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REVIEW PAPER
Deciding when to lend a helping hand: a decision-making
framework for seabird island restoration
Rachel T. Buxton
1,2
•Christopher J. Jones
3
•Philip O’Brien Lyver
3
•
David R. Towns
4,5
•Stephanie B. Borrelle
4
Received: 21 August 2015 / Revised: 22 February 2016 / Accepted: 3 March 2016 /
Published online: 10 March 2016
ÓSpringer Science+Business Media Dordrecht 2016
Abstract Following the removal of an introduced species, island restoration can follow
two general approaches: passive, where no further intervention occurs and the island is
assumed to recover naturally, and; active, where recovery of key taxa (e.g. seabirds) is
enhanced by manipulating movement and demography. Steps for deciding between these
techniques are: (1) outlining an explicit restoration goal; (2) building a conceptual model of
the system; (3) identifying the most effective management approach; and (4) implementing
and monitoring outcomes. After decades of island restoration initiatives, retrospective
analysis of species’ responses to active and passive management approaches is now fea-
sible. We summarize the advantages of incorporating these analyses of past restoration
results as an initial step in the decision-making process. We illustrate this process using
lessons learned from the restoration of seabird-driven island ecosystems after introduced
vertebrate eradication in New Zealand. Throughout seven decades of successful vertebrate
eradication projects, the goals of island restoration have shifted from passive to active
enhancement of island communities, which are heavily dependent on burrow-nesting petrel
Communicated by Stephen Garnett.
Electronic supplementary material The online version of this article (doi:10.1007/s10531-016-1079-9)
contains supplementary material, which is available to authorized users.
&Rachel T. Buxton
r.buxton@mun.ca
1
Department of Zoology and Centre for Sustainability: Agriculture, Food, Energy, and Environment,
University of Otago, PO Box 56, Dunedin 9054, New Zealand
2
Present Address: Department of Fish, Wildlife and Conservation Biology, Colorado State
University, 1474 Campus Delivery, Fort Collins, CO 80523, USA
3
Landcare Research, P.O. Box 69040, Gerald Street, Lincoln 7640, New Zealand
4
School of Applied Sciences, Institute for Applied Ecology, Auckland University of Technology,
Private Bag 92006, Auckland 1142, New Zealand
5
Department of Conservation, Private Bag 68908 Newton, Auckland 1145, New Zealand
123
Biodivers Conserv (2016) 25:467–484
DOI 10.1007/s10531-016-1079-9
population recovery. Using a comparative analysis of petrel response to past predator
eradications we built a conceptual model of petrel recovery dynamics and defined key site
and species characteristics for use in a stepwise decision tree to select between active or
passive seabird population management. Active restoration techniques should be imple-
mented when seabird populations are absent or declining; and on islands with no nearby
source colony, small remnant colonies, highly altered habitat with shallow soil and slopes,
and with competitive species pairs. As we continue to restore complex island communities,
decision-making tools using a logical, step-wise framework informed by previous
restoration successes and failures can aid in increasing understanding of ecosystem
response.
Keywords Adaptive management Decision tree Burrowing seabirds Eradication
New Zealand Prioritization Recovery
Introduction
Islands hold a disproportionately large percentage by area of global biodiversity and are
increasingly important repositories for populations or ecosystems eliminated from the
mainland (Daugherty et al. 1990; Mittermeier et al. 1998; Kier et al. 2009). Unfortunately,
island ecosystems are also more susceptible to disturbance and have experienced high
extinction rates—primarily due to the introduction of alien species (King 1985; Cour-
champ et al. 2003). Although advancements have been made in alien species eradications
and ecosystem restoration, islands’ remote locations and limited infrastructure make
conservation efforts expensive (Donlan 2007; Helmstedt et al. 2016). Priority setting and
decision-making tools are thus especially important in allocating limited resources while
maximizing successful outcomes of island restoration.
