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1Glen et al.: Wide-scale predator control for biodiversity
Predator control on farmland for biodiversity conservation: a case study from
Hawke’s Bay, New Zealand
Alistair S. Glen1* , Mike Perry2, Ivor Yockney2, Sam Cave2, Andrew M. Gormley2 ,
Campbell Leckie3, Rod Dickson3, Wendy Rakete-Stones3, Pouri Rakete-Stones3, Grant L. Norbury4
and Wendy A. Ruscoe2,5
1Manaaki Whenua - Landcare Research, Private Bag 92170, Auckland 1142, New Zealand
2Manaaki Whenua - Landcare Research, PO Box 69040, Lincoln 7640, New Zealand
3Hawke’s Bay Regional Council, 159 Dalton St, Napier 4110, New Zealand
4Manaaki Whenua - Landcare Research, PO Box 282, Alexandra 9340, New Zealand
5Institute for Applied Ecology, University of Canberra, ACT 2601, Australia
*Author for correspondence (Email: glena@landcareresearch.co.nz)
Published online: 26 November 2018
New Zealand Journal of Ecology (2019) 43(1): 3358 © 2018 New Zealand Ecological Society.
DOI: 10.20417/nzjecol.43.8
RESEARCH
Abstract: Invasive predator control to protect native fauna usually takes place in native habitat. We investigated
the effects of predator control across 6000 ha of multi-tenure, pastoral landscape in Hawke’s Bay, North Island,
New Zealand. Since 2011, low-cost predator control has been conducted using a network of kill traps for mustelids
(Mustela spp.), and live trapping for feral cats (Felis catus). Although not deliberately targeted, other invasive
mammals (particularly hedgehogs Erinaceus europaeus) were also trapped. We monitored predators and native
prey in the predator-removal area and an adjacent non-treatment area. Predator populations were monitored
using large tracking tunnels, which also detected native lizards. Invertebrates were monitored using articial
shelters (weta houses). Occupancy modelling showed that site use by cats and hedgehogs was signicantly lower
in the predator-removal area than in the non-treatment area. Site use by mustelids also appeared to be lower in
the treatment area, although sample sizes were too small to allow rm conclusions. Site use by invasive rats
(Rattus spp.) was higher in the treatment area, while that of house mice (Mus musculus) showed no difference
between treatments. There was evidence of positive responses of some native biodiversity, with site use by
native lizards increasing signicantly in the treatment area, but not in the non-treatment area. Counts of native
cockroaches were higher in the treatment area, but other invertebrates were detected in similar numbers in both
areas. Our results show that low-cost predator control in a pastoral landscape can reduce invasive predator
populations, with apparent benets for some, but not all, native fauna.
Keywords: feral cat, invasive predators, invertebrates, landscape-scale, lizards, mustelids, rodents
Introduction
Invasive predators are controlled to protect native fauna in many
parts of New Zealand (e.g. Innes et al. 1999; Reardon et al.
2012; Russell et al. 2015). However, predator control is usually
in conservation reserves, wildlife sanctuaries or remnants of
native habitat. Few published studies have investigated the
effects of controlling predators for conservation purposes in
multi-tenure, pastoral landscapes.
Although landscape-scale predator control may be
desirable, nancial and logistical challenges often prevent it.
The tools and techniques used to control predators at localised
scales (e.g. exclusion fencing; Innes et al. 2012; Hayward et al.
2014) may be prohibitively expensive at the landscape scale
(Norbury et al. 2014). Managing wildlife across different land
tenures can also be challenging, both logistically and socially.
Access to private property may not always be feasible, and
landholders may vary in their attitudes towards proposed
management activities (Epanchin-Niell et al. 2009; Glen et
al. 2017). Practical and affordable methods are needed to
reduce the impacts of invasive predators across multi-tenure,
pastoral landscapes.
