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Journal of InsectConservationAn international journal devotedto the conservation of insectsand related invertebrates ISSN 1366-638X J Insect ConservDOI 10.1007/s10841-011-9422-7

A preliminary assessment of changes inplant-dwelling insects when threatenedplants are translocated

Melinda L. Moir, Peter A. Vesk, KarlE. C. Brennan, Lesley Hughes, DavidA. Keith, Michael A. McCarthy, DavidJ. Coates & Sarah Barrett

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ORIGINAL PAPER

A preliminary assessment of changes in plant-dwelling insectswhen threatened plants are translocated

Melinda L. Moir • Peter A. Vesk • Karl E. C. Brennan •

Lesley Hughes • David A. Keith • Michael A. McCarthy •

David J. Coates • Sarah Barrett

Received: 28 March 2011 / Accepted: 7 July 2011

� Springer Science+Business Media B.V. 2011

Abstract Translocation of threatened species is a tool

used increasingly to conserve biodiversity, but the suite of

co-dependent species that use the threatened taxa as hosts

can be overlooked. We investigate the preliminary impact

of translocating three threatened plant species on insect

species and the integrity of insect assemblages that depend

on these plants as their hosts. We compare the insect

assemblages between natural populations of the threatened

species, related non-threatened plant species growing wild

near the threatened plants, and threatened plants translo-

cated to another site approximately 40 km away. We used

host breadth models and a coextinction risk protocol to

determine which insect species are potentially host-specific

on the threatened plants, and then assessed these insects’

potential presence at the translocation site. We found that

insect assemblages on naturally-occurring threatened plants

had more individuals, higher species density and higher

species richness than assemblages on translocated plants.

For one plant species, Leucopogon gnaphalioides, species

composition differed significantly between wild and

translocated populations (P \ 0.001). Furthermore, four

insect species that were host-specific to Banksia brownii

and B. montana were not detected on the translocated

plants. Instead, translocated plants supported insect assem-

blages more similar to those of related plant species from the

surrounding area. We conclude that threatened plant trans-

locations that involve seed collection and propagation may

have limited benefit for individual dependent species or the

supported insect assemblage. Additional conservation

actions will be required to maintain the diversity of insect

assemblages and host-dependent relationships.

Keywords Assisted colonization �Conservation planning � Coextinction risk �Host specificity � Plant-insect interactions � Translocation

Electronic supplementary material The online version of thisarticle (doi:10.1007/s10841-011-9422-7) contains supplementarymaterial, which is available to authorized users.

M. L. Moir (&) � P. A. Vesk � M. A. McCarthy

School of Botany, University of Melbourne, Parkville,

VIC 3010, Australia

e-mail: mmoir@unimelb.edu.au

K. E. C. Brennan

Western Australian Department of Environment &

Conservation, PO Box 10173, Kalgoorlie, WA 6430, Australia

L. Hughes

Department of Biological Sciences, Macquarie University,

North Ryde, NSW 2109, Australia

D. A. Keith

NSW National Parks & Wildlife Service, PO Box 1967,

Hurstville, NSW 2220, Australia

D. A. Keith

School of Biological, Earth and Environmental Sciences,

University of New South Wales, Sydney, NSW 2052, Australia

D. J. Coates

Western Australian Department of Environment &

Conservation, Locked Bag 104, Bentley DC,

WA 6983, Australia

S. Barrett

Western Australian Department of Environment &

Conservation, 120 Albany Hwy, Albany, WA 6330, Australia

123

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DOI 10.1007/s10841-011-9422-7

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Introduction

Disturbances such as habitat fragmentation and incursion

by exotic species are occurring at a rapid rate, and threaten

many species with extinction. Rapid transformation in

climate over the next 60–90 years will add to existing

environmental stresses and may extinguish or severely

reduce the populations of many threatened species (e.g.,

McClean et al. 2005; Thuiller et al. 2005; Williams et al.

2007; Wilson and Maclean 2011). The potential for loss of

global biodiversity through co-extinction is high because

many of the organisms currently at risk of extinction act as

hosts to numerous dependent species (Koh et al. 2004;

Moir et al. 2010a, 2011).

As one of the 25 biodiversity hotspots identified by

Myers et al. (2000), the south-west of Australia is partic-

ularly susceptible to loss of species. The south-west is

home to an estimated 8,000 plant species (Hopper and

Gioia 2004), 1,909 of which are listed by the State Gov-

ernment as threatened or in need of conservation actions

(Western Australian Herbarium 2011). Translocation trials

of 16 threatened plant species in the south coast region by

the Western Australian Department of Environment and

Conservation have aimed to alleviate the immediate risk of

extinction primarily due to Phytophthora dieback caused

by Phytophthora cinnamomi and inappropriate fire regimes

(Barrett et al. 2008). The ultimate goal of these translo-

cations is to establish viable self-sustaining populations

and to produce seed to establish further populations in

Phytophthora free areas. For this study, we selected three

of the most threatened species found in the montane

regions of the Stirling Range, which also demonstrated

rapid growth when translocated. These are Leucopogon

gnaphalioides Stschegl. (Ericaceae), Banksia brownii R.Br.

