Extraordinary micro-endemism in Australian desert spring amphipods

Molecular Phylogenetics and Evolution 66 (2013) 645–653

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Extraordinary micro-endemism in Australian desert spring amphipods N.P. Murphy a,⇑, M. Adams b,c, M.T. Guzik c, A.D. Austin c a

Department of Genetics, La Trobe University, Bundoora, Victoria 3086, Australia Evolutionary Biology Unit, South Australian Museum, North Terrace, Adelaide, South Australia 5000, Australia c Australian Centre for Evolutionary Biology and Biodiversity, School of Earth and Environmental Science, The University of Adelaide, South Australia 5005, Australia b

a r t i c l e

i n f o

Article history: Received 20 January 2012 Revised 12 October 2012 Accepted 14 October 2012 Available online 7 November 2012 Keywords: Aquatic invertebrate Conservation genetics Desert springs Endemism Evolutionarily significant unit Groundwater dependent ecosystem

a b s t r a c t Increasing pressure for water in the Australian arid zone is placing enormous stress on the diverse endemic communities inhabiting desert springs. Detailed information about the evolutionary processes occurring within and between individual endemic species will help to develop effective and biologically relevant management strategies this fragile ecosystem. To help determine conservation priorities, we documented the genetic structure of the endemic freshwater amphipod populations in springs fed by the Great Artesian Basin in central Australia. Phylogenetic and phylogeographic history and genetic diversity measures were examined using nuclear and mitochondrial DNA from approximately 500 chiltoniid amphipods across an entire group of springs. Pronounced genetic diversity was identified, demonstrating that levels of endemism have been grossly underestimated in these amphipods. Using the GMYC model, 13 genetically divergent lineages were recognized as Evolutionarily Significant Units (ESUs), all of which could be considered as separate species. The results show that due to the highly fragmented ecosystem, these taxa have highly restricted distributions. Many of the identified ESUs are endemic to a very small number of already degraded springs, with the rarest existing in single springs. Despite their extraordinarily small ranges, most ESUs showed relative demographic stability and high levels of genetic diversity, and genetic diversity was not directly linked to habitat extent. The relatively robust genetic health of ESUs does not preclude them from endangerment, as their limited distributions ensure they will be highly vulnerable to future water extraction. Ó 2012 Elsevier Inc. All rights reserved.

1. Introduction Due to their isolated island-like nature, desert springs often consist of unique assemblages that exhibit local specialization, very high levels of endemism, plus highly restricted and genetically structured distributions (Fensham et al., 2010). Worldwide these unique ecosystems are under increasing threat due to landscape modification and increasing use of groundwater in arid regions (Kodric-Brown and Brown, 2007). The desert springs fed by the Great Artesian Basin (GAB springs) represent one of Australia’s most threatened habitats. In Australia the use of groundwater has increased exponentially within the last 150 years, with extraction in some areas exceeding recharge and drastically impacting on the groundwater dependent ecosystems (Nevill et al., 2010). The GAB springs support a suite of endemic plants, and surface dwelling invertebrates and fish. However, many springs are degraded or no longer flow, and, as a consequence, the springs are now listed as an endangered community under the Australian Environment Protection and Biodiversity Protection Act. Despite this listing, only a very small portion of GAB springs are protected in National Parks ⇑ Corresponding author. E-mail address: [email protected] (N.P. Murphy). 1055-7903/$ - see front matter Ó 2012 Elsevier Inc. All rights reserved. http://dx.doi.org/10.1016/j.ympev.2012.10.013

or reserves. The remainder are on pastoral land, where many springs are afforded little to no protection (Fensham and Price, 2004). Most aspects of the biological ecosystems supported by the GAB springs are not currently well understood. Even at the level of species diversity, information is lacking for the composition of threatened GAB spring communities, let alone the conservation status of individual spring species. This lack of information is critical, since the extinction of a single species could have a serious impact on any ecological community comprised of only a small number of species (Purvis and Hector, 2000). Due to the fragmented nature of groundwater dependent ecosystems, populations are often highly geographically structured and cryptic species are very common (Juan et al., 2010). Genetic studies are particularly useful for the conservation management of fragmented systems, since they can provide pivotal insights into the geographic and temporal structure of genetic diversity both within and between species (Echelle et al., 1989; Juan et al., 2010; Witt et al., 2006). Management decisions are likely to have greater biological relevance when they incorporate genetic information such as the existence of unique evolutionarily significant units (ESUs), the drivers of genetic diversity within a species or population, and how migration and dispersal occurs across fragmented landscapes.

