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This is the author’s final version of the work, as accepted for publication following peer review but without the publisher’s layout or pagination.

The definitive version is available at http://dx.doi.org/10.1111/gcb.12444

Beatty, S.J., Morgan, D.L. and Lymbery, A.J. (2014) Implications of climate change for potamodromous fishes. Global Change

Biology, 20 (6). pp. 1794-1807.

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Copyright: © 2013 John Wiley & Sons Ltd.

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http://dx.doi.org/10.1111/gcb.12444

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This article has been accepted for publication and undergone full peer review but has not been through the copyediting, typesetting, pagination and proofreading process, which may lead to differences between this version and the Version of Record. Please cite this article as doi: 10.1111/gcb.12444 This article is protected by copyright. All rights reserved.

Received Date: 09-May-2013 Revised Date : 25-Sep-2013 Accepted Date: 03-Oct-2013 Article type : Primary Research Articles Implications of climate change for potamodromous fishes

Running head: Potamodromous fishes and climate change

STEPHEN J. BEATTY*, DAVID L. MORGAN and ALAN J. LYMBERY

Freshwater Fish Group & Fish Health Unit, School of Veterinary and Life Sciences, Murdoch

University, South St Murdoch, Western Australia

*Correspondence: Stephen J. Beatty, tel. +618 9360 2813, fax +618 9360 7512, e-mail:

[email protected]

Keywords: freshwater fishes, Mediterranean climate, surface flow decline, groundwater

reduction, secondary salinisation, south-western Australia, aquatic refuge

Primary Research Article

Abstract

There is little understanding of how climate change will impact potamodromous freshwater

fishes. Since the mid-1970’s, a decline in annual rainfall in south-western Australia (a

globally recognised biodiversity hotspot) has resulted in the rivers of the region undergoing

severe reductions in surface flows (~50%). There is universal agreement amongst Global

Climate Models that rainfall will continue to decline in this region. Limited data are

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This article is protected by copyright. All rights reserved.

available on the movement patterns of the endemic freshwater fishes of south-western

Australia or on the relationship between their life-histories and hydrology. We used this

region as a model to determine how dramatic hydrological change may impact

potamodromous freshwater fishes. Migration patterns of fishes in the largest river in south-

western Australia were quantified over a four year period and were related to a number of key

environmental variables including discharge, temperature, pH, conductivity and dissolved

oxygen. Most of the endemic freshwater fishes were potamodromous, displaying lateral

seasonal spawning migrations from the main channel into tributaries, and there were

significant temporal differences in movement patterns between species. Using a model

averaging approach, amount of discharge was clearly the best predictor of upstream and

downstream movement for most species. Given past and projected reductions in surface flow

and groundwater, the findings have major implications for future recruitment rates and

population viabilities of potamodromous fishes. Freshwater ecosystems in drying climatic

regions can only be managed effectively if such hydro-ecological relationships are

considered. Proactive management and addressing existing anthropogenic stressors on

aquatic ecosystems associated with the development of surface and groundwater resources

and land use is required to increase the resistance and resilience of potamodromous fishes to

ongoing flow reductions.

Introduction

Freshwater ecosystems are in crisis throughout the world. Aquatic habitats associated with

65% of global river flow are moderately or highly threatened, principally by water resource

development and pollution (Dudgeon et al., 2006; Vörösmarty et al., 2010). Freshwater

fishes are one of the most threatened faunal groups and are expected to be among the most

severely impacted by climate change, due to it adding to a wide range of current and future

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anthropogenic stressors (Ficke et al., 2007; Palmer et al., 2008; Olden et al., 2010; Comte et

al., 2013). Freshwater fishes in Mediterranean climatic regions are particularly imperilled

and while flow reduction due to water extraction and climate change is recognised as a major

threat (Hermoso & Clavero, 2011; Maceda-Veiga, 2013), its impacts are difficult to quantify

due to the complexity of abiotic and biotic interactions (Clavero et al., 2010).

The challenge of conserving critical habitat for freshwater species is exacerbated by life-

cycles that often involve movement between habitats for different life-history stages

(Magoulick & Kobza, 2003) and the linear nature of rivers, whereby upstream activities

influence downstream habitat quality and availability (e.g. Gorman & Karr, 1978; Allan et

al., 1997). Seasonal migrations in potamodromous fishes are critical for completing life-

cycles; particularly in accessing spawning and nursery habitats and refugia (Reynolds, 1983;

O’Connor et al., 2005; Munz & Higgins, 2013). Understanding refuge use and species

mobility is important for enhancing conservation efforts; particularly under drought or flow

reduction scenarios (Magoulick & Kobza, 2003; Morán-López et al., 2012; Chessman, 2013).

In comparison to the amount of research on diadromous fishes (e.g. Crozier et al., 2008;

Lassalle & Rochard, 2009; Finstad & Hein, 2012; Piou & Prévost, 2013), there is little

information on how flow reductions due to climate change will impact potamodromous

species (but see Buisson et al., 2008; Clews et al., 2010). The vast majority of research

examining the impacts of climate change on freshwater fishes has focussed on salmonids and

other cold water species, those with commercial or recreational value, and those in the

Northern Hemisphere (e.g. Comte et al., 2013). Broadening our understanding of these

impacts to other regions and to non-diadromous fishes with little commercial or recreational

value is needed for a more complete understanding of the vulnerability of freshwater

ecosystems to climate change. As predicted globally (e.g. Thieme et al., 2010), climate

change is expected to exacerbate the current anthropogenic stressors on Australian freshwater

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fishes (Morrongiello et al., 2011a, b) and cause major range contractions of several species

(Balcombe et al. 2011; Bond et al. 2011).

