials in these
areas. Indeed, dingoes reach their highest densities in places with abundant rabbits (Bird
1994; Allen 2012a), suggesting that invasive species are supporting carnivores rather than
large carnivores suppressing invasive species. Dingoes may even provide net benefits to
invasive rabbits through mesopredator suppression, just as dingoes putatively benefit rabbit-
sized native mammals (Cooke and Soriguer 2017; Gordon et al. 2017a ). In concert with
habitat changes (be these caused by livestock, fire or rabbits), dingo predation has been
identified as a key driver of native mammal decline independent of foxes or cats (e.g. Kerle
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et al. 1992 ; Corbett 2001; Lundie-Jenkins and Lowry 2005; Barnes et al. 2008 ; Allen 2011;
Allen and Fleming 2012; Allen and Leung 2012). Yet dingoes are typically considered part of
the solution to Australia’s fauna extinction crisis, when they are also part of the problem.
Continuing to ignore this and other alternative hypotheses wastes precious time in our
collective efforts to conserve native fauna under real threat of extinction.
In complex carnivore communities (where a wide variety of individual large carnivores
utilise a range of hunting strategies, resulting in increased heterogeneity in predator-prey
interactions), even manipulative experiments still struggle to tease apart the relative influence
of top-down and bottom-up processes (e.g. Gasaway et al. 1983 ; Boertje et al. 1996 ; Maron
and Pearson 2011; Sinclair et al. 2013 ; Ford et al. 2015b ; Riginos 2015). In fact, Riginos
(2015) goes as far as to suggest that behaviourally-mediated trophic cascades are either weak
or non-existent in African savanna systems because of the large sizes of many of the
herbivores (elephants, Loxodonta africana, in particular) and the over-riding effect of
climate. Predator diversity is known to dampen trophic cascade effects in model systems
(Finke and Denno 2004), and top-down forcing is also known to attenuate down through
trophic levels more rapidly than previously thought (Schmitz et al. 2000 ; Brodie et al. 2014 ).
One characteristic of overemphasising the current robustness of large carnivore science is
ignoring, suppressing or omitting reference to alternative hypotheses and contrary data
(Claridge 2013; Winnie 2014). This is easy for authors to do given the vast pool of citations
to choose from (e.g. Tables 1–3) and the limited number of references a journal will typically
accept. When accused of selective referencing, the plea of ‘not enough room’ (e.g. see Marris
2014 for examples) does not promote objectivity and transparency. Rather, it disregards the
legitimate scientific criticisms available and only widens the creeping cracks of bias
described by Sarewitz (2012), who argued that research is riddled with systematic errors (see
also Ioannidis 2005, 2014) and that the ensuing debate then erodes public confidence in
science itself (see also Fleming et al. 2012 ; Middleton 2014). Although large-scale and
observational ‘natural experiments’ have great value when their results are ‘consistent with’
or ‘inconsistent with’ a given hypothesis, plausible alternative explanations nonetheless
require thorough exploration and ranking before reported results from ‘natural experiments’
become the basis for changes in practice or policy (Barley and Meeuwig 2016). Investigating
alternative hypotheses should be a greater priority in future research on large carnivore
ecology.
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4. There is a dearth of applied-science studies
Some research questions are largely academic (e.g. do species and A and B have overlapping
diets?), whereas applied studies have direct and immediate relevance to land and fauna
managers (e.g. do interventions X and Y produce the same outcome for species A and B?).
The importance of understanding the ecological roles of large carnivores has implications for
the conservation and management of threatened carnivores and other fauna, such as livestock,
game, or threatened wildlife prey species (e.g. Boertje et al. 2010 ). Managers need
information that considers both the pros and cons of various management interventions, and
this is best achieved through manipulative experiments or adaptive-management studies that
investigate applied-science issues (Glen et al. 2007 ; Hone 2007; Hone et al. 2015 ). Questions
about the conservation utility of large carnivores as tools to restore biodiversity across the
landscape are answered much faster when truly applied questions are investigated.