We controlled invasive predators over 6000 ha of
farmland with fragments of native bush adjacent to an 800-ha
conservation reserve where intensive predator control had been
in place since 1996. The primary targets of the trapping were
feral cats (Felis catus) and mustelids (stoats Mustela erminea,
ferrets M. furo and weasels M. nivalis); however, large numbers
of other invasive mammals, particularly hedgehogs (Erinaceus
europaeus), were also trapped. By removing invasive predators
from a pastoral landscape with fragments of native forest,
we aimed to facilitate recovery of native fauna. We predicted
that predator control would lead to increased abundance and
distribution of native prey species. For example, many of New
Zealand’s lizard and invertebrate taxa have declined due to
the impacts of mammalian predators (Hitchmough et al. 2010;
2 New Zealand Journal of Ecology, Vol. 43, No. 1, 2019
Stringer & Hitchmough 2012). Here we describe trends in
predator populations and native biodiversity following this
landscape-scale intervention.
Methods
Study area
Our study took place on four adjacent pastoral properties in
Hawke’s Bay, North Island, New Zealand: Opouahi, Rangiora,
Toronui and Rimu Stations (39⁰10′ S; 176⁰46′ E). These sheep
and cattle stations are mainly covered by introduced pasture
grass with fragments of native beech forest (Fuscospora
solandri). Beech forest fragments range in size from about 10
to 100 ha. Adjoining the study area to the north is Boundary
Stream Mainland Island, an area of mixed broadleaf and
podocarp forest managed by the Department of Conservation
(DOC). Elevation in the study area ranges from about 300
to 1000 m a.s.l. Invasive predators have been controlled in
Boundary Stream since 1996 as part of DOC’s Mainland
Island programme (Saunders & Norton 2001; Abbott et al.
2013). There was no recent history of predator control on the
adjacent pastoral properties. Predator control was implemented
on Opouahi and Rangiora Stations while Toronui and Rimu
Stations were non-treatment areas (Fig. 1). The treatment and
non-treatment areas were similar in terms of habitat, although
the treatment area was directly adjacent to Boundary Stream,
where invasive predators are also suppressed.
Predator control
Predator control was conducted by Hawke’s Bay Regional
Council (HBRC). In November 2011, 680 kill traps were
deployed across an area of 6000 ha. These included 550
DOC-250 traps (DOC, Wellington) for mustelids, and 130
Timms traps (KBL Rotational Moulders, Palmerston North)
for cats. Traps were spaced 100 m apart in bush fragments
or 200 m apart on cleared farmland, based on the assumption
that predators were more likely to be found in bush fragments
(e.g. Alterio et al. 1998; King & Murphy 2005; Harper 2007;
Garvey 2016). Traps were baited with various combinations of
fresh rabbit meat, a rabbit-based paste (Erayz®, Connovation
Ltd, Auckland) or a synthetic, rat-scented lure (Goodnature
Ltd, Wellington). To minimise labour costs, traps were set in
locations that were easily accessible by an all-terrain vehicle
(ATV). The DOC-250 traps were modied to include a built-
in handle for quick re-setting. The position of the handle also
served as a visual signal to indicate whether the trap had been
triggered, eliminating the need to inspect each trap closely.
These modications were rened during the course of the
project, and will be described in detail in a separate publication.
Traps were initially checked every 3 weeks at an annual cost of
$5.53 ha-1; however, from November 2014, they were checked
four times a year (January, April, June and November), which
cost $2.30 ha-1 (HBRC, unpubl. data).
The DOC-250 traps were left in place for the duration of
the study. The Timms traps were left in place for the rst year,
after which cat control was conducted in two annual pulses
(May and August each year). The pulsed cat control was
carried out using a combination of live traps (cage (Havahart
Traps, Lititz, Pennsylvania), leg-hold (Victor #11/2 soft-catch,
Oneida Victor, Cleveland, Ohio)), kill traps (Timms and
Possum Master traps (Possum Master Industries, Tauranga)),
and opportunistic shooting. Live traps were checked daily
and captured predators were euthanased. Cat control targeted
areas of high rabbit (Oryctolagus cuniculus) activity as rabbit
abundance is a strong predictor of cat abundance (Norbury
& McGlinchy 1996; Norbury et al. 2002; Cruz et al. 2013).