South-western Australia, a global biodiversity hotspot (Myers et al., 2000), has a

Mediterranean climate that is isolated from high rainfall temperate and tropical regions by

extensive arid zones. Although the region has only 11 native freshwater fish species, nine are

regionally endemic (Morgan et al., 1998; Allen et al., 2002). These species are highly

imperilled, principally due to the synergistic impacts of a number of stressors such as

secondary salinisation (Morgan et al., 2003; Beatty et al., 2011), instream barriers (Morgan &

Beatty, 2006; Beatty et al., 2007), and the impacts of introduced species (Morgan et al.,

2004; Marr et al., 2010; Beatty & Morgan, 2013).

Since the mid-1970’s, a considerable climatic shift has occurred in south-western

Australia, resulting in a reduction of ~16% in annual rainfall corresponding to an ~50%

reduction in the annual volume of stream flow (Suppiah et al., 2007; Petrone et al., 2010;

Silberstein et al., 2012). There is an unusually uniform consensus across global climatic

models (GCMs) for this region as they all project this drying trend to continue, with median

reductions of an additional ~8% in annual rainfall and ~25% (range of 10-42%) in annual

stream flow expected by 2030 (Suppiah et al., 2007; Silberstein et al., 2012). Furthermore,

reductions in fresh groundwater levels of >10m are predicted in many areas in the region by

2030 (Barron et al., 2012). Quantifying the potential impacts of this dramatic hydrological

change on freshwater fishes of the region is currently hampered by a lack of ecological data

on these species (Beatty et al., 2010, Morrongiello et al., 2011a).

The two non-endemic native freshwater fishes in south-western Australia are the

galaxiids Galaxias maculatus and Galaxias truttaceus that may have diadromous (marine

larval phase) or land-locked populations; the latter utilising coastal lacustrine environments

for larval development (Morgan 2003; Chapman et al., 2006). While almost nothing is

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known on the migration patterns of the nine endemic freshwater fishes, they are not

diadromous and are known to breed during the major rainfall and surface flow periods (Pen

& Potter, 1991; Morgan et al., 1998). Given the timing of their breeding period, they may be

highly reliant on seasonal surface discharge and could be potamodromous in order to breed

within intermittent lotic systems that are common to the region.

Secondary salinisation is a major global threat that is expected to be amplified by climate

change (Cañedo-Argüelles et al., 2013). This process is severe in south-western Australia,

with most of the major rivers salinised due to the mobilisation of saline groundwater from

large scale land-clearing in their upper catchments (Mayer et al., 2005). While rising saline

groundwater is a major cause of secondary salinisation (Cañedo-Argüelles et al., 2013), fresh

groundwater aquifers can have a dilution effect in salinised rivers. The fresh south-west

Yarragadee Aquifer (SWYA) underlies a considerable area of south-western Australia and

can contribute to between 30-100% of the baseflow riverine discharge of the Blackwood

River (the largest of the region) during the prolonged dry-season (Strategen, 2006). This

contribution significantly dilutes the main channel during baseflow, when most tributaries fail

to flow or dry completely (Del Borrello, 2008; Beatty et al., 2010).

Freshwater fishes of the region are stenohaline and have undergone major downstream

range contractions away from the most salinised inland habitats towards those regions

underlain by fresh aquifers (Morgan et al., 2003; Beatty et al., 2011). Fresh groundwater

discharge maintains habitat connectivity for at least one species (Beatty et al., 2010) and is

likely to play a key role in maintaining fresh baseflow refuge for other species. As the

groundwater level and surface discharge during the annual dry season are positively related to

the previous year’s rainfall in the Blackwood River (Golder & Associates, 2008), both will be

affected by projected rainfall reductions (Barron et al., 2012; Silberstein et al., 2012).

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Understanding the movement patterns and habitat utilisation of freshwater fishes and

their relationships to hydrology and physicochemical variables will be vital in understanding

how hydrological change will affect future population viabilities. The current study aimed to

elucidate the spatial and temporal movement patterns of freshwater fishes, their life-histories,

and their relationships with environmental variables in a drying climatic region, using the

Blackwood River as a model system. The study hypothesised that: 1) south-western

Australian freshwater fishes are potamodromous, based on existing information on their

reproductive biology; 2) both fresh surface water and groundwater discharge are important in

maintaining remnant populations; and 3) ongoing reductions in rainfall and discharge due to

climate change will negatively impact potamodromous fishes.

Materials and methods

Sampling regime

Spatial and temporal patterns in the movement and demographics of native freshwater fishes

were determined in the Blackwood River main channel and four of its tributaries between

October 2005 and December 2009 (Fig. 1). Three main channel sites included one

downstream of the major zone of groundwater discharge from the SWYA (Denny Rd), one at

the upstream margin of the SWYA discharge zone (Milyeannup Pool) and one upstream of

the SWYA discharge zone (Jalbarragup) (Fig. 1). The tributary sites included one

perennially flowing system (Milyeannup Brook) that is maintained by groundwater from the

SWYA, and three seasonally intermittent systems (Layman Brook, Rosa Brook and McAtee

Brook) (Fig. 1).

The seven sites were sampled on a regular basis between October 2005 and December

2009. In each sampling event (n = 30 within the 50 month period), all sites that had flowing

water at the time were sampled. All seven sites were sampled during the major flow period

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between August and December (i.e. late winter through to early summer). However, as three

of the four tributaries naturally dry during late summer through autumn (i.e. February to

May), sampling did not occur in those systems during that period due to a lack of flow. For

each sampling event, two fyke nets (each net having a width of 11.2 m, including two 5 m

wings and a 1.2 m wide mouth, depth of 0.8 m, length of 5 m with two funnels all comprised

of 2 mm woven mesh) were simultaneously set for three consecutive 24 hour periods. One

net faced upstream, to determine downstream movements of fishes, and the other faced

downstream, to determine upstream movements of fishes (following Beatty et al., 2010).

