ORIGINAL PAPER

Reinvasion by ship rats (Rattus rattus) of forest fragmentsafter eradication

Carolyn M. King • John G. Innes • Dianne Gleeson •

Neil Fitzgerald • Tom Winstanley • Barry O’Brien •

Lucy Bridgman • Neil Cox

Received: 16 December 2010 / Accepted: 25 June 2011 / Published online: 9 August 2011

� Springer Science+Business Media B.V. 2011

Abstract Reinvasions provide prime examples of

source-sink population dynamics, and are a major

reason for failure of eradications of invasive rats from

protected areas. Yet little is known about the origins

and population structure of the replacement population

compared with the original one. We eradicated eight

populations of ship rats from separate podo-

carp-broadleaved forest fragments surrounded by

open grassland (averaging 5.3 ha, scattered across

20,000 ha) in rural landscapes of Waikato, New

Zealand, and monitored the- re-establishment of new

populations. Rats were kill-trapped to extinction

during January to April 2008, and then again after

reinvasion in April–May (total n = 517). Rats carrying

Rhodamine B dye (n = 94), available only in baits

placed 1–2 months in advance in adjacent source areas

located 170–380 m (average 228 m edge to edge)

away, appeared in 7 of the 8 fragments from the first

day of the first eradication. The distribution of age

groups, genders and proportions of reproductively

mature adults (more immature juvenile males and

fewer fully mature old females) was different among

marked rats compared with all other rats (P = 0.001,

n = 509); in all rats caught on days 7? of the first

eradication compared with on days 1–6 (P = 0.000);

and in the total sample collected in fragments by

trapping to and after local extinction compared with in

brief, fixed-schedule sampling of populations in

continuous forests (P = 0.000). Genotyping of 493

carcases found no significant population-level differ-

entiation among the 8 fragments, confirming that the

C. M. King (&) � B. O’Brien � L. Bridgman

Department of Biological Sciences,

University of Waikato, Private Bag 3105,

Hamilton 3240, New Zealand

e-mail: c.king@waikato.ac.nz

B. O’Brien

e-mail: b.obrien@waikato.ac.nz

L. Bridgman

e-mail: bridgmanl@landcareresearch.co.nz

J. G. Innes � N. Fitzgerald � L. Bridgman

Landcare Research, Private Bag 3127,

Hamilton 3240, New Zealand

e-mail: innesj@landcareresearch.co.nz

N. Fitzgerald

e-mail: fitzgeraldn@landcareresearch.co.nz

D. Gleeson � T. Winstanley

Ecological Genetics Laboratory, Landcare Research,

Private Bag 92170, Auckland 1142, New Zealand

e-mail: gleesond@landcareresearch.co.nz

T. Winstanley

e-mail: tom.winstanley@gmail.com

N. Cox

Invermay Statistics Group, AgResearch,

Invermay Research Centre, Puddle Alley,

Private Bag 50034, Mosgiel, New Zealand

e-mail: neil.cox@agresearch.co.nz

123

Biol Invasions (2011) 13:2391–2408

DOI 10.1007/s10530-011-0051-6

rats in all fragments belonged to a single dynamic

metapopulation. Marked rats of both genders travelled

up to 600 m in a few days. Conservation of forest

fragments is compromised by the problem that ship rats

cannot be prevented from rapidly reinvading any

cleared area after eradication.

Keywords Invasive predators � Reinvasion �Roof rat � Black rat � Forest fragment �Genetic differentiation � Eradication units

Introduction

In a fragmented landscape, the survival of functional

populations of native species in the remaining

patches of preferred habitat, and the rate at which

invasive species colonise those patches, depends on

the species-specific dispersal abilities and inter-patch

movements of individuals (Bowne and Bowers 2004).

Relatively little is known about how dispersal affects

the important early stages of either colonisation or

reinvasion (Puth and Post 2005), when population

structure and behaviour might be different from those

of established populations (Russell et al. 2009a).

Therefore, understanding how the spatial configura-

tion of habitats influences dispersal distances and

population dynamics is essential for wildlife man-

agement, both to protect native species affected by

forest fragmentation and to control introduced species

(With 2002; Hastings et al. 2005).

Eradication programmes on islands are now

clearing more and larger areas of invasive species,

with huge conservation benefits (Brooke et al. 2007;

Towns 2009). Forest fragments share certain charac-

ters of islands (Watling and Donnelly 2006), and

some aspects of conservation management developed

on islands could potentially be applied to managing

mainland fragments. The main limitation for manag-

ers of forest fragments is that the primary condition

required to match the successes of island pest

eradication programmes (isolation from reinvasion)

is hard to meet on the mainland.

Lowland rural landscapes in most settled countries

include many scattered remnants of original mainland

ecosystems, most in private ownership, supporting a

disproportionately large repository of threatened

native ecosystems and species (Walker et al. 2006).

In New Zealand, as elsewhere, it is especially

important to protect such remnants and their biota

from invasive predators, because forest fragments on

private land represent the principal opportunity to

protect biodiversity in lowland environments

(Department of Conservation and Ministry for the

Environment 2000; Green and Clarkson 2006).

Ship rats (Rattus rattus, also called black or roof

rats) are widespread and abundant pests world-wide.

Their depredations on the fauna of native forests on

the New Zealand mainland and offshore islands were

recently reviewed by Innes et al. (2010a). As arboreal,

nocturnal omnivores, ship rats are key predators of

small forest passerines (Innes et al. 1999; Armstrong

et al. 2006a, b; Brown et al. 1998), lizards, inverte-

brates (Gibbs 2009; Towns et al. 2006) and seeds.

They can search the canopy for vulnerable native

fauna with great thoroughness and agility, especially

in forests that offer a complex, three-dimensional

habitat (Foster et al. 2011).

Within continuous mainland forests of the North

Island, ship rats are very common, and vitally neces-

sary control operations are undone within months by

rapid repopulation of cleared areas (Innes et al. 1995).

Small and isolated forest fragments are much more

numerous than large ones, and yet the times required to

achieve a specified conservation outcome in them, and

conditions favouring or preventing it, have seldom

been analysed. The commonest and simplest form of

protection is stock fencing, but, unexpectedly, we have

found that invasive rats may be significantly more

abundant in fenced compared with unfenced fragments

(Innes et al. 2010b). The demographic characters of

invaders are almost unknown, largely because they

cannot be determined from the results of control

programmes (the majority) using toxic baits, and

reinvasion is rarely monitored by carcase analysis.

In short stature habitats of various kinds, such as

alpine tussock, very young pine forests and grassland,

ship rats are much less common than in forest (Innes

2005). One implication of this is that they might be

expected to avoid the open, short-grassed pasture that

mostly surrounds these fragments, where they would

be more vulnerable to attack from predators. Hence,

one might expect that ship rats could be slow to

disperse to forest fragments surrounded by grazed

pasture, much as populations of the Australian native

rat R. fuscipes eradicated from forest patches could be

2392 C. M. King et al.

123

slow to recolonise across surrounding cleared regions

(Peakall and Lindenmayer 2006). Confirmation of

any factor inhibiting reinvasion would greatly

encourage the implementation of pest control oper-

ations in privately-owned forest fragments, with

potentially valuable consequences for biodiversity

conservation in rural areas.