Monitoring
In October 2011, we established 15 monitoring lines in the
treatment area and 14 lines in the non-treatment area to assess
trends in populations of invasive predators and native prey.
However, due to access restrictions, the number of monitoring
lines in the non-treatment area was reduced to 12 from spring
2014 onwards. Each line consisted of ve large tracking
tunnels (see below) spaced 100 m apart, spanning the interface
between a native bush fragment and the adjacent pasture. The
rst point was inside the bush fragment, 200 m from the edge,
the third point was on the edge of the fragment, and the fth
point was in cleared pasture, 200 m outside the fragment.
Where possible, monitoring lines were at least 1 km apart to
improve spatial independence; however, steep topography
made this impracticable in some cases. The shortest distance
between any two monitoring lines was 500 m.
Figure 1. Map of the study area
showing the treatment and non-
treatment areas relative to Boundary
Stream Reserve in Hawke’s Bay,
North Island, New Zealand. The
locations of kill traps are indicated
by dots.
3Glen et al.: Wide-scale predator control for biodiversity
We used large tracking tunnels (20 × 20 × 100 cm) with a
removable oor, as described by Pickerell et al. (2014). Tracking
ink (Black Track, Pest Management Services, Wellington) was
applied to the oor in the middle of each tunnel, and sheets of
tracking paper (18 × 30 cm) were fastened to the tunnel oor
at each end with bulldog clips and drawing pins. Each tunnel
was baited with a cube of fresh rabbit meat in the middle of
the tracking ink. Tracking papers were retrieved after 3 days
and labelled with tunnel number and date; tunnels were left
in place year-round. Footprints left on the tracking papers
were identied using eld guides (Gillies & Williams 2002;
Agnew 2009; NPCA 2014).
The rst and third point on each monitoring line also had
an articial shelter (wētā house) for monitoring invertebrates.
Wētā houses were 7.5 × 62 cm, with six galleries, a clear
Perspex cover and a wooden door (Fig. 2). These were attached
to tree trunks at approximately chest height and left in place
year-round. By opening the wooden door, we were able to
count and identify invertebrates through the Perspex cover.
Monitoring lines were checked twice per year (spring and
summer) from 2011–2014. In 2015 and 2016 we sampled only
once each year (in summer).
Data analysis
We analysed the tracking tunnel data using an occupancy
modelling approach (MacKenzie et al. 2006) in which each
monitoring line was treated as a site. Although we placed
monitoring lines as far apart as practicable, we cannot rule
out the possibility that individual predators were detected on
more than one line. Occupancy models usually assume spatial
independence between sites; when this assumption is relaxed
the models estimate the proportion of the area used by the
target species during the sampling period, which we refer to
as ‘site use’ (MacKenzie & Royle 2005). Within a monitoring
line, each tracking tunnel was treated as a separate survey so
that each monitoring line yielded a detection history with ve
‘occasions’ per season. For example, if a species was detected
in the rst and last tunnel in a line, this yielded a detection
history of ‘10001’. A detection history need not comprise data
on detection / non-detection at different times; it can also be
made up of data collected at different points within each site
(MacKenzie et al. 2006).
We used a multi-season dynamic occupancy model
(MacKenzie et al. 2003) to estimate site use for cats, hedgehogs,
mustelids, rats, mice and skinks (Scincidae) in each area and
sampling season. This model estimates the proportion of
sites used by each species, as well as the probability that a
species will disappear from a site where it previously occurred
(‘extinction’), or appear at a site where it had been absent
(‘colonisation’). Probabilities of colonisation, extinction
and initial site use were allowed to vary between treatment
and non-treatment. Our model allowed detection probability
to vary between tracking tunnels within a monitoring line.