As it was not always possible to set the nets across the entire wetted channel, the proportion

of the total stream cross-section that was covered by each fyke was determined in order to

standardise captures (see statistical analysis). All fish captured were identified to species,

enumerated, measured to the nearest 1-mm total length (TL) and released either upstream or

downstream of the capture point, depending on their original movement direction.

Environmental variables

One of the most important steps in modelling species distributions is careful variable

selection, particularly when using predictors when there is a degree of a priori expectation of

ecological relevance (Elith & Leathwick, 2009; Burnham & Anderson, 2010). Many

physical, geographical and ecological characteristics of the study streams were very similar.

The catchments were all effectively completely forested by natural vegetation and the

streamlines had qualitatively similar, near-pristine riparian zones (CENRM, 2005; Strategen,

2006). Furthermore, all streams were highly consistent in terms of overall channel

morphology, slope, altitude, dissection and soil type (CENRM, 2005; Strategen, 2006).

Finally, all four study sites were situated within 14 km of each other and each was located at

a similar distance from the main channel of the Blackwood River with no artificial instream

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barriers existing in the streams (Figure 1). Therefore, given the considerable spatial and

temporal sampling effort undertaken, we focussed on other environmental variables that are

commonly associated with freshwater fish distributions and movement. These were

temperature, conductivity, pH, dissolved oxygen, and discharge (Matthews, 1998). In order

to determine the spatial and temporal patterns in water temperature and their relationship to

air temperature, data loggers (TinytagTM Gemini Data Loggers, Chichester, West Sussex,

UK) were placed in situ in all the tributary sites and two main channel sites and were

programmed to log water temperature every three hours. Temperature loggers were

downloaded periodically and the water temperature regimes of the various sites were

graphically compared to maximum daily air temperature obtained from the nearest Australian

Bureau of Meteorology weather station (Bridgetown). On each sampling event at each site,

instantaneous temperature, conductivity, pH, and dissolved oxygen were measured using a

handheld water quality meter (Oakton PCD650) in the middle of the water column at three

locations and a mean (± 1 SE) determined.

Flow velocity was measured using a hand held flow meter (Global F101) and

instantaneous discharge measurements were determined in all tributary sites during each

sampling event by undertaking cross-sectional depth-flow profiles (measured every 0.5m)

and applying the formula: Dinst.=VxDxW where Dinst. = instantaneous discharge m3.sec-1, V =

mean water velocity (m.sec-1) over the cross-section, D = mean depth (m) over the cross-

section, W = width (m) of the cross-section.

Statistical analysis

In order to determine the spatial and temporal patterns of movements and demographics of

the five most commonly captured native freshwater fishes in the tributary sites, i.e. the

western minnow Galaxias occidentalis, western mud minnow Galaxiella munda, western

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pygmy perch Nannoperca vittata, Balston’s pygmy perch Nannatherina balstoni, and

nightfish Bostockia porosa (see Results), the mean (±1 SE) numbers of each species captured

moving upstream and downstream over the three 24 hour periods (numbers were standardised

from the proportion of cross-section of the stream covered by each fyke to assume 100%

coverage) in each sampling event at each site were determined. To elucidate population

structure of each species and best illustrate movement patterns in relation to life-history stage,

the monthly length-frequency distribution (separated on movement direction) of the above

freshwater fishes throughout the study were plotted separately for Milyeannup Brook (G.

occidentalis, N. balstoni, N. vittata, and G. munda), Rosa Brook (G. occidentalis, N. vittata,

and B. porosa) and Layman Brook (B. porosa).

To test the effects of each species captured and the sampling month on fish captures, we

used general linear models (GLMs) with a repeated measures design. The log10(x+1)

transformed number of fish captured in the fykes moving in either an upstream or

downstream direction at each sampling event was used as a response variable, with species as

a between-subject fixed effect, months as a within-subject fixed effect and tributary nested

within species as a random effect.

The relationships between the upstream and downstream movement of the five native

species in tributaries and the prevailing environmental variables in those systems were

explored by first determining the mean of each environmental variable and the overall mean

upstream and downstream movements of each species in each system over the major flow

period in each year (i.e. the period that all four tributaries were flowing between August to

December). As excessive flooding of the tributary sampling sites occurred in August 2009,

they were excluded from the above analysis.

Bivariate correlations (using Pearson’s correlation coefficient) were calculated for each

pair of environmental variables (all except pH log10 transformed) to examine possible inter-

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relationships. Mean velocity was consistently highly correlated with discharge and was

therefore excluded from subsequent analyses; no other variable pairs showed correlations >

0.6 (Table S1). To determine which environmental variables explained most of the variation

in upstream and downstream migrations of each fish species during the major flow period, we

used GLMs, with a multimodel inference approach (Burnham & Anderson, 2010). For each

fish species and for each movement direction, a global model of number of fish captured

(log10(x+1) transformed) against all environmental predictor variables was fitted using the R

program (R Development Core Team, 2013). The global model was then used to generate a

set of models containing all combinations of predictor variables, using the R package MuMIn

(Bartoñ, 2013). Models were ranked by the Akaike Information Criterion, corrected for small

sample size (AICc) and model averaging performed across those models that were within four

AICc values of the best model (Burnham & Anderson, 2010), using MuMIn. The relative

importance of each environmental predictor variable was determined by summing the Akaike

likelihood weights across all models within the top-ranked set; this gives the selection

probability that a given variable will appear in the AIC-best model (Burnham & Anderson,

2010). For each fish species and each movement direction, we report the relative importance

of all of the environmental predictor variables, along with the model-averaged parameter

estimates and their standard errors.

