Frog breeding in rain-fed wetlands after a period of severe drought: implications for predicting the impacts
of climate change
Wassens, S.a*, Walcott, Ab., Wilson, A.a and Freire, R.b
a Institute of Land Water and Society, School of Environmental Sciences, Charles Sturt University, LMB 588,
Wagga Wagga, NSW, 2678, Australia
b School of Animal and Veterinary Sciences, Charles Sturt University, LMB 588,Wagga Wagga, NSW, 2678,
Dr Skye Wassens
School of Environmental Sciences
Charles Sturt University
LMB 588 , Wagga Wagga, NSW, 2678, Australia
Tel: (02) 69 332843
Word count: 6672
Globally, rain-fed wetlands provide critical habitat for a wide range of amphibian species, however information on
the use of rain-fed wetlands by Australian frog species is extremely limited. This study examined the distribution of
frog breeding in rain-fed wetlands following the first significant rain event after a period of severe drought (2002-
2009) in order to predict how frog communities may be affected in the future by changed climate. Tadpole
communities along with vegetation and water quality variables were described in 35 rain-fed wetlands across the
South West Slopes and Riverina bioregions of inland south-eastern Australia. In addition, weekly tadpole surveys
were conducted in a subset of these wetlands to describe temporal patterns of occupancy. Despite the protracted dry
period prior to this study 50% of the rain-fed wetlands surveyed contained tadpoles. However frog communities
were species poor with only five species recorded. The majority of wetlands were dominated be a single species,
Limnodynastes tasmaniensis which is also common within permanent waterbodies such as farm dams and irrigation
infrastructure in both bioregions. Tadpoles of two burrowing species Limnodynastes interioris and Neobatrachus
sudelli were restricted to a small number of wetlands mostly in the South West Slopes. The composition of tadpole
communities changed over time, and Crinia parinsignifera was the only species that continued to breed over winter.
The dominance of generalist species within rain-fed wetlands indicates that characteristics such as dispersal
capability, flexibility in breeding times and the ability to utilise created habitats may be more important than
burrowing ability and longevity when predicting vulnerability to climate change.
Key words: amphibians; tadpoles; rain-fed wetlands; climate change; drought
Rain-fed wetlands are of major economic and ecological importance, providing many benefits such as surface water
storage, floodwater protection, nutrient cycling and water quality maintenance. In addition these wetland ecosystems
provide vital habitat for unique biota in an otherwise dry environment where rainfall is highly variable (Jolly et al.,
2008). Temporary rain-fed wetlands support a wide range of species including non-vascular and vascular plants
(Casanova & Brock, 2000), microinvertebrates (Boix et al., 2001; Gleason et al., 2004) , macroinvertebrates (Lee
Foote & Rice Hornung, 2005) and amphibians (Homan et al., 2004; Trauth et al., 2006; Baldwin & deMaynadier,
Rain-fed wetlands are vulnerable to changes in rainfall and evaporation rates, including those driven by increases in
atmospheric CO2 and enhanced green house effect, because their hydrological regime is driven largely by local
weather patterns (Acreman et al., 2009; Brooks, 2009) . The impacts of climate change on rain-fed wetlands and
their dependent faunal communities are difficult to quantify due to the high level of uncertainty regarding future
changes to local weather patterns, particularly changes in precipitation. However, a number of climate models for
Australia predict a long-term shift to higher temperatures, increased evaporation rates and more extreme
hydrological events (Smith & Chandler, 2010). For example, the drought experienced between 2002 and 2009 may
have been more severe than previous drought events because reduced rainfall has also been accompanied by higher
temperatures and evapotranspiration rates (Murphy & Timbal, 2008).