Evidence for the effects of carnivore removal is also not the same thing as evidence for the
effects of their recovery (e.g. ansiotropic vs isotropic effects; sensu Ford and Goheen 2015).
Simply re-establishing or bolstering large carnivores may not fix the many environmental
problems that occurred as a result of (and/or in addition to) carnivore extirpation (Marshall et
al. 2013 ; Marshall et al. 2014 ; Wikenros et al. 2015 ). In some cases, food web structure and
ecological context may have changed irreversibly (for whatever reason), some niches may no
longer exist, and a carnivore’s function in the new ecosystem might now be different from
their previous function. Changes in the physical environment caused by the removal of large
carnivores may make the system resistant to complete restoration after large carnivores are
restored. This ‘change resistant’ hypothesis was tested against the existing TCH in a
replicated, randomized, manipulative experiment conducted over a decade. The hypothesis
that wolf restoration had caused ecosystem reorganization was rejected (Marshall et al.
2014), yet subsequent literature ignored it and instead repeated the story (i.e. Sustainable
Human 2014) that the ecosystems of Yellowstone have been dramatically restored by wolves
following their reintroduction. Restoring carnivore populations “to areas greatly modified by
human disturbance may not restore systems to their former state” (Glen et al. 2007 ; pg. 498)
and these new carnivore functions may not be viewed as desirable or produce net benefits to
novel and still-changing ecosystems (Fleming et al. 2012 ; Flagel et al. 2016 ).
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Large carnivore studies often report a negative relationship between larger carnivores and
smaller or mesocarnivores, and are then quick to recommend wholesale changes to the way
large carnivores are managed without first measuring any actual effect of carnivore
management (e.g. hunting, removal, restoration) on large or small carnivores, herbivores or
prey (for examples, see Letnic et al. 2009 ; Wallach et al. 2010 ; Colman et al. 2014 ; Gordon
et al. 2017a ; Gordon et al. 2017b ). Equally, perceived negative impacts of carnivores on
livestock have historically been addressed by wholesale lethal control without any
recognition of the positive impacts that carnivores may have on the herbivores that compete
with livestock or the consequences of lethal control on livestock losses (e.g. Wicks and Allen
2012; Allen 2014; Allen 2015a; Prowse et al. 2015 ; Allen 2017). Treves et al. (2016) and
others (e.g. Reddiex and Forsyth 2006; Doherty and Ritchie 2017) rightly point out that many
studies promoting predator control are badly designed, and we agree, but the same failing
exists in many studies condemning predator control and promoting predator conservation.
Unreliable science and poor science communication practices are a feature of literature
expressing both positive and negative views towards carnivores (Boertje et al. 2010 ).
To make ecological data useful for improving carnivore management and conservation,
researchers must provide managers with data they can apply. For example, when claiming
that large carnivore control (i.e. trapping, hunting, or poisoning) must be banned in order to
generate cascading, positive effects on biodiversity (e.g. Carwardine et al. 2012 ), information
on the actual effects of carnivore hunting or poisoning on biodiversity are needed, not just
information on how one carnivore species might interact with another (for examples, see
Fleming et al. 2012 ; Allen et al. 2015 ). Conversely, when claiming that large carnivore
control must be implemented to reduce livestock predation, information on actual carnivore
impacts and impact reduction is required to ethically justify carnivore control (Braysher
1993; Allen et al. 2014b ; Allen 2017). The paucity of applied ecological data in the wider
large carnivore literature means that much of the presently available information on the
MRH, TCH and BMTCH is not as useful to managers as it could be. This paucity also means
that, in most cases, we do not yet have a solid understanding of the actual cascading effects, if
any, of carnivore reintroduction, population control or manipulation (Ripple et al. 2014b ;
Newsome et al. 2015 ). This issue contributes to a significant knowledge-mobilization and
implementation gap for large carnivore science.