Reinvasion is a serious threat to all pest control

programs, because it is the most likely cause of

failure to achieve protection of core values (Pascal

et al. 2004; Abdelkrim et al. 2007; Russell et al.

2010). The sink effect, induced by eradication of

local populations connected by gene flow to a wider

meta-population (Russell et al. 2009b), makes rein-

vasion inevitable. Eradication operations are bound to

fail unless confined to ‘‘eradication units’’ isolated

from reinvasion, which may be much larger than their

obvious geographical components. For example, off

the coast of France, some eradication units for

Norway rats comprise whole archipelagos, not indi-

vidual islands (Abdelkrim et al. 2005).

The limited data available for ship rats suggests

that effective eradication units are likely to comprise

archipelagos of forest fragments. For example, Rus-

sell et al. (2009b) found no genetic distinction

between samples of ship rats from two mainland

fragments separated by 400 m of grazed pasture.

Likewise, Abdelkrim et al. (2010) found no signif-

icant genetic isolation between ship rats living in

Puketi Forest, Northland, and those in two adjacent

fragments each 3.5 km away.

Therefore, documenting the extent of potential

gene flow between local pest populations is vital

information for pest managers, for two reasons. (1)

Local populations that are not complete eradication

units must be eradicated simultaneously (Abdelkrim

et al. 2010) and (2) Genetic mixing across broad

geographical areas reduces the chances that post-

eradication samples could distinguish between invad-

ers and survivors at the population level (Russell

et al. 2009b, 2010).

In this study we aimed to integrate demographic

and genetic data, gathered during the course of a

closely monitored eradication programme, that could

document the processes of recovery of eight ship rat

populations eradicated from separate forest fragments

scattered over a large area (20,000 km2), and to

produce results of general significance for the man-

agement of rodent pests on mainland New Zealand.

Objectives

This paper describes part of a wider study of the

interactive effects of routine management (stock

fencing and pest control) on key ecological processes

on pastoral land around rural Waikato, New Zealand.

Elsewhere we have described the vegetation (Burns

et al. 2010) and abundance of invertebrates (Didham

et al. 2009) in 53 forest fragments managed under

different combinations of strategies (Dodd et al.

2011), and the abundance of ship rats in eight (Innes

et al. 2010b).

Our primary hypotheses were as follows.

(1) After an eradication, immigrant rats might be

expected to differ from the previous residents in

their age structure and breeding potential. To

test this hypothesis, we identified rats known to

have visited or arrived from outside as those

marked with vital dye obtainable only in

potential source areas, and searched for detect-

able differences in age, gender and breeding

status between marked and unmarked rats.

(2) The processes of eradication and replacement of

rat populations in all eight fragments could be

influenced by the nature of the fragments

(isolated or connected to source areas, fenced

or grazed by livestock); by precise timing

(fragments 1–4 and 5–8 differed in trapping

dates); and by longer-term seasonal effects. To

test this hypothesis we applied a generalised

linear mixed model with binomial distribution

and logit link to the complete dataset, searching

for possible inter-relationships between the

proportion of marked rats collected at each site

and for correlations suggesting how the above

factors might influence it.

(3) The replacement populations appearing in the

eradicated fragments could represent a slightly

different gene pool from the original population.

To test this hypothesis, we genotyped all rats

collected throughout the study, and checked for

genetic differentiation between fragments and

between original and replacement populations.

Study areas

The eight fragments chosen are scattered along a range

of hills 20–30 km SE of Hamilton City in the central

Eradication and reinvasion by Rattus rattus in forest fragments 2393

123

Waikato Region, North Island (Fig. 1), as described in

Table 1 and in further detail by Innes et al. (2010b).

All the fragments were small (\10 ha) cutover rem-

nants of previously continuous conifer/broadleaved

evergreen native forest invaded by ship rats after about

1860 (Atkinson 1973). Forest clearance began centu-

ries ago after Maori settlement, but these fragments

probably achieved their current form by about 1900

(Clarkson et al. 2002). All are now surrounded by

grazed pasture. The northern-most five fragments

surrounding Te Miro, a 400 ha forested reserve, are all

within 5 km of each other and 12–16 km from the

southern-most three in the Whitehall area, which are

all within 2 km of each other.

The fragments averaged 5.3 ha in size (range

2.4–9.9 ha); four were ‘‘isolated’’ (separated from

potential source areas in adjacent forests by 55, 100,

105, and 250 m, measured edge to edge) and four

connected by woody vegetation (e.g. strips of forest or

hedgerows) to source areas. Fenced fragments pro-

tected from grazing by livestock supported much

higher densities of rats before (though not after)

eradication, but connectedness to an adjacent source

of immigrants did not influence either original density

or the chances of reinvasion (Innes et al. 2010b).

The forest canopies of all the fragments were

dominated by tawa Beilschmiedia tawa, mangeao

Litsea calicaris, mahoe Melicytus ramiflorus, pukatea

Laurelia novae-zelandiae, and rewarewa Knightia

excelsa and treeferns. Data on regeneration of key

canopy and understorey vegetation, on litter mass and

decomposition rates, and on the composition of the

detritivore fauna, in relation to management (fencing,

pest control etc.) regimes, are reported elsewhere. None

of the fragments had been subject to effective pest

control programmes (trapping, poisoning) targeting

brush-tail possums and ship rats in the 2 years before

our field sampling began. For comparison between the

populations of ship rats in the fragments trapped to

extinction, and a population sampled only lightly at

1–3 month intervals, we used data from Pureora Forest

Park, a large forest in the central North Island support-

ing comparable forest vegetation (Innes et al. 2001).

Methods

We used a combination of data collection methods,

each with different advantages and all linked to

precisely defined locations by GIS, in order to build

an integrated picture of the physical, demographic

and genetic characteristics of the original and

replacement populations. Footprint tunnels detected

the presence of rats in a fragment, and the arrival of

replacements, without disturbing them; bait markers

detected the previous presence of some individuals in

a source area; snap trapping implemented the erad-

ication programme and provided samples for nec-

ropsy and genetic analysis; hair tubes supplemented

the other detection methods and also provided

information on movements of those individuals that

entered both hair tubes and traps; and genotyping

identified the extent of the wider population sampled

in the eight fragments. This combination of ecolog-

ical and molecular techniques applied over such a

large geographic area is unusual (Peakall and

Lindenmayer 2006) but very informative.

Eradications

Tracking tunnels containing Black TrakkaTM tracking

cards were set on a 50 9 50 ± 8 m grid throughout

each fragment, operated for one night immediately

before the first eradication and at monthly intervals

thereafter. The resulting tracking indices, calculated

according to standard methods, were reported previ-

ously (Innes et al. 2010b).

