Home Uncategorized 650 words 3 hrs

Uncategorized

650 words 3 hrs

Posted 2 reading stuff, choose one to read and do it Aim for 650 words; your mark will be 0% if fewer than 625 words or more than 675 words

Assignment #3: Writing for the public (7% of mark):

Most of the writing that students in university science do follows a style and rules
that are consistent with the scientific literature—regimented, concise and
technical. Even in non-sciences, the writing of scholarship is formal and reference-
based. However, it is important to be able to communicate scientific or scholarly
discoveries in a less intimidating and technical manner.

In this assignment, you are required to choose a recent paper from a scientific
journal from a selection of four papers that are posted on Moodle. These papers
deal with specific issues in conservation biology:

• Choice 1 (Brandt et al 2019): is ecotourism good for biodiversity? An
evaluation of the impact of ecotourism on forest loss/recovery in different
countries.

• Choice 2 (Hare et al 2019): is captive breeding a good way to save
endangered species? An assessment of whether captively-bred animals
(lizards) have the traits they need for survival in the wild.

• Choice 3 (Harris et al 2019): why are insect populations declining? An
investigation into the role of climate change.

• Choice 4 (Price et al 2019): are the impacts of overharvesting being
underestimated? Using DNA from 100-year-old fish scales to reconstruct
historical changes in salmon populations.

You are required to reduce the paper to its essence, including the research area or
context of which it is a part, into a magazine style piece that your magazine editor
requires to fit a space 625 to 675 words, without exception. (Aim for 650 words;
your mark will be 0% if fewer than 625 words or more than 675 words). It has to
be a non-illustrated piece.

The writing must catch the reader’s attention immediately, and it must set up the
context almost as fast. The writing should be precise, but not burdened by any
jargon. Imagine that you are writing for well-informed adults, but who are not
formally trained in environmental science. Use “magazine rules”. It is not an essay,
so referencing is NOT part of this exercise and your writing should be such that
the reader is led through your article. Do not provide statistics! Use a catchy title.

There are several publications that do this kind of scientific journalism. For
examples, you may wish to browse a recent edition of one or more or the following:

• American Scientist
• Discover

• Natural History
• Bioscience
• Scientific American
• National Geographic
• Smithsonian

Your article is due on Thursday, February 13th, 2020. You will be submitting it
through Turnitin.com (instructions will follow).

Predictors of translocation success of captive-reared
lizards: implications for their captive management

K. M. Hare1, N. Schumann2, A. J. Hoskins3, C. H. Daugherty1, D. R. Towns4,5 & D. G. Chapple2

1 Centre for Biodiversity and Restoration Ecology, Victoria University of Wellington, Wellington, New Zealand

2 School of Biological Sciences, Monash University, Clayton, VIC, Australia

3 CSIRO Townsville, Australian Tropical Science & Innovation Precinct, James Cook University, Townsville, QLD, Australia

4 New Zealand Department of Conservation, Auckland, New Zealand

5 Institute for Applied Ecology New Zealand, School of Sciences, Auckland University of Technology, Auckland, New Zealand

Keywords

locomotor performance; egg-laying skink;

Suter’s skink; life history; incubation

temperature; New Zealand; translocations;

captive rearing.

Correspondence

David G. Chapple, School of Biological

Sciences, Monash University, Wellington Rd,

Clayton 3800, Australia.

Email: david.chapple@monash.edu

Present address

School of Graduate Research, University of

Waikato, Hamilton, New Zealand

Editor: John Ewen

Associate Editor: Carolyn Hogg

Received 07 May 2019; accepted 24

September 2019

doi:10.1111/acv.12544

Abstract

Islands are biodiversity hotspots, but their native inhabitants are vulnerable to pre-
dation from exotic predators. Conservation of island endemics has often involved
translocating captive-reared populations to predator-free refugia. However, the
long-term success of these translocations has rarely been assessed. We investigated
the traits that maximize post-translocation survival in a cohort of captive-reared
Suter’s skinks Oligosoma suteri and compared traits normally associated with sur-
vival and persistence of lizards (body condition, speed and overall size) to that of
wild-born skinks 6 years after release onto a predator-free island in the north-east
of New Zealand. Our models showed that larger lizards, and lizards with lower
body condition, had improved survival. While sprint speed of captive-reared lizards
did not differ significantly to that of wild skinks, diving ability of captive-reared
skinks was poor, with only female captive-reared lizards diving during trials. Our
results indicate that the traits associated with higher survival after release are not
necessarily obvious and may be influenced by adaptation to captive conditions.
Long-term monitoring post-translocation is therefore vital to determine the success
of the translocation.

Introduction

Globally, islands represent important biodiversity hotspots
(Myers et al., 2000). They support a large number of endemic
taxa, and often contain species or lineages that are predomi-
nantly, or entirely, represented on islands (Cowie & Holland,
2003; Whittaker & Fern�andez-Palacios, 2007). Thus, when
adjusted for area, islands collectively support a disproportionate
number of the world’s species (Myers et al., 2000; Whittaker &
Fern�andez-Palacios, 2007). Insularity has favoured selection for
low reproductive output and high longevity (Cree, 2005; Whit-
taker & Fern�andez-Palacios, 2007; Covas, 2012), while release
from predation pressure on islands has resulted in a loss or modi-
fication of anti-predator behaviours (Blumstein & Daniel, 2005;
Whittaker & Fern�andez-Palacios, 2007). Combined with their
restricted geographic ranges, these island-specific behavioural
traits mean that island species are especially vulnerable to
anthropogenic threats (IUCN, 2018).

A key threat to island fauna is the introduction of exotic
species, which compete with island biota, alter habitat and
prey upon native species (Towns et al., 2006; Doherty et al.,
2016). Introduced vertebrate predators, especially rodents,
pose a particular threat. These have established on numerous
offshore islands where they prey upon na€ıve island fauna,
reducing reproductive success and recruitment and displacing
and suppressing island populations (reviewed in Towns
et al., 2006). The consequences for island endemics are dev-
astating (Towns et al., 2006; Stolzenburg, 2011), with intro-
duced predators implicated in a considerable number of
extinctions and declines. Indeed, a disproportionate number
of island species have been lost to extinction (reviewed in
Doherty et al., 2016). Accordingly, islands are also hotspots
of extinction (Myers et al., 2000; Whittaker & Fern�andez-
Palacios, 2007).

New Zealand is a world leader in conservation manage-
ment of species threatened by introduced predators

Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London 1

Animal Conservation. Print ISSN 1367-9430

https://orcid.org/0000-0002-7720-6280

mailto:

(Armstrong & McClean, 1995; Romijn & Hartley, 2016;
Towns et al., 2016a). Management has focused on eradicat-
ing mammalian predators from islands before translocating
native species to these island refugia, either to reinforce
existing populations or to found new populations (Towns
et al., 2016a,b, Hitchmough et al., 2016a). Translocated indi-
viduals may be bred in captivity (Seddon et al., 2007; Maran
et al., 2009) and are often ‘head-started’ prior to their release
(Connolly & Cree, 2008; Jarvie et al., 2015). While the suc-
cess of wildlife translocations has traditionally been difficult
to quantify, Miller et al. (2014) proposed a staged approach
to define translocation success using four standardized crite-
ria: survival and growth of individuals, evidence of repro-
duction, population growth and population viability. Based
on this approach, a translocation is considered a success
when released populations meet or exceed one of the four
stages during monitoring (Miller et al., 2014).

Miller et al.’s (2014) standardized criteria for translocation
success were developed using data from the New Zealand
herpetofauna, of which lizards are the major components.
New Zealand has a diverse, endemic lizard fauna comprising
~104 species (61 Eugongylinae skinks, 43 Diplodactylidae
geckos; van Winkel et al., 2018; Hitchmough, 2016). How-
ever, 83% of species are regarded as threatened or at risk,
with introduced mammalian predators representing the pri-
mary threat (Tingley et al., 2013; Hitchmough et al., 2016a;
Towns et al., 2016b). Consequently, many New Zealand
lizard species have declined on the main islands (North
Island, South Island) and primarily persist on smaller, off-
shore islands (Chapple & Hitchmough, 2016; Towns et al.,
2016b).

Conservation attempts have involved translocating lizards
to offshore islands or mainland refugia, and in New Zeal-
and, captive management is increasingly being used to
found populations for translocation to new sites. However,
the success of such translocations has generally been unde-
termined or assessed over relatively short timescales; longer
term monitoring (>4 years) is required to direct future
translocation efforts (Romijn & Hartley, 2016; Towns
et al., 2016a). We investigated the success of a transloca-
tion of a population of captive-reared Suter’s skinks on an
offshore island of New Zealand as well as the factors
likely to predict success. To this end, we used as our case
study a cohort of captive-reared Suter’s skinks (Oligosoma
suteri) that were hatched from eggs of females collected
from Green Island off the north-east coast of New Zealand
and head-started to 18 months of age before being translo-
cated to nearby Korupuki Island in 2001. This represents
the only translocated population of captive-reared lizards in
New Zealand for which long-term data are available. Loco-
motor performance of recaptured individuals from this pop-
ulation was compared to wild individuals captured from
Korapuki Island in 2006 (Miller et al., 2010). The wild
population is the result of a successful translocation of 10
male and 20 gravid female founders successfully translo-
cated from Green Island to Korapuki Island in 1992 (see
Towns & Ferreira, 2001 for details of this translocation).
The translocation sites were similar in appearance, with

similar sized rocks. However, the Towns & Ferreira (2001)
translocation of adults to the northern beach was north fac-
ing, and likely received more sunshine hours, and had a
gentle slope up to the forest; the second translocation of
juveniles was to a southern facing beach with a cliff face
up to the forest.

The nocturnal Suter’s skink is the only oviparous lizard
species endemic to New Zealand (Cree & Hare, 2005). It is
a medium-sized (maximum snout-vent length (SVL) of c.
108 mm, Hardy, 1977) skink that is restricted to rocky bea-
ches of northern New Zealand (Hare et al., 2008a; Chapple
& Hitchmough, 2017), presumably due to cooler tempera-
tures at more southerly latitudes (Hare et al., 2004). It is
vulnerable to introduced predators (Towns et al., 2003) and
has experienced significant range reductions in recent years,
with the species now largely limited to offshore islands
(Towns & Ferreira, 2001; Towns et al., 2016b). Individuals
are long-lived, reaching at least 12 y in the wild (Towns &
Ferreira, 2001), and typically deposit their first clutch at
~33 months after hatching (Towns, 1975). The breeding sea-
son is broadly mid-Spring to Summer, with ovulation occur-
ring in October–November and oviposition in late
December–early January (Towns, 1975).

Hare et al. (2012) reported the short-term (~12–18 months
pre-release) outcomes of captive rearing prior to release of
Suter’s skinks to Korupuki Island. Here we report on the
longer-term outcomes (~5.5 years) of the translocated Suter’s
skink population to assess the success of this translocation
as per Miller et al. (2014), and explore the factors that influ-
ence translocation success. Specifically, we examine:

1 the key factors that can be measured prior to release that
impact survival of captive-reared lizards in the wild (sex,
body condition, mass at hatching and at 12 months, and
sprint speed at 4 months, selected because these are com-
monly measured during translocation programmes and are
therefore relevant to other programmes attempting to
gauge translocation success), and

2 how locomotor performance of translocated captive-reared
adults compares to wild individuals post-release.

We expected that over the timeframe of the study, evi-
dence of survival, growth and reproduction (Miller et al.,
2014) would be detected if this translocation has been suc-
cessful, but longer term data would be necessary to evalu-
ate whether this population has met Miller et al.’s (2014)
remaining criteria for translocation success (population
growth and viability). Our study will help determine
whether captive rearing programmes represent a viable con-
servation technique for Suter’s skinks, and will facilitate
consideration of the potential impacts of captive rearing on
the likelihood of translocation success in lizards. In addi-
tion to furthering our understanding of what influences
translocation success, much of the information revealed in
this study will assist managers to predict the potential qual-
ity of captive-reared founders (i.e. individuals that are
likely to produce successful translocation outcomes). This
will enable the most suitable candidates for release into the
wild to be identified.

2 Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London

Translocation success in lizards K. M. Hare et al.

Materials and methods

Egg collection & incubation

Details of egg collection and incubation are described in full
in Hare et al. (2002), and in Hare et al. (2004, 2008b).
Briefly, eggs were collected from 58 female Suter’s skinks,
sourced from Green Island, Mercury Island group, New
Zealand (36°38’S, 175°51’E, 1–2 m a.s.l.) in Summer, 1999.
The eggs were randomly assigned to one of three incubation
temperatures (18°C, 22°C and 26°C) and one of two water
potentials (�120 and �270 kPa). Each clutch (two to five
eggs per clutch, Hare et al., 2002) was (where possible) rep-
resented at all three temperatures, but at only one water
potential per temperature. Water potential of the vermiculite
incubation medium had no influence on any variables by
12 months of age (Hare et al., 2004; 2008b) and is not con-
sidered further.