However, this model failed to converge for cats and mustelids;
therefore, we estimated site use by these predators assuming
constant detection probability for all tracking tunnels within a
line. Analyses were conducted using the ‘unmarked’ package
(Version 0.11-0) in R (Version 3.4.4; Fiske & Chandler 2011).
We tested for differences between treatments by bootstrapping
the 95% condence intervals to estimate a P-value for the null
hypothesis that site use was the same between the treatment
and non-treatment area.
We tested for differences in the numbers of detections of
animals inside bush fragments, at the edge of fragments, or in
pasture. Pooling data from all sites and sampling occasions,
we used chi-squared contingency tests with the null hypothesis
that detections in each habitat would be proportional to the
number of tracking tunnels in each habitat (i.e. 40% bush;
20% edge; 40% pasture).
For invertebrates, we calculated the mean number per
monitoring line of each taxon counted in the wētā houses in
each sampling season. Values for each season were compared
between the treatment and non-treatment areas using paired
t-tests after using Levene’s test to conrm homogeneity of
variances. These tests were performed using Microsoft Excel
2010.
Figure 2. Schematic diagram of a wētā house used to monitor invertebrates. Invertebrates enter through the holes in the side and shelter
in the hollow galleries. When the door is open, invertebrates can be identied and counted through the Perspex cover.
4 New Zealand Journal of Ecology, Vol. 43, No. 1, 2019
Results
The kill traps captured cats, mustelids, hedgehogs, ship rats
(Rattus rattus) and rabbits. The pulsed cat control removed
134 cats and 21 ferrets (Table 1).
The tracking tunnels detected a range of invasive mammals,
including cats (n = 53 detections), mustelids (Mustela spp.; n
= 15), hedgehogs (n = 218), rats (Rattus spp.; n = 148), mice
(Mus musculus; n = 202) and possums (Trichosurus vulpecula;
n = 47). Tracking tunnels also detected skinks (n = 54). We
were unable to identify individual skink species; however,
species likely to be present in the area include the common
skink Oligosoma polychroma, spotted skink O. lineoocellatum
and small-scaled skink O. microlepis (Bell 2012; Abbott et
al. 2013; DOC 2018).
Site use estimates for cats (Fig. 3a) and hedgehogs (Fig.
3b) were similar in both areas during the rst sampling season,
before predator removal began. After predator removal, site
use estimates for these species were consistently lower in the
treatment area than in the non-treatment area. Bootstrapping of
the 95% condence intervals for the equilibrium probability of
occupancy (i.e. the long-run probabilities based on the function
of probabilities of site colonisation and extinction) showed that
these differences were statistically signicant (cats: treatment
(t) = 0.4, non-treatment (n.t.) = 0.99, P = 0.02; hedgehogs:
t = 0.49, n.t. = 0.99, P < 0.001). Mustelids were detected in
low numbers, leading to a high level of uncertainty in site use
estimates (Fig. 3c). Thus, although the mean estimate was
lower in the treatment area (0.09) than in the non-treatment
area (0.28), the difference was not signicant.
Site use by rats was initially higher in the treatment area,
and remained so for the duration of the study (t = 0.67, n.t.
= 0.39, P = 0.006; Fig. 3d). Mice showed no difference in
site use between the treatment and non-treatment area (0.58
for both treatments, P = 0.48; Fig. 3e). Skinks (Fig. 3f) were
not detected in either area before predator removal began.
However, site use by skinks increased rapidly in the treatment
area, while remaining near zero in the non-treatment area (t =
0.43, n.t. = 0.01, P < 0.001).
Cats were detected more often at the edge of bush fragments
than expected based on sampling effort (χ2 = 6.83, d.f. = 2, P
= 0.03; Fig. 4a). Mice were detected more often than expected
in edge or bush habitats (χ2 = 21.6, d.f. = 2, P < 0.0001; Fig.