Victor snap traps (passed by NAWAC standards as

humane kill traps for rats) were set in tunnels at and

in the spaces between tracking stations (i.e., on a

50 9 25 m grid) in numbers and on dates shown in

Table 1. The traps were set inside tunnels just large

enough to accommodate the trap, to discourage non-

target species, guide target rats across the treadle, and

to provide public safety. They were baited with

peanut butter, checked and re-baited daily until no

rats were caught. Then, the tracking stations were

reset, and both trapping and tracking continued

together.

Our aim was first to eradicate the entire population

of rats in each fragment, and then to monitor how

long the fragment would remain clear of replace-

ments. For the first eradication, we set a rigorous end

point for any given fragment, defined as the day on

which no rats had been trapped or tracked for the

three previous consecutive nights. At monthly inter-

vals thereafter, we set tracking tunnels, checked

daily, to test for reinvasion. If rats had reappeared,

2394 C. M. King et al.

123

we immediately began snap-trapping again. For the

second eradication, we defined our end point in all

fragments as 10 days since the start of trapping.

We divided the eight fragments into two groups,

numbers 1–4 and 5–8, arranged to enable the paired

comparisons of fencing and isolation that we

expected to have primary influence on the results.

Because it was logistically impossible to treat all the

fragments simultaneously, both eradications in frag-

ments 1–4 were done earlier than in fragments 5–8

(Table 1). Although the difference between the two

groups of fragments was analysed under the heading

‘‘Location’’, the distinction between them was pri-

marily a matter of timing. We did not expect any

geographical effect, so did not attempt to control for

it: the northern group included Fragments 1, 2, 4, 7

Lake Karapiro

3

8

6

2

5

4 1

7

Lake Karapiro

170°E

40°S

Whitehall

Te Miro

Study fragmentOther forest

mk50

Fig. 1 Map of the study

area in the Waikato

province east of Hamilton,

showing the relative

positions of the eight forest

fragments. Lake Karapiro

occupies part of the course

of the Waikato River

southeast of Cambridge

Eradication and reinvasion by Rattus rattus in forest fragments 2395

123

and 8, and the southern group numbers 3, 5 and 6

(Fig. 1).

Rat population data

All rats caught were returned to the lab for dissection.

We removed all the whiskers and half an ear from

one side, and sent them for further analysis. We

recorded gender, colour morph, whole body weight,

head and body length, tail length and reproductive

condition.

Sexual maturity

Sexual maturity was classified as a simple binary

category, yes or no, regardless of whether the

individuals so classified were in active reproductive

condition at the time of capture. Extensive breeding

data for ship rats analysed by Innes et al. (2001) from

various North Island forest types at Pureora Forest

Park have previously shown that no indicator of

reproductive activity varied significantly with habitat,

including % females pregnant; mean number of live

embryos; % with uterine scars; mean number of

uterine scars; % lactating; % pregnant females with

resorbing embryos; % males with scrotal testes (Innes

et al. 2001: tables 6 and 7). For females, these

characters also did not vary between summer and

autumn, which were the only two seasons sampled

here. The proportion of fertile males did not vary with

habitat, but declined slightly in autumn.

For the present analysis, we therefore defined

sexually mature females as those with active uteri

(‘‘string’’ or ‘‘cord’’ condition, as opposed to the

‘‘thread’’ condition of juveniles) or pregnant or with

uterine scars. Sexually mature and fertile males were

Table 1 Study areas, all located 20–30 km SE of Hamilton, central Waikato, where rats were trapped between December 2007 and

June 2008

Fragment 1 2 3 4 5 6 7 8

Area (ha) 3.5 4.9 9.9 4.3 4.8 5.7 2.4 7.0

Fenced or Grazed G F G F G F G F

Connected to source, or isolated C C I I C I I C

Distance from source (m) 115 210 100 250 40 55 105 10

Number of traps 25 44 75 42 48 48 23 54

RhB baits placed or renewed 11 Dec,

10 Jan,

21 Feb,

28 Mar,

10 Ap

17 Dec,

10 Jan,

27 Feb,

27 Mar,

10 Ap

10 Dec,

10 Jan,

27 Feb,

28 Mar,

28 Ap

11 Dec,

10 Jan,

26 Feb,

31 Mar,

28 Ap

10 Dec,

17 Jan,

26 Feb,

26 Mar,

23 Ap

17 Dec,

17 Jan,

27 Feb,

27 Mar,

28 Ap

12 Dec,

17 Jan,

26 Feb,

28 Mar,

23 Ap

13 Dec,

17 Jan,

26 Feb,

27 Mar,

24 Ap

Distance bait station to

nearest fragment edge (m)

215 250 170 380 245 200 195 170

Traps set for 1st eradication 10 Jan 10 Jan 10 Jan 10 Jan 14 Feb 14 Feb 14 Feb 14 Feb

Traps set for 2nd eradication 11 Ap 11 Ap 29 Ap 29 Ap 21 May 8 May 21 May 21 May

Number of hair tubesa 10/6 21/10 33/10 19/9 22/6 20/10 10/10 24/10

Dates hair tubes set 9 Jan

21 Feb

17 Mar

10 Ap

9 Jan

11 Mar

10 Ap

9 Jan

11 Mar

28 Ap

9 Jan

17 Mar

28 Ap

13 Feb

20 Mar

23 Ap

20 May

13 Feb

7 May

13 Feb

23 Ap

20 May

13 Feb

20 May

Genotypes/matchesb 0 12/4 0 0 0 0 5/1 12/3

During the first eradication, trapping ended when no rats had been trapped or tracked for 3 consecutive nights, and during the second,

after 10 days. For density indices derived from footprint tracking, taken immediately before each eradication started, see the

matching Table 1 in Innes et al. (2010b)a Number of hair tubes set in eradication and reinvasion source areas respectivelyb Number of hair samples genotyped, and the number of genotypes matching other hair or ear samples

2396 C. M. King et al.

123

defined as those with scrotal, active testes containing

visible tubules in the epididymes. The lack of any

correlation between these classifications and habitat

in the Pureora sample (n = 547, summer and autumn

captures only) showed that the same classifications

could be applied with confidence to the entire sample

of rats collected from our forest fragments that were

in good enough condition to be classified by age,

gender and reproductive condition (n = 509).

Age groups

Innes et al. (2001) also showed that body weight

controlled for season, year and gender increases

predictably with age estimated from tooth wear,

regardless of habitat. For this study we divided the

samples into three age-related groups by weight. The

lower boundary of the middle group, the young

adults, was defined as the upper limit of body weights

of all individuals still not sexually mature, which was

found by inspection of the ranked data from the

current (forest fragments) samples (109 g in males

and 79 g in females). The upper boundary of this

group was based on the average whole body weight

of rats from Pureora (males 141 g, females 121 g).

We confirmed that these groups represented real

categories distinctly different from each other in age

by applying the same weight-class definitions retro-

spectively to the Pureora data set (summer-autumn

samples only) (Table 2).