Islands are particularly important for seabirds, typically providing terrestrial predator-
free nesting habitat in proximity to pelagic feeding areas. Similarly, seabirds are partic-
ularly important members of island ecosystems, where their presence drives ecological
processes (Mulder et al. 2011a). On islands where seabirds nest (henceforth ‘‘seabird
islands’’), seabird guano enriches the soil with marine-derived nutrients and nest-building
disturbs vegetation and aerates the soil (Mulder et al. 2011b; Smith et al. 2011). Fur-
thermore, seabirds play a special cultural role on islands around the world, providing
people with food (Circumpolar Seabird Working Group 2001), income (Duffy 1994), and
cultural cohesion and identity (Lyver et al. 2008). Because most seabirds have nested for
millennia on islands free of mammalian predators, they lack behavioural and life-history
adaptations to avoid ground-based predation (Milberg and Tyrberg 1993). Thus, seabirds
are vulnerable to alien predators, leading to species extirpation or severe population
reduction on invaded islands (Towns et al. 2011). Seabird population declines have
transformed entire island socio-ecological systems and their recovery is often integral to
successful restoration (Croll et al. 2005; Fukami et al. 2006; Young 2014).
Predator eradication has become a prevalent seabird island restoration technique
worldwide, with over 900 islands cleared of predators as of 2011 (Keitt et al. 2011; DIISE
2015). Following eradication of alien predators, restoration efforts can follow two basic
approaches: passive, where no further intervention occurs and seabird populations and
islands are assumed to recover naturally, and; active, where recovery of seabirds is
468 Biodivers Conserv (2016) 25:467–484
123
enhanced by manipulating distribution and demography (Hobbs et al. 2011). Passive
management operates under the ‘‘Field of Dreams’’ hypothesis, assuming that removing a
stressor is sufficient to restore habitat and thus the capacity of seabirds to recover (‘‘if you
build it, they will come’’; Palmer et al. 1997). It is generally cheaper than active man-
agement, but relies exclusively on the unpredictable capacity of populations and ecosys-
tems to recover on their own (Scott et al. 2001; Jones and Schmitz 2009). Passive seabird
population recovery is assumed to be slow because of k-selected life history characteristics,
seemingly high philopatry, and intermittent breeding (Kappes and Jones 2014). Active
management can help overcome impediments to population recovery, but often has high
logistical and financial demands and variable success rates (Holl and Aide 2011; Jones and
Kress 2012). Despite their logical interdependency, passive and active seabird island
restoration techniques have evolved largely independently, so that few guidelines are
available to help managers decide when to support passive recovery with active techniques.
Like most restoration endeavors, seabird island restoration is complex, operating within
a dynamic network of ecological, cultural, social, and economic ideologies. Thus, deciding
whether to employ active techniques would benefit from using a structured decision-
making framework (Noss et al. 2009). Structured decision-making frameworks organize
information, define objectives, and identify outcome alternatives and uncertainty, allowing
restoration options to be weighted objectively (Wyant et al. 1995). A typical decision-
making process involves: (1) outlining an explicit restoration goal; (2) building a con-
ceptual model of the system; (3) identifying the most effective management approach; and
(4) implementing and monitoring outcomes (Possingham et al. 2001). Seabird island
restoration is unique in that there is an abundance of pre-existing passive and active
restoration projects (Jones and Kress 2012; DIISE 2015). This represents an extraordinary
opportunity to incorporate large-scale retrospective analyses of restoration outcomes
directly into the decision-making framework. In this way, the amount of information about
the recovery process is maximized, allowing for a more complete conceptual model of the
system, and specific decision criteria to guide the prioritization of passive versus active
management approaches before a project is implemented.
The objective of this study was to summarize how retrospective analyses of existing
restoration data enhance a typical decision-making process, in particular for guiding when
to invest additional time and money into active management. To illustrate this approach in
a seabird island restoration context, we use the New Zealand archipelago as a case study.
We focus on New Zealand due to the prevalence of burrowing seabirds, seabird-driven
island systems, the country’s rich history of predator eradications and island restoration,
and the cultural significance of seabirds to Ma¯ori, New Zealand’s indigenous peoples
(Moller et al. 2000; Taylor 2000; Mulder et al. 2011a; Towns 2011; Buxton et al. 2014).