As estimates of parameters and variable importance from model averaging can be

adversely affected by collinearity among predictors (Murray & Conner, 2009; Freckleton,

2011), we attempted to account for this in three ways. First, only predictor variables with

bivariate correlations ≤ 0.6 with all other predictors were used in the analysis (see Table S1).

Second, for each predictor we calculated the variance inflation factor (VIF), which represents

the overall correlation with all other predictors, using the R package HH (Heiberger, 2013).

As a general rule of thumb, VIF’s > 10 are thought to indicate excessive collinearity (Myers,

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1990; although see O’Brien, 2007 for a contrary view). Finally, we also used a hierarchical

partitioning approach to determine variable importance, as this has been shown in some

simulations to outperform model averaging when predictor variables are correlated (Murray

& Conner, 2009; although see Smith et al., 2009). R2 values were hierarchically partitioned

to determine the proportion of variance explained independently and jointly by each variable

(Chevan & Sutherland, 1991; Mac Nally, 2002), and the independent effect of each predictor

was calculated as the percentage of total independent contributions, using the R package

hier.part (Walsh & Mac Nally, 2013). The significance of independent effects was assessed

with Z scores from a comparison of observed effects with those obtained after 500

randomisations of the data, using hier.part.

Results

Environmental variables

Water and air temperatures were highly associated at all sites (Fig. 2). Temperatures in the

main channel sites were much higher than tributary sites during summer and autumn, with the

greatest differences occurring between main channel sites and the perennial (groundwater

maintained) Milyeannup Brook, which experienced a less variable temperature regime than

other tributaries. For example, between late December and March (i.e. summer to early

autumn) in 2006 and 2008, Milyeannup Brook was generally >1°C cooler than the non-

perennial Rosa Brook and McAtee Brook, and >6°C cooler than the main channel sites.

Between May and August (i.e. late autumn to late winter) 2006, Milyeannup Brook was ~2°C

warmer than the non-perennial systems and was similar in temperature to the main channel

sites (Fig. 2).

Mean conductivity in the main channel sites increased in an upstream direction, ranging

from 3564 to 4955 µS.cm-1 at the Denny Rd and Jalbarragup sites, respectively (Table S2).

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In contrast, the tributary sites were always fresher with mean values that did not exceed 500

µS.cm-1 (Table S2). Mean dissolved oxygen at all sites was >6.00 mg.l-1 and pH in the

tributary sites was slightly acidic (means from 6.04-6.67), whereas pH was near neutral in the

main channel sites (means from 7.18-7.33) (Table S2).

Species captures and spatial distributions

Six of the nine endemic freshwater fishes of south-western Australia, along with four

estuarine fishes, three introduced fishes and one native anadromous agnathan were captured

during the study. The six endemic freshwater fish species were the galaxiids G. occidentalis

and G. munda, the percichthyids N. vittata, N. balstoni and B. porosa and the plotosid

Tandanus bostocki (Table S3).

Galaxias occidentalis was the most commonly recorded species, present in 92.2% of all

sites sampled in all months, with N. vittata, B. porosa, and T. bostocki recorded at similar

frequencies (~66-67%), N. balstoni at 38.5%, and G. munda (16.5%) being the least

encountered (Table S3). Stark contrasts existed in captures of the various species between

main channel and tributary sites, with five of the six endemic freshwater species more

commonly encountered in the tributary sites (Table S3). The most difference in habitat usage

was recorded for G. munda (32.4 cf 2.7% in tributary and main channel samples,

respectively) (Table S3). Galaxiella munda was recorded in tributary sites that were cooler,

fresher, and had lower mean discharge than habitats occupied by all other species (Table S3).

Galaxias occidentalis was recorded in similar frequency between the two habitat types (i.e.

95.2 and 90.3%, in tributary and main channel sites, respectively) whereas T. bostocki was

much more frequently recorded in main channel (99.0%) compared with tributary (30.5%)

sites (Table S3).

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Patterns of species movements

There were clear spatial and temporal patterns of movements of the five endemic freshwater

fish species that were most commonly encountered in the tributaries (i.e. G. occidentalis, G.

munda, N. balstoni, N. vittata and B. porosa). Upstream fish movements were significantly

affected by species (F = 8.25, P = 0.0003) and month of sampling (F = 6.14, P < 0.0001),

with a significant interaction between species and month (F = 2.51, P < 0.0001).

Downstream movement did not quite differ significantly among species (F = 2.79, P = 0.05),

but was significantly affected by the month of sampling (F = 8.63, P < 0.0001), with a

significant interaction between species and month (F = 3.27, P < 0.0001).

Details of the habitat utilisation, spatial and temporal movements and population

structure of the five fish species in each sampling month at each site are shown in the

supporting information (Table S3, Figs S1-S13). Schematic diagrams of the movement

patterns of adults and juveniles of the two species of galaxiids and three species of

percichthyids (Figs 3 and 4) are based on the data presented in the Supporting Information.

While interspecific variation existed in migration patterns (spatially and temporally), all but

G. munda moved laterally between the main channel of the Blackwood River and its

tributaries (Figs 3 and 4).

Downstream movements of G. occidentalis occurred predominately within tributary sites

between August and December (i.e. late winter to early summer), with an upstream

directional movement occurring in main channel sites between June and August (i.e. winter)

(Figs 3 and S1). Downstream movement of juveniles (<40 mm TL) generally occurred

annually between October to December, with a notable one to two month earlier movement

occurring in the perennial Milyeannup Brook compared to Rosa Brook (Figs 3, S2 and S3).

Larger individuals (>120 mm TL) were also much more prevalent in Milyeannup Brook

compared with the seasonally flowing systems. Unlike all other species, there were no strong

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movement patterns associated with G. munda within the tributaries (Figs 3 and S4).