(Proteaceae: Spicigerae), and Banksia montana (A.

S. George) (Proteaceae: Spathulatae), all of which are lis-

ted nationally as Endangered under the Environmental

Protection and Biodiversity Conservation Act 1999, and

are also listed as Critically Endangered under the Western

Australian Wildlife Conservation Act 1953. Global climate

change may detrimentally affect these predominantly

montane species through drying of their habitat (e.g.,

Bureau of Meteorology—Australia (BoM) 2011). We

examined the experimental translocation of these plant

species and assessed the preliminary changes in the insect

assemblages between their in situ and ex situ plant popu-

lations. This is a preliminary study because the transloca-

tion experiments have only been underway for a few years

with some plants only recently reaching reproductive

maturity. However, this investigation is warranted given

that insect species associated with the threatened plants are

potentially under immediate risk of coextinction. We

addressed the following questions:

• How distinctive are the assemblages of insects on these

threatened plants in the wild?

• How similar are the species compositions of insect

assemblages found on wild threatened plants and

translocated individuals of the same plant species?

• Are the insect species that use these threatened plants

also found elsewhere, and are the most host-specific

insects also found on translocated individuals?

• Which co-occurring plant species have insect assem-

blages that most resemble those of translocated host

plants?

Methods

Study area

The study areas of the Stirling Range National Park and

Torndirrup Peninsula (Albany city reserve), plus the

translocation site between the two native sites, are located

within the South-west Australian Floristic Region of

Western Australia (Fig. 1). The Stirling Range National

Park covers 115,920 hectares of remnant vegetation, with

sampling occurring in approximately half the area.

Torndirrup Peninsula is remnant vegetation which is joined

to the 3,868 ha Torndirrup National Park, although sam-

pling only occurred within 1 ha of the Peninsula itself.

Banksia montana and L. gnaphalioides are restricted to

areas of high elevation in the Stirling Range ([950 m

above sea level). Banksia brownii occurs in areas of high

elevation in the Stirling Range, but it also persists in sev-

eral disjunct populations within 80 km to the south and

south-east, including at Torndirrup Peninsula. Although

there is no historical record of B. brownii occurring within

the translocation site or the surrounding area, recent pop-

ulation genetic studies suggest that it was probably dis-

tributed throughout the region during at least the mid

Pleistocene (Coates and McArthur 2010). These studies

also suggest that the current disjunct distribution is the

result of subsequent climatic events (Hopper 2009) and

significantly predates any land clearing over the last

150 years (Coates and McArthur 2010). Individuals of

each of the threatened plant species, L. gnaphalioides,

B. brownii, and B. montana, were grown from seed or

cuttings, and transplanted as seedlings approximately

50 km from their nearest current natural populations

(herein termed ‘in situ’) in the Stirling Range National Park

(see Cochrane et al. 2011; Fig. 1). Developed in 2003, the

site to which the plants were translocated is in a 2.8 km2

remnant of woodland on lateritic soils surrounded by

farmland and other native remnants near Kamballup. This

contrasts with in situ sites which consist of thickets and

mallee-heath on rocky sand clay loam soils in the Stirling

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Range, and mallee heath and low woodland on sandy clay

soils over granite at the Torndirrup Penisula (Coates and

McArthur 2010). Climate of all three regions hasn’t been

recorded but 5 km north of Torndirrup, Albany has an

average annual rainfall of 929 mm, and 30 km west of

Kamballup, Mt Barker receives 729 mm annually (Bureau

of Meteorology—Australia (BoM) 2011). Approximately

50 km northeast of the Stirling Range, Ongerup receives

385 mm annually (Bureau of Meteorology—Australia

(BoM) 2011), which may be a close approximation of the

rainfall in the low-lying regions of the Stirling Range, but

montane conditions where the threatened plants occur are

wetter and cooler. The translocation site was selected for

this study due to its location between both in situ sites

(Fig. 1). The translocated plants (herein termed ‘ex situ’)

are being grown within this native vegetation in several

monospecific transects of two rows each, numbering

approximately 40 plants per transect.