General decision-making framework
Decision-making frameworks are not new to prioritizing restoration interventions, but to
our knowledge, are rarely used in island restoration (Helmstedt et al. 2016). Moreover, the
formal use of previous restoration outcomes in decision frameworks when evaluating
active versus passive management options is rare, notably for animal communities (but see
Richardson et al. 2009 for decision-making framework for assisted colonization). We
outline how initial analyses of existing restoration project outcomes can make each step of
the decision-making process more efficient:
Biodivers Conserv (2016) 25:467–484 469
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(1) Outline SMART project goals/objectives: Setting explicit goals and objectives is the
integral starting point of restoration planning. By defining ‘‘SMART’’ (specific,
measurable, attainable, relevant, and time-bound) goals and assessing the likelihood
of achieving them at the outset of a restoration project, planning becomes efficient
and step-wise, and allows measurement of success and return on investment (Hobbs
et al. 2011). Assessing the effectiveness of previous restoration alternatives can help
refine more robust and widely-accepted current restoration objectives. For example,
by basing current objectives on past restoration efforts with high success rates,
restoration plans are more likely to be cost-effective and thus receive financial and
public support (Tunstall et al. 1999; Wilson and Bruskotter 2009; Kettenring and
Adams 2011);
(2) Generate a conceptual systems model: In a restoration context, conceptual models
outline a set of factors influencing the restoration target (i.e. population, community,
or ecosystem), providing a simple visual or mathematical means to examine how a
system may respond to management interventions (Heemskerk et al. 2003;
Margoluis et al. 2009). Studies of chronosequences or comparative analyses of
previous restoration outcomes can be used to gain insight into the mechanics of the
recovery process, namely the natural ability of species and ecosystems to recover
after passive restoration management;
(3) Project design– identify and prioritize management approaches: Restoration
management interventions range from removal of the primary stressor to construc-
tion of novel ecosystems (Hobbs and Cramer 2008). Passive management may
consist of little more than monitoring, as defined by the life histories of the system
components being monitored. Active management is becoming more common at
highly disturbed sites where the approach is feasible. Where more than one approach
is available, the choice of which management intervention should be used on each
species, site, or ecosystem requires a structured analysis of the probability of success
of each alternative approach (Tear et al. 2005). Attributes affecting the probability
of natural recovery can be identified and tested empirically to determine their
relative contribution to a restoration approach successfully achieving the desired
outcome (LoSchiavo et al. 2013). In many cases, collecting field data to model how
a system may respond to a management intervention is not possible, either because
the response to restoration is beyond the timescale of a current project or there are
few comparable reference sites. Thus, a retrospective study of the successes and
failures of similar projects, if available, can be a valuable source of information
from which a set of key management decision criteria can be identified;
(4) Implement project and monitor outcomes: Monitoring outcomes of a restoration
approach allows informed revision of project goals, conceptual models, and
prioritization of future activities (i.e. adaptive management; Westgate et al. 2013). A
project which is established as adaptive will be subject to fewer confounding factors
than one that uses retrospective analyses of restoration outcomes (Armstrong et al.
2007; Lindenmayer et al. 2011). However, given the time frames associated with
ecosystem recovery, adaptive management is likely to be less demonstrably
effective in the short-term, highlighting the value of examining data from previous
restoration projects (Lewis 2000; Jones and Schmitz 2009).
470 Biodivers Conserv (2016) 25:467–484
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Applying the framework: New Zealand seabird island restoration
We illustrate a decision-making framework that incorporates retrospective analyses using
the restoration of seabird-dominated island ecosystems in New Zealand. With an abun-
dance of pre-existing restoration projects and a clear connection between seabird popu-
lations and island ecosystem functioning, a comparative analysis of seabird population
response to eradication around the New Zealand archipelago can provide information to
guide subsequent seabird island restoration management decisions.
New Zealand has over 700 islands larger than 1 ha and no native land mammals other
than two extant bat species (Parkes and Murphy 2003). Since human colonization, over
three-quarters of these islands have been modified substantially through burning, clearing,
and the introduction of alien vertebrate species (Parkes and Murphy 2003; Bellingham
et al. 2010). Seabirds were once abundant and widespread around New Zealand, which still
has the highest diversity of seabird species in the world (Taylor 2000). However, predator
introduction has resulted in the extinction, extirpation, or severe reduction of seabird
populations (Holdaway and Worthy 1994; Taylor 2000; Veitch et al. 2004). Marine
nutrient subsidies and in particular soil bioturbation by seabirds are central to island
ecosystem functioning, and the loss of seabird populations due to predation has trans-
formed ecosystem structure (Fig. 1; Mulder et al. 2011a; Orwin et al. 2015).
New Zealand is acknowledged internationally as a leader in island conservation and pest
eradication with over 100 islands around the archipelago cleared of all alien animals
(Towns et al. 2013). The country also has a long history of active restoration, including
nearly 260 species transfers since the 1960 s, 16 of which were seabird species (Craig et al.