5. Logical fallacies underpin much of the literature
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Most research about the ecological roles of large carnivores is also grounded in two logical
fallacies, post hoc ergo propter hoc and cum hoc ergo propter hoc. Post hoc ergo propter hoc
is the notion that if X occurred before Y, then X caused Y. When X is undesirable, this pattern
is often extended in reverse as: avoiding X will prevent Y. Cum hoc ergo propter hoc is the
notion that if X changed similarly to Y, then X and Y are linked. The fallacies lie in coming to
a conclusion based on the order or pattern of events, rather than accounting for other factors
that might rule out a proposed connection.
Examples of post hoc ergo propter hoc in the large carnivore literature are rife and include,
for example, conclusions to the effect that ‘the ecological changes observed in Yellowstone
National Park occurred after wolves were reintroduced, so wolves must have caused these
ecological changes’ (epitomised in Sustainable Human 2014; see Table 2). Or alternatively,
‘the last population of highly endangered mammals went extinct after predator control, so
predator control must have caused the extinction through trophic cascade effects’ (discussed
in Fleming et al. 2013 ). There are also many examples of cum hoc ergo propter hoc,
including almost all the relevant literature on dingoes’ ecological roles (see Allen et al.
2013b; see Table 1). That wolves may not have been the cause of all the observed ecological
changes in Yellowstone since the mid-1990s is argued by Kauffman et al. (2010), Mech
(2012) and others (e.g.Creel and Christianson 2009; Winnie 2012; Marshall et al. 2013 ;
Marshall et al. 2014 ; Middleton 2014; Peterson et al. 2014 ; see Table 2). The long term study
of wolf–moose (Alces americanus)–habitat–climate relationships on Isle Royale illustrate the
difficulties of attributing cause and effect even in very simple ecosystems (Vucetich and
Peterson 2004). This case study stands out because researchers have explored multiple factors
at the same time, have been excessively cautious in the language they use to attribute
causality, and have constantly updated their views concerning the functioning of the
ecosystem as new data becomes available. Shifting the research focus from ‘trophic cascades’
to ‘food webs’ in this way can help overcome the subtle yet troublesome overreliance on
logical fallacies in studies of carnivores’ ecological roles (Eisenberg et al. 2013 ).
6. Most of the ‘best evidence’ comes from ecosystems that do not represent the majority
of the earth’s surface or species
Although there are still some large tracts of relatively intact land in some places, the reality is
that the majority of the earth’s surface has been substantially altered by humans, and
continues to be altered, in a modern epoch now labelled as the Anthropocene (Zalasiewicz et
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al. 2008 ; Kueffer and Kaiser-Bunbury 2013). Modern, human-dominated ecosystems
typically comprise mixed land-uses including urbanisation, forestry, mining, hunting,
recreation, agriculture (crops and/or livestock production) or other areas fragmented by roads,
railways and fences, and containing exotic plant and animal species and artificial water
sources (Linnell 2011; Fleming et al. 2012 ; Mech 2012). Most tests of the MRH, TCH and
BMTCH have occurred in relatively intact ecosystems with relatively minor human
footprints, such as the National Parks of Canada and the United States (Hebblewhite et al.
2005; Ray et al. 2005 ; Hayward and Somers 2009; Eisenberg 2011; Kuijper et al. In press ).
Where studied, however, the strength and utility of carnivore effects on food webs in human-
modified systems appear dissimilar to those in less modified ecosystems (e.g. Elmhagen et
al. 2010 ; Muhly et al. 2013 ; Meadows et al. 2017 ; Morgan et al. 2017 ).
For example, the recolonization of wolves in Sweden resulted in widespread behaviour
change by humans in their moose (Alces alces) hunting practices that precluded, or at least
reduced, the anticipated numerical effects of wolves on moose (Wikenros et al. 2015 ).