However, downstream movement of juveniles (<35 mm TL) were recorded during late spring

(October and November) in most years in Milyeannup Brook (Fig. S5). Notably, G. munda

was the only species found to be effectively absent from the main channel (Fig. 3 and Table

S3).

Nannatherina balstoni only utilised Milyeannup Brook as a breeding and nursery area

and was recorded very infrequently in all other tributaries (Figs 4 and S6). A clear annual

downstream movement of juveniles (<50 mm TL) occurred in that system between

November and December (i.e. late spring to early summer) (Fig. S7). Although this species

was not abundant in the main channel sites (Fig. 4 and Table S3), it was recorded

occasionally at the Denny Rd and Milyeannup Pool sites; both located downstream of the

confluence of Milyeannup Brook (Figs 1 and S6). Nannoperca vittata migrated within all

tributaries (aside from Layman Brook), with both upstream and downstream directional

movements of the species occurring between August and December (Figs 4 and S8). The

species was also recorded to lesser degree moving within the main channel sites, with the vast

majority moving in a downstream direction (Fig. S8). Juvenile (<30 mm TL) N. vittata were

recorded moving downstream within tributaries and the timing of that movement varied

between systems. For example, it occurred mostly in March (with some November and

December movement of <20 mm TL fish) in the perennial Milyeannup Brook, compared

with August to November in the ephemerally flowing Rosa Brook (Figs S9 and S10). Large

downstream movements of juvenile (<30 mm TL) B. porosa occurred in a number of

tributaries between November and December (Figs 4 and S11) and were most prevalent in

Milyeannup, Rosa and Layman Brooks, where it was recorded in all years aside from 2009 in

Layman Brook (Figs S11, S12, S13).

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Relationships between environmental variables and movement patterns

The influence of environmental variables on the average upstream and downstream

movements of the five fish species over the major flow periods in the four tributaries between

2005 and 2009 are shown in Tables 1 and 2. Variance inflation factors were relatively low

for all variables (temperature 3.11; conductivity 3.31; pH 1.95; dissolved oxygen 3.29;

discharge 1.83) and both model averaging and hierarchical partitioning analyses gave very

similar indications of variable importance (Tables 1 and 2). This suggests that these

measures were not biased by collinearity among predictors. Discharge ranked as the most

important predictor variable in four of the five species for both upstream and downstream

movement, being included in between 43-100% of the best fitting models (Tables 1 and 2).

In particular, discharge explained a substantial proportion of the upstream movements of the

three most widespread and abundant species, N. vittata, B. porosa and G. occidentalis (Table

1). With the exception of one case (i.e., downstream movement in B. porosa), more fish were

captured with greater discharge levels. Temperature also figured prominently as an important

predictor variable, particularly for N. vittata, with more fish captured at lower temperatures.

Of the other variables, pH was positively associated with upstream movements, and dissolved

oxygen was negatively associated with downstream movements of B. porosa.

Discussion

There has been limited research on potamodromous fishes that migrate laterally between

main channel and low order streams, compared to those that undertake longitudinal

migrations within the main streams of large river networks (e.g. O’Connor et al. 2005; Dugan

et al., 2010) or between lentic and lotic systems (e.g. Barthel et al., 2008; Conallin et al.,

2011). The new recruits of five of the six endemic freshwater fishes found within the study

area were recorded almost exclusively within tributaries, and their downstream movement

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into main channel habitats occurred for four of the five species during periods of declining

surface flows (i.e., late spring and early summer). As with other climatic regions with

pronounced dry seasons, most lower order tributaries in south-western Australia either dry

completely or consist of separated, often isolated refuge pools during summer and autumn.

Therefore, our findings strongly suggest that most species are potamodromous and undergo

lateral spawning migrations into lower order tributaries during the seasonal flow period.

Discharge was clearly the most important environmental variable related to the upstream and

downstream movement of most species and therefore our findings have considerable

implications in understanding how life-cycles of potamodromous fishes in drying temperate

regions may be impacted by climate change.

Influence of hydrology on life-histories

Hydrologic regime has been shown to significantly influence the life-history strategies of

freshwater fishes in lotic systems over broad biogeographic scales (Mims & Olden, 2012;

Chessman, 2013; Sternberg & Kennard, 2013). Due to the highly seasonal and predictable

flow regimes of aquatic systems in south-western Australia, its freshwater fishes would be

expected to have periodic life history strategies (Mims & Olden, 2012). Consistent with this

expectation is that most species have annual spawning periods that coincide with peak flow in

winter and spring (e.g. Pen & Potter, 1991; Morgan et al., 1998; this study). Periodic species

are also characterised by having large body size, late maturation, and high fecundity. Most of

the freshwater fishes of south-western Australia, however, have many traits that would

classify them as opportunistic (r strategists), such as small body sizes (<190 mm TL) and

early maturation (most in the first or second year of life) (Morgan et al., 1998). This may be

due to these species having evolved in relatively small, seasonally intermittent, naturally

unproductive aquatic systems typical of this region (Bunn & Davies, 1990), which would

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limit their maximum sizes and reproductive investment. Despite having several opportunistic

life-history traits, the movements of most species were undoubtedly associated with spawning

and recruitment and they would be therefore vulnerable to reductions in amount of surface

discharge.

Rivers that are free-flowing in relatively undeveloped basins are predicted to be resilient

to climate change compared with those that are more highly regulated (Palmer et al., 2008).

Artificial instream barriers can have profound, wide-ranging impacts on aquatic ecosystems,

particularly on migratory fishes (Bunn & Arthington, 2002; Katopodis & Aadland, 2006).

Most other studies relating to anthropogenic impacts on the migrations of potamodromous

fishes have focussed on the impacts of barriers and river connectivity (e.g. Baumgartner &

Harris, 2007; Mallen-Cooper & Brand, 2007; Branco et al., 2012). While our study systems

are free from artificial barriers, our findings suggest that the numerous natural instream

barriers that exist elsewhere throughout the region may also be disrupting the life-cycles of

these potamodromous fishes by restricting access to upstream spawning habitats exacerbating

the impacts of past and projected declines in surface discharge.