2000; Miskelly and Powlesland 2013).
Because of the widely recognized damage caused by alien mammals in New Zealand,
the public is highly aware of conservation issues and community groups are increasingly
involved in eradication and restoration (Forgie et al. 2001). However, general under-
standing of the role and diversity of seabirds is low (Seabrook-Davidson and Brunton
2014). Moreover, increasing co-governance of natural resources by Ma¯ori in New Zealand
has led to greater consideration of their values, customary approaches, and practices within
island management systems (Newman and Moller 2005; Lyver et al. 2009; Lyver et al.
2015a). Some Ma¯ori tribes harvest the chicks of sooty shearwaters (Puffinus griseus)or
grey-faced petrels (Pterodroma gouldi), which are an important source of food, trade,
cultural identity, and social cohesion (Taiepa et al. 1997; Lyver et al. 2008; Moller 2009).
Historical relationships with the Pacific rat or kiore (Rattus exulans) add further com-
plexity to island restoration planning. Kiore were brought to New Zealand by Polynesians
c. 1280 (Wilmshurst et al. 2008), used by Ma¯ori as a food source, and feature prominently
in their traditions, proverbs, and prayer (Haami 1993). Evidence suggests that kiore sup-
press native flora and fauna and that their eradication results in recovery of threatened
species (e.g. Towns 2009). Yet for some Ma¯ori, the extinction of kiore would represent a
the loss of a cultural treasure (Chanwai and Richardson 1998).
New Zealand’s Department of Conservation (DOC) is charged with managing the
country’s biological and historic heritage. DOC’s mandates include active management
interventions, including restoration of island ecosystems; advocacy; and integrating the
rights of local Ma¯ori (Towns et al. 1990b). However, DOC has undergone repeated large-
scale restructures ever since its inception. The most recent, and most profound, was in
2011–2013, which included a structural model aimed at increased participation from
business, non-government agencies, and community groups (Bushnell and Pratt 2014). As
Biodivers Conserv (2016) 25:467–484 471
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a result, DOC may delegate restoration projects to outside agencies more frequently than in
the past. Given the high cost of seabird island projects, this could have consequences for
the future financial security of island restoration.
Restoration goals
As the number of New Zealand islands from which alien predators have been eradicated
has increased, island restoration goal setting has evolved. To illustrate this shift, we col-
lated restoration goals from islands with predators eradicated around New Zealand from
1934 to 2011 (Table 1; Supplementary material S1). Using an updated list of predator
eradication projects (Buxton et al. 2014), we searched for goals in published literature,
restoration plans, threatened species plans, and eradication plans. The most common early
(prior to 2000) goals of eradication were to improve eradication techniques, protect key
species, or enable recovery of native species by removing predation pressure. Within the
past decade, goals such as working with Ma¯ori to promote and support cultural values and
aspirations (i.e. ‘bi-cultural management’ and ‘protect and conserve cultural aspects’;
Table 1), education and outreach, restoring ecosystem functioning, and creating self-
Fig. 1 Seabird island restoration in New Zealand to illustrate use of decision framework. The ecological,
social and cultural contexts of the proposed project are summarized at the outset, generating a long-term
SMART goal. A conceptual model of burrow-nesting seabird population recovery, constructed using models
of metapopulation dynamics and information on drivers of intrinsic population growth, guides key decision
criteria. To choose management approaches, these criteria predict whether natural recovery of burrow-
nesting seabirds is likely and, if not, which active management methods should be used. Alternative
approaches can then be compared for likely cost-effectiveness using a simple prioritisation metric. Both
short- and long-term outcomes should be monitored, using a robust sampling scheme to report progress and
to manage the project adaptively over time. (TEK =traditional ecological knowledge)
472 Biodivers Conserv (2016) 25:467–484
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sustaining populations of rare plants and animals have become more prevalent. Generally,
we found few goals that conformed to SMART criteria.