“Because most of the worlds’ habitat that will be available for future colonization by large
predators are likely to be strongly influenced by humans…, human response behaviour may
constitute an important factor that ultimately may govern the impact of large predators on
their prey and thus on potential trophic cascades” (Wikenros et al. 2015 ; pg. 18). This point is
further underscored by the situation in South Africa, where the introduction or removal of
large carnivores has largely been driven by economic incentives (Lindsey et al. 2007 ), and
the long term ecological effects have been overlooked. In Kenya, the indirect effect of
carnivores on tree communities was mediated by ranching practices and the spatial
distribution of cattle corrals (Ford et al. 2014 ). Comparative analyses of mammalian food
webs in protected areas versus human-dominated areas of Canada concluded that ‘human
influence on vegetation may strengthen bottom-up predominance and weaken top-down
trophic cascades in ecosystems’ and that ‘human influences on ecosystems may usurp top-
down and bottom-up effects’ (Muhly et al. 2013 ).
Theories about the effects of large carnivores on food webs, as developed in relatively
pristine areas, may not be readily transferable or applicable to the human-modified
landscapes that make up the majority of the Earth’s surface (Haswell et al. 2017 ; Morgan et
al. 2017 ). This is because the direct and indirect effects of humans on all trophic levels may
simply overshadow any carnivore effects (Muhly et al. 2013 ; Darimont et al. 2015 ; Clinchy
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et al. 2016 ; Kuijper et al. In press ). Carnivores are but one potential causal factor in a
multicausal world (Vucetich and Peterson 2004; Peterson et al. 2014 ; MacNulty et al. 2016 ;
Engeman et al. 2017 ), and restoring large carnivores into these human-modified systems
without removing the many other, more important causal factors influencing biodiversity loss
is unlikely to succeed in reversing the situation (Allen and Fleming 2012; Fleming et al.
2012). This is not to say that carnivore restoration efforts are unnecessary or should be
avoided (Chapron et al. 2014 ), but that we should more carefully consider the anticipated
benefits of these actions against the biophysical and anthropogenic factors that mediate the
top-down effects of carnivores.
Implications for large carnivore science and management
The prevalence of these six aforementioned issues in the literature on large carnivores (Tables
1–3) underscores our contention that evidence for the MRH, TCH and BMTCH is undeniably
weaker than is often claimed in journal articles or public discourse. Syntheses and literature
reviews of large carnivores’ ecological roles should identify these issues, but they usually do
not, instead routinely failing to assess the internal validity of the original studies reviewed, as
described by Bilotta et al. (2014). When the individual empirical studies that make-up the
content of these reviews are judged against Platt’s (1964) criteria for strong inference, Hone’s
(2007) deconstruction of experimental design capabilities, or Sutherland and colleagues’
(2013) 20 tips for interpreting scientific claims, it is clear that even literature reviews (e.g.
Ritchie and Johnson 2009; Estes et al. 2011 ; Ripple et al. 2014b ) seldom offer reliable
guidance on the state of the literature addressing the MRH, TCH and BMTCH. These remain
intriguing hypotheses, but they are each inadequately tested and not yet demonstrated for
almost all large carnivores and contexts.
We fear that the debates about the issues we raise here (Tables 1–3) are heading towards the
type of science denialism that plague medicine or climate science (see Diethelm and McKee
2009). In a growing number of cases, strong evidence against MRH, TCH and BMTCH is
denied while promoting these hypotheses using tactics common to science denial in other
disciplines, such as selectivity, use of logical fallacies, disregard of experimental work, and
deference to correlations (for examples, seeLetnic et al. 2011 ; Ripple et al. 2011 ; Beschta et
al. 2014 ; Forsyth et al. 2014 ; Johnson et al. 2014 ; for responses, see Hodges 2012; Squires et
al. 2012 ; Fleming et al. 2013 ; Allen et al. 2014a ; Winnie 2014; Allen and West 2015).