Rain-fed wetlands provide important habitat for a range of native frogs (Williams, 1985; Beja & Alcazar, 2003;
Vignoli et al., 2007b; Babbitt et al., 2009; Gómez-Rodríguez et al., 2009). Unlike more persistent water bodies, rain-
fed wetlands are typically free of large predators, particularly fish, making them important breeding habitat for
species with predator-sensitive tadpoles (Gillespie & Hero, 1999; Adams, 2000; Baber & Babbitt, 2003; Goldingay,
2008). However, water level fluctuations result in variable physio-chemical conditions and frog species utilising
rain-fed wetlands must trade-off the reduced risk of predation with the increased risk of breeding failure due to
desiccation and declining water quality (Vignoli et al., 2007a; Gahl et al., 2009; Gómez-Rodríguez et al., 2009).
Life-cycle adaptations such as plastic developmental times and bet-hedging in tadpoles (Dziminski & Alford,
2005), burrowing and aestivation (Penman et al., 2004; Tracy et al., 2007; Kayes et al., 2009), and emigration under
adverse conditions (Wassens et al., 2008) are common characteristics of Australian frog species in temperate, semi-
arid and arid environments. Dispersal capability may be a critical factor influencing the ability of both burrowing
and non-burrowing species to recover following periods of drought because it enables them to rapidly recolonise
newly created rain-fed wetlands. The spatial distribution of permanent waterbodies and soil types suitable for
burrowing may also influence recolonisation of rain-fed wetlands (Mazerolle & Villard, 1999) although the latter
issue has received little attention.
Despite their many adaptations to the use of variable resources, amphibians that preferentially breed in rain-fed and
temporary pools are considered to be extremely sensitive to changes in temperature and precipitation occurring as a
result of climate change (Blaustein et al., 2010). While long-term shifts in precipitation are difficult to predict, a
number of recent studies have demonstrated the negative impact of severe drought on amphibian community
composition and breeding success (Rohr & Madison, 2003; Piha et al., 2007; McMenamin et al., 2008; Mac Nally et
al., 2009; Rohr & Raffel, 2010). Increased climatic variability and increased frequency of extreme weather events
can have a number of direct and indirect impacts on amphibian populations, including increased mortality during
extreme dry or extreme cold events (Blaustein et al., 2010), reduced breeding success and recruitment (Babbitt &
Tanner, 2000; Piha et al., 2007) and increased susceptibility to disease (Rohr & Raffel, 2010). It can however, be
difficult to separate the normal inter-annual variability in amphibian population dynamics and distributions from
potential long-term shifts caused by climate change. For example some population declines attributed to extreme
drought and climate change may have been a result of normal variability in occupancy patterns caused by
individuals retreating back to areas of core habitat during extended drought, only to recolonise formerly occupied
wetlands with a return to favourable conditions (Patla et al., 2009).
Much of our understanding of the impacts of climate change on amphibians in rain-fed waterbodies is derived from
studies in the northern hemisphere where vernal pools are important habitat for a wide range of species e.g.
(Blaustein et al., 2010; Blaustein et al., 2011). Knowledge of the importance of rain-fed wetlands for frogs in
Australian temperate and semi-arid regions is extremely limited making predictions on the impacts of future climate
change difficult. Frog communities in the dry temperate and semi-arid regions of south-eastern Australia are
typically species poor compared with the wetter temperate, tropical and monsoonal regions of Australia (Caughley
& Gall, 1985). The Riverina and South West Slopes Bioregions support 15 frog species, the majority of which occur
across a range of aquatic habitats including seasonally flooded wetlands associated with the Murrumbidgee River
system (Wassens, 2006). Only two species Notaden bennettii and Pseudophryne bibronii are considered to be rain-
fed wetland specialists and are active only after heavy rain (Cogger, 2000).
The use of rain–fed wetlands by resident frog species is primarily driven by the timing of filling and drying
(seasonality of rainfall) and the hydrological regime (duration of inundation, and the length time between filling
events). A large number of frog species have restricted activity periods which are often linked to temperature
(Carroll et al., 2009) and as a result tadpole communities within the same rain-fed wetland can vary considerably
between years depending on the timing of inundation (Paton & Crouch, 2002; Jakob et al., 2003). Those species
which follow specific breeding times may suffer more from altered seasonality of rainfall under climate change, than
those with long breeding periods or opportunistic breeding.