In the earlier days of ecological restoration planning in New Zealand, goals were
dominated by a Eurocentric historical perspective, where the endpoint of a restoration
program was a biotic community that was assumed to represent the pre-human past
(Atkinson and Towns 1990; Towns et al. 1990b). However, palaeoecological investigations
indicate that, before human colonization, some offshore islands were dominated by
podocarp forests which have no modern analogue (Wilmshurst et al. 2014). This evidence,
and the cultural significance of islands for Ma¯ori, suggests that pre-human island
restoration targets may be neither feasible nor desirable (Bellingham et al. 2010; Lyver
et al. 2015b). Thus, contemporary post-eradication island restoration goals in New Zealand
include: maintaining ecosystem processes, preventing extinction, improving species and
functional diversity, forming partnerships with local Ma¯ori, preserving historic and cultural
Table 1 Summary of reported post-predator eradication goals of island restoration initiatives in New
Zealand in published literature and technical plans between 1936 and 2011. Among all years and islands
where the goal was reported, the ‘mean report date’ represents the average
Context Goal Total
reports
Earliest
report
Most recent
report
Mean report
date
Ecological None 10
Test eradication techniques 37 1978 2012 1994
Protect specific species 51 1970 2008 1995
Determine the impacts of predation 8 1979 2005 1996
Seabird recovery/restoration 21 1985 2014 1996
Reintroduce natural flora and fauna 44 1970 2014 1996
Restore communities 11 1990 2012 1997
Reintroduce functional groups 1 1999 1999 1999
Enable recovery by removing
predation pressure
63 1946 2014 2000
Research reintroductions 3 1996 2007 2000
Restore pre-human state 3 1999 2003 2000
Create a refuge for threatened
species
63 1970 2000
Reforestation 14 1982 2014 2003
Create self-sustaining population of
rare species
4 1999 2012 2007
Economic Ecotourism 6 1982 2012 2001
Restore ecosystem functioning 21 1990 2014 2003
Social Community participation 10 1990 2014 2001
Education outreach 29 1982 2013 2004
Bi-cultural management 22 1978 2013 2005
Reinstate sustainable cultural
harvest of seabirds
2 1998 2012 2005
Protect and conserve historic sites 22 1990 2014 2006
Protect and conserve cultural aspects 11 1978 2012 2009
Biodivers Conserv (2016) 25:467–484 473
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values, and fostering community engagement, enjoyment, and recreational use (Table 1;
Lee et al. 2005; Department of Conservation 2010).
Although seabird population recovery is likely to be a significant component of
attaining these modern island restoration goals (Moller 2010; Mulder et al. 2011a), few
early restoration initiatives acknowledged its importance (Table 1). Appropriate baseline
data were therefore rarely collected to allow seabird recovery outcomes to be assessed.
Because seabirds, especially petrels and shearwaters, exert a large influence on island
ecosystems in New Zealand, we propose that seabird re-colonization and colony growth
may be beneficial SMART post-eradication outcomes for many offshore islands (e.g. see
Imber et al. 2003). A project may select a particular species, the recovery of which is
measurable within the respective species’ generation time (approximately 15 years for
petrels and shearwaters; IUCN 2012). Longer-term (e.g. [30 years) outcomes for seabird-
driven island restoration could include re-establishing island ecosystem functioning and
community structure, and, on some islands, the re-instatement of cultural harvest (Fig. 1).
Conceptual systems model for the restoration of NZ’s seabird islands
Assuming that the restoration goal of seabird recovery is selected, reliable prediction of the
response of seabird populations to management interventions at breeding sites requires a
basic understanding of local seabird population dynamics (Margoluis et al. 2009). In this
way, a conceptual model of seabird recovery can be constructed by identifying the drivers
of population growth (Buxton et al. 2014).
Generally, seabirds form metapopulations, with breeding sites separated by water
barriers and where local population growth rates are affected by both intrinsic and extrinsic
dynamics (McCullough 1996; Matthiopoulos et al. 2005). Intrinsically, seabirds have low
annual reproductive output, fecundity is low, and intermittent breeding is common,
resulting in low rates of per capita growth (Warham 1990; Cubaynes et al. 2011). Intrinsic
negative density-dependence may result from limitations in breeding sites and food
(Croxall and Rothery 1991; Sandvik et al. 2012), while positive density dependence may
be associated with coloniality (Kildaw et al. 2005). Moreover, despite being highly mobile,
behavioral mechanisms associated with coloniality (e.g. philopatry and social attraction)
mean that the number of immigrants recruiting from other colonies is thought to be low
(Hamer et al. 2002). Thus, population dynamics are likely to be slow and characterized by
traits that may not conform to metapopulation theory (Matthiopoulos et al. 2005). When
predators are removed from an island, local seabird population growth (i.e. passive
recovery) will depend on a number of factors; for example, the size of the remnant
population at the time of eradication, species-specific life history traits (such as age of first
reproduction), local habitat quality, and density-mediated immigration rates. However,
because of seabirds’ slow population growth, evaluating how these parameters drive
population recovery is problematic within the time constraints of a restoration project.