Sampling

There were three study sites in total: the translocation site at

Kamballup, and two in situ (or natural/native) sites at

Torndirrup Peninsula and Stirling Range National Park

(Fig. 1). Sampling occurred during September 2007 for the

translocation site, October 2008 at one in situ site (Tornd-

irrup) and from September—December 2007, plus October

2008, within the second in situ site (Stirling Range). The

length of time taken was due to the number of plants sampled

(3,026 individuals, 104 species) for this, and a larger study

on the coextinction of insects on threatened plants. Sampling

was also stratified according to altitude, to account for

delayed responses of insect assemblages at higher altitudes

due to slower temperature increases in these areas. For

example, Fielding et al. (1999) found that for every 100 m in

elevation, the appearance of adults of a Cercopidae bug was

delayed by 5.6 days. Therefore, lowland sites (15–300 m

above sea level: Translocation, Torndirrup and some Stirling

Range areas) were sampled in early spring, while higher

mountain altitudes (350–650 m above sea level) were

sampled later in spring, and the highest altitudes

(700–1,039 m above sea level) sampled in very late spring.

At in situ sites (Stirling Range and Torndirrup), insects

were collected from both threatened and non-threatened,

co-occurring plant species by beating and vacuuming.

When used in combination, these techniques have been

shown to be efficient for sampling insects within this type

of vegetation (Moir et al. 2005b; McCarthy et al. 2010). All

plant individuals were sampled with similar sampling

intensity regardless of site, with all above-ground plant

structures vacuumed with the samples frozen and sorted

under a microscope (see Moir et al. 2005a), or plants were

beaten and insects fell into a modified butterfly net

(*70 cm in diameter) which had the netting replaced with

white calico material. Each plant that was beaten or vac-

uumed was also searched by hand to collect cryptic insect

fauna, such as galls and lerps, not readily captured by the

other sampling methods. Fifteen individual plants of

B. brownii and L. gnaphalioides were vacuumed and

another 15 beaten, while only seven individuals of

B. montana were vacuumed on Bluff Knoll due to the low

numbers of plants remaining here (\15 mature plants). An

additional 14 individuals of 15–22 confamilial species and

genera for each plant species were sampled (seven indi-

viduals vacuumed, seven beaten) to determine which

insects were restricted in host breadth to the focal plant

Fig. 1 Location of study sites

within the south-west of

Western Australia showing the

plant species and number of

individual plants sampled in

square brackets (adapted from

Google Earth, 2010)

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species (Appendix 1—see Electronic supplementary

material; Fig. 1). Confamilial taxa for each threatened

plant species were selected on the basis that they 1.)

occurred in the same area (within 1 km), or 2.) were sister

species and were sampled up to 20 km away.

At the translocation site, 15 individuals of B. montana,

B. brownii (originating from Stirling Range, not Torndir-

rup) and L. gnaphalioides (39, 3, and 28 months-old

respectively) were sampled by vacuuming. We were not

granted permission to beat translocated individuals due to

the low numbers of plants available and concerns that

beating might damage plants. To account for the potential

difference in assemblages captured using the different

sampling techniques, vacuum samples were analysed sep-

arately, as well as in combination with all other samples.

The heights of translocated plants were generally smaller

than those plants sampled at in situ sites due to the dif-

ferences in age (i.e. B. montana translocated 76.2 cm, in

situ *150 cm; B. brownii translocated *70 cm, in situ

*120–280 cm, L. gnaphalioides translocated 31.4 cm, in

situ *56 cm), however, translocated plants grew more

vigorously than the same aged plants at in situ sites (e.g.,

translocated B. montana were on average 76.2 cm and

same aged plants in situ, 6.8 cm; 2-year old translocated

L. gnaphalioides were on average 31.4 cm while 6-year old

plants in situ were 4.7 cm due to grazing: Barrett unpub-

lished data). Three to four confamilial species, and seven to

eight unrelated species, were also sampled at the translo-

cation site by all methods (Appendix 1—see Electronic

supplementary material; Fig. 1). These surrounding plant

species were within 10–100 m of the translocated plants.

Identification of specimens

We preserved all plant-dwelling insects, and identified the

beetles (Coleoptera) and bugs (Hemiptera) to species level,

with validation from taxonomic experts. The smaller orders

(representing 10% of total individuals collected), Orthop-

tera, Phasmatodea, Hymenoptera, Thysanoptera and larval

Lepidoptera, were identified to morpho-species. Introduced

insect species of Coleoptera and Hemiptera (e.g., aphids)

were removed from the analysis as the objective of the

study was to assess native insect assemblages.

Data analysis

Insect assemblages

Differences in insect species density (sensu Gotelli and

Colwell 2001) and insect abundance between sites were

compared for each threatened plant species using one-way

analysis of variance (ANOVA) using SPSS (SPSS Inc.