Science denialism is often characterised by downplaying the scope of a threat (Russell and
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Blackburn 2017). In the field of large carnivore science, this is clearly manifest in claims that
carnivores are not a major problem for livestock producers or game ranchers (e.g. Forsyth et
al. 2014 ). It is also manifest in claims that native large carnivores will suppress unwanted
exotic species while denying that the same native carnivores can also suppress the threatened
native species they are assumed to provide protection for (see Fleming et al. 2013 or Allen
and Fleming 2012 for discussion). Dismissing or downplaying the legitimacy of scientific
criticisms as mere ‘controversy’ or ‘debate’ (e.g. Ritchie et al. 2014 ; Newsome et al. 2015 ) is
also a form of passive science denialism. In truth, carnivores can have direct and indirect
positive, negative or neutral impacts on social, economic and environmental values, and these
impacts can change from time to time and place to place (Chamberlain et al. 2014 ; Haswell
et al. 2017 ). But emphasizing ‘the good’ while downplaying ‘the bad’ only produces ‘the
ugly’ literature on carnivore science, while also fostering the rise of invasive species science
denialism (Russell and Blackburn 2017). Such post-truth incredulities over evidence risks
reversing progress in a field that is tackling some of the most important and engaging
questions in modern ecology – namely, how does society restore and coexist with large fauna
in human-occupied landscapes (LaRue et al. 2012 ; Chapron et al. 2014 ) and what may be the
ecological outcomes of this restoration effort?
Debates about the scientific understanding of, and appropriate management response to, large
carnivore impacts are not new. For example, in Alaska and northern Canada there has been an
ongoing debate about the impact of wolf and grizzly bear (Ursus arctos) predation on moose
and caribou (Rangifer tarandus) populations for decades (e.g. Orians et al. 1997 ; Kennedy
and Fiorino 2011). The discourse has centred on the extent to which lethal control of wolf and
bear populations will lead to an increase in the harvestable surplus of moose and caribou. An
enormous amount of intensive research, of both descriptive and experimental types (reviewed
by Boertje et al. 2010 ), has been conducted in the region since the 1970’s with the aim of
understanding predator-prey relationships. But just like the Yellowstone region (MacNulty et
al. 2016 ), there is still huge uncertainty and controversy about the nature of these trophic
interactions and their consequences for management despite this considerable research
investment (e.g. Van Ballenberghe 2006; Boertje et al. 2010 ; Kennedy and Fiorino 2011).
Lessons that can be extracted from this ongoing saga include: (1) even with massive
investment in research over many decades in relatively simple ecosystems it can still be a
challenge to understand the nature of interactions between predators and prey, let alone the
wider ecosystem impacts of human intervention on lower trophic levels; (2) valuable insights
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can be obtained by exploring such relationships through the lens of predator-prey theory and
demographic models, an approach which has been almost absent from the recent generation
of trophic cascade studies (Tables 1–3); and (3) competing scientific results can rapidly be
included into what are essentially value debates about different worldviews. The maturation
of this controversy clearly shows how important it is to be aware of the intrinsic uncertainty
and context-dependence (in time and space) of any research results, and of the need to clearly
distinguish science from values in policy debates.
There are, of course, studies that are not encumbered by the six issues we raise, studies that
do indeed provide strong support for the MRH, TCH and BMTCH. Much of this can be
found in literature from marine, aquatic and invertebrate systems (Heath et al. 2014 ;
Meadows et al. 2017 ), or systems and models where bottom-up processes are relatively
predictable, stable and controllable. Reliable work on MRH, TCH and BMTCH in terrestrial
systems is only beginning to catch up to these disciplines. Literature reviews and syntheses
are important as the field develops, but as described above, most of the reviews presently
available are inadequate. There is, therefore, an urgent need for a systematic review (sensu
Pullin and Knight 2009) of terrestrial studies that have used only manipulative experiments to
investigate these hypotheses– experiments inclusive of paired treated and non-treated areas,
sampled before and after treatments (e.g. carnivore removal or addition) over sufficient
temporal and spatial scales to detect cascading responses of predators, prey and plants. A
systematic review of such experimental studies, which excludes low-inference studies and
summarises the results of only those with the actual capacity to assess causal processes, may
produce useful insights into underlying ecological processes and be of great value to
carnivore managers (Pullin and Knight 2009; e.g. Boertje et al. 2010 ). It would also yield
lessons on how to do more such research on different species, and in different contexts.