Tadpole development times are species specific and this is reflected in the hydroperiod of the wetlands that these
species inhabit. Species with rapid tadpole development times typically have a higher breeding success rate in
temporary waterbodies, while species with long development times may be excluded (Lane & Mahony, 2002;
Popescu & Gibbs, 2009). Tadpole mortalities due to desiccation are a common result of hydroperiods which fail to
extend over the given species complete developmental time frame (Williamson & Bull, 1999; Babbitt & Tanner,
2000; Brodman et al., 2003; Jakob et al., 2003). Predicted increases in potential evapotranspiration, especially
during summer may contribute to reduced hydroperiods within rain-fed wetlands and increased frequency of
recruitment failure due to tadpole desiccation.
The aims of this study were to:
1) identify which frog species utilise rain-fed wetlands after a period of severe and extended droughts and
2) describe temporal patterns of breeding activity by frogs within rain-fed wetlands in order to better
understand the impacts of potential shifts in the timing of significant rain events.
Regional scale surveys were conducted at a total of 35 rain-fed wetlands, located in the South West Slopes bioregion
(n=18) and in the Riverina bioregion centred on the Hay Plain (n=17) in south west New South Wales, Australia
(Figure 1). Vegetation of the South West Slopes is diverse, with the original vegetation dominated by open
woodlands containing white box (Eucalyptus albens, Benth), yellow box (E. melliodora, Cunn. ex Schauer) and
blakelys red gum (E. blakelyi, Maiden) the vast majority of which has been cleared. Black box (Eucalyptus
largiflorens, Muell) communities occur along intermittent creeks and wetlands and river red gum forests (E.
camaldulensis, Dehnh) occur close to permanent rivers and seasonally flooded wetlands in both bioregions. The
Riverina bioregion is a flat, largely treeless plain dominated by Atriplex communities and native grasslands. The
dominant land use is pastoral grazing and along with extensive areas of irrigated crop production.
Climate and rainfall patterns
The annual mean rainfall in the South West Slopes is 572 mm, summer temperatures are hot, averaging between 29
and 32 degrees Celsius and winter temperatures range from overnight minimums of three to daily maximums of 12
to 14 degrees Celsius (Bureau of Meteorology 2010). The climate in the Riverina is semi-arid (mean annual rainfall
365 mm), with very hot summers, mean maximum temperatures of 33 degrees Celsius, and cool winters (average
min. 3.5 to average max. of 15 degrees Celsius) (Bureau of Meteorology 2010). Rainfall is slightly winter dominant,
although high intensity summer thunderstorms of short duration are common during wetter years.
Inter-annual variability in precipitation is extremely high in both bioregions, between 1941 and 2010 the annual
precipitation ranged from 164.9 mm to 836.8 mm in the Riverina (variance 21094.9) measured at Hay and between
165.7 mm and 1019.2 mm in the South West Slopes (variance = 28345.1) measured at the nearby town of Wagga
Wagga (Bureau of Meteorology 2010). In both bioregions, 2002, 2006 and 2008 were in the top 12 driest years
between 1941 and 2010, while 2010 was the wettest year for that period in Wagga Wagga and the 4th wettest year in
Hay (Figure 2) (Bureau of Meteorology 2010).
As the distribution of rain-fed wetlands is poorly documented, air photos and consultation with property owners and
Murrumbidgee Catchment Management Authority staff were initially employed to identify potential wetlands for
investigation. In February 2010 a low pressure system brought significant rainfall to the Riverina and the South
West Slopes, this was followed by an inland trough in early March which brought a second significant rain event
(record highest rainfall in Wagga Wagga) to both regions (Bureau of Meteorology 2010) (Figure 3). Further
significant rain events occurred in May, July and August. All wetlands included in this study first filled in February
2010 with subsequent top-up falls in March and May after which time the smaller shallower wetlands dried out,
before refilling in June. Larger wetlands retained water throughout the study period for six to nine months. All
wetlands were dry during the severe drought conditions experienced through 2009.