Juvenile maintenance

Immediately after hatching, we attempted to identify sex,
and individuals were weighed (�1 mg), measured (SVL and
tail length, �0.5 mm) and permanently marked by toe clip-
ping. Skinks were initially housed in groups of three under
the same conditions that gravid females were provided, in
transparent boxes (215 9 330 9 110 mm, 7 L) with a lid
containing 1 mm2 wire mesh (165 9 120 mm) for ventila-
tion. At least 20 mm of moist, non-fertilized potting mix
was used as the substrate and pieces of bark were provided
for shelter. A heat strip at one end of each box provided a
temperature range of 14–30°C. At 5–8 months of age, juve-
niles were re-sorted by size into groups of 6–12 in metal
enclosures (700 9 580 9 350 mm) with 1 mm2 mesh lids.
The metal enclosures were lined with 50 mm of sand, with
driftwood, dried seaweed and small stones for shelter. Juve-
niles continued to be periodically re-sorted by size.

Juveniles were given free access to water and fed ad libi-
tum with live food items and fish-based cat food. Live food
was regularly supplemented with calcium powder and vita-
min drops (Avi-CalTM; calcium borogluconate 200 g/L,
cholecalciferol (Vitamin D3) 12 500 IU/L and magnesium
sulphate 5 g/L). Day/night was simulated using UV bulbs
(Duro-test� True-lite� power twist fluorescent tubes) on a
12L:12D light cycle.

Release, monitoring, survival and
reproduction post-release

Of the 136 hatchlings, 94 (69.1%) survived the 18 months
until release in early October 2001 (Hare et al., 2004; this
study). These were released en masse on a south-facing
beach (hereafter Release Beach) on Korapuki Island, Mer-
cury Island group, New Zealand (36°39’S, 175°50’E 36, 1-
2 m a.s.l), ~1.75 km from the source location of Green
Island. The release site is separated from other beaches on
the island by steep cliffs; while Suter’s skinks have

previously been translocated to Korapuki Island (Towns &
Ferreira, 2001), those skinks in previous translocations were
released on a north-west facing beach ~200 m (coastal dis-
tance) from Release Beach. At that time there was no evi-
dence of existing Suter’s skink individuals occupying
Release Beach (Towns unpubl. data). Both translocated pop-
ulations of lizards were exposed to a suite of native preda-
tors, and for New Zealand lizards it included birds, bigger
reptiles and invertebrates (Hare et al., 2016).

We conducted a mark–recapture exercise in 2005 and 2007,
approximately 4–6 years after release. Due to the cost and per-
mits required to undertake research on protected islands in New
Zealand, the timing and effort involved in the monitoring trips
was opportunistic, and part of much larger scientific events. In
2006, we measured the locomotor performance of recaptured
individuals, and were therefore unable to do a mark–recapture
study as individuals were captured for the experiments. For
mark–recapture studies, the site was monitored over 5 days on
18–22 November 2005 and 28 February–4 March 2007. Five
pitfall trapping sites were established: three on Release Beach,
and one each on beaches to the east and west of Release Beach
(to determine if skinks were expanding around the island). On
Release Beach, trapping sites were located: (A) at the initial
release site, (B) 20 m west of the initial release site and (C)
40 m west of the initial release site. At Sites A and B, pitfall
traps were set at 2-m intervals in three lines of five, positioned
from the storm line (bottom of the cliff face) to the high-tide
mark. At Site C and the western beach, we placed two sets of
four traps midway between the storm line and the high-tide
mark at 2-m intervals. At the eastern beach, we used a perma-
nent lizard monitoring transect of 20 traps in five stations, each
20 m apart along the storm line with four traps set on the cor-
ners of each 2 9 2 m station. All traps comprised 4l plastic
paint pails that were provided with cover and baited daily with
canned cat food. Traps close to the high-tide mark were cov-
ered with galvanized netting to stop entry by large crabs, which
are known to prey upon trapped lizards (Hare et al., 2016) and
indeed even wild-ranging lizards (Bell & Bauer, 2008). We
placed large flat rocks upon the traps, with gaps beneath, to pre-
vent captured lizards from overheating in the sun. On the
south-eastern beach, each of the four traps at each station was
baited with either fish- or meat-based cat food, canned pear, or
left unbaited. Except for traps placed on the south-eastern
beach, all traps were set forc. 12 h, and checked in the early
morning and late evening for 5 days. The south-eastern beach
traps were checked at intervals of 48 h and set three times over
6 days.

Sex, mass (�0.5 g), SVL (�0.5 mm) and for females,
gravidity status were recorded for recaptured individuals dur-
ing each year of trapping. Sex was determined by eversion
of the hemipenes in males. Gravidity status and an estimate
of clutch size were obtained via abdominal palpation by a
trained expert (KMH).

Locomotor performance

We used sprint speed and dive duration as measures of loco-
motor performance post-release in 2006. The methods used

Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London 3

K. M. Hare et al. Translocation success in lizards

to measure these locomotor performance traits in other
O. suteri individuals are described in detail in Miller et al.
(2010). Briefly, both speed and diving trials were conducted
under natural photoperiods (14:10 light:dark cycle, with sun-
rise at ~0600 h) and at ambient air temperatures averaging
17.5 � 0.05°C. Three trials were conducted within 1 day for
both performance measures, and all individuals were at least
2 days post-absorptive. Dive trials were conducted after com-
pletion of sprint trials, with a minimum of 120 min rest
between trials of the two performance measures.

Sprint speed was measured between 0930 h and 1230 h
using a field portable timer (Huey et al., 1981). The order of
individuals was randomized over the three trials, but in all
cases, individuals were given at least 15 min rest between
trials. A plastic racetrack (1.5 9 0.8 m), with five paired
infrared lights (0.25 m apart and 4 mm high) over 1 m, was
used to test sprint speed. The lights transmitted and received
an infrared beam horizontally across the track, and the inter-
ruption of each successive infrared beam stopped its paired
timer. The fastest speed over 0.25 m was used in analyses
since burst speed is likely to be a more ecologically relevant
measure for species that live on rocky shores where long
sprints are not necessary to reach cover. To encourage
sprinting, we gently touched the tails of lizards with a paint-
brush. We also recorded the number of pauses, scored when
an individual ran forwards after its tail was touched before
stopping and requiring another touch to continue, over 1 m.
Performance in captive-reared individuals was compared with
that of wild-born individuals, captured from the northern
release beach; these data were published in Miller et al.
(2010).

Voluntary diving trials were undertaken in an artificial
rock pool between 1530 h and 1930 h. The order of individ-
uals was randomized over the three trials and animals were
given at least 60 min rest between each dive trial. The artifi-
cial rock pool comprised a darkened round plastic bin (base
340 mm diameter; 20 L to fill line) filled with rocks (in a
centre spire with one emergent rock), seaweed and intertidal
seawater (mean temperature 16.7 � 0.07°C; modified from
Hare & Miller, 2009). Skinks were placed on the emergent
rock facing away from the researcher and diving was encour-
aged by tapping on the tail. Some skinks did not dive, and
instead swam about on the surface of the water; if these ani-
mals did not dive after 10 sec, they were removed. If an ani-
mal dived, we recorded dive duration (i.e. the time between
submergence and emergence) using a manual stopwatch (ac-
curate to 1 sec). Dive duration of captive-reared skinks was
compared to that of wild-born individuals, using data from
Miller et al. (2010).

Statistical analyses

In order to determine the factors that influenced long-term
survival of captive-reared Suter’s skinks, we fitted a Baye-
sian Cormack-Jolly-Seiber (CJS) survival model (hereafter,
full model). To account for differences in trap days, detec-
tion probability was included as a fixed effect with a random
effect of year (eq. 1).

logit observationð Þ � 1 þ randomeffect year of surveyð Þ ((1))

Survival probability was modelled as a function of sex,
mass at 12 months, mass at hatching and body condition
(eq. 2). In order to account for the 5-year gap between
release and the first trapping period, year since release was
treated as a two level factor (i.e. Year 1 and the following
years) within the survival component of the model (eq. 2).
We calculated body condition according to whether individu-
als were heavier or lighter than predicted for their length
based on the residuals from the estimated mass using log/log
allometric equations.

logit survivalð Þ � mass12months þ masshatching
þ body cond12months þ factor yr since releaseð Þ (2)

The influence of sprint speed at 4 months on survival
post-release (Hare et al, 2008b) was incorporated into a sep-
arate model, following the same structure, using a reduced
dataset to include only those individuals that were run in
sprint speed trials upon recapture (hereafter, reduced model).
Snout–vent length (SVL) was omitted from all models due
to its high correlation with mass. Analyses were carried out
using JAGS (Plummer, 2003) through R (R Core Team,
2017) using the r2hags package (Su & Yajima, 2015).

Sprint speeds of captive-reared individuals were compared
to those of 63 wild-caught individuals, all males, from Miller
et al. (2010) using a linear regression that modelled sprint
speed against the interaction of snout–vent-length and type
(captive-reared or wild-caught) (eq. 3). To enable a fair com-
parison with adult wild-caught animals, sprint speeds of cap-
tive reared individuals were those measured upon recapture
as adults and not the juvenile sprint speed used in the sur-
vival analysis.

sprint speed � snout � vent � length � factor typeð Þ (3)

Results

Survival, size and reproduction

In total, 21 captive-reared individuals were recaptured over
the 3 years, with an additional three unmarked individuals
captured in 2006 and 2007 (SVL range = 55–80 mm). Of
the marked individuals, 17 captures of 10 individuals were
made in 2005, 10 individuals were trapped in 2006 and 25
captures of 17 individuals were recorded in 2007. Recap-
tured individuals were incubated at either 22°C (n = 12) or
26°C (n = 9); no individuals incubated at 18°C were recap-
tured. All recaptures were from Release Beach, and most
were from the original release site (transect A; 78% of cap-
tures), with some from transect B (20 m south of the release
site; 22% of captures), and none from transect C (40 m
south of the release site). No Suter’s skinks were captured
from beaches directly north or south of Release Beach.

Size of individuals increased during the course of the 7-
year study, with growth rate averaging 7.1 mm per year. At

4 Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London

Translocation success in lizards K. M. Hare et al.

hatching, mean SVL and mass of the 21 individuals was
34.9 � 0.16 mm and 0.77 � 0.01 g, respectively, and by
November 2005, all recaptured individuals had attained size-
based maturity (Towns, 1975, Table 1). The average size
and weight of individuals continued to increase during the
course of the study (Table 1).

In November 2005, three of five females were gravid with
1–2 eggs, and in November 2006, all five females captured
were gravid with 1–3 eggs. Two females were captured in
both 2005 and 2006, and of these, one was gravid with one
egg in both years, and the other was only gravid in 2006.

Factors influencing survival probability

The full model indicated strong support (at the 95% confi-
dence level) for a positive influence of mass at 12 months
on lizard survival, with larger lizards experiencing greater
survival, but no influence of mass at hatching (Table 2,
Fig. 1). In contrast, body condition negatively impacted sur-
vival. This suggests that individuals who are heavier, but
lighter than predicted for their length, are more likely to sur-
vive (Fig. 2). These findings were replicated in the reduced
model, which also did not show support (at the 95% confi-
dence level) for a relationship between sprint speed and sur-
vival (Table 3).