Identifying the sources of re-established

populations

Rhodamine B marking

In forest adjacent to each fragment, located

170–380 m (average 228 m edge to edge) from each

fragment, we set out up to ten possum-proof bait

stations at 40 m spacing a month before trapping was

due to begin. Each initially contained 200 g of bait

pellets (RS5 cereal prefeed pellets, made by Animal

Control Products, Wanganui) surface-coated with

Rhodamine B (RhB) biomarker at 0.1% concentra-

tion. RhB dye can be detected in external keratinous

tissue such as hair and claws (Fisher 1999). Baits

were replaced with 100 g of fresh stock at monthly

intervals, on the dates shown in Table 1.

During necropsy, all rats were checked visually for

RhB staining on the fur of the paws, face and belly.

We then selected 6–7 medium to long whiskers from

each carcase, preferably those with an intact follicle

and a fine tip, for examination under flourescence

microscopy. Whiskers were washed and dry-

mounted, at least 6 to a glass microscope slide, under

a 60 mm #2 coverslip, and examined at 1009 under a

compound microscope fitted with a 50 W mercury

vapour arc lamp and a Leitz Ploemopak epifluores-

cence unit. Green excitation was produced by an M2

filter set (excitation BP 530–560 nm, dichroic mirror

580 nm, barrier filter LP 580 nm) which passes all

wavelengths in the yellow, orange and red range of

the spectrum.

Autofluorescence, a natural property of some

biological materials, can be distinguished from a true

RhB signal by adding an I2 block. The resultant blue

excitation gives a stronger signal from the autofluo-

rescent sources and a substantially reduced rhoda-

mine excitation. We scored whiskers showing this

relative change of intensity as negative for RhB.

We examined how the proportions of RhB-marked

rats collected at each site and in each trapping session

might vary with isolation (connected to a source or not),

habitat disturbance (grazed or not), trapping history

(first eradication or second), timing of the first eradi-

cation (earlier in fragment locations 1–4 than in 5–8)

and day of operation (days 1–6 or day 7?). Connect-

edness was included because, although we have already

shown it did not affect the chances that some rats would

reappear after the first eradication (Innes et al. 2010b), it

Table 2 Definitions of weight groups used as a proxy for age

classifications of 509 Fragments rats, compared with the actual

toothwear classes of 547 Pureora rats allocated to the same

weight groups

Age

group

Fragments Pureora

Body weight

range (g)

n Mean

toothwear class

SD

Males

1 \109 52 2.6 0.64

2 110–140 76 3.4 0.87

3 [141 186 4.3 0.70

Females

1 \79 18 2.5 0.92

2 80–120 80 3.4 0.84

3 [121 135 4.3 0.78

Totals 547 3.9 1.00

Eradication and reinvasion by Rattus rattus in forest fragments 2397

123

might affect the age, gender, reproductive or genetic

distributions of which rats arrived.

We checked for possible inter-relationships

between all these factors by applying a generalised

linear mixed model with binomial distribution and

logit link to the complete dataset. The model included

random terms (hence extra-binomial variation) for the

individual fragments and repeats within these. It

adjusted the estimated percentage frequencies for each

variable by controlling for imbalances in the others.

To check for a possible seasonal variation in the

supply of immigrants, we divided the total sample

into four periods of time (roughly a month each)

defined from the dates on which the Rhodamine B

baits were renewed in each area (dates given in

Table 1). Period 2 started on the date of the first

renewal of marked baits in mid January and ran until

the second renewal at the end of February, and so on

to period 5 running from the last bait renewal in late

April until the end of the study in late May.

We assumed that any rats we caught within a

fragment that were clearly marked with RhB (but not

only these) could be classified as definite replace-

ments. We do not call them ‘‘reinvaders’’ because

some could have been wandering residents.

Hair tubes

Before and during the eradications, the rat popula-

tions in the fragments, and in adjacent reinvasion

source areas, were monitored with hair tubes (Glee-

son et al. 2010) baited with peanut butter and set at

trap sites (on a 50 9 50 m grid) 2–4 times through

the study as shown in Table 1. GPS positions of all

hair tubes were recorded to ±8 m. Hair samples were

collected after 1 day, and stored dry pending DNA

extraction. We searched for exact matches between

genotypes derived from hair samples collected from

more than one hair tube, and between hair samples

and ear tissue from carcases, assuming that an exact

match represents the same animal sampled on

separate occasions. Distances between matching

sample sites indicate minimum movements by indi-

viduals between or within fragments.

DNA extraction and genotyping

We extracted genomic DNA from ear samples stored

in 95% ethanol, and from dry hair samples including

at least five clearly visible follicles, using the Corbett

X-tractor Gene automated DNA solid tissue extraction

system following the manufacturer’s recommendation

with external digest (Corbett Robotics, Brisbane,

Australia). We used a total of nine microsatellite

markers, previously developed from R. norvegicus

genome mapping (Jacob et al. 1995), to genotype each

sample: D2Rat234, D5Rat83, D7Rat13, D10Rat20,

D11Mgh5, D15Rat77, D16Rat81, D18Rat96,

D19Mit2. We performed PCR amplification in 10ul

volumes, following Abdelkrim et al. (2010), and ran

the PCR products as two multiplex runs on an ABI

prism 3130 capillary electrophoresis system (Applied

Biosystems). We derived genotypes using GeneMap-

per software (Applied Biosystems). From the tissue

genotypes, we calculated an overall Psib score using

GenAlEx v 6.2 (Peakall and Smouse 2006) in order to

determine the probability that a sibling would possess

an identical genotype. From the hair samples, we

made three replicates per sample, and did not accept a

genotype as complete unless there was consensus

from at least two of the three replicates.

We assessed evidence for allelic drop-out, scoring

error due to stutter, and the presence and frequency of

any null alleles with MICRO-CHECKER (Oosterhout

et al. 2004) using a Bonferroni adjusted 95% confi-

dence interval and 10,000 repetitions. We calculated

genetic diversity indices and pairwise genetic differ-

ences, and conducted Principal Components Analysis

(PCA) using GENALEX v 6.2 (Peakall and Smouse

2006), and conducted tests for Hardy–Weinberg

Equilibrium (HWE) and linkage disequilibrium using

GENEPOP v 4.0 (Raymond and Rousset 1995).

The fixation index (FST) is calculated from the

correlation of randomly chosen alleles within one

sub-population relative to that found in the entire

population. Population differentiation can then be

estimated from a series of pairwise comparisons.

Values for FST between 0.0–0.05 are consistent with

little genetic differentiation (near panmixis);

0.05–0.15 indicates moderate, and 0.15–0.25 high

differentiation between populations. We calculated

pair-wise FST estimates using FSTAT (version

2.9.3.2) (Goudet 1995), and determined their signif-

icance by Fisher’s method (Manly 1985: 432–433;

Raymond and Rousset 1995).