2001). Data for insect abundance on all plants, and species

density on the banksias, were square-root transformed to

improve normality of the data and homogeneity of variance;

F-ratios were considered significant at P \ 0.05. Smoothed

species accumulation curves (50,000 randomizations)

against observed number of insect individuals for each

threatened plant species at each site were constructed. The

number of species that remain undetected on each plant

species was estimated with the Abundance-based Coverage

Estimator (ACE). Curves were constructed using EstimateS

8.2 (Colwell 2006).

We used analysis-of-similarities (ANOSIM) (Clarke

1993) to assess differences in insect assemblages between

the focal threatened plant species and co-occurring species

(see Appendix 1—see Electronic supplementary material).

We also used ANOSIM to determine the differences in

composition between insect assemblages on in situ and

translocated populations of the three focal plant species.

ANOSIM R-statistics generally range between 0 and 1; a

value of 0 indicates that two assemblages are identical

whereas a value of 1 indicates that two assemblages are

entirely different (see Clarke 1993). Similarity matrices were

constructed using the Bray–Curtis measure for the abun-

dance of insect species on each plant individual. Square-root

transformation down-weights the importance of abundant

species, and increases the contribution from rarer species

(Clarke 1993), which was desirable in this study to prevent

common species from dominating analyses and overlooking

rarer but potentially host-specific insects on each plant spe-

cies. As a precautionary approach, we reduced the chance of

a Type II error (failing to detect dissimilarity when it exists)

by considering ANOSIM R-statistics significant at P \ 0.10.

Similarity percentages (SIMPER) identified the contribution

of each insect species to the dissimilarity between plant

species to help identify insect species that may be under- or

over-represented on translocated hosts relative to in situ

hosts. Non-metric multi-dimensional scaling (MDS) was

performed (1,000 restarts) on the Bray–Curtis matrix to

produce ordinations. These analyses were conducted in

PRIMER-E version 6.1.11 (PRIMER-E Ltd 2008).

Insect species

The taxonomy and/or ecology of the majority of insects

comprising the assemblages on the plants at the study sites

are unknown. To determine the insect species most likely

to be host-specific to the focal plant species we used a host

breadth model using hierarchical Bayesian Zero-Inflated

Poisson modelling. Details of this model are presented

elsewhere (Vesk et al. 2010). Briefly, the model includes

sub-models for host use, insect occupancy and abundance

and ‘tourist’ insect occupancy. The model that we use here

is a modification of those presented in Vesk et al. (2010)

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with the additional aim of estimating occupancy of insect

species at different sites—the in situ and translocated

populations of the focal plant species.

The data Yi are observed counts of insects on individual

plants. We model the data as having three components—

counts of zero, counts of one and other counts. Zeros arise

from either the plant species not being used by the insect

species or from the insect species using the plant species as a

host, but being absent from the sample. Absence from the

sample can be either because the insect truly did not occupy

the individual plant sampled or it occupied the plant but

evaded capture. We do not distinguish between these two

processes in terms of their relative contribution to detection.

Counts of one may arise either from true occupancy of

insects on individual plants that they use, or from ‘tourist’

occupancy on plants that the insects do not actually use.

Counts greater than one, we assumed, would only result from

detections of insects on plants that they truly use as hosts, i.e.

we explicitly assumed that tourist occupancy would not

result in two or more individuals of an insect species on a

plant from a species that they did not actually use.

In contrast with Vesk et al. (2010), here we model the

likelihood of individual observations. We have found that

this approach has better Markov Chain Monte Carlo sam-

pling properties, and while more computationally

demanding and thus slower, provides a more reliable route

for analysis (PA Vesk unpublished data).

The likelihoods are as follows.

LðYi ¼ yjhi;-i;ui;kiÞ

¼hið1�-iÞ þ ð1� hiÞð1�uiÞ; y¼ 0

ð1� hiÞuiþ hi-iki expð�kiÞ=ð1� expð�kiÞÞ; y¼ 1

hi-ikyi expð�kiÞ=ðy!ð1� expð�kiÞÞÞ; y[1

8><

>:

here, pi indicates whether the invertebrate species actually

uses the plant species on which it was recorded in obser-

vation i (hi = 1) or not (hi = 0). The parameter xi is the

probability that the individual plant represented by obser-

vation i is occupied by the insect species represented by

observation i conditional on the plant species being used by

that invertebrate. /i is the probability that the observation

is that of a tourist individual when the plant species is not

used, and ki is the mean abundance of the insect species on

the individual plant when present (but not as a tourist).

Being counts, abundance was assumed to be Poisson dis-

tributed, and truncated to values above 1 because absence

was modelled separately.