Locomotor performance comparisons

Gravidity in female skinks has been shown to affect sprint
speed (Shine, 1980). Unfortunately, we only had sprint speed
information for gravid captive-reared females (present study)

Table 1. Mean snout–vent length (SVL) and weight of Suter’s

skinks Oligosoma suteri throughout the course of the study. Values

in parentheses represent range

Date (age) SVL (mm) Mass (g)

Feb–Jun 2000

(hatching)

34.9 � 0.16 (28-38) 0.76 � 0.01 (0.33–0.92)

Feb–Jun 2001

(12 months)

49.9 � 0.3 (40–56) 2.07 � 0.04 (0.77–2.82

November 2005

(<6 years)

80.56 � 0.85 (75–84) 10.17 � 0.41 (8.5–13)

November 2006

(<7 years)

83.4 � 0.48 (82–87) 10.58 � 0.19 (9.75–11.8)

March 2007

(7 years)

84.81 � 0.59 (81–89) 10.81 � 0.38 (8.25–12.8)

Table 2. Results of Bayesian BJS survival model for the survival of

captive raised skinks released into the wild. Italicized and

underlined values indicate those where a strong level of support

(95% CIs do not cross 0) for their influence on survival were

detected

Mean SE 2.50% 97.50%

Mass (12 mo) 1.056 0.486 0.248 2.132

Mass (0 mo) 0.013 0.332 �0.590 0.707
Condition �0.945 0.489 �2.074 �0.157
Year(1) �1.313 0.320 �1.958 �0.711
Years(2:4) 82.229 58.697 7.444 222.100

Intercept (obs. probability) �1.402 0.562 �2.160 �0.635
Sigma 0.467 0.785 0.012 2.096

Figure 1 Estimated probability of survival from year one showing

individual influential responses for body condition (blue) and mass

(red) in Suter’s skinks (Oligosoma suteri) [Colour figure can be

viewed at wileyonlinelibrary.com]

Figure 2 Observed (points) and estimated (dashed line) mass of

Suter’s skinks (Oligosoma suteri) by snout–vent length. Estimated

line indicates allometric equation (log[mass] ~ log[length]) gener-

ated from the observed data. Point colour indicates the estimated

probability of survival based on fitted survival model. Note: X-axis

values have a precision s mm but have been ‘jittered’ to improve

visibility [Colour figure can be viewed at wileyonlinelibrary.com]

Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London 5

K. M. Hare et al. Translocation success in lizards

and non-gravid, wild females, caught from the northern
Release beach (Miller et al., 2010). Therefore, to avoid the
confounding effect of reproductive status, we did not com-
pare speed in these. Hence, only males were included in sta-
tistical comparisons of sprint speed. These comparisons
indicated no significant differences in maximum sprint
speeds between captive-reared and wild-born males
(P > 0.05 for all covariates), with average sprint speeds of
1.05 � 0.06 m/s (range 0.89–1.21) and 1.12 � 0.03 (range
0.79–2.21) attained by captive-reared and wild-caught male,
respectively. Captive-reared (gravid) and wild-caught (non-
gravid) females achieved mean speeds of 0.74 � 0.13 (range
0.45–1.05) and 1.05 (0.87–1.27) respectively.

Captive-reared lizards rarely dived; of the 10 captive-
reared individuals (five males and five gravid females) used
in dive trials, only four (40%) dived in at least one trial.
Two skinks (20%) dived in all three trials and six (60%) did
not dive in any of the trials. Instead, for the majority of dive
trials, lizards swam about the water surface. All individuals
who dived were female, with dive durations averaging
124 � 35.5 s (range 22.2–185.8 s).

Discussion

Larger, leaner lizards have a higher
probability of survival following
translocation

Unlike previous studies on survival of this species, which
investigated the factors influencing survival in captivity (e.g.
Hare et al., 2004, 2008b), our study presents novel data on
long-term post-release outcomes for captive-reared Suter’s
skinks. These indicate that the best predictors of survival of
Suter’s skinks post-translocation are size and body condition;
both larger lizards and lizards with lower body condition at
12 months experienced greater survival. The magnitude of
the size difference between skinks is large enough to be bio-
logically significant, with a 16-mm and 2.05 g difference in
SVL and weight, respectively, between the smallest and lar-
gest individuals about 12 months old. These results have

implications for the increasing use of captive management
for threatened species and suggest that larger, leaner lizards
should be prioritized for translocations.

Higher survival might be expected in larger lizards for
several reasons. Firstly, bigger individuals typically have lar-
ger heads, with positive follow-on effects on bite force (Her-
rel et al., 1999; Herrel et al., 2001; Verwaijen et al., 2002),
prey handling efficiency and the ability to eat bigger and
harder-bodied prey (Verwaijen et al., 2002). In addition, lar-
ger lizards may have fewer predators than their smaller
counterparts, with some predators preferentially selecting, or
limited to, smaller prey individuals (e.g. Hasegawa, 1990;
Webb & Shine, 1993). Finally, greater size confers a com-
petitive advantage, with larger-bodied individuals more likely
to win agonistic interactions (Tokarz, 1985; Cooper & Vitt,
1987; Sacchi et al., 2009), thereby maximizing their ability
to secure key resources, such as territories (Sacchi et al.,
2009), favourable perching sites (Tokarz, 1985), food
(Stamps, 1977) and optimal basking and refuge sites (Car-
others, 1981).

Large size also has reproductive benefits. Body size in
skinks has a positive relationship with clutch and litter size
(Cree & Hare, 2005; Chapple, 2006, 1994; Chapple et al.,
2014), and Suter’s skinks are no exception, with females of
a greater SVL producing significantly more eggs (Hare
et al., 2002). Thus, not only did larger individuals experi-
ence higher survivorship, but they likely also have greater
reproductive output. This could potentially increase the over-
all likelihood of translocation success (Towns et al., 2016a).

The reduced survival of skinks with higher body condition
seems counter-intuitive. Lipid reserves of lizards provide an
energy store during lean periods (Avery, 1970; Derickson,
1976), and support reproduction, growth and maintenance
(Derickson, 1976), so it is logical that heavier translocated
lizards, with their presumably greater fat reserves to draw
on, should have improved survival. So why would Suter’s
skinks experience the opposite following translocation? The
answer may simply be that they are overly acclimated to
captive conditions. These can influence phenotype, including
social and anti-predator behaviour (Snyder et al., 1996; Con-
nolly & Cree, 2008), and produce inactive, overweight indi-
viduals (Connolly & Cree, 2008) that are less adept at
capturing prey in the wild; when the fat reserves of these
individuals are depleted, their survival is compromised. In
contrast, Hare et al. (2012) found no influence of body con-
dition index or size on survival post-release in captive-reared
Otago skinks (O. otagense), potentially a function of differ-
ences in sample size, species and/or the amount of time
post-release over which monitoring was conducted. Nonethe-
less, our results suggest that captive management can result
in selection for unfavourable traits in lizards.

It is also possible that underlying factors that affect phe-
notype, such as the rearing environment and maternal effects,
are driving survival post-release. For example, maternal
effects, which influence traits such as body size and condi-
tion (e.g. Shine & Harlow, 1993; Wapstra, 2000), may be
the causal drivers behind the relationships observed here
rather than body size and condition per se. Regardless of the

Table 3. Results of Bayesian BJS survival model, on reduced

dataset including only those individuals where sprint speed

informations were detected, for the survival of captive raised

skinks released into the wild. Italicized and underlined values

indicate those where a strong level of support (95% CIs do not

cross 0) for their influence on survival were detected

Mean SE 2.50% 97.50%

Mass (12 mo) 1.805 0.702 0.621 3.402

Mass (0 mo) 0.005 0.533 �1.053 1.107
Condition �1.617 0.597 �2.883 �0.549
Sprint Speed �0.520 0.528 �1.656 0.441
Year (1) �1.867 0.438 �2.793 �1.085
Years (2:4) 83.890 60.963 7.400 229.463

Intercept (obs. probability) �1.369 0.450 �2.190 �0.551
Sigma 0.455 0.656 0.011 2.054

6 Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London

Translocation success in lizards K. M. Hare et al.

proximate mechanism, it is reasonable to use phenotypic
traits, which can be easily and routinely measured by captive
managers, as predictors of survival since phenotype is an
expression of these underlying factors.

Lizards that were incubated at 18°C were not recaptured 4
or more years after the translocation, supporting Hare et al.’s
(2004) suggestion that individuals incubated at this tempera-
ture would not survive to maturity. Such low survival rates
may be due to the smaller body sizes reached by these indi-
viduals at 12 months (Hare et al., 2004) or a consequence of
the increased rate of abnormalities from this incubation treat-
ment (Hare et al., 2002). This result is congruent with previ-
ous research, which has demonstrated a negative influence of
incubation temperature extremes on survival and phenotype
on reptiles (Noble et al., 2018), highlighting the importance
of considering incubation regimes when designing transloca-
tion programmes; individuals should not be incubated at tem-
perature extremes, and if they are, should not be released
due to reduced likelihood of translocation success.

As we predicted, the captive-reared Suter’s skinks translo-
cated to Korapuki Island have met two of Miller et al.’s
(2014) four criteria that indicate translocation success: sur-
vival and growth, and evidence of reproduction. Individuals
increased in body size throughout the course of our seven
year study, reaching maturity by 2005 when the first recap-
tures commenced. By this stage, most females were gravid
with one or more eggs. Thus, this translocation can be tenta-
tively deemed successful. Wild individuals translocated from
Green Island to Korapuki Island have met all of Miller
et al.’s (2014) criteria for success (Towns & Ferreira, 2001;
Miller et al., 2011, 2014), with population growth detected
within 3 years of release (Towns & Ferreira, 2001), and
juveniles representing approximately 25% of the consistently
high number of individuals captured since 2000 (e.g. 168
captured in 2006 from 30 founders released in 1992) during
biennial monitoring (Miller et al., 2011, 2014). However,
Towns & Ferreira (2001) cautioned that at least 20 years
may be required before translocation success can be
declared. Nonetheless, the success rate of translocations for
captive-reared lizards has rarely been monitored and evalu-
ated (e.g. Towns et al., 2016a). As such, our case study pro-
vides valuable information that can inform future
translocation efforts.

Captive-reared and wild individuals
sprinted at equivalent speeds

Our study presents important data on the performance of cap-
tive-reared skinks in the wild compared to that of wild-born
individuals (Miller et al., 2010). Sprint speed of captive-
reared males 5 years post-release was similar to that of wild
males. Our result was unexpected in light of previous studies
on Oligosoma skink species in which sprint speeds of cap-
tive-reared lizards were considerably slower than those of
wild individuals (Connolly & Cree, 2008; Hare et al., 2012).
The slower speed has previously been attributed to heavier
body mass of captive-reared skinks (Connolly & Cree, 2008;

Hare et al., 2012), reduced fear of humans or physiological
changes owing to the restricted size of enclosures (Hare et al.,
2012). These factors are unlikely to have influenced lizards in
our study since they were not captive at the time of trials and
had 5 years in the wild to develop their sprinting ability. The
similarity between captive-reared and wild-born males in our
study could represent a bias towards faster captive-reared
adults if 1. sprint speed at 4months, which did not affect sur-
vival probability, is not predictive of adult speed, and 2. adult
survival is influenced by locomotor ability, resulting in only
the fastest captive-reared males surviving to recapture. Alter-
natively, maximum sprint speed in this species may simply
not be a fitness-related trait. Instead, the proportion of maxi-
mum speed used in nature, or ecological performance, could
be more biologically relevant, and the trait upon which selec-
tion acts (Irschick & Garland, 2001; Irschick, 2003; Husak,
2006). For example, Husak (2006) found that survival of
yearling and adult collared lizards (Crotaphytus collaris) was
dependent on the speed used by individuals to escape preda-
tors, irrespective of their maximal sprinting capacity. Like-
wise, Hoskins et al. (2017) proposed that the low
repeatabilities of sprint speed in the closely related shore
skinks (O. smithi) could indicate that skinks were not per-
forming at their maximum capacity in field trials.

Diving was inhibited in captive-reared
males

Only 40% (all female) of captive-reared skinks dived in one
or more dive trials (cf 89% in Korapuki Island wild-born
skinks: Miller et al., 2010), and while the proportion of
females that dived was similar to the wild lizard average
(89% of skinks dived in at least one trial), their dive dura-
tions (22–186 s) were shorter than, or in the lower range of,
dive durations by wild-caught lizards (71–1229 s: Miller
et al., 2010). The willingness to dive may reflect predator
escape behaviour in gravid females (Miller et al., 2010)
since gravidity impedes sprint speed in lizards (Shine, 1980;
Cooper et al., 1990). While the effects of small sample size
cannot be discounted, it is possible that avoidance of diving
by males and shorter dive durations by females were due to
a lack of opportunity to acquire diving skills during crucial
stages of development. Whatever the reasons for the diving
behaviour observed here, our results imply a generally nega-
tive impact of captivity on diving behaviour in Suter’s
skinks, though additional research is needed to elucidate this
further.

Conclusions

Our study has shown that the traits associated with post-
translocation success may not always be intuitive, and could
potentially be influenced by adaptation to captivity. We
found that larger, leaner skinks had greater survival, suggest-
ing that these should be preferentially selected for transloca-
tions of this species, and that sprint speed of captive-reared
individuals was equivalent to that of wild-born lizards. Our

Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London 7

K. M. Hare et al. Translocation success in lizards

study demonstrates the importance of long-term monitoring
of translocated populations in order to ascertain success rates
and identify key traits that positively influence translocation
success. In turn, the most suitable candidates for future
release can be determined. Similar quantitative assessments
should be a fundamental component of future translocations
globally.