Several species in the current study are known to also occupy natural and artificial lentic

systems (Morgan et al., 1998; Beatty et al. 2013) and this suggests they have a degree of

habitat flexibility. This could help at least some populations survive the expected ongoing

reductions in flow period and surface discharge due to climate change by utilising main

channel lotic habitats or artificial lentic habitats for spawning and recruitment. Nonetheless,

these projected hydrological changes will undoubtedly reduce recruitment to some degree

through reductions or loss of spawning habitats in lower order streams.

Galaxiids are generally known to migrate for spawning purposes (McDowall &

Frankenberg, 1981; Baker & Montgomery, 2001; Baker & Hicks, 2003; Morgan, 2003) and

G. occidentalis in the current study showed the strongest movement patterns; migrating

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laterally into all tributaries to spawn, with this movement positively associated with amount

of stream discharge during the annual flow period. As opposed to galaxiids, pygmy perches

are generally not known to be strong migrators (e.g. Hammer et al., 2009), although

movement between water bodies has been linked to flooding events (Morgan et al., 1995;

Hughes et al., 1999; Knight & Arthington, 2008). We demonstrated that the two pygmy

perches were also potamodromous and that the upstream spawning movements of N. vittata

and the other percichthyid, B. porosa, were positively associated with amount of surface

discharge. Furthermore, hydrological seasonality would explain why N. balstoni, the most

threatened species in the study, only breeds within the groundwater maintained perennial

Milyeannup Brook, despite other tributaries appearing to provide equally suitable habitat

when they are flowing. The onset of breeding by N. balstoni is earlier than others in this

study (i.e. from early winter cf late winter and spring (Morgan et al., 1995, 1998)), requiring

its spawning habitat to be available at a time when seasonal tributaries are yet to flow.

Ongoing changes in both the amount and timing of discharge may therefore have serious

implications for this and other potamodromous species.

Declines in refuge habitats due to climate change

Understanding refuge habitat requirements is crucial in assessing impacts of climate change

on freshwater fishes in seasonal systems (Hodges & Magoulick, 2011; Morán-López et al.,

2012; Davis et al., 2013). System size and the volume and periodicity of discharge, along

with a number of other abiotic characteristics, such as anthropogenic pollution, have been

demonstrated to influence refuge availability and use by freshwater fishes (e.g. Magoulick &

Kobza, 2003; Hodges & Magoulick, 2011; Morán-López et al., 2012). Groundwater

maintained aquatic refuges can be thought of as ‘evolutionary refugia’ and are extremely

important in maintaining endemic aquatic species (Davis et al., 2013). Our study identified

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that the key refuge habitats for freshwater fishes during the dry season in the Blackwood

River were the secondarily salinised main channel and a fresh perennial tributary; both reliant

on fresh groundwater.

Pollution caused by human activities is one of the major global stressors on riverine

ecosystems (Vörösmarty et al., 2010). Secondary salinisation of rivers is a major threat to

aquatic ecosystem health and climate change is expected to amplify the problem, particularly

in climates that experience prolonged periods of drought (Cañedo-Argüelles et al., 2013).

Salinisation in south-western Australia has caused fundamental ecosystem change, including

widespread range contractions of many stenohaline freshwater fishes and concomitant

colonisation by estuarine euryhaline species (Morgan et al., 2003; Mayer et al., 2005; Beatty

et al., 2011). Reduced rainfall is projected to cause severe reductions in surface flow and

fresh groundwater levels in many catchments (Barron et al., 2012; Silberstein et al., 2012),

which may cause remnant fresh perennial tributaries to be lost and further increase the

salinity in main channel refuges to levels that may exceed the tolerances of several species

(Beatty et al., 2011). In order to model future distributions and identify viable refuges in this

region and elsewhere, robust analyses of physicochemical thresholds of freshwater fishes are

required (e.g. Segurado et al., 2011).

The effect of interactions between pollution and other stressors on aquatic ecosystems

can be complex and difficult to predict (Cañedo-Argüelles et al., 2013). For example,

although reductions in rainfall and fresh surface flow may reduce the dilution capacity of

rivers in south-western Australia, this may be partially offset by concomitant reductions in

saline groundwater levels (Mayer et al., 2005). Specifically, the disconnection of saline

groundwater with low order streams due to rainfall decline results in a disproportionally

greater reduction in streamflow and also a reduction in the stream salinity (Kinal &

Stoneman, 2012). Given the vital importance of dry-season refuge in the life-cycles of

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potamodromous fishes in temperate environments, the impact of climate change on its

availability and quality warrants much greater research attention.

Impacts of climate change on migrations

Globally, reductions in the discharge from rivers due to water extraction and climate change

is predicted to cause a loss of up to 75% of freshwater fish diversity by 2070 (Xenopoulos et

al., 2005; Xenopoulos & Lodge, 2006). However, the threat of climate change to freshwater

ecosystems on the Australian continent, particularly freshwater fishes, has only recently

begun to be investigated (Koehn, 2011; Morrongiello et al., 2011a,b; Chessman, 2013).

There is universal agreement amongst GCMs that rainfall will continue to decline in south-

western Australia (Suppiah et al., 2007; Silberstein et al., 2012). The dramatic past and

projected climatic shift that has occurred in this region renders it a ‘living experiment’ in

assessing the ecological impacts of climate change in temperate regions.