We used a hierarchical structure to write models for

each of the parameters hi, xi and ki. The actual use of an

individual plant by an invertebrate species hi was modelled

as a Bernoulli realization with the specified probability pi

for that plant family and invertebrate species combination,

to account for appropriate dependencies in the data.

Variation between plant species within the family was

unmodelled ‘noise’ (Vesk et al. 2010). Occupancy by an

invertebrate species was modelled at the level of species-

by-site combinations. This enabled occupancy of the par-

ticular invertebrate species to be compared between sites,

e.g. in situ and translocated populations of the three focal

threatened plant species. The average abundance ki was

modelled at the level of insect orders. This is quite coarse,

but our focus here is on occupancy. The logit of the

probabilities of use, occupancy and tourist occurrence were

assumed to have normal distributions with unknown means

and standard deviations that were estimated from the data.

We used Bayesian inference, employing vague priors

throughout. Parameters representing means were drawn

from normal distributions with means equal to 0 and stan-

dard deviations equal to H1000. Standard deviations were

specified as drawn from uniform distributions between 0

and 20. The prior distribution for the probability of a tourist

occurrence /i was specified as uniform between 0 and 1.

We used Markov chain Monte Carlo sampling to esti-

mate posterior probability distributions of parameters. The

model was implemented in the freeware OpenBUGS 3.11

(Thomas et al. 2006). The code is presented in Appendix

2—see Electronic supplementary material.

We routinely inspected chain traces and found that

convergence usually occurred within 2,000 iterations. We

discarded the first 5,000 iterations as a burn in and sampled

at least the next 10,000 samples from the Markov chain.

Using the host breadth model, we estimated: the prob-

ability of an insect using each of the three focal plant

species B. montana, B. brownii or L. gnaphalioides; the

host breadth of those insect species and; the probability of

occupancy of those insect species at each site. Insects

identified as having [40% probability of using each

threatened host plant species and a 25% chance that the

host breadth was three plant species or fewer were then

subjected to the decision protocol (Fig. 1 of Moir et al.

2011) to assess their potential as co-threatened taxa. We

used 25% as an example lower bound or threshold to

reduce the risk of failing to recognise a host-restricted

invertebrate (Type II error) below that which might be

expected from use of means or medians. Because host

breadth can only be positive and is akin to a count, there is

a mean–variance relationship that results in higher mean

host breadth with greater uncertainty. The actual threshold

used would be a matter for managers, and would depend

upon their attitude to risk.

Results

Our database of plants and their insect affiliates comprised

1,240 plants representing 44 plant species (11 genera from

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5 families, Appendix 1—see Electronic supplementary

material) and 7,512 insects, representing 607 species (76

families), from the orders Hemiptera, Coleoptera, Orthop-

tera, Lepidoptera, Hymenoptera and Thysanoptera.

Insect assemblages

Translocated individuals of the three focal plant species

harboured on average 84% fewer insect species and 87%

fewer individuals per plant than their in situ counterparts

(Fig. 2). Translocated L. gnaphalioides and B. montana also

had 28% and 36% fewer insect species respectively than in

situ plants, standardised by number of individuals (Fig. 3).

There also remains a larger proportion of insect species

unsampled for on in situ L. gnaphalioides (56%) and

B. montana (62%), than on translocated plants (L. gnapha-

lioides: 55%; B. montana: 22%), according to the estimated

total species richness on each plant species as calculated by

ACE. On the 15 translocated plants sampled of B. brownii,

only three insect species represented by three individuals

were collected, resulting in a steep species accumulation

curve (Fig. 3). The lack of fauna on translocated B. brownii

plants suggests that these three insects were tourists. In situ

B. brownii plants were also estimated to host undetected

insect species (Torndirrup: 28%; Stirlings: 53%).

The insect assemblage of in situ L. gnaphalioides differed

significantly from the 15 other Ericaceae species examined

(listed in Appendix 1—see Electronic supplementary

material; Fig. 4, R = 0.202–0.790, P = 0.001–0.011). In

contrast, the insect assemblages on in situ B. brownii and

B. montana were not significantly different to other Banksia

species. Insect assemblages on B. brownii at Torndirrup

were similar to those on B. quercifolia at the same location

(R = 0.055, P = 0.327). Assemblages on B. brownii in the

Stirling Range were not significantly different from those on

B. concinna (R = 0.024, P = 0.402) at the same location.

Similarly, B. montana had insect assemblages which were

not significantly different from those on B. plumosa

(R = 0.081, P = 0.187), B. anatona (R = 0.029, P =

0.315), or B. oreophila (R = 0.170, P = 0.100) in the

Stirling Range (results for B. plumosa and B. anatona

remained non-significant when only vacuum samples were

analysed). Interestingly, insect assemblages on B. brownii

differed significantly between the Stirling Range and

Torndirrup populations (R = 0.327, P \ 0.001). Similar

results were obtained when the data were analyzed using

presence/absence transformation, which means that there

were actual differences in species, rather than just in the

abundances of species (Moir unpublished data).