Rearing animals in captivity until they attain a size where
they are less likely to succumb to predation or starvation
through the critical neonatal period (Ferguson et al., 1982)
can improve survival post-release (Ferguson et al., 1982;
Escobar et al., 2010). However, captive conditions can influ-
ence phenotype in ecologically relevant ways (Burghardt
(2013); Snyder et al., 1996; Connolly & Cree, 2008), and
this must be addressed when establishing captive manage-
ment programmes. Certainly, captive rearing should be con-
sidered a last resort option only, and not all species are
suitable for captive conditions. For example, western popula-
tions of grand (O. grande) and Otago skinks (O. otagense)
experienced poor survival and breeding in captivity (Hare
et al., 2019). Therefore, where captive management is
deemed appropriate, breeding facilities should provide ample
opportunity for lizards to acquire the skills necessary to max-
imize survival post-release. To that end, large enclosures
comprising a variety of structurally complex microhabitats
that closely mimic wild conditions, and allow individuals to
display natural foraging, antipredator and social behaviours,
should be established. For example, semiaquatic species
would benefit from access to deep water to acquire diving
skills, whereas arboreal species would benefit from climbing
apparatus of different structures.

Acknowledgements

We thank all those who provided expertise, time and assis-
tance, in particular Ian Atkinson, Iris Broekema, Chris
Green, Sue Keall, Chris Longson, Kim Miller, Richard
Moore, and Nicky Nelson. Funding was provided by the
New Zealand Department of Conservation (DOC) and Victo-
ria University of Wellington (VUW). The research was car-
ried out with DOC approval, VUW Animal Ethics
Committee approval, and after consultation with Ng�ati Hei,
Ng�ati Maru and Ng�ati Whanaunga.

References

Armstrong, D.P. & McClean, I. (1995). New Zealand
translocations: theory and practice. Pac. Conserv. Biol. 2,
39–54.

Avery, R.A. (1970). Utilisation of caudal fat by hibernating
common lizards, Lacerta vivipara. Comp. Biochem. Physiol.
37, 119–121.

Bell, T. & Bauer, A.M. (2017). Predation on a free-ranging
Raukawa gecko (Woodworthia maculata) by a purple rock
crab (Leptograpsus variegatus). Biogecko Issue 4, 20–35.

Blumstein, D.T. & Daniel, J.C. (2005). The loss of anti-
predator behaviour following isolation on islands. Proc.

Roy. Soc. B: Biol. Sci. 272, 1663–1668. https://doi.org/10.
1098/rspb.2005.3147

Burghardt, G.M. (2013). Environment enrichment and
cognitive complexity in reptiles and amphibians: concepts,
review, and implications for captive populations. Appl.
Anim. Behav. Sci. 147, 286–298.

Carothers, J.H. (1981). Dominance and competition in a
herbivorous lizard. Behav. Ecol. Sociobiol. 8, 261–266.

Chapple, D.G. (2003). Ecology, life-history, and behavior in
the Australian Scincid genus Egernia, with comments on
the evolution of complex sociality in lizards. Herpetol.
Monogr. 17, 145–180.

Chapple, D.G. (2005). Life history and reproductive
ecology of white’s skink, Egernia whitii. Aust. J. Zool.
53, 353–360.

Chapple, D.G. & Hitchmough, R.A. (2016). Biogeography of
New Zealand lizards. In New Zealand lizards: 109–131.
Chapple, D.G. (Ed.). Switzerland: Springer.

Chapple, D.G., Miller, K.A., Chaplin, K., Barnett, L.,
Thompson, M.B. & Bray, R.D. (2014). Biology of the
invasive delicate skink (Lampropholis delicata) on Lord
Howe Island. Aust. J. Zool. 62, 498–506.

Connolly, J.D. & Cree, A. (2008). Risks of a late start to
captive management for conservation: phenotypic
differences between wild and captive individuals of a
viviparous endangered skink (oligosoma otagense). Biol.
Conserv. 141, 1283–1292.

Cooper, W.E. & Vitt, L.J. (1987). Deferred agonistic behavior
in a long-lived scincid lizard Eumeces laticeps. Field and
laboratory data on the roles of body size and residence in
agonistic strategy. Oecologia 72, 321–326.

Cooper, W.E., Vitt, L.J., Hedges, R. & Huey, R.B. (1990).
Locomotor impairment and defense in gravid lizards
(Eumeces laticeps): Behavioral shift in activity may offset
costs of reproduction in an active forager. Behav. Ecol.
Sociobiol. 27, 153–157.

Covas, R. (2012). Evolution of reproductive life histories in
island birds worldwide. Proc. Roy. Soc. B: Biol. Sci. 279,
1531–1537. https://doi.org/10.1098/rspb.2011.1785.

Cowie, R.H. & Holland, B.S. (2006). Dispersal is fundamental
to biogeography and the evolution of biodiversity on
oceanic islands. J. Biogeogr. 33, 193–198.

Cree, A. (1994). Low annual reproductive output in female
reptiles from New Zealand. New Zeal. J. Zool. 21, 351–372.

Derickson, W.K. (1976). Lipid storage and utilization in
reptiles. Am. Zool. 16, 711–723.

Doherty, T.S., Glen, A.S., Nimmo, D.G., Ritchie, E.G. &
Dickman, C.R. (2016). Invasive predators and global
biodiversity loss. PNAS 113, 11261–11265.

Escobar, R.A., Besier, E. & Hayes, W.K. (2010). Evaluating
headstarting as a management tool: Postrelease success of
green iguanas (Iguana iguana) in Costa Rica. J. Biodiver.
Conserv. 2, 204–214.

Ferguson, G.W., Brown, K.L. & DeMarco, V.G. (1982).
Selective basis for the evolution of variable egg and

8 Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London

Translocation success in lizards K. M. Hare et al.

https://doi.org/10.1098/rspb.2005.3147

https://doi.org/10.1098/rspb.2011.1785

hatchling size in some iguanid lizards. Herpetologica 38,
178–188.

Hardy, G.S. (1977). The New Zealand Scincidae (Reptilia:
Lacertilia); a taxonomic and zoogeographic study. New
Zeal. J. Zool. 4, 221–325.

Hare, K.M. & Miller, K.A. (2009). What dives beneath:
diving as a measure of performance in lizards.
Herpetologica 65, 227–236.

Hare, K.M., Daugherty, C.H. & Cree, A. (2002). Incubation
regime affects juvenile morphology and hatching success,
but not sex, of the oviparous lizard Oligosoma suteri
(Lacertilia: Scincidae). New Zeal. J. Zool. 29, 221–229.

Hare, K.M., Longson, C.G., Pledger, S. & Daugherty, C.H.
(2004). Size, growth, and survival are reduced at cool
incubation temperatures in the temperate lizard Oligosoma
suteri (lacertilia: Scincidae). Copeia 2, 382–390.

Hare, K.M., Daugherty, C.H. & Chapple, D.G. (2008a).
Comparative phylogeography of three skink species
(Oligosoma moco, O. smithi, O. suteri; Reptilia: Scincidae)
in northeastern New Zealand. Mol. Phylogenet. Evol. 46,
303–315.

Hare, K.M., Pledger, S. & Daugherty, C.H. (2008b). Low
incubation temperatures negatively influence locomotor
performance and behavior of the nocturnal lizard Oligosoma
suteri (Lacertidae: Scincidae). Copeia 2008, 16–22.

Hare, K.M., Norbury, G., Judd, L.M. & Cree, A. (2012).
Survival of captive-bred skinks following reintroduction to
the wild is not explained by variation in speed or body
condition index. New Zeal. J. Zool. 39, 319–328.

Hare, K.M., Chapple, D.G., Towns, D.R. & van Winkel, D.
(2016). The ecology of New Zealand’s lizards. In New
Zealand lizards: 133–168. Chapple, D.G. (Ed). Switzerland:
Springer.

Hare, K.M., Borrelle, S.B., Buckley, H.L., Collier, K.J.,
Constantine, R., Perrott, J.K., Watts, C.H. & Towns, D.R.
(2019). Intractable: Species in New Zealand that continue to
decline despite conservation efforts. J. Roy. Soc. New Zeal.
49, 301–319. https://doi.org/10.1080/03036758.2019.
1599967.

Hasegawa, M. (1990). The thrush Trudus celaenops, as an
avian predator of juvenile Eumeces okadae on Miyake-Jima,
Izu islands. Jpn. J. Herpetol. 13, 65–69.

Herrel, A., Spithoven, L., Van Damme, R. & De Vree, F.
(1999). Sexual dimorphism of head size in Gallotia galloti:
Testing the niche divergence hypothesis by functional
analyses. Funct. Ecol. 13, 289–297.

Herrel, A., De Grauw, E. & Lemos-Espinal, J.A. (2001). Head
shape and bite performance in xenosaurid lizards. J. Exp.
Zool. 290, 101–107.

Hitchmough, R.A., Adams, L.K., Reardon, J.T. & Monk, J.M.
(2016a). Current challenges and the future directions in
lizard conservation in New Zealand. J. R. Soc. N. Z. 46,
29–39.

Hitchmough, R.A., Patterson, G.B. & Chapple, D.G. (2016).
Putting a name to diversity: taxonomy of the New Zealand

lizard fauna. In New Zealand Lizards: 87–108. Chapple,
D.G. (Ed.). Switzerland: Springer.

Hoskins, A.J., Hare, K.M., Miller, K.A., Schumann, N. &
Chapple, D.G. (2017). Repeatability, locomotor performance
and trade-offs between performance traits in two lizard
species, Oligosoma alani and O. Smithi. Biol. J. Lin. Soc.
122, 850–859.

Huey, R.B., Schneider, W., Erie, G.L. & Stevenson, R.D.
(1981). A field-portable racetrack and timer for measuring
acceleration and speed of small cursorial animals.
Experientia 37, 1356–1357.

Husak, J.F. (2006). Does survival depend on how fast you
can run or how fast you do run? Funct. Ecol. 20, 1080–
1086.

Irschick, D.J. (2003). Measuring performance in nature:
Implications for studies of fitness within populations. Integr.
Comp. Biol. 43, 396–407.

Irschick, D.J. & Garland, T. Jr (2001). Integrating function
and ecology in studies of adaptation: investigations of
locomotor capacity as a model system. Annu. Rev. Ecol.
Syst. 32, 367–396.

IUCN (2018). The IUCN red list of threatened species.
Version 2018–2.

Jarvie, S., Senior, A.M., Adolph, S.C., Seddon, S.C. &
Cree, A. (2015). Captive rearing affects growth but not
survival in translocated juvenile tuatara. J. Zool. 297,
184–197.

Maran, T., P~odra, M., P~olma, M.O. & Macdonald, D.W.
(2009). The survival of captive-born animals in restoration
programmes – case study of the endangered European mink
Mustela lutreola. Biol. Conserv. 142, 1685–1692.

Miller, K.A., Hare, K.M. & Nelson, N.J. (2010). Do alternate
escape tactics provide a means of compensation for
impaired performance ability? Biol. J. Lin. Soc. 99, 241–
249.

Miller, K.A., Towns, D.R., Allendorf, F.W., Ritchie, P.A. &
Nelson, N.J. (2011). Genetic structure and individual
performance following a recent founding event in a small
lizard. Conserv. Genet. 12, 461–473.

Miller, K.A., Bell, T.P. & Germano, J.M. (2014).
Understanding publication bias in reintroduction biology by
assessing translocations of New Zealand’s herpetofauna.
Conserv. Biol. 28, 1045–1056.

Myers, N., Mittermeier, R.A., Mittermeier, C.G., da Fonseca,
G.A.B. & Kent, J. (2000). Biodiversity hotspots for
conservation priorities. Nature 403, 853–858.

Noble, D.W.A., Stenhouse, V. & Schwanz, L.E. (2018).
Developmental temperatures and phenotypic plasticity in
reptiles: a systematic review and meta-analysis. Biol. Rev.
93, 72–97.

Plummer, M. (2003). JAGS: a program for analysis of
Bayesian graphical models using Gibbs sampling. In
Proceedings of the 3rd International Workshop on
Distributed Statistical Computing (DSC 2003)). Vienna,
Austria.

Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London 9

K. M. Hare et al. Translocation success in lizards

https://doi.org/10.1080/03036758.2019.1599967

R Core Team (2017). R: a language and environment for
statistical computing. Vienna, Austria: R Foundation for
Statistical Computing.

Romijn, R.L. & Hartley, S. (2016). Trends in lizard
translocations in New Zealand between 1988 and 2013.
New Zeal. J. Zool. 43, 191–210.

Sacchi, R., Pupin, F., Gentilli, A., Rubolini, D., Scali, S.,
Fasola, M. & Galeotti, P. (2009). Male-male combats in a
polymorphic lizard: residence and size, but not color, affect
fighting rules and contest outcome. Aggress. Behav. 35,
274–283.