We found that spawning migrations of potamodromous fishes in the current study were

almost always positively related to discharge. Our study streams were remarkably similar in

most geophysical and many ecological variables and we therefore examined only a modest

number of potential predictors of migrations. A study with a wider geographical range may

reveal other environmental variables that also influence migrations in these potamodromous

fishes. Nevertheless, the projected reductions in annual surface water discharge and stream

flow periods of the region (Silberstein et al. 2012) will reduce spawning habitat availability

and connectivity for freshwater fishes on both a temporal and spatial scale. The amount of

river discharge has previously been shown to be positively associated with migration,

reproduction and recruitment of fishes elsewhere (e.g. Reynolds, 1983; O’Connor et al.,

2005; Beatty et al., 2010; Crook et al., 2010; Walsh et al., 2012; Munz & Higgins, 2013).

Our findings suggest that potamodromous fishes may be particularly impacted by flow

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reductions via reduced levels of recruitment and increased resource competition, with

potentially serious implications for population dynamics and viabilities (e.g. Morán-López et

al., 2012).

Along with causing flow reductions, climate change is projected to increase mean air

temperatures and the number of extreme hot days (>35°C) in south-western Australia

(Suppiah et al., 2007). As expected, water temperatures in the current study were closely

linked to air temperatures. As temperature affects the aerobic scope of fishes, which dictates

their swimming performance (Farrell et al., 2008), these increases may further restrict the

migratory capacity and population viability of potamodromous species in this region. Hague

et al. (2011) modelled the effects of climate change on sockeye salmon populations and

concluded that increasing temperatures would result in most being forced to migrate at sub-

optimal aerobic scopes. The geographical location of south-western Australia (i.e. the

southern edge of the Australian continent) means that aquatic species including freshwater

fishes cannot migrate southward in response to increasing temperatures and this greatly

increases their vulnerability to climate change (Davies, 2010).

In a recent assessment of how 39 species of south-eastern Australian fishes fared under a

‘Millennium’ drought, the most impacted species had a small maximum size, were

invertivorous (cf omnivorous), had a low age at maturity, spawned in low temperatures, had a

prolonged spawning period, low fecundity, demersal eggs and a low upper thermal limit

(Chessman, 2013). The species in the current study have almost all of these vulnerable traits

(Morgan et al. 1998). There is a paucity of research into how Australian freshwater fishes

affect aquatic food webs (Stoffels, 2013). The lack of any comprehensive food-web studies

that have incorporated freshwater fishes in south-western Australia precludes predicting what

impact their decline will have on the prevailing aquatic ecosystems. However, the loss of

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such highly abundant top-order consumers (Morgan et al., 1998) may cause considerable

ecological change through shifts in prey populations and trophic cascades.

Protecting freshwater biodiversity, particularly migrating fauna, is extremely challenging

as freshwater ecosystems are impacted by activities that occur on terrestrial and aquatic

landscapes and there is a fundamental ‘thirst’ by humans for fresh water (Dudgeon et al.

2006). Altered flow regimes caused by climatically induced reductions in runoff and water

abstraction coupled with pollution, particularly salinisation, are acting synergistically in

south-western Australia. Setting thresholds of declines in discharge in order to guide

management intervention has been undertaken globally but can also be undertaken on a more

local scale to increase sensitivity (Palmer et al., 2008). Our study suggests that spawning

migrations of potamodromous fishes will inevitably decline, leading to reduced recruitment

and threatening the viability of populations and species. Such hydro-ecological research is

urgently required elsewhere to better understand the resistance and resilience of

potamodromous fishes to climate change (see Moilanen et al., 2008; Williams et al., 2011).

Given the recent climatic shift and consensus among GCMs of continued severe drying

in south-western Australia, there is little excuse not to undertake proactive management

strategies and adapt our responses to buffer the impacts on aquatic ecosystems (see Palmer et

al., 2009). Remediation strategies to re-connect aquatic habitats may help mitigate the

impact of flow declines (Palmer et al., 2008, 2009; Beatty et al., 2013). Addressing existing

stressors on freshwater ecosystems in drying climatic regions at their source, such as

undertaking more sustainable land use practices and approaches to both surface water and

groundwater resource development (Xenopoulos et al., 2005; Xenopoulos & Lodge, 2006;

Palmer et al., 2008, 2009; Vörösmarty et al., 2010; Davis et al., 2013) must also be a priority

if the current and projected impacts of climate change are to be reduced.

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Acknowledgements

This project was jointly funded by the Department of Water Government, of Western

Australia and the Australian Government under its Water for the Future Plan. Special thanks

to Adrian Goodreid, Natasha Del Borrello and Rob Donohue (Department of Water,

Government of Western Australia) for project management. Many thanks also to the

Department of Water Bunbury staff including Ash Ramsay, Richard Pickett and Andrew

Bland for hydrological data. We are grateful for the field assistance that was provided by

Fiona and Andrew Rowland, Travis Fazeldean, Simon Visser, Mark Allen, David John,

Joshua Johnston and Trine Beatty. Thanks also to Adrian Gleiss for reviewing earlier drafts

of the manuscript, and to three anonymous referees and Rhea Kressman, who made many

constructive suggestions which substantially improved the manuscript. The study complied

with permits provided by the Murdoch University Animal Ethics Committee, the Department

of Environment and Conservation and the Department of Fisheries, Government of Western

Australia.

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SUPPORTING INFORMATION Table S1 Matrix of Pearson product-moment correlations among environmental predictor

variables (all except pH log10 transformed).

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Table S2 Water quality at sites sampled in the Blackwood River.

Table S3 Prevalence of species and the water quality at their capture sites in the Blackwood

River.

Fig. S1 Mean number of Galaxias occidentalis moving in the tributary and main channel

sites of the Blackwood River. NF indicates that the system was not flowing and therefore

was not sampled.

Fig. S2 Length-frequency distributions of Galaxias occidentalis in Milyeannup Brook.