The insect assemblage on in situ L. gnaphalioides dif-

fered significantly from those assemblages on translocated

plants (R = 0.798, P \ 0.001), including samples obtained

only by vacuuming (R = 0.738, P \ 0.001), because insect

species from in situ sites were absent at the translocated

site, and vice versa. For example, the beetles Corticaria

australis, Corticaria sp. 124, and Cudnellia sp. 228 were

present on in situ L. gnaphalioides (Appendix 3—see

Electronic supplementary material), but were not found on

Fig. 2 Insect species density (top graphs) and abundance (bottomgraphs) across different treatments (means with different letters are

significantly different from each other at P \ 0.05 from post-hoc

tests following ANOVA) for the plant species L. gnaphalioides,

B. montana, and B. brownii

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Fig. 3 Species accumulation curves for the focal threatened plant

species L. gnaphalioides, B. montana, and B. brownii. The graphs on

the left provide curves for all individuals, graphs on the right provide

curves for the same plant species, but with fewer individuals.

Standardisation shown as dotted lines: for L. gnaphalioides at 17

individuals translocated plants had 5 species, Stirling plants had 7

species; for B. montana at 25 individuals translocated plants had 8

species, Stirling plants had 12.5 species. Error bars are one standard

deviation from the mean

Fig. 4 Ordinations of insect

assemblages on the threatened

plant L. gnaphalioides and

seven plant species within the

same family (selected from

ANOSIMs). Outliers have been

excluded from the figure for

clarity (2 outliers from each of

Leucopogon lasiophyllus,

Leucopogon atherolepis; 3 from

Leucopogon sp. ‘short style’)

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translocated plants. Instead, translocated L. gnaphalioides

had an insect assemblage more similar to a closely related

plant species within the surrounding area, Astroloma

epacridis (R = 0.006, P = 0.376, Fig. 4). The dominant

insect in the assemblage on both Astroloma epacridis and

L. gnaphalioides at the translocation site was the Achilidae

bug Plectoderini sp. 12 (Appendix 3—see Electronic sup-

plementary material).

Insect species

After assessing the complete insect assemblage with host

breadth models and the decision protocol, no insect species

were identified as potentially host-specific to L. gnapha-

lioides in the wild (Table 1). Interestingly, some beetles

(Corticaria australis, Corticaria sp. 124, and Cudnellia sp.

228), highlighted by the community assessment above as

important components of the distinctive assemblage of in

situ L. gnaphalioides, were not host-specific to this plant

species, but were also found on other epacrids in the Stir-

ling Range. For Corticaria, this is most likely due to its

non-specific mycetophagous (fungal) feeding habitats

(Matthews 1992). In contrast, Cudnellia sp. 228 is a her-

bivore and potentially a specialist at family-level as it was

only found on epacrids. Moreover, it is flightless and was

restricted to the summit of Bluff Knoll, suggesting that it is

characteristic of that plant community. These results indi-

cate that, rather than consisting of multiple host-specific

insect species, the assemblage on in situ L. gnaphalioides

was an unusual combination of insect species, which were

not seen together on other epacrids (similar to assemblages

consisting of Threatened Ecological Communities which

may not feature a threatened species; Nicholson et al.

2009).

Host-breadth models and the decision protocol identified

one insect species on in situ B. montana and three insect

species from in situ B. brownii that appeared host-specific

to these plant species (Table 1). Of these four species, the

three hemipterans are highly likely to be host-specific as all

feed exclusively on the sap of the plants. Although the

fourth insect, beetle Bruchinae sp. 282, was not found on

any other plant species, its host-specificity on B. brownii is

uncertain as the Australian members of this subfamily of

insects are thought to be seed feeders as larvae and pollen

feeders as adults (Hangay and Zborowski 2010), and plants

were not in flower when sampled. In any case, the proba-

bility that any of these four insects were present at the

translocated site was very low (\0.6%, Table 2). These

insects were not found on translocated individuals of

B. montana or B. brownii, despite one, the plant-louse

Trioza sp. 03, occurring on in situ populations of

B. brownii at both Torndirrup and Stirling Range.

Discussion

This study suggests that plant translocations that involve

seed collection and propagation, may have limited benefit

for individual host-specific dependent species or the sup-

ported insect assemblage within the first few years, unless

the insects are also assisted in colonising the plants at the

new location. The accumulation of new insect species on

translocated plants was limited in terms of number of

species per individuals collected, but was similar to the

insect colonisation of plant species introduced to countries

or continents outside their natural range (Brandle et al.