Towns (2002) and Towns et al. (2012) recommend long term environmental monitoring
of islands that escaped predator invasion to build conceptual models of recovery. In the
absence of robust long-term monitoring data, chronosquence analyses or ‘space-for-time
substitutions’ have been used to compare ecosystem function or population dynamics
among islands with different periods since pest eradication (Fukami et al. 2006; Jones
2010a; Jones 2010b). Although caution must be taken when examining successional
processes using chronosequences (Johnson and Miyanishi 2008), if results can be validated
from data using other methods (e.g. monitoring before and after eradication), they could
represent a useful proxy for temporal recovery dynamics (Towns 2009). Given the
474 Biodivers Conserv (2016) 25:467–484
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abundance of eradication projects around New Zealand, chronosequences have been useful
in identifying factors driving post-eradication seabird response (Buxton et al. 2014).
Generally, we have a poor understanding of how seabirds and ecosystems respond to
introduced predator eradication and active restoration (Mulder et al. 2009). In the past,
threatened species conservation has assumed greater urgency than consideration of the
impact of such species on the island’s existing biota. For example, the reintroduction of
natural seabird predators (e.g. tuatara Sphenodon sp.) at too early a stage in recovery may
result in reduced probability of the re-establishment of small seabird species (Atkinson and
Towns 1990; Towns et al. 1990a). Moreover, if seabird predators are introduced before the
recovery of their seabird prey, this may result in lower probability of establishing the
predator itself.
Management techniques
The most widespread and effective passive seabird island restoration technique in New
Zealand is the eradication of alien predators (Towns and Broome 2003; Clout and Russell
2006; Broome 2009). In some cases, abundant native avian predators may also prey on
breeding seabirds, and populations must be controlled to reduce their impact on threatened
seabirds (Miskelly 2013). The temporary cessation, ra
¯hui, of traditional harvest from
seabird islands where Ma¯ori are owners or have customary rights may also be employed to
assist population maintenance or recovery (Moller 2006; Kitson and Moller 2008; Lyver
et al. 2015a).
After eradication, a number of active techniques can be used to restore habitat and
encourage seabird population recovery. Both exotic weed control and revegetation of
indigenous plant communities can facilitate seabird recovery actively through habitat
enhancement (Towns et al. 1997; Forbes and Craig 2013). Chick translocation to a
restoration site, before the age where they imprint to their natal site, has been successful for
a number of seabird species (Jones and Kress 2012). However, because of the high cost and
labor requirements of transportation and feeding, plus an average lag time of five years
until birds return to a restoration site to breed, translocation is expensive and outcomes are
difficult to predict (Miskelly et al. 2009). Social attraction (e.g. playback of recorded
vocalizations or decoys; Kress 1998), where birds are lured to formerly-occupied or new
breeding habitat by mimicking social cues is emerging as a cost-effective alternative
technique on some islands in New Zealand (Sawyer and Fogle 2010). Because remnant
colonies of seabirds may have skewed species composition, resulting inter-specific com-
petition may require active management (Buxton 2014). For example, widespread inter-
ference competition between broad-billed prions (Pachyptila vittata) and endangered
Chatham petrels (Pterodroma axillaris) on South-east Island, New Zealand, poses the most
serious threat to the latter (Was et al. 2000). Restoration managers have intervened by
culling broad-billed prions and constructing Chatham petrel nest-boxes that exclude broad-
billed prions (Gummer et al. 2015). Similarly, an active technique used by some Ma¯ori
tribes has involved the splitting of existing burrows or drilling of new ones to increase
breeding habitat for burrowing petrels (Lyver et al. 2008).
Prioritizing management approaches
Assuming the goal of restoration is to encourage recovery of seabird populations, a series
of decisions will need to be made relating to how population recovery can best be
achieved. Managers will first need to consider a conceptual model of the key factors
Biodivers Conserv (2016) 25:467–484 475
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influencing natural seabird population recovery. This includes factors influencing the
recruitment of individuals from other islands to the newly predator-free space as well as the
intrinsic rate of growth of any remnant colony. We used a previously published compar-
ative analysis of over 100 seabird population responses to predator eradication on 41
islands (Buxton et al. 2014) to identify key decision criteria in choosing between active and
passive management alternatives (Fig. 2). These criteria are:
Distance to source population
Colony growth and re-colonization after eradication is reliant on the size, distance, and
connectivity to nearby source populations (Hanski 1998). Evidence suggests that recruit-
ment probability decreases exponentially with distance from a source population, and that
re-colonization is unlikely without a colony within *25 km (Oro and Pradel 1999; Buxton
et al. 2014).