Fig. S3 Length-frequency distributions of Galaxias occidentalis in Rosa Brook.

Fig. S4 Mean number of Galaxiella munda moving in the tributary sites of the Blackwood

River. NF indicates that the system was not flowing and therefore was not sampled.

Fig. S5 Length-frequency distributions of Galaxiella munda in Milyeannup Brook.

Fig. S6 Mean number of Nannatherina balstoni moving in the tributary and main channel

sites of the Blackwood River. NF indicates that the system was not flowing and therefore

was not sampled.

Fig. S7 Length-frequency distributions of Nannatherina balstoni in Milyeannup Brook.

Fig. S8 Mean number of Nannoperca vittata moving in the tributary and main channel sites

of the Blackwood River. NF indicates that the system was not flowing and therefore was not

sampled.

Fig. S9 Length-frequency distributions of Nannoperca vittata in Milyeannup Brook.

Fig. S10 Length-frequency distributions of Nannoperca vittata in Rosa Brook.

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Fig. S11 Mean number of Bostockia porosa moving in the tributary and main channel sites

of the Blackwood River. NF indicates that the system was not flowing and therefore was not

sampled.

Fig. S12 Length-frequency distributions of Bostockia porosa in Rosa Brook.

Fig. S13 Length-frequency distributions of Bostockia porosa in Layman Brook.

Table 1 Association between environmental variables and number of fish moving upstream for five freshwater fish species over the major flow period in four tributaries of the Blackwood River. All variables except pH were log10 transformed. Variables are ranked by relative importance, with coefficient estimates and standard errors from model averaging. The percentage independent contribution of each environmental variable to total explained variance was determined by hierarchical partitioning of R2 values. Significant predictors (P < 0.05) from model averaging are shown in bold and from hierarchical partitioning with an asterix. Response Environmental

variable Importance Estimate SE Independent

effect (%)

N. balstoni Discharge 0.44 0.44 0.29 38.3

Conductivity 0.39 -2.49 1.70 35.2

pH 0.19 -0.27 0.42 11.7

Dissolved oxygen 0.16 -0.93 2.52 9.1

Temperature 0.15 -1.97 5.10 5.7

N. vittata Discharge 0.98 1.00 0.27 49.8

Temperature 0.81 -12.02 4.48 25.3

Conductivity 0.20 1.90 1.71 14.0


pH 0.10 0.14 0.41 5.4

B porosa Discharge 0.99 0.46 0.12 63.0*

pH 0.57 0.29 0.13 23.9

Conductivity 0.34 -1.13 0.68 4.6

Temperature 0.13 -1.40 1.99 2.2

Dissolved oxygen 0.10 0.02 1.01 6.3

G. occidentalis Discharge 0.86 0.69 0.26 61.8

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Conductivity 0.54 -2.89 1.49 23.5

pH 0.17 --0.23 0.41 5.6


Temperature 0.13 -2.45 2.56 2.9

G. munda Temperature 0.62 -1.04 0.51 42.5*

Discharge 0.34 0.07 0.03 43.1*

Conductivity 0.16 0.11 0.19 5.5


pH 0.13 0.01 0.04 2.1

Table 2 Association between environmental variables and number of fish moving upstream for five freshwater fish species over the major flow period in four tributaries of the Blackwood River. All variables except pH were log10 transformed. Variables are ranked by relative importance, with coefficient estimates and standard errors from model averaging. The percentage independent contribution of each environmental variable to total explained variance was determined by hierarchical partitioning of R2 values. Significant predictors (P < 0.05) from model averaging are shown in bold and from hierarchical partitioning with an asterix.

Response Environmental variable

Importance Estimate SE Independent effect (%)

N. balstoni Discharge 0.63 0.52 0.55 26.3

Conductivity 0.33 -4.60 3.32 37.9

pH 0.22 -0.77 0.78 20.7

Temperature 0.16 -6.10 9.69 6.3


N. vittata Discharge 1.00 1.15 0.22 51.1*

Temperature 0.98 -14.84 3.81 31.7*

Conductivity 0.15 1.33 1.39 8.8


pH 0.08 0.00 0.30 2.6

B porosa Dissolved oxygen 0.89 3.87 1.31 62.7*

Discharge 0.84 -0.19 0.17 21.2

Temperature 0.11 -0.02 4.21 9.9

Conductivity 0.10 0.26 1.15 4.5

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pH 0.09 0.02 0.23 1.7

G. occidentalis Discharge 0.55 0.59 0.32 38.6*

Conductivity 0.43 -3.35 2.03 22.3

pH 0.41 -0.69 0.43 23.5


Temperature 0.15 -4.00 5.72 6.8

G. munda Discharge 0.99 0.13 0.10 38.8

Conductivity 0.42 -1.01 0.58 33.9

pH 0.33 -0.20 0.13 18.7

Temperature 0.17 -1.58 1.60 5.0


FIGURE LEGENDS

Fig. 1 Sites sampled in the Blackwood River and its tributaries to determine the movement

patterns of south-western Australian freshwater fishes.

Fig. 2 Water temperatures at the sites sampled in the Blackwood River and its tributaries.

Note the close association between the water temperatures in all sites and the maximum daily

air temperature.

Fig. 3 Schematic diagram of the temporal and spatial movement patterns of two species of

Galaxiidae in the main channel of the Blackwood River and in a seasonal and a perennial

tributary, south-western Australia. Direction of movement (upstream or downstream) is

indicated by the orientation of fish. New recruits are indicated by smaller sized fish.

Fig. 4 Schematic diagram of the temporal and spatial movement patterns of three species of

Percichthyidae in the main channel of the Blackwood River and in a seasonal and a perennial

tributary, south-western Australia. Direction of movement (upstream or downstream) is

indicated by the orientation of fish. New recruits are indicated by smaller sized fish.

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