Regional scale surveys of tadpole communities, vegetation characteristics and water quality were conducted
between the 15th and 31st of March 2010. In addition to regional scale surveys, repeat weekly sampling was
conducted at a subset of wetlands between the 31st of March and 10th of August, 2010. Wetland type and size were
classified according to (Semeniuk & Semeniuk, 1995) as Playa (n = 20)(intermittently flooded basin), Sumpland (n
= 7)(seasonally inundated basin) or Wadi (n = 8) (intermittently flooded channel). Wetlands fell into one of two
size classes, Leptoscale (n = 24) less than 100 x 100m, and mesoscale (n = 3) between 500 x 500m to 1000 x
1000m. Wadi channels were all classified as microscale (n = 8) less than 10m across and 100m long (Semeniuk &
Three replicate measures of water quality (turbidity, conductivity, pH, dissolved oxygen and temperature) were
taken on each survey occasion using a hand held YSI 6820 V2 multi-parameter sonde at approximately one metre
from the water’s edge. Vegetation was assessed by visual estimation of the proportion of microhabitats (fringing
vegetation, tall emergent vegetation, short emergent vegetation, submerged vegetation, floating vegetation and
timber within the water body) comprising five metre wide bank to bank sections of the wetland. Fringing vegetation
was considered to include all plant classes present within two meters of the shoreline, or where the there was an
abrupt change in vegetation communities from semi-aquatic and riparian vegetation to terrestrial grassland and
Atriplex communities. Tall emergent vegetation was classified as vegetation greater than 20 cm above the water
level (e.g. Typha sp.), whilst short emergent vegetation included sedges and short rushes less than 20 cm above the
water level. Submerged vegetation included all plant species present below the water surface, including vegetation
recently inundated due to rising water levels. Floating vegetation was any vegetation that floated on the water’s
surface and included both attached and free floating species.
Tadpoles were surveyed using sweep nets, targeting the range of water depths and habitats within the wetland. This
is a common and effective technique for sampling tadpoles within rain-fed water bodies (Anstis, 2002). Sweeping
was conducted at each wetland for 20 minutes. Visual searches for tadpoles and metamorphs were also conducted by
walking the entire wetland perimeter, or in the case of larger systems, walking a 50 m transect. Tadpoles were
collected in a soft mesh sweep net ( mesh size 2 mm, 40 cm wide), and immediately transferred to a container where
they were identified to species, measured and their development stage described based on the extent of limb
development (Grosner, 1960) before being released at the point of capture.
Temporal occupancy patterns
Change in tadpole communities over time was assessed at a subset of five wetlands on 14 occasions (weekly), from
the 31st of March to the 10th of August, 2010. Wetlands were selected on the basis of their tadpole communities with
the aim of including all species likely to occur in both the South West Slopes and Riverina bioregion sites. In order
to conduct intensive surveys of wetlands on a weekly basis and to minimise potential variability in tadpole responses
caused by regional variation in temperature and rainfall, we aimed to select wetlands within a 15 km radius of one
another which contained all five species. Three key species, Crina parinsignifera Main, Litoria peronii Tschudi and
Neobatrachus sudelli Lamb, were restricted to a small number of wetlands in the South West Slopes, as a result
wetlands containing these three species formed the basis for our search area with the remaining wetlands selected
from within the 15 km radius that contained the more common and widespread species.
In order to limit potential confounding effects caused by land use disturbance, all wetlands were selected from
within nature reserves. Vegetation and water quality data were collected on each survey occasion, as described for
the broad-scale occupancy surveys, along with collection of additional tadpole data for body size. Tadpoles were
identified to species and stage of development via visual assessment (Anstis, 2002).
General Linear Models (GLMs) and Bonferroni post-hoc tests were employed to analyse univariate data using SPSS
(SPSS for Windows Release 17.0 SPSS Inc 2005). The mean values of temperature, salinity, turbidity, dissolved
oxygen, conductivity and pH were compared between regions and among wetland types. Differences in the
composition of tadpole communities between bioregions and between seasons, autumn (n = 7) and winter (n = 7)
(temporal occupancy patterns) were tested using Analysis of similarities (ANOSIM) with a Bray Curtis Similarity
matrix. Data were square-root transformed prior to analysis (Clarke & Warwick, 2001).