2008). Furthermore, insect assemblages on translocated

plants were most similar to those from surrounding, related

plant species at the translocated site, a finding consistent

with what is known about colonisation patterns of intro-

duced exotic species (e.g., Strong et al. 1977; Brandle et al.

2008). Guild structure of the insect assemblages between in

Table 1 Summary of insects collected from targeted threatened plant species, with the insect species potentially restricted to each plant

identified, following analysis by host breadth models (Vesk et al. 2010) and the decision protocol (Moir et al. 2011)

Plant species Location Total insect

species

Total insect

abundance

Insect species restricted to each plant species

Leucopogon gnaphalioides In situ (Stirling) 35 207 0

Translocated 5 17 NA

Banksia montana In situ (Stirling) 19 45 Pseudococcus sp. 15 (Pseudococcidae)

Translocated 8 25 NA

Banksia brownii In situ (Torndirrup) 23 63 Bruchinae sp. 282 (Chrysomelidae)

Ceronema sp. 18 (Coccidae)

Trioza sp. 03 (Triozidae)

In situ (Stirling) 19 118 Trioza sp. 03 (Triozidae)

Translocated 3 3 NA

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situ and translocated plants remain to be investigated, and

may give different results as to the efficacy of host trans-

location on insect assemblages (e.g., Andrew and Hughes

2007). Similarly, as we only examined predominantly one

trophic level (herbivores) present on threatened plants, the

impact of translocation on other trophic levels, as well as

on the cascade of interactions between trophic levels,

remains unknown.

Natural reassembly of insect assemblages on individual

plants that have been reintroduced or translocated into an

area is possible, but depends on several factors including

dispersal capabilities of the insects (i.e. Moir et al. 2010b),

competition from insects already established on the plants

(e.g., Gonzalez-Megıas and Gomez 2003), parasitism and

predation (e.g., Denno et al. 2003), and the physiological

requirements of the insect species themselves (e.g., New

2008). Time lags are expected in the dispersal and estab-

lishment of insect populations on translocated host popu-

lations (Moir et al. 2010b), although reassembly can occur

quickly when source insect populations are located nearby

(e.g., Gratton and Denno 2005). Insect species with poor

dispersal abilities, such as those without wings, may not be

capable of reaching new sites and thus require assistance to

establish (Moir et al. 2005b; Watts et al. 2008). Even

winged insects can be poor dispersers, such as Mitchell’s

satyr butterfly, which does not travel further than

420 m year-1 (Szymanski et al. 2004). In the case pre-

sented here, the translocation site was [40 km away,

located within a landscape matrix fragmented by agricul-

tural development and occurred within different vegetation

types to that of the in situ sites (montane and coastal heath,

versus jarrah/marri woodland), perhaps making dispersal

more difficult. Competition and predation may not inhibit

insect species establishment in this study as most of the

translocated plants were host to very few other insects,

particularly B. brownii. Finally, the physiological require-

ments of the insects examined here may possibly contribute

to their absence from translocated plants as the majority

occupy cooler, wetter areas than the habitats at the trans-

location site. Elucidating the mechanisms as to why insect

species or assemblages are not present on translocated

plants requires further study. In any case, to maintain the

diversity of insect assemblages and host-dependent rela-

tionships that occur on the threatened plant species in the

wild, the next step should involve further conservation

actions, and in particular, overcoming taxonomic impedi-

ments by species descriptions of all host-specific insects

(note that all host-specific species identified here are

undescribed).

Our study has significant implications for assisted

migration or other movements of threatened host species,

which often occur over much greater distances than those

described within this study. Namely, that dependent spe-

cies and associated assemblages are highly unlikely to

re-establish naturally. Rather than a single species

approach when translocating for conservation purposes,

land managers should consider moving suites of species

(Seddon 2010), including dependent herbivores and other

dependent invertebrates with the host taxa (see Moir et al.

in review). This broader method of translocation may be

moving towards the novel ecosystem approach (see Hobbs

et al. 2006), whereby assemblages of species under threat

of extinction through climate change are translocated to an

area where they do not currently occur, thereby creating a

novel ecosystem in the new location (Lindenmayer et al.

2010). Such an approach may carry risks if the translocated

species spread and displace indigenous species in the

vicinity of the translocation site, a question likely to require

significant experimentation to resolve. According to Hunter

(2007), our translocation site would provide a good initial

test of insect translocation as it ‘‘was well connected his-

torically, but is currently surrounded by human-dominated

landscapes that might be a barrier if the translocated spe-

cies had unacceptable effects’’.