Fig. 2 Ecological decision tree to guide management interventions for burrow-nesting seabirds based on
the probability of natural recovery following predator eradication. Active (green) versus passive
management (yellow) can be chosen based on site- or species-specific characteristics. Asterisks indicate
that the majority of criteria are met (‘‘yes’’) or not met (‘‘no’’). Further considerations include cost and
stochastic events. (Color figure online)
476 Biodivers Conserv (2016) 25:467–484
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Remnant colony size
seabird colony growth is regulated by both positive density dependence, where birds are
attracted to settle in larger pre-existing colonies, and negative density dependence, where
crowding decreases habitat quality at higher densities (Kildaw et al. 2005; Buxton et al.
2016b). Thus recovery is more likely in a mid-sized remnant colony (*25–100 breeding
pairs; Buxton et al. 2014).
Metapopulation status
Growth of a local population after eradication is more likely if the species’ overall
metapopulation is stable or increasing in size (Oro 2003).
Presence of other breeding seabird species
Not only are seabirds more likely to settle amongst colonies of conspecifics, colonies of
other species may also be attractive; re-colonization is more likely on islands with higher
seabird species richness (Parejo et al. 2005; Buxton 2014). This relationship may reflect
habitat quality, where recovery is more likely on islands with less habitat alteration (e.g. no
grazing by cattle), or heterospecific habitat copying, where prospecting individuals of one
species respond to the presence of other species with similar ecological needs (Wagner
et al. 2000).
If a potential restoration site does not meet these criteria for the focal species, active
methods should be considered. Some species and sites may be better candidates for a less
logistically demanding active approach (e.g. social attraction; Fig. 2). For example, grey-
faced petrels readily disperse from their natal site and will settle and breed at sites with
call-playback (Sawyer and Fogle 2010; Lawrence et al. 2014; Buxton et al. 2015b). For
grey-faced petrels and fluttering shearwaters (Puffinus gavia), responses to playback
increases if larger, denser colonies are closer to the restoration site. In instances where
species exhibit strong site fidelity, the restoration site is far removed from a source pop-
ulation, or species are critically threatened and require immediate intervention to prevent
extinction, chick translocation is more likely to achieve recovery objectives (Fig. 2;
Miskelly et al. 2009). Conversely, the biology of some species may inhibit the success of
active restoration (e.g. unfeasibly complex post-fledgling care or diet) and must also be
considered during prioritization.
The ability of a remnant colony to grow will also depend on local habitat suitability.
Although burrow-nesting seabirds’ nesting habitat selectivity can weaken as a colony
grows, birds use deeper soils, steeper slopes facing prevailing winds, and later-succession
vegetation types more readily than other habitat characteristics (Whitehead et al. 2014;
Buxton et al. 2015a). Some features of a proposed restoration site might preclude con-
sideration of restoration or demand intensive preparation (e.g. revegetation) prior to any
attempts at species manipulation. Once a species is established, the nature and intensity of
interactions between sympatric species in a community can vary, particularly if selective
predation pressure has skewed community structure (Buxton 2014). Some burrow-nesting
seabird species may competitively exclude others, especially when communities are re-
assembling after eradication (e.g. grey-faced petrel and little shearwater Puffinus assimilis
Pierce 2002; Buxton 2014). In this case, some species may remain rare because they cannot
withstand inter-specific competition, especially if the population of the dominant species
Biodivers Conserv (2016) 25:467–484 477
123
continues to increase (Oro et al. 2009). These factors must be considered in advance of any
decision to embark on a project (Fig. 2).
At this point in the planning process, the probability of restoration success should be
analyzed in a bioeconomic framework to weigh cost and benefit (see Joseph et al. 2009;
Jones and McNamara 2014 for details of accessible methods).