Seddon, P.J., Armstrong, D.P. & Maloney, R.F. (2007).
Developing the science of reintroduction biology. Conserv.
Biol. 21, 303–312.

Shine, R. (1980). “Costs” of reproduction in reptiles.
Oecologia 46, 92–100.

Shine, R. & Harlow, P. (1993). Maternal thermoregulation
influences offspring viability in a viviparous lizard.
Oecologia 96, 122–127.

Snyder, N.F.R., Derrickson, S.R., Beissinger, S.R., Wiley,
J.W., Smith, T.B., Toone, W.D. & Miller, B. (1996).
Limitations of captive breeding in endangered species
recovery. Conserv. Biol. 10, 338–348.

Stamps, J.A. (1977). The relationship between resource
competition, risk, and aggression in a tropical territorial
lizard. Ecology 58, 349–358.

Stolzenburg, W. (2011). Rat island: predators in paradise and
the world’s greatest wildlife rescue. Sydney: Bloomsbury.

Su, Y.-S. & Yajima, M. (2015). R2jags: using r to run
‘JAGS’. R package version 0.5-7.

Tingley, R., Hitchmough, R.A. & Chapple, D.G. (2013). Life
history traits and extrinsic threats determine extinction risk
in New Zealand Lizards. Biol. Conserv. 165, 62–68.

Tokarz, R.R. (1985). Body size as a factor determining
dominance in staged agonistic encounters between
male brown anoles (Anolis sagrei). Anim. Behav. 33,
746–753.

Towns, D.R. (1975). Reproduction and growth of the black
shore skink, Leiolopisma suteri (lacertilia: Scincidae), in
north-eastern New Zealand. New Zeal. J. Zool. 2, 409–423.

Towns, D.R. & Ferreira, S.M. (2001). Conservation of New
Zealand lizards (Lacertilia: Scincidae) by translocation of
small populations. Biol. Conserv. 98, 211–222.

Towns, D.R., Parrish, G.R. & Westbrooke, I. (2003). Inferring
vulnerability to introduced predators without experimental
demonstration: case study of Suter’s skink in New Zealand.
Conserv. Biol. 17, 1361–1371.

Towns, D.R., Atkinson, I.A.E. & Daugherty, C.H. (2006).
Have the harmful effects of introduced rats on islands been
exaggerated? Biol. Invasions 8, 863–891.

Towns, D.R., Miller, K.A., Nelson, N.J. & Chapple, D.G.
(2016a). Can translocations to islands reduce extinction risk
for reptiles? Case studies from New Zealand. Biol. Conserv.
204, 120–127.

Towns, D.R., Hitchmough, R.A. & Perrott, J. (2016b).
Conservation of New Zealand lizards: a fauna not forgotten
but undervalued? In New Zealand lizards: 293–320.
Chapple, D.G. (Ed.). Switzerland: Springer.

Verwaijen, D., Van Damme, R. & Herrel, A. (2002).
Relationships between head size, bite force, prey handling
efficiency and diet in two sympatric lacertid lizards. Funct.
Ecol. 16, 842–850.

Wapstra, E. (2000). Maternal basking opportunity affects
juvenile phenotype in a viviparous lizard. Funct. Ecol. 14,
345–352.

Webb, J.K. & Shine, R. (1993). Prey-size selection, gape
limitation and predator vulnerability in Australian
blindsnakes (Typhlophidae). Anim. Behav. 45, 1117–1126.

Whittaker, R.J. & Fern�andez-Palacios, J.M. (2007). Island
biogeography: ecology, evolution and conservation. 2nd
edn. Oxford: Oxford University Press.

van Winkel, D., Baling, M. & Hitchmough, R. (2018).
Reptiles and amphibians of New Zealand: a field guide.
Auckland: Auckland University Press.

10 Animal Conservation �� (2019) ��–�� ª 2019 The Zoological Society of London

Translocation success in lizards K. M. Hare et al.

Contributed Paper

Effects of ecotourism on forest loss in the Himalayan
biodiversity hotspot based on counterfactual analyses

Jodi S. Brandt ,1 ∗ Volker Radeloff,2 Teri Allendorf,2 Van Butsic,3 and Anand Roopsind1
1Human-Environment Systems Center, Boise State University, 1910 University Drive, Boise, ID 83725, U.S.A.
2Forest and Wildlife Ecology, University of Wisconsin-Madison, 1630 Linden Drive, Madison, WI 53706, U.S.A.
3Department of Environmental Science, Policy and Management, University of California, Berkeley, Mulford Hall, 130 Hilgard Way,
Berkeley, CA 94720, U.S.A.

Abstract: Ecotourism is developing rapidly in biodiversity hotspots worldwide, but there is limited and mixed
empirical evidence that ecotourism achieves positive biodiversity outcomes. We assessed whether ecotourism
influenced forest loss rates and trajectories from 2000 to 2017 in Himalayan temperate forests. We compared forest
loss in 15 ecotourism hubs with nonecotourism areas in 4 Himalayan countries. We used matching statistics to
control for local-level determinants of forest loss, for example, population density, market access, and topography.
None of the ecotourism hubs was free of forest loss, and we found limited evidence that forest-loss trajectories in
ecotourism hubs were different from those in nonecotourism areas. In Nepal and Bhutan, differences in forest loss
rates between ecotourism hubs and matched nonecotourism areas did not differ significantly, and the magnitude
of the estimated effect was small. In India, where overall forest loss rates were the lowest of any country in
our analysis, forest loss rates were higher in ecotourism hubs than in matched nonecotourism areas. In contrast,
in China, where overall forest loss rates were highest, forest loss rates were lower in ecotourism hubs than
where there was no ecotourism. Our results suggest that the success of ecotourism as a forest conservation
strategy, as it is currently practiced in the Himalaya, is context dependent. In a region with high deforestation
pressures, ecotourism may be a relatively environmentally friendly form of economic development relative to
other development strategies. However, ecotourism may stimulate forest loss in regions where deforestation rates
are low.

Keywords: community-based forestry, environmental policy, Mahalanobis matching, quasi-experimental, sus-
tainable development

Efectos del Ecoturismo sobre la Pérdida de Bosques en el Punto Caliente de Biodiversidad en el Himalaya con base
en Análisis Contrafactuales

Resumen: El ecoturismo está desarrollándose rápidamente en los puntos calientes de biodiversidad en todo el
mundo, pero existe evidencia emṕırica mixta y limitada de los resultados positivos que se logran con el ecoturismo.
Valoramos si el ecoturismo influyó sobre las tasas de pérdida forestal y sus trayectorias entre el 2000 y el 2017
en los bosques templados del Himalaya. Comparamos la pérdida forestal en quince focos ecotuŕısticos con la
pérdida forestal en las áreas sin ecoturismo de cuatro páıses del Himalaya. Utilizamos estad́ıstica correspondiente
para controlar las determinantes a nivel local de la pérdida del bosque, por ejemplo, la densidad poblacional, el
acceso al mercado y la topograf́ıa. Ninguno de los focos ecotuŕısticos estaba libre de pérdida forestal, además
de que encontramos evidencia limitada de que las trayectorias de la pérdida forestal en los focos ecotuŕısticos
eran diferentes a las trayectorias en las áreas sin ecoturismo. En Nepal y en Bután, las diferencias en la pérdida
forestal entre los focos ecotuŕısticos y las áreas sin ecoturismo correspondidas no difirieron significativamente y la
magnitud del efecto estimado fue menor. En la India, donde las tasas generales de pérdida forestal fueron las más
bajas de cualquier páıs en nuestro análisis, las tasas de pérdida forestal fueron más altas en los focos ecotuŕısticos

∗email jodibrandt@boisestate.edu
Article impact statement: Whether ecotourism protects forests is context dependent, and an important factor is the deforestation pressure
overall on the landscape.
Paper submitted December 13, 2018; revised manuscript accepted March 13, 2019.

1318
Conservation Biology, Volume 33, No. 6, 1318–1328
C© 2019 Society for Conservation Biology
DOI: 10.1111/cobi.13341

https://orcid.org/0000-0002-1954-5997

Brandt et al. 1319

que en las áreas sin ecoturismo correspondidas. Como contraste, en China, donde las tasas generales de pérdida
forestal fueron más altas, las tasas de pérdida forestal fueron más bajas en los focos ecotuŕısticos que en donde no
existe el ecoturismo. Nuestros resultados sugieren que el éxito del ecoturismo como estrategia de conservación
del bosque, a como se práctica actualmente en el Himalaya, depende del contexto. En una región con presiones
altas de deforestación, el ecoturismo puede ser una forma de desarrollo económico relativamente amigable con el
ambiente comparado con otras estrategias de desarrollo. Sin embargo, el ecoturismo puede estimular la pérdida
forestal en regiones en las que las tasas de deforestación son bajas.

Palabras Clave: correspondencia Mahalanobis, cuasiexperimental, desarrollo sustentable, poĺıtica ambiental,
silvicultura basada en la comunidad

��: ��, ��, ��
�� 2000–2017 ��
��
��, � ��, ��
��,
��, ��
��
, ��, ��
��, ��, ��
��, ��, ��
��, ��, ��
��
��, ��
�� , ��
��, ��
��: ��; ��: ��

��: ��, ��, ��,
��, ��

Introduction

Ecotourism is proliferating in biodiversity hotspots, and
its proponents claim it can achieve conservation and
economic development goals. Ecotourism has become
a major driver of economic growth and socioeconomic
transformation in many areas. The amount spent on eco-
tourism is estimated to be 10 times more than that spent
by official aid agencies and the UN Global Environment
Facility on conservation projects (Kirkby et al. 2011; Wal-
dron et al. 2017). Ecotourism accounts for as much as
40% of gross domestic product (GDP) in some countries
and is growing 10% per year in other countries (WTTC
2014). Despite this major investment, there is limited
empirical evidence that ecotourism achieves biodiversity
conservation goals in the long term and at the landscape
scale.

It is difficult to ascertain whether ecotourism actually
achieves biodiversity goals. In developing regions, eco-
tourism and its cumulative effects on biodiversity are
unclear. Ecotourism may generate the same or more
income in an area than the consumption of natural re-
sources (Kirkby et al. 2010, 2011). Thus, ecotourism
can provide an economic incentive to protect ecosys-
tems and species tourists visit. For example, governments
may establish protected areas or enforce wildlife protec-
tion to ensure revenue from international tourists (Buck-
ley 2011). Similarly, community ecotourism projects
in unprotected landscapes may dedicate a portion of
ecotourism proceeds into conservation efforts to pro-
tect their natural assets (Nagendra et al. 2005; Buck-

ley 2009; Wyman & Stein 2010). A major justification
for ecotourism investments by developed countries is
the assertion that ecotourism may lead to biodiversity
conservation because it provides economic rewards for
doing so.

In contrast, ecotourism may lead to biodiversity loss
because it can require or encourages economic develop-
ment, which often entails strong, negative environmen-
tal outcomes (Mather et al. 1999). Ecotourism usually
requires improved transportation networks (e.g., roads
and airports), which can result in intensive natural re-
source exploitation, such as logging and poaching, be-
cause of increased accessibility to the area (Laurance
et al. 2014; MoCTCA 2015b; Shui & Xu 2016). Increased
local wealth can change residents’ consumption pat-
terns, adding pressure on local forest resources (Liu et al.
2001; Brandt et al. 2012). Tourism also stimulates popu-
lation growth, in the form of seasonal tourists and eco-
nomic immigrants, which can raise demand for forest re-
sources (Hall & Lew 2009). Tourists essentially represent
a form of population growth, and they typically consume
more resources per capita than local residents (Buckley
2011). Finally, tourism inherently leads to an integration
of local and regional markets, another factor strongly
associated with increased resource extraction (Hall &
Lew 2009; Wang & Buckley 2010; Lambin & Meyfroidt
2011).

Incentives for protection and economic development
thus may either result in positive or negative biodiversity

Conservation Biology
Volume 33, No. 6, 2019

1320 Himalayan Ecotourism

Figure 1. Ecotourism hubs across 4 countries in the
Himalayan temperate forest zone (as defined by
Olson et al. 2001) and with forest-cover data from
Hansen et al. (2013).

outcomes. However, even if the net result of ecotourism
is negative, it may still be beneficial as long as eco-
tourism leads to less biodiversity loss than would have
occurred if an alternative economic development strat-
egy had been implemented instead (i.e., ecotourism may
not completely stop biodiversity loss, but it may be better
than alternatives). Developing nations typically rely on
extraction-based land uses for economic development,
including the production of raw goods (e.g., mining tim-
ber) or the conversion of natural ecosystems to more
economically productive uses (e.g., agriculture). There-
fore, even if ecotourism stimulates economic develop-
ment that leads to environmental degradation, it may
lead to less biodiversity loss than more extractive models,
such as palm oil production. Quantifying whether or not
that is the case requires the application of a counterfac-
tual approach to estimate rates of negative biodiversity
outcomes had there not been ecotourism (Andam et al.
2008; Butsic et al. 2017).