Many authors have called for such large-scale, long-term manipulative experiments
investigating the removal or addition of large carnivores (e.g. Glen et al. 2007 ; Ritchie et al.
2012; Newsome et al. 2015 ). Although such experiments are expensive and difficult to
achieve because of the logistical challenges arising from the massive scales that large
carnivores utilise, they can and have been done in some places (e.g. Eldridge et al. 2002 ;
Hayes et al. 2003 ; Hebblewhite et al. 2005 ; Allen et al. 2013a ; Marshall et al. 2013 ; Allen et
al. 2014b ; Christianson and Creel 2014; Ford et al. 2014 ; Hervieux et al. 2014 ; Ford et al.
2015b; Mitchell et al. 2015 ). These have often, but not always, shown support for elements of
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the MRH, TCH and BMTCH; less so for dingoes (Allen et al. 2014b ) but more so for wolves
(Winnie and Creel 2017). It is unlikely that many large carnivores will be subject to
experimental studies like these, or like the famous Kluane project on the Canada lynx (Lynx
canadensis) and snowshoe hare (Lepus americanus) system (Krebs et al. 2001 ). As a
consequence, it is highly unlikely that we will ever have access to knowledge from such
experiments for most large carnivores. Thus, a systematic review of studies testing the MRH,
TCH and BMTCH with only strongly-inferential methods will be all the more valuable. It
must also be remembered that while well-designed and implemented experiments will greatly
advance our understanding of theoretical ecological principles (Engeman et al. 2017 ), the
portability of their results may still be limited (Schmitz et al. 2000 ; Haswell et al. 2017 ;
Morgan et al. 2017 ).
Our focus on improving research rigour is not intended to imply that observational or
correlative ecological studies are not useful. Such studies are absolutely crucial to capture the
broad spatial and temporal dynamics over which large carnivores and their prey interact
(Barley and Meeuwig 2016). However, we argue that researchers need to exercise a greater
degree of caution in the interpretation and communication of studies on the MRH, TCH and
BMTCH, no matter how they are designed and conducted, and especially when they are used
as the basis for radical changes in carnivore management and policy – including cases where
lethal control and reintroduction are used. The associated biases, uncertainties, and ability to
make inferences need to become ever more central parts of the communication of research
results (Johnson et al. 2015 ). While we hope that scientists should manage this within the
pages of peer-reviewed journals, additional challenges arise when trying to communicate
uncertainty to the wider public (Dixon and Clarke 2013). In such contexts it is normally
impossible to successfully communicate such intrinsic limitations, making it all the more
important that authors take extreme care to not oversell the generality of their findings, nor
allow others to do so, and clearly separate between scientific findings and the various
normative policy or management action contexts within which these findings might be
operationalised.
The reality is that the knowledge available to wildlife managers will at best be limited to a
solid understanding of the natural history and ecology of the predators, their prey, and the
ecosystem, and based largely on data derived from time series, cross-site comparisons,
‘natural experiments’ or other correlative studies (Barley and Meeuwig 2016; MacNulty et al.
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2016). A good understanding of species ecology can serve to exclude spurious or
unreasonable interpretations of correlative data, and such studies can also exclude certain
hypotheses or provide indirect support for other hypotheses for which experiments could be
designed to provide a definitive test (e.g. Platt 1964; Kershaw 1969; Underwood 1997;
Fairweather and Quinn 2006). While these types of lower-inference studies may not
overcome all the aforementioned issues we describe, they do have the advantage of being far
cheaper and faster to conduct under a wide range of different ecological conditions, which
can address problems associated with the transferability of knowledge between contexts.
Ideally, conservation actions should be monitored within an adaptive management system
that can be used to permit the study of system responses to specific management
interventions (Fleming et al. 2014 ; Johnson et al. 2015 ). This provides insights into how the
system functions and how management actions produce outcomes. Certain forms of carefully
designed adaptive management exercises can even be viewed as quasi-experiments (Williams
and Brown 2014; Johnson et al. 2015 ).