Implementation and monitoring
Although monitoring is an integral component of management activity, the outcomes of
eradication are rarely measured in New Zealand (Jones et al. 2011). If SMART project
outcomes are defined clearly, indicators of success at each stage should be easily identi-
fiable and monitoring can be included as part of the project planning. Without appropriate
monitoring it is impossible to determine whether or not the project has been a success or to
manage the project adaptively through time. Moreover, monitoring data can help validate
predictions made by analyzing recovery across chronosequences.
For seabirds, there are numerous logistical challenges that preclude simple monitoring
strategies, including cryptic nesting behavior (below-ground nesting and nocturnal colony
attendance; Warham 1990), high costs of getting to island breeding grounds, and high
interannual variability in breeding participation (Newman et al. 2009). However, new
technologies are being applied to seabird monitoring, including infra-red burrow camera
surveys (Hamilton 2000), automated acoustic sensors (Buxton and Jones 2012; Borker
et al. 2014), and radar (Gauthreaux and Belser 2003), while modern simulation techniques
allow greater refinement of monitoring data (Buxton et al. 2016a). Moreover, there is a
well-established tradition of community involvement in conservation in New Zealand
(Peters et al. 2015) and an increasingly strong recognition of the role of Ma¯ori traditional
knowledge in monitoring seabird populations (Kitson 2004; Moller et al. 2004). Partici-
patory approaches are also valuable, as public involvement enhances restoration accep-
tance, increases the capacity for active restoration through volunteering, and builds
capacity through education and outreach (Parkes and Panetta 2009; Hardie-Boys 2010).
Discussion
Deciding which restoration techniques to implement on islands is complicated: restoration
goals are steeped in conflicting socio-political and cultural contexts, recovery trajectories
of species are difficult to predict, restoration is expensive, and budgets are limited.
Decision-making tools are well-suited to help managers decide objectively between
restoration alternatives under complex circumstances. When it comes to decision-making,
island systems are at a unique advantage– a wealth of previous restoration outcome
information exists that can be incorporated readily into decision-making frameworks.
Accordingly, the formal inclusion of retrospective analyses of species response to past
restoration approaches can contribute to several stages of a decision-making framework:
successes can increase the socio-economic acceptance of a project, while comparative
analyses can identify factors driving species recovery and determine the likelihood of the
success of alternative restoration approaches. Identifying a set of key criteria based on
previous restoration data can help create step-wise decision process to select among
restoration approaches, without the need for further field data to be collected.
478 Biodivers Conserv (2016) 25:467–484
123
Due to the prevalence and success of predator eradication around New Zealand, a
comparative analysis of seabird population responses after eradication revealed drivers of
recovery, site characteristics, and species where natural recovery is more likely. Con-
versely, this indicated situations where active intervention should be prioritized. Retro-
spective analyses require results from previous projects to be made available, even if those
projects were unsuccessful. Accordingly, this approach depended on the ‘grey’ literature,
given the bias in publishing only success stories (Csada et al. 1996). Decision-making
within the seabird island restoration process would be based, ideally, on assessment of
passive and active restoration outcomes planned initially as tests of alternative approaches
(i.e. adaptive management; Williams 2011). However, the timeframe required for this
adaptive approach may be unrealistic for restoration decisions involving long-lived sea-
birds. The existence of numerous pre-existing passive and active restoration projects means
that their outcomes can be incorporated into prioritizing current restoration approaches,
while monitoring outcomes of the current project can validate or refine predictions made
using previous projects. The seabird island decision framework outlined in this paper can
guide managers’ choice of tools to facilitate seabird recovery.
Since its inception as a discipline in conservation biology 60 years ago, island
restoration has evolved from a field that was dominated by removing threats and assuming
that populations will eventually recover, to one where the recovery of ecosystems and
wildlife can be facilitated actively and outcomes monitored. As we move forward with
restoring complex island communities, decision-making tools using a logical, step-wise
framework informed by previous restoration successes and failures will aid in increasing
understanding and reducing uncertainty of ecosystem response.
Acknowledgments We thank C. Stone, D. Hamon, and H. Moller for helpful discussions and editing.
P. Gaze, J. Greenman, E. Dunning for contributing restoration goals and unpublished data. This project was
funded by the New Zealand Ministry of Business, Innovation and Employment (Te Hiringa Tangata Ki Te
Tai Timu Ki Te Tai Pari—facilitating bicultural restoration of coastal forests using seabirds as ecosystem
engineers contract C09X0908) and the University of Otago.
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