Table 2 Estimated probability

of occurrence at in situ (Stirling

Range [Stirl] or Torndirrup

[Torn]) and translocation sites

of insect species identified as

host specific to the targeted

threatened plant species in the

wild (see Table 1)

Plant host Insect species (family) Location Estimated probability

of occurrence (%)

Banksia montana Pseudococcus sp. 15 (Pseudococcidae) Translocation 0.4

In situ (Stirl) 6.8

Banksia brownii Bruchinae sp. 282 (Chrysomelidae) Translocation 0.5

In situ (Stirl) 0.3

In situ (Torn) 5.6

Banksia brownii Ceronema sp. 18 (Coccidae) Translocation 0.3

In situ (Stirl) 0.3

In situ (Torn) 5.2

Banksia brownii Trioza sp. 03 (Triozidae) Translocation 0.5

In situ (Stirl) 10.7

In situ (Torn) 2.8

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Limitations and management needs

This preliminary study has two limitations; lack of site

replication and the relatively limited time since transloca-

tion. Both factors were imposed by logistic constraints on

the availability of both in situ and translocated sites for all

three threatened plant species, but suggest important

directions for future adaptive management. The lack of

replication for both native and translocation sites means

that we were unable to assess the influence of site variation

on the results. While some sources of site variation cannot

be accounted for, differences in plant growth rates, local

richness of invertebrate species, and sizes of plant popu-

lations appear not to explain the observed differences in

invertebrate assemblages between the translocated site and

the in situ sites. Plants in the translocated site grew faster

compared to similar aged plants within in situ sites and, in

the case of B. montana, reaching sexual maturity and

flowering far earlier than in the source population (Barrett

pers. obs.). The translocation site also had as many indi-

viduals of the focal plant species as the in situ sites. There

was a rich diversity of insects in the translocation site (144

insect species in total, excluding introduced species), and

when a plant species occurring naturally both within in situ

and ex situ sites was examined (Banksia sessilis), plants in

the translocated site had more insects than those situated

within in situ sites (translocated site: 110 individual insects,

32 insect species; Stirling Range ‘northeast’ population: 42

individuals, 20 species; Stirling Range ‘southeast’ popu-

lation: 21 individuals, 7 species). Despite these findings, it

is likely that some ex situ sites will produce better trans-

location outcomes than others. Establishment and sampling

of additional translocation sites would improve under-

standing about factors that influence success for both plants

and dependent invertebrates, providing a more informed

basis for planning future translocation projects. Increased

replication of in situ locations is greatly constrained by the

fact that some of the plant species such as B. montana only

remain in one in situ location.

The temporal dynamics of invertebrate occupation of

translocated host plants is unknown (e.g., Moir et al.

2005b, 2010b). Our sampling began relatively soon after

plants were translocated (3–39 months later), and pre-

sumably recorded the early stages of a trajectory in which

increasing numbers of invertebrate species find and colo-

nise newly available host plants. Whether this trajectory

continues to increase in subsequent years requires further

work to resolve. Thus, an important adaptive action is the

continued monitoring of invertebrate assemblages on

translocated plants, relative to in situ populations of the

hosts, as well as the surrounding ecosystem, to determine

the needs for supplementary translocations of invertebrate

species that fail to colonise their translocated hosts. The

Stirling Range has been identified as a hotspot for inver-

tebrate endemism in the south-west (e.g., Department of

Environment & Conservation 2008; Moir et al. 2009), with

invertebrate fauna restricted to this area due to the onset of

aridity in the late Miocene-Pliocene (Main 1999; Moir

et al. 2009; Cooper et al. 2011). Given this, and also that

the majority of insects identified here are wingless and/or

small, together with the long distance of the translocation

site from native areas, we expect that most species will not

recolonise for many decades, if at all.

Given the immediate threats to these host plants in the

wild (see Barrett et al. 2008), and the fact that populations

of host-specific insects may decline faster than that of their

hosts (e.g., Taylor and Moir 2009; Moir et al. 2010a),

conservation actions to assist in the insects survival are

required urgently. This study provides the first step in the

development of an adaptive management strategy that

guides a structured process of learning by doing (e.g., Keith

et al. 2011) to improve translocation outcomes for depen-

dent invertebrate fauna as well as their threatened plant

hosts.

Acknowledgments Grants from the Australian Research Council

(DP0772057) and Australia & Pacific Science Foundation (APSF

07/3) supported this work. We thank P. Luscombe for his enthusiasm

in the translocation experiment conducted on his property, and

assistance with plant identifications. We gratefully thank B. Hanich,

R. Oberprieler, G.S. Taylor, A. Szito, P.J. Gullan, S. Barker and

C.A.M. Reid for taxonomic assistance. Finally, two anonymous

reviewers and Editor T. Shreeve are thanked for their time and useful

suggestions on this paper.

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