We sought to measure environmental outcomes of
ecotourism by comparing ecotourism hubs with other
areas in which other development strategies had been
implemented. Specifically, we measured the association
between ecotourism and forest loss in the Himalayan
temperate forest zone (Fig. 1) with a counterfactual ap-
proach. We identified 15 ecotourism hubs (i.e., areas
where ecotourism is the primary strategy for economic
development) across 4 countries with diverse tourism
strategies: India, Nepal, Bhutan, and China. Our objec-
tives were to characterize the type of ecotourism strat-
egy implemented in each country; identify whether eco-
tourism hubs have rates and trajectories of forest loss
distinct from nonecotourism areas; and quantify differ-
ences in forest loss rates between ecotourism hubs and
nonecotourism areas.

Methods

Study Area

The Himalayan temperate forest zone extends 3000 km
from southern Afghanistan to southwest China (Olson
et al. 2001). It contains 2 of Earth’s biodiversity hotspots
(Myers et al. 2000), an extraordinary array of ecological
niches in a relatively small area, and globally the highest
fractions of endemic and threatened species in the world
(Grenyer et al. 2006). Himalayan temperate forests have
been used for thousands of years to support subsistence-
based livelihoods. Forests are the primary source of fuel
for cooking, heating, and construction; are intensively
used for livestock grazing, hunting, food gathering, and
traditional medicines; and provide raw materials for eco-
nomic development. These same forests contain highly
threatened, endemic biodiversity and provide essential
ecosytem services, including climate and water-cycle reg-
ulation. Since the 1980s, demand for timber and fuel-
wood increased, resulting in forest loss and degradation
(Pandit et al. 2014). Even though forest protection is a
primary conservation target across Himalayan countries,
forest loss has continued (Brandt et al. 2017).

The Himalayan region provides opportunity for a nat-
ural experiment to investigate ecotourism impacts be-
cause it contains countries in very different stages of
economic and tourism development. Tourism has prolif-
erated across the Himalaya as a way to balance economic
development and forest conservation (Pandit et al. 2014).

We analyzed 4 Himalayan countries with active eco-
tourism industries: India, Nepal, Bhutan, and China.
Nepal and Bhutan are relatively small and located primar-
ily in the Himalaya and its foothills. India and China are
mostly outside the Himalayas and have regional adminis-
trative units with distinct policies and contexts. Because
of their large size, we focused on single administrative
units located primarily in the Himalaya: Himachal Pradesh
State, India, and Yunnan Province, China.

Forest Change Data

We defined forest loss as stand-replacing disturbance or a
change from forest to nonforest. Forest cover and change
data for 2000–2017 were derived from a publicly available
data set of global forest dynamics (Hansen et al. 2013).
The forest-cover data set contains canopy cover of each
30-m pixel in the baseline year of 2000. Each pixel is
classified from 0 to 100 (0, no canopy; 100, 100% canopy
cover). We considered a pixel with >50% canopy cover
as forested.

Site Identification and Characterization of Ecotourism
Strategies

To identify ecotourism sites and characterize ecotourism
strategies and contexts, including the number, origin, and

Conservation Biology
Volume 33, No. 6, 2019

Brandt et al. 1321

Figure 2. Ecotourism hubs with 15-, 25-, 35-, 45-, and
55-km circular zones surrounding each hub. The zone
with the boldest outline is the 35-km zone, which was
used for the counterfactual analysis. All areas within
the 35-km circular zone are ecotourism hubs, and all
areas outside this zone were nonecotourism areas.

purpose of tourists and economic wealth, we reviewed
the peer-reviewed and gray literature on ecotourism for
each country (Supporting Information). We compiled
statistics when possible, but comparable and consistent
statistics across sites were typically not available; thus,
we relied on province-, state-, or country-level data. We
obtained information about the overarching forest gover-
nance strategy and forest change in each country (Brandt
et al. 2017). We also used information from studies pub-
lished in the literature about tourism impacts on forests
(see Brandt & Buckley 2018).

Tourism areas, unlike protected areas or administra-
tive units, do not have delineated boundaries. However,
tourism tends to be concentrated, and tourists typically
visit an ecotourism hub (i.e., a population center where
tourists concentrate for accommodation, food, guides,
and other amenities). To identify our ecotourism hubs,
we compiled a list of the most popular general tourism
hubs based on official tourism statistics and other litera-
ture reviewed for each country; searched Google (search
terms such as ecotourism in Bhutan and ecotourism
in Yunnan) to identify hubs that advertised ecotourism;
narrowed the list by asking regional experts to identify
ecotourism hubs they considered the most popular; and
overlaid our map of potential hubs on an ecoregion map
(Olson et al. 2001) to identify sites that included forest
in the Himalayan temperate zone (Fig. 1). Our selection
of ecotourism sites was not designed to be a represen-
tative sample of all ecotourism hubs; rather, it was an
attempt to select the most important ones. We identi-
fied 15 ecotourism hubs across 4 administrative units:
Himachal Pradesh (n = 4), Nepal (n = 5), Bhutan (n =
3), and Yunnan (n = 3).

To determine the appropriate spatial boundary, we
demarcated circular zones with 15-, 25-, 35-, 45-, and
55-km radii surrounding each ecotourism hub (Fig. 2).
We summarized forest loss rates in these zones and their

respective nonecotourism areas (Fig. 3). For example,
the 15-km ecotourism zone represented all forests 0–15
km from the hub, and the corresponding nonecotourism
areas included all forests within the same country that
were not within 0–15 km of any ecotourism hub in that
country. When calculating deforestation at the hub level,
forests included in >1 hub boundary were attributed to
both hubs. When calculating deforestation at the country
level, forests in areas of overlapping boundaries were
included only once in the analysis to avoid overestimat-
ing deforestation rates at the aggregate scale. In all 4
countries, forest-loss rates in ecotourism and nontourism
zones came close to convergence by the 35-km bound-
ary and fully converged by the 55-km boundary. Thus,
we used the 35-km zone for subsequent analyses. We
designated areas within the 35-km boundaries as eco-
tourism hubs and areas beyond the 35-km boundaries as
nonecotourism areas.

Trajectories of Forest Loss Area

We plotted trajectories of the area of forest loss for
all forests in each ecotourism hub (35 km) and for all
forests in the nonecotourism areas of each country. We
calculated trend lines showing 2-year moving averages to
smooth out errors in the annual forest loss measurements
due to cloud compositing during remote-sensing analyses
(Hansen et al. 2013). We also fitted a linear trend line
for the entire trajectory to visualize whether each hub
increased, decreased, or was stable from 2000 to 2017.

Counterfactual Analysis of Forest Loss Rate

An increasingly common counterfactual approach to de-
termine the impact of conservation policies is quasi-
experimental counterfactual matching analysis (Andam
et al. 2008). Matching has been used, for example, to
assess the effectiveness of certification policies (Miteva
et al. 2015), national forest management regimes (Brandt
et al. 2017), protected areas (e.g., Nolte et al. 2013),
community forests (e.g., Brandt et al. 2015), and logging
concessions (e.g., Brandt et al. 2016). To our knowledge,
matching has not been used in the context of ecotourism.

Ecotourism hubs are typically located in remote places
that have relatively few people, are less accessible, and
have retained more forests than the country as a whole.
The goal of our matching analysis was to find areas with
similar population density, accessibility, and forest cover
that only differed in whether they were associated with a
major ecotourism hub or not. To do so, we matched treat-
ment units (e.g., forested cells influenced by ecotourism)
with control units (e.g., forest cells not influenced by eco-
tourism). With matched samples, it is possible to predict
what outcomes would have been observed in forests with
ecotourism had they not been subjected to ecotourism
(Abadie & Imbens 2006). In essence, we asked: What

Conservation Biology
Volume 33, No. 6, 2019

1322 Himalayan Ecotourism

Figure 3. Forest-loss rates (2000–2017) within 15, 25, 35, 45, and 55 km of ecotourism hubs and nonecotourism
areas for each country.

would be the rate of forest loss in ecotourism zones of
country A if ecotourism had not been adopted and if that
area had been developed like other nonecotourism areas?
We also performed the opposite comparison by asking:
What would be the rate of forest loss in nontourism zones
of country A if ecotourism had been adopted?

We performed pair-wise comparisons of the forest-loss
rate in ecotourism hubs and nonecotourism areas in each
country by applying Mahalanobis matching with replace-
ment and bias adjustment (Sekhon 2011). We aggregated
annual forest cover and forest change data into 1-km cells
to achieve a sample size that was computationally feasible
and consistent with similar analyses (Ferraro et al. 2013;
Nolte et al. 2013; Brandt et al. 2017). We calculated an
adjusted forest-loss rate, which is the total area of forest
loss divided by the forested area in that cell in 2000. Cells
that did not have any forest were excluded from the
analyses. We matched treatment and control units based
on 7 covariates (i.e., factors that influence forest loss):
distance to market, population density, slope, elevation,
precipitation, temperature, and percent forest cover in
2000. For each pairwise comparison, we randomly sam-
pled 20% of the treatment parcels and matched them
with control parcels. To determine the validity of the
matches, we calculated balance statistics, which indicate

the extent to which the pool of potential controls con-
tains units that are sufficiently comparable to treatment
units. We dropped treatment parcels for which no com-
parable control parcel could be found within 0.5 SD
of each covariate. To compute the reverse estimate,
we switched control and treatment group (i.e., assigned
nonecotourism areas as the treatment and found matches
from the ecotourism hubs). We repeated this procedure
for pairwise comparisons in each country for a total of 8
different pairwise comparisons. Matched treatment and
control units were always from the same country. See
Supporting Information for the full results of the match-
ing analysis and balance statistics.

Results

Ecotourism Strategies

The 4 countries we analyzed differed considerably in
terms of the types of ecotourism they implemented (Sup-
porting Information). India is one of the most populous
and rapidly developing countries in the world. Himachal
Pradesh, the focus of our analysis, had the second-lowest
GDP/capita of any unit in our analysis (US$2200), and the

Conservation Biology
Volume 33, No. 6, 2019

Brandt et al. 1323

lowest overall forest-loss rate from 2000–2017 (0.40%).
Himachal Pradesh has a unique forest governance sys-
tem of traditional, community-based forestry that is of-
ficially recognized by the federal government (Brandt
et al. 2017). Tourism in Himachal Pradesh started in
the colonial era in the form of seasonal vacation centers
(i.e., Hill Stations), created by the British during the 19th
century (Ahluwalia & Little 1998). Since the 1970s, the
state government has implemented policies to encourage
the development of both corporate and leisure tourism
(Pandey & Wells 1997), and the number of tourist visi-
tors grew from 8 million in 2006 to 15 million in 2011
(KPMG 2012), the vast majority of which were domes-
tic (ACNielsen 2012). Nature-based, adventure, and reli-
gious tourism dominated (Singh 2002; Donovan 2013).
We found 2 empirical articles about ecotourism and de-
forestation in the Indian Himalaya, both of which link
ecotourism to deforestation and forest degradation be-
cause increased demand for timber and fuelwood was
sourced from local forests (Singh et al. 2009; Mahapatra
et al. 2012).

Nepal is among the poorest countries in the world
with primarily community-based sustainable forest man-
agement (Brandt et al. 2017). Nepal had the lowest
GDP/capita (US$835) and the second-highest rate of tem-
perate forest loss (1.0%) of our study units. Nepal has
been a popular international ecotourism destination since
the 1970s, and tourism has been important for economic
development (Schroeder & Sproule-Jones 2012). Nepal’s
tourism policy is designed to maximize the number of
tourists and offers relatively inexpensive visas and few
restrictions on travel and the length of time tourists may
stay in the country (Schroeder & Sproule-Jones 2012). In
2014, Nepal hosted over 790,000 tourists, a 58% increase
from approximately 500,000 in 2000 (MoCTCA 2015a,
2015b), and has a national goal of 2 million tourists
per year by 2020 (Nepal & Karst 2017). The majority
of tourists in Nepal are international and come to visit
national parks and to trek (MoCTCA 2015b). Similar
to Himachal Pradesh, we found 2 empirical articles for
the Nepal Himalaya, both of which report more local
deforestation and forest degradation due to ecotourism
because of increased demand for timber and fuelwood
(Stevens 2003; Garrard et al. 2016).