4b), while rats (χ2 = 98.5, d.f. = 2, P < 0.0001) and possums
(χ2 = 20.87, d.f. = 2, P < 0.0001) were detected more often
in bush habitat (Fig. 4c, d). Skinks were detected more often
than expected at the edge of fragments or in pasture (χ2 = 16.4,
Table 1. Numbers of animals removed by kill trapping and
pulsed cat control on pastoral properties, November 2011
to November 2015.
____________________________________________________________________________
Species Number removed
____________________________________________________________________________
Kill trapping Pulsed cat
control
____________________________________________________________________________
Cat (Felis catus) 111 134
Ferret (Mustela furo) 51 21
Stoat (Mustela erminea) 90
Weasel (Mustela nivalis) 2
Hedgehog (Erinaceus europaeus) 748
Rabbit (Oryctolagus cuniculus) 431
Ship rat (Rattus rattus) 463
____________________________________________________________________________
d.f. = 2, P = 0.0003; Fig. 4e). Hedgehog detections showed no
signicant difference between habitats (χ2 = 4.19, d.f. = 2, P =
0.12; Fig. 4f). There was a trend for mustelids to be detected
more often in edge or bush habitats than in pasture (Fig. 4g),
but low sample size precluded statistical testing.
Taxa observed in wētā houses included tree wētā
(Hemideina spp.), cave wētā (Rhaphidodophoridae),
cockroaches (Blattodea), spiders (Araneae) and slaters
(Isopoda). No non-native invertebrates were recorded. During
the pre-treatment period, no invertebrates had occupied the
wētā houses. During subsequent seasons, counts of cockroaches
were higher in the treatment area (P = 0.001). No differences
were observed between treatments for any other invertebrate
taxon (Table 2).
Discussion
Our results show that extensive, low-cost trapping in a pastoral
landscape was associated with lower site use by invasive
predators (feral cats and hedgehogs), with apparent benets
for some native fauna (skinks and cockroaches). Although cat
control was not continuous year-round, we believe that a large
proportion of the cat population was removed. The total of 245
cats captured (Table 1) equates to one cat per km2 per year.
This density is roughly twice the estimated population density
of 0.49 cats per km2 for Hawke’s Bay farmland (Langham &
Porter 1991).
Table 2. Mean numbers of invertebrates recorded per monitoring line in wētā houses in the treatment and non-treatment
area during eight sampling sessions, February 2012 to November 2015. All t-values and P-values are for two-tailed, paired
t-tests with 7 degrees of freedom. Each monitoring line had two wētā houses; one 200 m inside a bush fragment and one
near the bush-pasture margin.
__________________________________________________________________________________________________________________________________________________________________
Taxon Mean count (±SD) per monitoring line t P
Treatment Non-treatment
__________________________________________________________________________________________________________________________________________________________________
Cockroaches (Blattodea) 1.3 ± 0.7 0.3 ± 0.3 -3.84 0.001
Spiders (Araneae) 1.5 ± 0.4 1.8 ± 0.6 1.32 0.21
Cave wētā (Rhaphidodophoridae) 1.5 ± 0.7 1.0 ± 0.6 -1.6 0.14
Tree wētā (Hemideina spp.) 1.5 ± 0.5 1.9 ± 0.9 1.1 0.3
Slaters (Isopoda) 0.1 ± 0.2 0 ± 0 n/a* n/a*
__________________________________________________________________________________________________________________________________________________________________
*Slaters were recorded in the treatment area in only one sampling season; therefore, numbers were too low for statistical analysis.
5Glen et al.: Wide-scale predator control for biodiversity
Figure 3. Site use during each sampling season (with 95% condence intervals indicated by grey shading) of (a) cats, (b) hedgehogs,
(c) mustelids, (d) rats, (e) mice and (f) skinks in the treatment (predator removal) and non-treatment areas. Predator removal began in
the treatment area after the rst sampling season.