We used the Bayesian clustering method imple-

mented in STRUCTURE 2.3.2 (Hubisz et al. 2009) to

estimate the best-supported number, K, of genetic

2398 C. M. King et al.

123

clusters. We chose the admixture model with corre-

lated allele frequencies, and conducted 10 replicates

for each value of K varying from 1–8 for each run,

consisting of a burn-in period of 100,000 MCMC

(Markov Chain Monte Carlo) steps followed by 106

iterations.

Results

Altogether, 517 ship rats were collected from the 8

fragments over two eradication attempts (mean 46,

range 0–110 each; Table 3), of which 509 were

classified by age and reproductive condition, and 465

by RhB status. Sample and sub-sample sizes are

given in context as appropriate in Tables 2, 3, 4, 5, 6,

7, 8, 9, 10. We also caught 11 Norway rats (in two

fragments), one weasel and two mice.

Summer eradication

In January-April 2008, the mean time needed to meet

our definition of eradication in each of the eight

fragments was 33 nights (range 3–67). Summer rat

density estimated from the first six nights of trapping

was significantly higher in fenced (6.5 rats ha–1, SE

1.4) than in stock-grazed fragments (mean 0.5 rats

ha–1, SE 0.4; P = 0.02, and the time to complete the

first eradication averaged 47 nights in fenced and 19

nights in grazed fragments (Table 3). Ship rats

replaced the eradicated populations in all fragments

within a month, regardless of fencing or isolation

(Innes et al. 2010b).

Replacement (autumn) rat populations

The mean time between initial eradication and the

second tracking index (Table 1) was 63 nights, range

30–92 nights. The relative abundances of the re-

established rat populations were often different from

those measured in the same fragments 4 months

previously (Table 3). Rats were no longer signifi-

cantly more abundant in the fenced fragments, at least

over the 10 days we observed them (Innes et al.

2010b). The autumn populations were not pursued to

local extinction.

Distribution of rats marked with RhB

Rhodamine B dye had been available since December

only in adjacent areas outside the fragments

(Table 1). From January onwards, only one of the

eight fragments did not produce at least one marked

rat from among the previously untrapped populations.

Of the 517 rats examined, the RhB status was

ambiguous or unknown in 52, and definite in 465, of

which 94 were marked (Table 3) The overall mean

minimum distance they had moved, from the centre

Table 3 Proportions of marked rats caught in eight forest fragments (total 94 of 465 rats with definite RhB status)

Fragment 1 2 3 4 5 6 7 8

First eradication (summer)

Days to complete 6 27 29 41 3 54 39 67

Rats/ha (days 1–6) 0.3 6.3 1.3 4.2 0 11.1 1.3 4.5

Rats with clear RhB status 1 45 28 52 0 98 27 72

Rats marked n (total 79) 0 4 3 8 0 26 17 21

Rats marked % of 323 0 8.9 10.7 15.4 – 26.0 63.0 29.2

Total rats trapped 1 49 33 60 0 110 30 83

Second eradication (autumn)

Rats/ha (days 1–6) 4.6 3.1 2.3 4.0 2.3 3.7 0.8 0.9

Rats with clear RhB status 23 13 34 26 13 21 2 10

Rats marked n (total 15) 0 3 2 2 3 1 0 4

Rats marked % of 142 0 23.1 5.9 7.7 23.1 4.3 0 40.0

Total rats trapped 23 17 35 26 15 23 2 10

Pre-eradication tracking indices are given by Innes et al. (2010b). Density estimates are given only for the first 6 days of trapping

(Watkins et al. 2010), after which immigration is expected to increase the total catch to many more rats than could live in each

fragment at one time. The second eradication attempt was stopped after 10 days

Eradication and reinvasion by Rattus rattus in forest fragments 2399

123

of a minimum convex polygon enclosing the loca-

tions of the set of bait stations nearest the fragment to

the point of capture within it, was 387 m (range

190–640 m, n = 88); means for marked rats col-

lected from individual fragments varied from 282 to

547 m.

The distribution of marked rats through the

samples was uneven, but they continued to appear

in all fragments throughout the eradication process.

Of the 94 marked rats collected, 79 were caught

during the first eradication (Table 3), and 10 of these

were caught on the first day of trapping (12 January),

whereas only 15 marked rats were caught during the

second eradication. After accounting for other factors

influencing the samples, our GLM model estimated

these proportions as 19.9 and 13.9% respectively

(P = 0.413, Table 4).

The proportion of marked rats was significantly

higher (P = 0.004) in fragment numbers 5–8

(Table 4), where the first eradication was done a

month later than in fragments 1–4 (Table 1). The

proportion of marked rats was also higher after the

first 6 days of trapping (P = 0.003), as would be

expected if they represent the immigrant group, but it

was not higher in connected versus isolated frag-

ments, or in fenced versus grazed fragments, or

during the first versus the second eradications. Visual

inspection of the plotted capture positions of the

marked rats collected within each fragment showed

no perceptible clustering along the edge habitats

closest to the sources of RhB-marked bait.

We compared the proportion of marked rats

caught during four successive monthly periods,

defined in Table 4, to check for seasonal effects

whilst controlling for isolation, location and other

factors as before. This comparison confirmed a

significant variation in proportion of marked rats

through the trial, but almost all of that effect was due

to the large influx of marked rats (40%) observed

from mid January to the end of February. The total

number of rats caught increased slightly in May, but

only 12% of these were marked, implying no

increase in the supply of immigrants after the end

of the breeding season in autumn.

Table 4 Distribution of 465 rats with determined Rhodamine B status

Factor SD Significance

Isolation Connected Isolated

% marked 13.7% 18.1% 6.9% P = 0.522

n 177 288

Eradication 1st 2nd

% marked 17.9% 13.9% 4.9% P = 0.413

n 323 142

Location 1–4 5–8

% marked 8.4% 27.8% 6.8% P = 0.004

n 222 243

Disturbance Fenced Grazed

% marked 15.1% 16.5% 7.1% P = 0.842

n 337 128

Day of trapping 1–6 7?

% marked 10.9% 22.3% 3.8% P = 0.003

n 261 204

Period since renewal of RhB Jan–Feb Feb–Mar Mar–Ap Ap–May

% marked 16.3% 39.9% 5.8% 12.0% 9.2% P = 0.043

n 228 70 25 142

For each factor listed, the model calculates the percentage of rats marked in each of two conditions plus the standard error and

significance of the differences attributable to that factor, after allowing for all the other factors. The two groups comprising Fragments

1–4 and 5–8 (in sub-heading ‘‘Locations’’) are distinguished by the slightly different timings of the operations in each (Table 1), not

their geographic locations (Fig. 1). The exact significance values are given for all comparisons, in bold type if P \ 0.005

2400 C. M. King et al.

123

Population structure (age and gender)

In the 509 rats collected in good enough condition to

be classified, the six groups defined from gender and

weight were distinctly different in age by, on average,

one toothwear class (Table 2). This difference in age

is consistent with the significant increase in the

proportions of rats in the top compared with the

middle weight classes of each gender that were

sexually mature (Table 5). There was no overall

difference in the proportions of males and females

that were mature (P = 0.144).