Two-tailed t-tests were used to compare the means for conductivity, pH, turbidity, percent cover of open water,
floating, submerged, fringing and short emergent vegetation, and total percent cover of aquatic vegetation, total
percent cover of aquatic vegetation for vacant and occupied wetlands from the broad-scale data.
Water quality was variable between regions and among wetlands, reflecting the wide geographic spread of wetlands
included in this study. Conductivity was high in playa wetlands (mean 1.088 mScm-1 SE 0.086) within the Riverina
compared to range of 0.054 to 0.426 mScm-1 for other wetland types resulting in significant differences among
wetland types (GLM F = 3.779, p = 0.039), however there was no significant difference in conductivity between
bioregions. In contrast pH differed significantly between bioregions, with mean pH significantly higher in the
Riverina (mean 10.20 SE 0.04) than in the South West Slopes (mean 8.38 SE 0.11) (GLM F = 15.711, p = 0.001).
Aquatic vegetation communities were variable across wetlands in both regions and there were no significant
differences in the structural diversity of aquatic vegetation, aquatic or fringing vegetation cover between bioregions
or among wetland types.
Broad scale patterns of tadpole occupancy
A total of 310 tadpoles of five species were collected from 17 of the 35 wetlands (49%) surveyed. The proportion of
wetlands containing tadpoles was similar across the two bioregions, seven out of 17 wetlands in the Riverina and
nine out of 18 wetlands in the South West Slopes contained tadpoles. Three species, Neobtrachus sudelli (n = 3),
Crinia parinsignifera (n = 2) and Litoria peronii (n = 1) occurred only in the South West Slopes (Figure 4). Of the
remaining two species Limnodynastes interioris Fry occurred in both bioregions but was rare (n = 4) while
Limnodynastes tasmaniensis Günther, was widespread and abundant in both the bioregions. Consequently, species
richness was on average higher in the South Western Slopes (0.94) than the Riverina (0.47) but this difference was
not significant (GLM; F = 1.680; P = 0.204). It was not possible to identify statistically significant differences in
community composition between the two bioregions due to lack of statistical power as the vast majority of wetlands
in both bioregions contained just one species (L. tasmaniensis) (ANOSIM; Global R= 0.011; P = 0.435).
As the majority of species were very rare, we were only able to analyse the relationship between presence/absence
and habitat and water quality variables for L. tasmaniensis. There were no significant relationships for mean water
quality and vegetation attributes measured in wetlands where L. tasmaniensis tadpoles were present and where they
Temporal patterns in tadpole occupancy
Significant rain events occurred from the 13th and 16th of February (60mm) and the 6th to the 8th of March (164 mm),
the smallest of the five wetlands dried out in late may and was refilled in June (see Figure 3). Water temperature
decreased over time at the five weekly monitoring sites from a mean of 25oC in March to 10oC in late July and
August. While conductivity also decreased from a mean high of 0.099mScm-1 in March to a low of 0.062 mScm-1 in
August. The percent cover of aquatic vegetation also decreased over time mainly due to fluctuating water levels,
ranging from a mean of 37 % in March to 6% in August.
Tadpoles of five frog species (L. peronii, C. parinsignifera, L. interioris, N. sudelli and L. tasmaniensis) were
monitored over the duration of the study within the five wetlands. The tadpole communities differed significantly
between autumn and winter (ANOSIM Global R=0.278, p=0.01) and according to location (discrete wetlands)
(ANOSIM Global R= 0.683, p = 0.01). Tadpole communities in autumn were typically more diverse, with four
species, C. parinsignifera, N. sudelli, L. tasmaniensis and L. interioris present, while the winter tadpole
communities were generally dominated by C. parinsignifera (Figure 5).