(a) (b)
(c) (d)
(e) (f)
Developing low-cost methods to remove predators from
large areas is an essential step towards the vision of a Predator-
Free New Zealand (Russell et al. 2015). Our predator control
did not reduce predators to zero-density, but did achieve
measurable reductions in predator populations at a relatively
low cost compared to intensive predator control. This cost-
effectiveness was due to the innovative approach used. The
spatial coverage of our trapping effort was made possible
by placing traps in accessible locations where they could be
checked rapidly by staff on an ATV. This design maximised
the number of traps that could be checked in a day, thereby
increasing the area that could be trapped with the available
budget. Our network of kill traps also used mechanical signals
that allowed the trapper to see whether a trap had been triggered
without dismounting the ATV, saving time and reducing labour
costs. Recent developments in wireless sensor networks (Jones
et al. 2015) may further reduce costs of trapping by alerting
managers when a trap is triggered.
Our study is among the rst to use large tracking tunnels
for detecting invasive predators such as cats, mustelids and
hedgehogs (see also Pickerell et al. 2014). However, tracking
tunnels detected few animals during the rst sampling season.
This effect may have been due to neophobia as the tunnels
had been in place for only a few days. Detection rates were
much higher after 3 months, suggesting that this was sufcient
time for animals to become habituated to the tracking tunnels.
It is likely that predator site use was underestimated in the
rst sampling session because of neophobia; the apparent
increase in predator site use in the non-treatment area may be
an artefact of this initial underestimation. We believe predator
6 New Zealand Journal of Ecology, Vol. 43, No. 1, 2019
Figure 4. Numbers of detections of (a) cats, (b) mice, (c) rats, (d) possums, (e) skinks, (f) hedgehogs and (g) mustelids by large tracking
tunnels in bush, edge and pasture habitats from October 2011 to February 2016. The rst tracking tunnel in each monitoring line was
inside a bush fragment, 200 m from the edge (bush 1). Tracking tunnels were 100 m apart, with the third tunnel being at the bush-pasture
margin (edge), and the last being in pasture, 200 m beyond the edge of the fragment (pasture 2). Asterisks indicate habitats where species
were detected signicantly more frequently than expected by chance.
(a) (b) (c) (d)
(e) (f) (g)
site use at both sites during the pre-treatment period was
likely much higher than our estimates suggest, and probably
declined in the treatment area while remaining relatively stable
in the non-treatment area. Future trials should compare the
efcacy of large tracking tunnels with other tools for detecting
predators, e.g. camera traps and wildlife detector dogs (Glen
et al. 2014, 2016; Glen & Veltman 2018). Studies using large
tracking tunnels should include a longer period of repeated
sampling in the pre-treatment period to test for the possible
effect of neophobia and generate more reliable estimates of
pre-treatment site use or abundance. A longer deployment
period has been shown to increase the probability of detecting
predators in tracking tunnels, although this may require the
bait to be replaced periodically (Pickerell et al. 2014).
Ideally our study would have included spatial replication
(Underwood 1994); however, this is often unaffordable for
large-scale adaptive management programmes such as ours.
One solution would be to apply a treatment reversal (e.g.
Innes et al. 1999) in which the treatment and non-treatment
areas are switched. However, stopping predator control in our
current treatment area would be contrary to the aims of this
conservation intervention. Another alternative may be to apply
a ‘treatment extension’ in which predator removal is applied
to both areas. If similar results and outcomes were observed in
the former non-treatment area, this would increase condence
that the observed changes were due to predator removal.
Acknowledgements
Sincere thanks to R. Pech and M. Scroggie for advice on
data analysis. We are grateful also to D. Schaw (Toronui
Station), C. Drysdale (Landcorp, Opouahi Station), G. & S.
Maxwell (Rangiora Station), and S. McNeil (Rimu Station)
who allowed us access to their properties for pest control and
monitoring. We also thank the Department of Conservation
– in particular M. Melville, P. Abbott and D. Carlton – for
providing accommodation in the eld. D. Anderson, D. Smith
and two anonymous referees provided helpful comments on
an earlier draft.
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Editorial board member: Des Smith
Received 3 October 2016; accepted 13 September 2018