The age and gender ratios among the 94 marked

rats differed significantly (P = 0.001) from those of

unmarked rats (Table 6). The largest single group of

marked rats comprised juvenile males (28%), and the

smallest, old breeding females (8%).

Impact of removal trapping on age

and gender ratios

The reference samples from Pureora were collected

according to a fixed, intermittent schedule in contin-

uous forest, in which traps were never set for more

than 12 days at a time (Innes et al. 2001), whereas the

first trapping programme continued until a robust

predetermined criterion of eradication had been met

in each fragment, which took up to 67 consecutive

Table 5 Distribution of

weight groups and sexual

maturity, total fragments

sample

Body weight

range (g)

n % sexually

mature

Difference between

weight groups in %

mature by gender

Males

1 Juveniles \109 99 0

2 Young 110–140 47 22.7 P \ 0.001

3 Old [141 96 77.1

Females

1 Juveniles \79 59 0

2 Young 80–120 84 32.5 P \ 0.001

3 Old [121 124 84.7

Totals 509 43.1

Table 6 Distribution of age/gender classes among rats with and without RhB marker dye, which they could have acquired only

outside the study fragments

Total classified by

age/gender

Marked rats (% marked

of total age group)

Unmarked, Eradication

1 only

Unmarked, Eradication

2 only

Males

1 Juveniles 99 28 (28.3) 49 18

2 Young 47 4 (8.5) 16 23

3 Old 96 20 (20.8) 44 20

Females

1 Juveniles 59 12 (20.3) 32 9

2 Young 84 20 (23.8) 23 36

3. Old 124 10 (8.1) 85 21

Totals 509 94 249* 129*

The age structure of marked rats was significantly different from that of all unmarked rats (P = 0.001), and also from that of the

replacement population caught during Eradication 2 (P = 0.0034)

* Excluding rats of uncertain RhB status

Eradication and reinvasion by Rattus rattus in forest fragments 2401

123

days (Table 3). These different sampling regimes

were associated with a highly significant difference

(P = 0.000) in the age structure of the rats collected

(Table 7). Many more juvenile males were collected

by trapping to extinction in the fragments than by the

repeated, brief, low-level sampling at Pureora. The

disproportionate representation of juvenile males in

the prolonged Fragments samples was associated

with a significant decline by 20.90 g, se = 3.52 g

(P \ 0.001) in the average male body weight of rats

collected from the Fragments compared with the

Pureora males (Table 7). The same difference did not

reappear in the female body weights.

Not all immigrant rats were marked, which raises

the question of whether the marked rats were typical

of the immigrant group as a whole. Our data suggest

that they were. During the first eradication in the

Fragments, the shift towards juvenile males in

distribution of age/gender classes was already evident

among the 206 rats caught after day 7 compared with

the 160 rats caught on days 1–6 (P = 0.000,

Table 8). If these marked rats do indeed represent

the unmarked immigrants during that period, there

should be no difference in age distribution between

the marked rats and the 127 unmarked rats caught on

days 7?; that was so (P = 0.168, Table 9).

By the second eradication, the proportion of

juvenile males caught was not different in the later

versus early days of trapping (P = 0.191, Table 8).

The marked rats no longer represented the immi-

grants, because total sample of rats collected during

the second eradication, conducted late in the season,

included fewer juveniles and more young adults than

in the marked group (P = 0.005).

Genetic diversity and population differentiation

A total of 493 ear samples from all fragments and 29

hair samples from three fragments (Table 1) were

successfully genotyped at nine microsatellite loci

identifying nine pairs of hair-hair or hair-ear matches.

Eight loci had between zero and \10% null alleles.

One locus, D10Rat20, had [10% null alleles at all

sites, so we removed it from subsequent population

analyses, as did Abdelkrim et al. (2010) for the same

reason.

Allele frequencies and Hardy–Weinberg Equilib-

rium (HWE) summary statistics for each fragment are

given in Table 10. The mean number of alleles per

locus varied by fragment from 5.9 to 12.1. Fragments

with private alleles had, on average, higher numbers

of alleles per locus, and these private alleles were all

low frequency (\0.05). All fragments had similar

observed heterozygosities ranging from 0.74–0.78.

Only Fragment 5, the area from which fewest rats

Table 7 Distribution of age/gender groups in samples of rats

collected by sampling to and after local extinction (Fragments

programme) and by fixed-period short-term sampling (Pure-

ora), summer/autumn data only

Group Fragments Pureora

n % n %

Males

1 Juveniles 99 41 52 17

2 Young 47 19 76 24

3 Old 96 40 186 59

Mean male 244 323

Body weight (g) 121.7 142.6

sd 46.88 33.19

Females

1 Juveniles 59 22 18 8

2 Young 84 31 80 34

3 Old 124 46 135 58

Mean female 267 343

Body weight (g) 117.5 121.8

sd 41.72 27.47

Totals 509 547

P 0.000

Table 8 Distribution of age/gender groups in samples of rats

collected on the first 6 days of trapping compared with sub-

sequent days, Fragments programme

Group Eradication 1 Eradication 2

Days 1–6 7? 1–6 7?

Males

1 Juveniles 23 51 13 12

2 Young 7 14 19 7

3 Old 35 38 19 4

Females

1 Juveniles 20 30 7 2

2 Young 6 35 33 10

3 Old 65 36 18 5

Totals 160 206 111 40

P 0.000 0.191

2402 C. M. King et al.

123

were collected (Table 3) showed a significant depar-

ture from HWE.

For our entire collection (n = 493 genotyped rats),

FST values ranged between 0.024 and 0.096, indicat-

ing low to moderate genetic differentiation between

fragments. Plotting pair-wise FST values against the

corresponding linear separations of the eight frag-

ments showed no significant correlation between

genetic and geographic distance (Fig. 2). Principal

Components Analysis based on pairwise genetic

differences showed no correlation between rats caught

during the first versus the second eradication attempts

(Fig. 3). The cumulative percentage of variation

explained by the first three PCA axes was very low

(24.13, 42.78 and 59.09%) and subsequent compo-

nents added progressively less, suggesting a ball-like

data cloud. Analysis using STRUCTURE 3.2 also

showed little genetic structure between fragments.

Models that subdivided the total dataset into K [ 1

were not supported when compared with the null

hypothesis of no substructure.

Hair/hair and hair/ear matches

We collected 78 hair samples from 230 hair tube-

nights, of which 29 (Table 1) from three fragments

returned genotypes of good enough quality for

subsequent analysis. Hair samples from the other

five fragments were not analysed.

The overall Psib score was 0.004, which is below

the 0.01 reported as the upper limit required for

individual genetic assignment and population esti-

mation (Mills et al. 2000). We therefore considered it

unlikely that matching hair and tissue genotypes were

from siblings, as opposed to two samples from the

same individual.