Bhutan is a small Buddhist kingdom, known for its gross
domestic happiness (GDH) policy, where environmen-
tal protection and economic growth are equally priori-
tized, a sustainable development approach that is unique
among developing nations (Brooks, 2010, 2013). Bhutan
had the second highest GDP/capita of any unit in our
study (US$3110) and the second lowest forest-loss rate
(0.9%). Bhutan’s national forest policy emphasizes forest
conservation (Brandt et al. 2017), and tourism in Bhutan is
a relatively recent phenomenon compared with India and
Nepal. Tourism was introduced as a means of attaining
foreign currency to help achieve economic development

and autonomy from donor aid (Nepal & Karst 2017), but,
in contrast to Nepal, Bhutan has pursued a controlled
approach. In 1974, Bhutan implemented a policy known
as ‘‘high value, low volume’’ (Schroeder & Sproule-Jones
2012). Tourist visas are expensive and short, and travel
permissions are tightly controlled, which limits both
tourism numbers and the activities tourists can engage
in. Although there is great potential for adventure-based
tourism, such as climbing and trekking, it is limited due
to the tight control (Gurung & Seeland 2008), and the
primary activity of most tourists is “cultural sight-seeing”
on designated tours to specific sites (Bhutan 2014;
TCB 2015). In 2014, Bhutan hosted 133,480 tourists,
about one-fifth as many as Nepal (Bhutan 2014). We
found no empirical studies about ecotourism and forests
in Bhutan.

China has had the fastest growing economy in the
world in recent decades. Strong economic development
policies for western China have stimulated high rates
of economic growth in Yunnan (Xu et al. 2006). Yun-
nan’s forest governance policy emphasizes for-profit use
of forests (Brandt et al. 2017). Yunnan had the highest
GDP/capita (US$5117) and the highest forest-loss rate
(2.9%) of all of our study units. The most common eco-
nomic development strategies include extractive-based
activities, including cash crops, mining, and hydropower,
except for specific areas where ecotourism has been im-
plemented (Li & Han 2000; Donaldson 2007; Wang &
Buckley 2010). The Himalayan region of Yunnan is des-
ignated as the premiere ecotourism destination in China
and aggressively marketed to the growing middle class
in eastern China (Nyaupane et al. 2006). Ecotourism has
grown exponentially since 1990 and the vast majority of
tourists are domestic (Jenkins 2009; Brandt et al. 2012;
HKTDC 2017). For example, tourist visitors in Diqing
Prefecture grew from 40,000 tourists in 1995 to 5.3 mil-
lion visitors in 2009. We found 2 empirical studies from
the Chinese Himalaya about the impacts of tourism on
forests, both of which reported that ecotourism led to
accelerated deforestation due to rapid economic devel-
opment and population growth (Liu et al. 2001; Brandt
et al. 2012).

Rates and Trajectories of Forest Loss in Ecotourism Hubs
and Nonecotourism Areas

In simple comparisons, we did not find that forest loss dif-
fered clearly between ecotourism hubs and nontourism
areas, but there were differences among countries
(Fig. 4). India had the lowest forest-loss rates, ranging
from 0.4% in Shimla and nonecotourism areas to 1.1% in
Manali. The highest hub-level forest-loss rates occurred in
China, ranging from 2.0% in Tacheng to 4.3% in Lijiang;
nonecotourism areas had an intermediate forest-loss rate
of 2.9%. Bhutan and Nepal’s forest-loss rates were be-
tween those in India and China, and forest-loss rates in

Conservation Biology
Volume 33, No. 6, 2019

1324 Himalayan Ecotourism

Figure 4. Forest-loss rates in
each ecotourism hub and
nonecotourism areas, and
the average of all
ecotourism hubs in each
country.

nonecotourism areas were within the range of individual
ecotourism hubs in the respective countries.

In terms of annual forest-loss trajectories, nontourism
areas and ecotourism hubs in India all had a decreasing
forest-loss trend from 2000 to 2017 (Fig. 5). Dalhousie
had the highest forest loss of all sites and years (1.5 km2

in 2002) but had no forest loss in the most recent years
(2014–2017). Other Indian hubs and the nontourism ar-
eas also showed high interannual variability; there was
no specific temporal pattern other than low or 0 forest
loss after 2013. In Nepal, nontourism areas showed a
stable to slightly increasing trend of forest loss, whereas
4 of 5 ecotourism hubs had a decreasing forest-loss tra-
jectory. Bhutan displayed a unique temporal pattern that
was consistent across the country; a spike in forest loss
occurred in all 3 hubs and the nontourism areas in 2010
and rates increased steadily from 2012 to 2017. China’s
nonecotourism areas, similar to Nepal’s, showed a slight
increase in forest loss over the entire period, and the
ecotourism hubs varied greatly: 1 (Lijiang) increased, 1
(Shangrila) decreased, and 1 (Tacheng) was stable.

Differences in Forest-Loss Rates in Ecotourism Hubs and
Nonecotourism Areas Based on Counterfactuals

According to our counterfactual analysis, forest-loss rates
in Nepal and Bhutan in ecotourism hubs were not signif-
icantly different (p < 0.05 threshold) relative to noneco- tourism areas (Fig. 6 & Supporting Information). The difference in loss rates between ecotourism hubs and nonecotourism areas was +0.08% in Nepal and +0.12% in Bhutan. In India, ecotourism hubs had higher forest-loss

rates than matched nonecotourism areas; effect size was
small but significant (+0.40%, p < 0.001). Similarly, when we matched cells in nonecotourism areas with those in ecotourism hubs, nonecotourism areas had less forest loss than ecotourism hubs; effect size was −0.74% (p < 0.001). China was the only country where ecotourism hubs had lower forest-loss rates than matched noneco- tourism areas (effect size of −1.73%, p < 0.001). The inverse comparison was also significant; cells in noneco- tourism areas matched with those in ecotourism hubs had +0.77% more forest loss (p < 0.001).

Discussion

We found little evidence that ecotourism reduces rates
of forest loss, but also little evidence that ecotourism
leads to higher forest-loss rates due to more rapid devel-
opment. At first glance, our results seemed to suggest
that ecotourism spurs forest loss because forest-loss rates
themselves were in most cases higher in ecotourism hubs
than in nonecotourism areas (i.e., Fig. 3). However, sim-
ple comparisons of forest-loss rates are only valid when
ensuring that sites with similar deforestation pressure are
compared. Indeed, when we controlled for deforestation
pressure, we found that the effects of ecotourism varied
among countries. We were surprised by our results be-
cause previous empirical studies in both India (Singh et al.
2009; Mahapatra et al. 2012) and Nepal (Stevens 2003;
Garrard et al. 2016) found more deforestation in eco-
tourism hubs. However, none of these prior studies used
counterfactual approaches to control for deforestation

Conservation Biology
Volume 33, No. 6, 2019

Brandt et al. 1325

Figure 5. Annual forest loss, 2-year moving averages, and linear trend-lines from 2000–2017 in each ecotourism
hub and nonecotourism areas in each country.

pressure, and their findings may reflect that ecotourism
hubs are often located in places that inherently have
higher deforestation pressure. Similarly, we expected
that Bhutan’s strategy of a tightly controlled ecotourism
industry would result in better forest conservation out-
comes compared with Nepal, which hosts many more
tourists and exerts less control on where and how they
travel (Schroeder & Sproule-Jones 2012). However, we
found no evidence that ecotourism in Bhutan reduced
deforestation pressures or differed in its impact from
ecotourism in Nepal.

Yunnan was the only study unit where ecotourism
hubs had lower forest-loss rates than nontourism areas,
and even when comparing areas with similar deforesta-
tion pressure, the effect size was considerable (1.73%).

This result surprised us because prior case studies re-
ported empirical evidence of accelerated deforestation
after ecotourism started in the Chinese Himalaya (Liu
et al. 2001; Brandt et al. 2012). However, our matching
results suggest that in the context of rapid development
in China as a whole, and in Yunnan in particular, eco-
tourism led to less forest loss than areas where tourism
was not prominent. It is important to note that while
forest-loss rates in Yunnan’s ecotourism zones were less
than that in China’s nonecotourism areas, they were
still 2 to 3 times higher than forest-loss rates in eco-
tourism hubs of Bhutan, Nepal, and India. Overall, forest-
loss rates in Yunnan were very high due to China’s
national-level forest management policy that encourages
for-profit use of forests instead of the sustainable use or

Conservation Biology
Volume 33, No. 6, 2019

1326 Himalayan Ecotourism

Figure 6. Matched comparisons of forest-loss rates between ecotourism hubs and nonecotourism areas in each
country (black, ecotourism hubs with higher forest-loss rates than nonecotourism areas; gray, ecotourism hubs
with lower forest-loss rates than nonecotourism areas; no shading, no significant difference; ∗estimates significant
at p < 0.001). See Supporting Information for full results.

conservation-oriented policies implemented in the other
3 countries (Brandt et al. 2017). China’s background
forest-loss rates were 3–5 times higher than in the other
3 countries, suggesting intense forest loss pressures due
to other economic development strategies in southwest
China, including timber extraction from forests, mining,
and hydropower development (Buckley 2010). When
compared with these other development strategies, eco-
tourism may be a relatively environmentally friendly form
of economic development in China.

Our empirical analysis highlights an urgent need for
more rigorous, empirical, and multiscale analysis of the
effects of ecotourism in biodiversity hotspots. To our
knowledge, this analysis is the only multinational and
the only counterfactual analysis that evaluates ecotourism
outcomes. Because economic development is also a goal
of ecotourism, there is an urgent need to analyze eco-
nomic benefits concurrently with forest change, for ex-
ample, by calculating a forest change per unit of eco-
nomic growth among different economic development
strategies. It is likely that, similar to other environmental
governance interventions, the effects of ecotourism vary
in space and time (Ostrom et al. 2007). Thus, case stud-
ies in diverse social-ecological contexts, and at different
spatial and temporal scales, followed by rigorous meta-
analyses will be essential to build a stronger knowledge
base.

Our findings have important implications for policy
makers because they highlight that forces of economic

development, even when stimulated by a nonextractive
development strategy like ecotourism, can lead to en-
vironmental degradation. Specifically, our results suggest
the rates of forest loss resulting from ecotourism are com-
parable to those resulting from other, more conventional,
development strategies. The exception to this rule ap-
pears to be areas where deforestation pressures are very
high. In these high deforestation areas, ecotourism may
slow forest loss. More research, at finer spatial scales, and
in other biodiversity hotspots, is necessary to build the
evidence base about under what conditions ecotourism
generates sustainable forest conservation outcomes.

Acknowledgments

J.S.B. acknowledges support from the National Science
Foundation (IIA-1301792). V.R. acknowledges support
from the NASA Land Cover and Land Use Change Pro-
gram.

Supporting Information

Covariates used in the analysis (Appendix S1), Full match-
ing results and balance statistics (Appendix S2), articles
reviewed for each country (Appendix S3) and summary
of ecotourism types and context in each country (Ap-
pendix S4) are available online. The authors are solely

Conservation Biology
Volume 33, No. 6, 2019

Brandt et al. 1327

responsible for the content and functionality of these
materials. Queries (other than the absence of material)
should be directed to the corresponding author.

Literature Cited

Abadie A, Imbens GW. 2006. Large sample properties of matching esti-
mators for average treatment effects. Econometrica 74:235–267.

ACNielsen. 2012. Final report for Himachal Pradesh—tourism survey.
Ministry of Tourism, Government of India, Himachal Pradesh, India.

Ahluwalia I, Little I, editors. 1998. India’s economic reforms and de-
velopment, essays for Manmohan Singh. Oxford University Press,
Delhi, India.

Andam KS, Ferraro PJ, Pfaff A, Sanchez-Azofeifa GA, Robalino JA. 2008.
Measuring the effectiveness of protected area networks in reducing
deforestation. Proceedings of the National Academy of Sciences of
the United States of America 105:16089–16094.

Bhutan. 2014. Bhutan tourism monitor annual report. Tourism Council
of Bhutan, Thimpu, Bhutan.

Brandt JS, Allendorf T, Radeloff V, Brooks J. 2017. Effects of national-
level forest management regimes on non-protected forests of the
Himalaya. Conservation Biology 31:1271–1282.

Brandt JS, Buckley RC. 2018. A global systematic review of empirical
evidence of ecotourism impacts on forests in biodiversity hotspots.
Current Opinion in Environmental Sustainability 32:112–118.

Brandt JS, Butsic V, Schwab B, Kuemmerle T, Radeloff VC. 2015. The
relative effectiveness of protected areas, a logging ban, and sacred
areas for old-growth forest protection in southwest China. Biological
Conservation 181:1–8.