Tadpole abundance for the majority of species declined over time with the exception of C. parinsignifera which
declined and then increased following winter rainfall. Neobatrachus sudelli was present for the shortest period of
time (no longer recorded after the 15th of May) and its abundance was initially high and steadily decreased over time
with late stage tadpoles ( grosner stages 37 and above) recorded in early May. Crinia parinsignifera was the only
species to undergo two breeding events, early stage tadpoles were recorded in March and April, with later stage
tadpoles (grosner stages 37 and above) were collected in mid-May. The site containing C. parinsignifera dried out
briefly and then refilled following significant rain events on 26 and 29 of May (26 and 20 mm respectively) (see
figure 4). Egg masses and very early stage tadpoles (stage 26) located from the 2nd of June through July, mid stage
tadpoles (grosner stages 32) (Grosner, 1960) were present when surveys ceased on the 10th August.
Tadpoles were relatively widespread in the rain-fed wetlands of both bioregions, but communities were species
poor. Only five of the 15 species known to occur in both bioregions were recorded within rain-fed wetlands in this
study (Wassens, 2006). The vast majority of wetlands contained only one species, the habitat generalist
Limnodynastes tasmaniensis. Despite the lower survey effort in this study, tadpoles occurred at a far higher
proportion of wetlands than reported by McNally et al (2009) in their study in south-eastern Victoria. This probably
reflects the different focus of that study on more persistent waterbodies which were present during the drought
period in 2006 and 2007.
Wetlands are subject to a wide range of disturbances and threatening processes including draining, damming,
grazing and cultivation (Finlayson & Rea, 1999). The spatial context of rain-fed wetlands has also been altered
within agricultural landscapes with dams and canals increasing the availability of permanent habitats, while the
availability of temporary waterbodies has decreased (Brock et al., 1999). These shifts in habitat availability have
undoubtedly altered frog communities (e.g McNally et al 2009), with a potential shift towards generalist species
capable of utilising more persistent waterbodies as well as the opportunistic use of rain-fed waterbodies when they
The results of this study suggest that dispersive generalist frog species, rather than burrowing frog species tend to
dominate rain-fed wetlands in modified agricultural landscapes after periods of extended drought. Limnodynastes
tasmaniensis has a number of adaptations that make it highly suited to the opportunistic use of rain-fed wetlands, it
has comparably good dispersal ability (Schauble, 2004 ), and utilises a wide range of aquatic habitats including
farm dams (Hazell et al., 2001) and irrigation infrastructure (Wassens, 2006) which provide a more constant source
of water during drought periods. Studies of other vertebrate taxa consider dispersal ability to be a critical factor in
determining the risk from climate change, however this typically relates to large scale movements which reflect
range shifts (Pearson & Dawson, 2003; Thomas et al., 2004). At local scales dispersal ability may be a key factor
influencing the ability of aquatic and semi-aquatic fauna to recover from severe climate perturbations such as
drought because it influences their ability to seek out permanent refuge sites and recolonise waterbodies following
local extinctions (Harrison, 1994; Smith & Green, 2005).
Burrowing ability is a specific adaptation of frogs to variable rainfall and allows individuals to persist for extended
dry periods (Tracy et al., 2007; Kayes et al., 2009). On this basis burrowing ability might be seen as lowering the
susceptibility of species to future climate changes, with large-bodied, long-lived species such as Limnodynastes
interioris being comparatively resilient to drought compared to the shorter lived, non-burrowing species. However
there have been few studies on the actual capacity of these species to aestivate for extended periods. Many
burrowing species are active throughout the year, but this activity is closely linked to meteorological conditions,
particularly rain and humidity (Penman et al., 2004). Extended drought may limit the ability of these species to
forage and this, along with their requirement for wetlands with extended hydroperiods to support breeding, may
explain their rarity in rain-fed wetlands in the drier Riverina bioregion. Non-cocoon forming species such as L.
interioris may also burrow deeper during extended dry periods in order to maintain moisture balance (Thompson et
al., 2005) and retreating water tables or increased groundwater salinity could potentially limit the survival of these
species during periods of extended drought.