We obtained pairs of matching hair samples

(identical at all nine loci) from three individual rats

that visited two hair tubes within a source area, 45

and 87 m apart in Fragment 8, and 32 m apart in

Fragment 7 (mean 54.5 m). The hair tubes were

Table 9 Distribution of rats classed by both RhB status and

age/gender group collected during the first eradication com-

pared with the unmarked rats collected after day 7

Group Marked Unmarked

Days All days 7?

Males

1 Juveniles 17 31

2 Young 3 12

3 Old 18 17

Females

1 Juveniles 12 20

2 Young 15 20

3 Old 9 27

Totals 74 127

P 0.168

Table 10 Summary statistics of genetic diversity for rats trapped within eight Waikato remnant forest fragments

Fragment 1 2 3 4 5 6 7 8

n 23 65 67 79 15 121 31 92

Na 6.6 8.1 10.1 9 5.9 9 8.8 12.1

Priv. – – 0.86 0.75 – 1 1 2

He 0.73 0.74 0.75 0.72 0.66 0.71 0.74 0.78

Ho 0.78 0.78 0.77 0.73 0.74* 0.74 0.76 0.77

n number of samples, Na mean number of alleles per locus, Priv. mean number of private alleles per locus, Ho mean observed

heterozygosity, He mean expected heterozygosity

* denotes a significant departure from HWE

0

0.02

0.04

0.06

0.08

0.1

0.12

0 5 10 15 20

F S

T v

alu

e

Distance between pairs, km

FST

Fig. 2 Pair-wise comparisons of all FST values plotted against

the cross-country distances between each pair of fragments

Eradication and reinvasion by Rattus rattus in forest fragments 2403

123

always set 1 day and cleared the next day, so all these

distances refer to net movements over 24 h. The

identities, ages and genders of these rats were

unknown because none of the genotypes identified

from hair tubes in source areas matched any of the rat

ears genotyped from the adjacent (or any) fragment.

Four hair samples returned identical matches with

an ear sample from four rats trapped; three were

caught 27, 28, and 55 m from the tube where the hair

was collected within Fragment 2 and one (matching

at 8 loci) at 40 m from the hair tube within Fragment

8. In every case, the hair sample was collected on the

day before the traps were set, and the rat was

collected 2 days later, on the first day of the first

eradication. All four pairs of hair-ear matches came

from old females that we sampled inside a fragment,

both before and after death, presumably within their

home ranges.

A fifth hair sample matched an ear sample from

another old female caught 99 m from the hair tube on

the first day of the second eradication in Fragment 2.

We cannot distinguish whether this rat was (1) a

survivor of the 27 days of the first eradication

attempt, or (2) a post-eradication colonist arriving

from or via the broad shelterbelt that connected this

fragment to the adjacent Te Miro forest reserve. This

rat was not detected by the hair tubes and RhB bait

stations along the Te Miro forest edge, and was not

genetically distinct from the rats caught during the

first eradication. The minimum overnight local

movements of these five old females averaged 50 m.

Only a single old male recorded a hair-ear match,

again within Fragment 2. This rat was caught on the

second day of the first eradication, in a trap 203 m

from the hair tube it had visited 3 days previously.

None of the marked rats trapped had left matching

samples in hair tubes in source areas. Hence none of

the nine hair-ear matches (n = 9) or the 94 carcases

marked with RhB identified individual animals

moving into a fragment from a source area.

Discussion

Ship rats are key pests throughout New Zealand.

Conservation managers working to protect endemic

species on the mainland have long known how

rapidly reinvasion can reverse the beneficial effects

of local ship rat eradication programmes (Innes et al.

1999), but the demographic characters of replacement

populations have remained obscure. Here we present

new data decisively demonstrating three key obser-

vations on the re-invasions we observed: (1) the

replacement rats were derived from a meta-popula-

tion so large that our samples, spread over

20,000 km2, could not define its boundaries; (2) gene

flow between forest fragments within the meta-

population was not inhibited by large areas of

intervening non-preferred habitat, so a few hundred

metres of pasture clearly could not protect our study

areas from reinvasion after eradication; and (3)

movements of individual rats towards the cleared

fragments from outside were not stimulated by

human interference–on the contrary, we detected

substantial numbers of marked rats on the very first

day of trapping, suggesting an ongoing dynamic

interaction between the rats living in the fragments

and in neighbouring areas that was already in

progress before our first eradications began to take

effect.

Demography of the replacement population

Marked rats could have picked up the RhB marker

dye if they were either (1) resident in a fragment with

home ranges extending into a source area, which had

wandered outside the fragment, picked up the bait

marker and then returned, or (2) immigrants which

had newly arrived from a different area outside the

fragment. We acknowledge that our methods could

not objectively distinguish between these two cate-

gories, so the rats collected from the original

population comprised, at least at first, an unknown

mixture of original residents and immigrants.

Component 1

E1E 2

Co

mp

on

ent

2

Fig. 3 Principal components analysis (PCA) performed with

GENALEX software using the nine microsatellite loci from rats

caught during the first eradication (E1) versus rats caught

during the second eradication (E2)

2404 C. M. King et al.

123

However, our criteria for achieving an eradication

were severe, and allowed for trap evasion (we

specified that no rats be detected in either traps or

footprint tunnels for three consecutive nights), so we

think it likely that few or no rats remained by the end

of the first eradication.

Dispersal is male-biased in most rodents, including

Rattus sp. (Calhoun 1963; Krebs et al. 2007) and is

most frequently undertaken by younger, non-breeding

individuals (Peakall and Lindenmayer 2006). For

example, Stokes et al. (2009) found higher propor-

tions of subadults, and no breeding females at all, in

replacement populations of R. rattus collected from

eradication sites. If our marked rats fairly represented

the immigrants, they would be expected to differ

significantly in age, gender ratio and reproductive

(but not genetic) characters from unmarked rats. This

was true (P = 0.001, Table 6); the marked rats

included more immature juvenile males than any

other age/gender group, and fewer old adult females

(Tables 5, 6), and the difference was established

relatively early (after day 7 of the first eradication:

Tables 7, 8, 9).

The implication is that, although rat populations

re-establish rapidly after an eradication operation, the

new group can be expected to start with a different

demography and a lower reproductive potential

compared with that of the original population. These

observations have some important implications for

management of invasive rats on the mainland,

because age distribution and sexual maturity are

two of the parameters most affecting population rate

of increase. This information gives some interesting

clues about the population dynamics of invaders,

although it does not help conservation in the short

term since rats of all ages and both genders are

equally unwelcome conservation pests.

In our study, the movements of individual rats

revealed by both genetic and RhB techniques were

consistent with previous knowledge of male and

female ship rat range lengths in New Zealand podo-

carp-broadleaved forest (summarised by Innes 2005).