Brandt JS, Kuemmerle T, Li H, Ren G, Zhu J, Radeloff VC. 2012. Us-
ing Landsat imagery to map forest change in southwest China in
response to the national logging ban and ecotourism development.
Remote Sensing of Environment 121:358–369.

Brandt JS, Nolte C, Agrawal A. 2016. Deforestation and timber produc-
tion in Congo after implementation of sustainable forest manage-
ment policy. Land Use Policy 52:15–22.

Brooks JS. 2010. Economic and social dimensions of environmental
behavior: balancing conservation and development in Bhutan. Con-
servation Biology 24:1499–1509.

Brooks JS. 2013. Avoiding the limits to growth: gross national happiness
in Bhutan as a model for sustainable development. Sustainability
5:3640–3664.

Buckley R. 2009. Ecotourism principles and practices. CAB Interna-
tional, Oxfordshire, United Kingdom.

Buckley R. 2010. Conservation tourism. CAB International, Oxfordshire,
United Kingdom.

Buckley R. 2011. Tourism and environment. Pages 397–416 in Gadgil A,
Liverman DM, editors. Annual review of environment and resources.
Volume 36. Annual Reviews, Palo Alto, California.

Butsic V, Lewis DJ, Radeloff VC, Baumann M, Kuemmerle T. 2017.
Quasi-experimental methods enable stronger inferences from ob-
servational data in ecology. Basic and Applied Ecology 19:1–10.

Donaldson JA. 2007. Tourism, development and poverty reduction in
Guizhou and Yunnan. The China Quarterly 190:333–351.

Donovan R. 2013. What have been the ecological and cul-
tural impacts due to tourism in the Himalayan region of
Dharamshala in India? Journal of Experiential Education and Learn-
ing https://doi.org/10.1111/cobi.13341.

Ferraro PJ, Merlin MH, Daniela AM, Gustavo Javier C-B, Subhrendu KP,
Katharine RES. 2013. More strictly protected areas are not necessar-
ily more protective: evidence from Bolivia, Costa Rica, Indonesia,
and Thailand. Environmental Research Letters 8:025011.

Garrard R, Kohler T, Price MF, Byers AC, Sherpa AR, Mahar GR. 2016.
Land use and land cover change in Sagarmatha National Park, a world
heritage site in the Himalayas of Eastern Nepal. Mountain Research
and Development 36:299–310.

Grenyer R, et al. 2006. Global distribution and conservation of rare and
threatened vertebrates. Nature 444:93–96.

Gurung DB, Seeland K. 2008. Ecotourism in Bhutan: extending its bene-
fits to rural communities. Annals of Tourism Research 35:489–508.

Hall C, Lew A. 2009. Understanding and managing tourism impacts.
Routledge, London, United Kingdom.

Hansen MC, et al. 2013. High-resolution global maps of 21st-century
forest cover change. Science 342:850–853.

HKTDC. 2017. Yunnan: market profile. HKTDC Research, Wanchai,
Hong Kong.

Jenkins M. 2009. Searching for Shangri-La. National Geographic, Wash-
ington, D.C.

Kirkby CA, Giudice-Granados R, Day B, Turner K, Velarde-Andrade LM,
Duenas-Duenas A, Lara-Rivas JC, Yu DW. 2010. The market triumph
of ecotourism: an economic investigation of the private and social
benefits of competing land uses in the Peruvian Amazon. PLOS ONE
5 (e13015). https://doi.org/10.1371/journal.pone.0013015.

Kirkby CA, Giudice R, Day B, Turner K, Soares-Filho BS, Oliveira-
Rodrigues H, Yu DW. 2011. Closing the ecotourism-conservation
loop in the Peruvian Amazon. Environmental Conservation 38:
6–17.

KPMG. 2012. Tourism in Himachal Pradesh and the way ahead. KPMG,
PHD Chamber of Commerce and Industry, Delhi, India.

Lambin EF, Meyfroidt P. 2011. Global land use change, economic global-
ization, and the looming land scarcity. Proceedings of the National
Academy of Sciences of the United States of America 108:3465–
3472.

Laurance WF, et al. 2014. A global strategy for road building (vol 513,
pg 229, 2014). Nature 514:262–262.

Li W, Han N. 2000. Ecotourism management in China’s nature reserves.
Ambio 30:62–63.

Liu J, Linderman M, Ouyang Z, An L, Yang J, Zhang H. 2001. Ecological
degradation in protected areas: the case of Wolong Nature Reserve
for Giant Pandas. Science 292:98–101.

Mahapatra PS, Pandey R, Pradhan S. 2012. River rafting in moun-
tainous regions of Uttarakhand: impacts, suggested mitigation
measures and sustainability. Journal of Mountain Science 9:511–
522.

Mather AS, Needle CL, Fairbairn J. 1999. Environmental kuznets curves
and forest trends. Geography 84:55–65.

Miteva DA, Loucks CJ, Pattanayak SK. 2015. Social and environmental
impacts of forest management certification in Indonesia. PLoS ONE
10 (e0129675). https://doi.org/10.1371/journal.pone.0129675.

MoCTCA (Ministry of Culture, Tourism and Civil Aviation). 2015a.
Nepal tourism statistics 2014. Ministry of Culture, Tourism and Civil
Aviation, Kathmandu, Nepal.

MoCTCA (Ministry of Culture, Tourism and Civil Aviation). 2015b.
Nepal Tourism statistics 2015. MoCTCA, Kathmandu, Nepal.

Myers N, Mittermeier R, Mittermeier C, da Fonseca GAB, Kent J. 2000.
Biodiversity hotspots for conservation priorities. Nature 403:853–
858.

Nagendra H, Karmacharya M, Karna B. 2005. Evaluating forest manage-
ment in Nepal: views across space and time. Ecology and Society
10:24.

Nepal SK, Karst H. 2017. Tourism in Bhutan and Nepal. Pages 187–197
in Hall CM, Page SJ, editors. The Routledge handbook of tourism in
Asia. Routledge, Oxon, United Kingdom.

Nolte C, Agrawal A, Silvius KM, Soares-Filho BS. 2013. Governance
regime and location influence avoided deforestation success of pro-
tected areas in the Brazilian Amazon. Proceedings of the National
Academy of Sciences 110:4956–4961.

Nyaupane GP, Morais DB, Dowler L. 2006. The role of community
involvement and number/type of visitors on tourism impacts: a con-
trolled comparison of Annapurna, Nepal and Northwest Yunnan,
China. Tourism Management 27:1373–1385.

Olson DM, et al. 2001. Terrestrial ecoregions of the worlds: a new map
of life on Earth. BioScience 51:933–938.

Conservation Biology
Volume 33, No. 6, 2019

https://doi.org/10.1111/cobi.13341

https://doi.org/10.1371/journal.pone.0013015

https://doi.org/10.1371/journal.pone.0129675

1328 Himalayan Ecotourism

Ostrom E, Janssen MA, Anderies JM. 2007. Going beyond panaceas.
Proceedings of the National Academy of Sciences of the United
States of America 104:15176–15178.

Pandey S, Wells M. 1997. Ecodevelopment planning at India’s Great
Himalayan National Park for biodiversity conservation and participa-
tory rural development. Biodiversity & Conservation 6:1277–1292.

Pandit MK, Manish K, Koh LP. 2014. Dancing on the roof of the world:
ecological transformation of the Himalayan landscape. BioScience
64:980–992.

Schroeder K, Sproule-Jones M. 2012. Culture and policies for sustainable
tourism: a South Asian comparison. Journal of Comparative Policy
Analysis 14:330–351.

Sekhon JS. 2011. Multivariate and propensity score matching software
with automated balance optimization: the matching package for R.
Journal of Statistical Software 42:1–52.

Shui W, Xu GW. 2016. Analysis of the influential factors for changes
to land use in China’s Xingwen Global Geopark against a tourism
development background. Geocarto International 31:22–41.

Singh RB, Mal S, Kala CP. 2009. Community responses to mountain
tourism: a case in Bhyundar Valley, Indian Himalaya. Journal of
Mountain Science 6:394–404.

Singh S. 2002. Managing the impacts of tourist and pilgrim mobility in
the Indian Himalayas. Revue de Geographie Alpine 90:25–36.

Stevens S. 2003. Tourism and deforestation in the Mt Everest region of
Nepal. Geographical Journal 169:255–277.

TCB (Tourism Council of Bhutan). 2015. Bhutan tourism monitor annual
report 2015. TCB, Thimpu, Bhutan.

Waldron A, Miller DC, Redding D, Mooers A, Kuhn TS, Nibbelink N,
Roberts JT, Tobias JA, Gittleman JL. 2017. Reductions in global biodi-
versity loss predicted from conservation spending. Nature 551:364–
367.

Wang CH, Buckley R. 2010. Shengtai Anquan: managing tourism
and environment in China’s forest parks. Ambio 39:451–
453.

WTTC (World Travel and Tourism Council). 2014. The economic im-
pact of travel and tourism. WTTC, London, United Kingdom.

Wyman MS, Stein TV. 2010. Modeling social and land-use/land-cover
change data to assess drivers of smallholder deforestation in Belize.
Applied Geography 30:329–342.

Xu JT, Yin RS, Li Z, Liu C. 2006. China’s ecological rehabilitation:
unprecedented efforts, dramatic impacts, and requisite policies.
Ecological Economics 57:595–607.

Conservation Biology
Volume 33, No. 6, 2019

RUNNING HEAD: MAGAZINE

MAGAZINE

The Travelling Insects

Submitted By:

Submitted To:

Date:

The Travelling Insects

Change is a very important factor for all living things out there, right from humans to animals to birds all living species travel through various distances because of several reasons. Similarly, insects are the of living species that travel to places far away and cover several miles of distances due to various factors involved. The insects that migrate the most from one region to another are the green darners. There is this annual cycle that describes their migrating behavior; the migration of the green darners depends on the generations of the insects. There are a total of two generations of green darners that form the annual migrating cycle, the first one comprises of two generations of migratory generation, and the second one comprises of the non migratory generations. So, their migration cycle has a pattern that is followed every year. The first generation makes the first spell of rounds from the southern to the northern range limits; here they lay eggs and then die. This prevents the species from getting extinct as the eggs prevent the breed and the new generations to survive. Then the pattern is continued by the second generation that is born from the eggs laid and they travel back to the southern range, from where their first generations had started the journey. Then the last round of spell starts as the resident generation of the green darners emerges, however they tend to reproduce locally and give rise to the population at the set location. This generation then again starts the annual cycle the next year where they travel to the northern range the next year. The same cycle continues every year and hence these insects reproduce, die, lay eggs and make sure the specie remains alive through migrating to and fro across the northern and southern range limits.

Comparing these ‘travelling’ insects to humans, one can deduce that migrating make these insects spread over several locations and they leave their foot prints on these location extending their reach all across the globe. A very interesting factor to note is that these insects migrate and carry out their annual cycle according to the seasons yearly. They make their first round of spell in the spring season of every year as in spring the insects with the southern origins arrive at the northern locations beforehand the species that have the origins from the mid latitudes. Similarly they thirds round of spells is most suitable in the autumn season. This happens when the individual insects are trapped in the southern part of the region. This migration has various positive effects most importantly the insect species become immune to all kinds of negative consequences, they learn to take advantage of the warm surface waters that helps them travel easily over the water bodies and this can cause them to travel immense long distances and this can cause the insect species to spread all over the world, making the insects a specie common all across the globe. This unique strategy and the cycle of nature make great revelations about how things work apart from the human and animal natural cycle. The migration of animals can be because of several reasons be it ecological and emigrational, however the migration has a very positive impact on humans, animals and insects.

Turn in your highest-quality paper
Get a qualified writer to help you with

“ 650 words 3 hrs ”

Get high-quality paper

NEW! AI matching with writer

Hire a Writer

Client Reviews

4.9

Sitejabber

4.6

Trustpilot

4.8

Our Guarantees

100% Confidentiality

Information about customers is confidential and never disclosed to third parties.

Original Writing

We complete all papers from scratch. You can get a plagiarism report.

Timely Delivery

No missed deadlines – 97% of assignments are completed in time.

Money Back

If you're confident that a writer didn't follow your order details, ask for a refund.

New to Your Trusted Assignment Help Service? Sign up & Save

Calculate the price of your order

Type of paper needed:

Pages:

You will get a personal manager and a discount.

Academic level:

We'll send you the first draft for approval by at

Total price:

$0.00

Power up Your Academic Success with the
Team of Professionals. We’ve Got Your Back.

Power up Your Study Success with Experts We’ve Got Your Back.

Order Now Order Now