Weekly monitoring of tadpole communities within selected wetlands was used in this study to assess change in
tadpole communities over time. This study was conducted outside the reported core calling periods of L.
tasmaniensis (Lemckert & Mahony, 2008) but despite this, tadpoles of this species were abundant, which
demonstrates the ability of this species to respond to unusual weather conditions. Crinia parinsignifera was the only
species able to respond to winter rain events and underwent two distinct breeding events between March and
August. This high level of flexibility and rapid breeding times may increase the resilience of this species to changes
in the seasonality of rain events. Studies conducted in the Northern Hemisphere have reported on changes in
breeding phenology, including earlier calling times in response to increased spring temperatures (Neveu, 2009;
Blaustein et al., 2010; Todd et al., 2010). These studies focused on areas where the availability of wetland habitat
was unchanged but where changes in temperature altered breeding cues. Assessing the impacts of altered rainfall
patterns on frog breeding is more complex because both the temporal patterns of habitat availability, and the
temperature driven breeding cues are altered. For frogs to respond to a rain event it must occur within their activity
window and for tadpoles to successfully develop, wetlands must retain water for a sufficient period of time. Shifts
towards summer dominated rainfall may allow breeding for many species but higher temperature and evaporation
rates may reduce wetland hydroperiod excluding those species with longer and less flexible development times, such
as Limnodynastes interioris.
Marked inter-annual variation in rainfall, wetland hydrology and subsequent variability in frog breeding and
community composition has also been widely reported for rain-fed wetlands systems (Jakob et al., 2003; Piha et al.,
2007; Gómez-Rodríguez et al., 2009). This high level of variability presents a significant challenge for predicting
impacts of future climate change on frogs within rain-fed wetland systems. In this respect predictive models that
consider species distributions with respect to climate variability, for example variability in the duration of wet-dry
periods, may be more informative when predicting the impacts of climate change on species in variable landscapes
than models based on temperature shifts alone. The long-term study of frog communities in rain-fed wetlands and
better assessment of their hydrological requirements and distributions will also greatly assist our understanding of
population processes in systems that already have a high level of inter-annual variability.
David Read (Wagga Wagga City Council) and Dr Patricia Murray (Murrumbidgee Catchment Management
Authority) greatly assisted in the identification and selection of survey wetlands. Two anonymous reviewers
provided valuable advice which greatly improved the quality of this paper. Vanessa Griese and Sarah Cordell
assisted with data collection. We thank land holders for giving us access to study sites on their property. This
research was conducted under a NSW National Parks and Wildlife Service Licence (S12393). Ethics approval was
granted by Charles Sturt University Animal Care and Ethics Committee (10/042).
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List of figures
Figure 1. Location of rain-fed wetland study sites within the Riverina (Riv) and South West Slopes (SWS).
Figure 2. Annual precipitation over a 50 year period (1940-2010) in the Riverina at Hay (75031) and South West
Slopes at Wagga Wagga (72150). Reference line shows median precipitation for the same period. Source:
Commonwealth Bureau of Meteorology 2010.
Figure 3. Total monthly rainfall (mm) and mean daily air temperature recorded at 3pm at Wagga Wagga (72150)
and Hay (75031) between February 1st 2010 and August 31st 2010. Source: Commonwealth Bureau of Meteorology
Figure 4. Proportion of wetlands occupied by tadpoles in the Riverina and South West Slopes for surveys conducted
between the 15th and 31st of March
Figure 5. Temporal changes in tadpole numbers of each species between March 31st and August 10th 2010 at five
wetlands within the South West Slopes near Wagga Wagga.
0 100 200 30050 Km
Figure 1. Location of rain-fed wetland study sites within the Riverina (RIV) and South West Slopes (SWS).
Figure 2. Annual precipitation from1940-2010in the Riverina at Hay (75031) and South West Slopes at Wagga
Wagga (72150). Reference line shows median precipitation for the same period. Source Commonwealth Bureau of
Figure 4. Proportion of wetlands occupied by tadpoles in the Riverina and South West Slopes.
Figure 5. Temporal changes in tadpole numbers of each species between March 31st and August 10th 2010 at five
wetlands within the South West Slopes near Wagga Wagga.