Given the perilous conservation situation of many large carnivore species, there is a clear
need to act based on the best available knowledge at any given time (Ripple et al. 2016a ).
However, manipulative experiments clearly trump anecdotal, observational and/or correlative
information for their informative value, and should therefore be valued more highly in the
decision making process (Platt 1964; Fleming et al. 2013 ). While the weight of evidence for
the general role of large carnivores in triggering trophic cascades is indeterminate at this time
(but we look forward to this potentially changing one day), we caution researchers and
science communicators to carefully consider the implications of simultaneously advocating
for both large carnivore conservation and the primacy of top-down trophic cascades. These
two forms of advocacy need not be linked – carnivore conservation can often be justified on a
number of moral, ethical, and existential grounds that have nothing to do with trophic
cascades. At one extreme, such advocacy may contribute towards specious reintroduction
efforts that divert funds from broader conservation goals and/or place the livelihoods of local
people at risk (Ford et al. 2017 ). On the other extreme, we recognize that there will be no
perfect study to ever ‘close the book’ on the prevalence of trophic cascades, regardless of
their occurrence in nature. Because strongly-inferential, long-term, manipulative studies will
be difficult to implement in a cost-effective and timely manner to support these decisions, we
argue that knowledge of trophic cascades must be considered in management deliberations
but should not necessarily determine their outcome.
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Whether or not society should or shouldn’t restore large carnivores is outside the scope of our
present analysis (but see Lewis et al. 2017 ), and in the end, how large carnivores are
managed is a judgement that society must make, and which will largely be based on which
species (predator or prey or human interest) is given priority over another. In the Canadian
case of Hervieux et al. (2014), for example, the immediate interests of ungulates were
ultimately favoured over those of the wolves. Whereas, in the familiar Yellowstone story (e.g.
Middleton 2014), the interests of wolves were ultimately favoured over those of the
ungulates. Whether large carnivores are viewed as a ‘good thing’ or a ‘bad thing’ for an
ecosystem largely rests on the attention given to which species (livestock, invasive pests,
game species or threatened native fauna) carnivores happen to be killing at the time (Allen et
al. 2011b ; Mech 2012). As carnivore conservationists ourselves, we relish any excuse to
promote their conservation and recovery where it is needed and possible. But as scientists, we
lament the lack of objectivity and critical thinking underpinning the current ‘parental
affection’(sensu Chamberlin 1890) towards the MRH, TCH, and BMTCH and the extent to
which this affection is used to legitimise selected views on carnivore management.
Upon reflection, we also observe that debates about large carnivore management (Tables 1–3)
are often not so much about differing beliefs or views about carnivores’ actual functional
roles, but more so about the quality of scientific evidence people are willing to accept. Large
carnivore conservation is a bold and historically-novel judgement which must inevitably be
made on incomplete ecological evidence. Ecological evidence alone is insufficient to make
decisions, which must also account for the ethical, cultural and socio-political factors that
shape decision making in society (e.g. Van Ballenberghe 2006; Mech 2010; Trouwborst 2010;
Fleming et al. 2014 ; Olson et al. 2015 ; Trouwborst 2015; Marshall et al. 2016 ; Lewis et al.
2017). We hope that the issues we raise here prompt deeper consideration of actual evidence,
leading to an improvement in both the rigour and communication of large carnivore science,
because the fates of many large carnivores and the integrity of associated ecological
processes are depending on it.
Acknowledgments
Carl Mitchell, John Winnie, Stewart Breck and Tom Hobbs contributed discussion points and
provided helpful comments and suggestions on earlier drafts of the manuscript. John Linnell
was funded by the Research Council of Norway (Grant #251112). Any use of trade, firm or
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product names is for descriptive purposes only and does not imply endorsement by the U.S.
Government, or any other author or affiliation.
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Tables
Table 1 – Some recent lines of debate discussing large carnivores’ roles in trophic cascades in
Australia, demonstrating that evidence for the ecological roles of dingoes is equivocal,
primarily due to the six issues described in the present article.