Five of the six matched hair and ear samples came from

old females caught on the first day of a trapping session,

and the sixth from an old male caught on the second

day. These data suggest that old rats freely ran through

both hair tubes and trap tunnels, and so were likely to be

caught as soon as the traps were set. Larger, older and

socially dominant rats are often the first to explore new

devices containing food, because fear of conspecific

aggression prevents smaller, younger rats exploring or

feeding in a confined space inside a tunnel or bait

station (Quy 2003). If the old females had often moved

outside the fragments, they might have been more often

marked with RhB, but, on the contrary, Table 6 shows

that they were the least likely group to be marked.

Since RhB dye was available only outside the

fragments, this confirms that the old females were the

most sedentary group (with smallest home ranges and

shortest local movements) resident within the frag-

ments (Innes 2005). Older adult males are much more

mobile and hold larger home ranges than old females

(Hooker and Innes 1995), perhaps because older males

are at greater risk of mating with their own offspring.

We did not document any larger movements that

could clearly be seen as natal dispersal. In other

species of Rattus, dispersal studies have revealed

typically small local movements, with occasionally

some much longer, e.g., \100 m and up to 1,750 m

in R. fuscipes (Bentley 2008) or c.62 m and 11 km in

R. norvegicus (Gardner-Santana et al. 2009). Such

comparisons depend on assuming no difference in

trappability between marked and unmarked rats, but

we have no way to test for this.

Our data illustrate incidentally the potential bias

inherent in using the results of continuous snap-

trapping for population analyses. Density estimates

calculated from samples of rats taken as by-catch in

permanently-set trap lines for carnivores (Dilks et al.

2003), assume that the probability of capture remains

constant over the long term, but it does not (Watkins

et al. 2010). Likewise, the proportions of age/gender

classes (and hence, the reproductive potential) will be

estimated differently from small rat populations

trapped to extinction than from samples of rats taken

from a large population sampled periodically.

Genetic differentiation among fragments

Table 10 may be compared directly with Abdelkrim

et al. (2010: table 1), who document the same

statistics for 274 rats from 5 km2 Puketi forest

(Northland) and 39 from two outlying fragments each

separated by 3.5 km from Puketi in opposite direc-

tions. The values for observed and expected mean

heterozygosity were comparable across all samples.

The lack of any significant correlation between

genetic and geographic distance among our 8

Eradication and reinvasion by Rattus rattus in forest fragments 2405

123

fragments separated by up to 16 km (Fig. 2) sup-

ported our RhB results in suggesting that all the rats

we sampled belonged to a single widely distributed

meta-population. Our data support the inference of a

continuous, dynamic interaction between rats living

in fragments and in the adjacent source areas

supplying immigrants, thereby ruling out the use of

genetic techniques to distinguish survivors from

reinvaders (Russell et al. 2010). Abdelkrim et al.

(2010) reached the same conclusion for Puketi and its

two adjacent fragments. No model dividing the

Puketi sample and its outliers into K [ 1 was

supported, and likewise, we found no clear genetic

distinction between rats collected during the first and

second eradications. Our estimates of FST are a little

higher for some pair-wise comparisons, as expected

considering the greater separation distances we

sampled. Private alleles appeared in five of our

fragments, but in neither of the two Puketi outliers.

All the fragments we sampled were situated along

a line of low hills east of the flood plain of the

Waikato River, which was until about 1900 com-

pletely covered with conifer/broadleaved forest

(Clarkson et al. 2002). Fragmentation of the original

forest has also fragmented the rat population, so in

theory our results could reflect either the former

contiguity and common ancestry of the present

isolates, or later gene flow between them. However,

if there had been a complete lack of gene exchange

across the cleared areas for more than a 100 years, we

would have expected to see some evidence of genetic

distinction between our sample areas. Our data

support the conclusion that gene flow among frag-

mented populations of non-commensal ship rats can

extend across substantial distances of non-forest

habitat.

Conclusions

Our analyses were done on the largest geographic

scale reported so far (up to 16 km between sample

areas), but we still could not find the boundaries of

the metapopulation of ship rats surrounding our eight

study areas. Pair-wise comparisons of FST between all

8 fragments returned low values, indicating that all

the individuals sampled belonged to interbreeding

sub-populations. Hence, we cannot define a suitable

‘eradication unit’ that might protect native biota in

our forest fragments in the sense proposed by

Abdelkrim et al. (2010). Eradication operations in

small fragments like those we studied are therefore

bound to fail, because they cannot be confined to an

isolated landscape unit not connected to extensive

source areas. On the contrary, we showed that

replacement populations arrived very quickly, and

were genetically indistinguishable from the previ-

ously resident populations, although including a large

proportion of young rats at the beginning of their

reproductive lives. Marked individual rats travelled

from at least 0.5 km away, reaching all but one of our

fragments in a few days, including those that we had

classed as isolated from potential reinvaders. Pest-

proof fencing (Connolly et al. 2009) is the only

effective way to prevent reinvasions.

MacQueen et al. (2008) also found no genetic

differentiation between populations of Rattus fuscipes

in fragmented Queensland landscapes. They sug-

gested that R. fuscipes was able to move between the

larger forest reserves using habitat corridors. It would

be valuable to know more about movement behaviour

of R. rattus in New Zealand farmed landscapes,

especially about how they typically move between

fragments. Certainly, a fragmented landscape is no

barrier to their movements, so is no help in control-

ling their spread (With 2002). Indeed, ship rats can

cross far more daunting barriers, such as the sea water

channels separating the 13 islets off the main island

of Lavezzu, between Corsica and Sardinia (Abdelk-

rim et al. 2009).

We could not exclude less obvious connections

between our fragments such as wire fences or

waterways, and if these features routinely provide

possible reinvasion routes, they could be targeted

with traps or poison in the future. We need more

detailed documentation of the use made by ship rats

of habitat corridors, and of their patterns of natal

dispersal. We suggest that these should be important

priorities for future research, not only in New

Zealand but in any of the large number of countries

where Rattus rattus is a pervasive conservation pest.

Acknowledgments We gratefully acknowledge the

landowners and managers who generously gave us free access

to their forest fragments: Lyndon Bergerson, Shane Blair, Paul

and Jo Bodle, Mr R.A. Boyte, Bill Boyte, Des McAllister, Allen

Marr, Greg and Jackie Nicholls, Brett and Narelle Pollock,

Deborah and John Stretton. Chris Floyd helped locate study sites,

and Penny Fisher gave valuable advice about Rhodamine B

2406 C. M. King et al.

123

marking. Scott Bartlam, Sam Cave, Toni Cornes, and

particularly Stacey Foster and Jordan Edgar helped with the

substantial fieldwork required for this project, including daily

trap-clearing that lasted several months. The manuscript was

greatly improved by the perceptive comments of two anonymous

referees. The research was funded by the Foundation for

Research, Science and Technology under UOW X0609 and the

2010 Strategic Investment Fund of Waikato University.

Approval for the Rhodamine B bait-marking work was granted

under Waikato University Animal Ethics Protocol No. 707.

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