Debated topic Chronological order Reference
Trophic cascades following dingo
control
1Wallach and O'Neill 2009
2Allen 2010
Ecological niche of dingoes 1 Fleming et al. 2012
2Johnson and Ritchie 2013
3Fleming et al. 2013
4Claridge 2013
Dingo predation risk to fauna 1 Dickman et al. 2009
2Allen and Fleming 2012
Methodological problems with
dingo studies
1Allen et al. 2011a
2Letnic et al. 2011
3Allen et al. 2011b
4Glen 2012
5Allen et al. 2013b
Cause of historical declines of
marsupials
1Johnson et al. 2007
2Allen 2011
Importance of dingo social
structure
1Wallach et al. 2009
2Allen 2012b
Trophic cascades following dingo
control
1Colman et al. 2014
2Allen 2015b
3Colman et al. 2015
Effects of dingoes on sheep 1 East and Foreman 2011
2Allen and West 2013
3Forsyth et al. 2014
4Allen and West 2015
Trophic cascades following dingo
control
1Allen et al. 2013a
2Johnson et al. 2014
3Allen et al. 2014a
4Allen et al. 2014b
5Hayward and Marlow 2014
6Nimmo et al. 2015
7Hayward et al. 2015
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1365
1366
1367
1368
1369
1370
1371
1372
1373
1375
1376
Table 2 – Some recent lines of debate discussing large carnivores’ roles in trophic cascades in
North America, demonstrating that evidence for the ecological roles of wolves is equivocal,
primarily due to the six issues described in the present article.
Debated topic Chronological order Reference
Wolf-induced
behaviourally-mediated
trophic cascades in
Yellowstone
1Ripple and Beschta 2004
2Kauffman et al. 2007
3Ripple and Beschta 2007
4Kauffman et al. 2010
5Kimble et al. 2011
6Winnie 2012
7Beschta and Ripple 2013
8Kauffman et al. 2013
9Middleton et al. 2013a
10 Beschta et al. 2014
11 Winnie 2014
12 Painter et al. 2015
Willow recovery in
Yellowstone following wolf
reintroduction
1Ripple and Beschta 2003
2Despain 2005
3Ripple and Beschta 2006
4Wolf et al. 2007
5Beyer et al. 2007
6Bilyeu et al. 2008
7Creel and Christianson 2009
8Tercek et al. 2010
9Johnston et al. 2011
10 Middleton et al. 2013a
11 Marshall et al. 2013
12 Marshall et al. 2014
13 Smith et al. 2016
Trophic cascades and
Mexican wolves
1Beschta and Ripple 2010
2Mech 2012
Wolf effects on lynx 1 Ripple et al. 2011
2Hodges 2012
3Squires et al. 2012
4Wirsing et al. 2012
Wolf effects on bears 1 Ripple et al. 2014a
2Barber-Meyer 2015
3Ripple et al. 2015
Ethics and effects of
predator control for moose
conservation in Alaska
1WMRC 1996
2Orians et al. 1997
3Van Ballenberghe 2006
4Boertje et al. 2010
5Kennedy and Fiorino 2011
43
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1378
1379
1380
1381
1382
44
1384
1385
Table 3 – Some recent lines of debate discussing large carnivores’ roles in trophic cascades in
Europe, demonstrating that evidence for the ecological roles of large carnivores is equivocal,
primarily due to the six issues described in the present article.
Debated topic Chronological order Reference
Human influence on trophic
cascades in Europe
1Melis et al. 2009
2Kuijper 2011
3Dorresteijn et al. 2015
4Kuijper et al. 2016
5Ritchie et al. 2016
Large carnivore impacts on
mesocarnivores
1Palomares et al. 1995
2Palomares et al. 1998
3Sunde et al. 1999
4Linnell and Strand 2002
5Helldin et al. 2006
6Elmhagen and Rushton 2007
7Kowalczyk et al. 2009
8Pasanen-Mortensen et al. 2013
9Wikenros et al. 2014
10 Pasanen-Mortensen and
Elmhagen 2015
45
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