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This study used DNA sequence data to examine the phylogenetic, phylogeographic and demographic structure of freshwater chiltoniid amphipods inhabiting the GAB springs surrounding Lake Eyre in central Australia (Fig. 1). Prior to 2009, it was thought that a single widespread species was distributed across approximately 2000 springs (Mitchell, 1985; Murphy et al., 2009) implying that these amphipods were efficient dispersers. As human impacts on springs generally occur at a local level only, such a species would not be extensively threatened by local habitat degradation. However, it is now apparent that GAB spring amphipod diversity has been drastically underestimated, with recent molecular and morphological studies finding that multiple amphipod genera and species are present (King, 2009; Murphy et al., 2009). Coupled with this increased species diversity, significant population sub-structure and poor dispersal is also evident between populations occupying springs less than 5 km apart in the one amphipod taxon studied in detail thus far (Murphy et al., 2010). Poor dispersal and confinement to a discontinuous habitat is a common feature of desert spring invertebrates (Hershler et al., 1999; Liu and Hershler, 2007; Ponder et al., 1995; Witt et al., 2006), leading to extreme isolation of spring populations, or the ‘‘Death Valley’’ model (DVM) of population structure (Meffe and

Vrijenhoek, 1988). The DVM predicts that there will be a high degree of very localized population structure, due to the complete absence of both long distance dispersal and isolation by distance (Meffe and Vrijenhoek, 1988). Moreover, by virtue of populations being completely isolated, the DVM predicts a drastic reduction in genetic diversity within populations, with smaller isolated populations being less genetically diverse than larger, less-isolated ones. This study tested two hypotheses related directly to the isolated GAB spring amphipods and the DVM. The patchy distribution of the GAB springs means that amphipods are present both in very small and isolated clusters of 2 or 3 springs and in large clusters of 100s of interconnected springs. Given this distribution of habitat, our two hypotheses are that (a) genetically distinct amphipod lineages should be completely restricted to geographically isolated groups of springs and (b) extremely restricted lineages will have reduced genetic diversity when compared with those occupying larger and more complex spring clusters. The expectation of a poorly dispersing and highly biodiverse species complex means that it is critical to examine the genetic diversity and structure of GAB spring amphipods across their entire range. Clearly, any species with reduced genetic diversity and restricted to a very small habitat is of immediate conservation

Fig. 1. Location of Lake Eyre GAB springs and spring complexes (shaded areas) sampled in this study. The letters in brackets identify the ESUs found in each spring group. Inset map of Australia shows location of Lake Eyre GAB springs (shaded area).

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concern, given the degradation of even a single spring could breakdown metapopulation processes and increase the extinction potential (Hess, 1996), and, in the most extreme cases, lead to the loss of an ESU or even an entire species (Miller et al., 1989). 2. Methods The Lake Eyre GAB springs are located throughout the western portion of the Lake Eyre Basin in central Australia (Fig. 1). These springs form around areas of geological weakness and as such often display large clusters of directly connected spring outlets, known as spring groups. These spring groups can be categorized further based on their hydro-geological location as spring complexes, which are generally located within a single surface drainage. Chiltoniid amphipods were sampled from across their entire distribution in the Lake Eyre springs. As there are 2000–3000 individual spring outlets, it was prohibitive to sample every single spring where amphipods are found. However, all spring groups containing amphipods were sampled for this study. Our sampling increased the number of spring groups from 15 in Murphy et al. (2009) to 30 and the number of individuals from 92 to 462 (Fig. 1, Table 1). Since Murphy et al. (2009), where all amphipods were referred to as Austrochiltonia sp., two new genera and species (viz. Wangiannachiltonia guzikae King, and Arrabunachiltonia murphyi King) have been described (King, 2009). 2.1. DNA methods DNA extraction, PCR conditions and Sanger sequencing of the mitochondrial cytochrome oxidase I (COI) gene was undertaken in accordance to previous studies (Murphy et al., 2009). In addition to this, selected individuals were also sequenced for the more conserved 28S nuclear ribosomal RNA gene using primers 28S-3311F GGGACTACCCCCTGAATTTAAGCAT, and 28S-4434R CCAGCTATCCTGAGGGAAACTTCG (Witt et al., 2006). PCR conditions for this gene were the same as those used to amplify COI. 2.2. Phylogenetic analysis Sequence data were analyzed using Geneious 5.3 (Drummond et al., 2010). Three separate datasets were subjected to phylogenetic analyses, namely the complete COI data set, the complete 28S dataset, and a combined dataset incorporating only individuals that were sequenced for both genes. Included in these analyses alongside the Lake Eyre GAB spring amphipods were representatives of the

other named Australian chiltoniid species, Austrochiltonia australis Sayce (1901) and A. subtenuis Sayce (1902) (both from non-spring stream habitats); A. dalhousiensis Zeidler, 1997 and Phreatochiltonia anophthalma Zeidler 1991 (from the Dalhousie Springs, north of the Lake Eyre springs); and an unnamed chiltoniid from the distant Edgebaston springs in Queensland (Australia). As the Lake Eyre amphipods are not monophyletic and the relationships amongst chiltoniids are unclear (Murphy et al., 2009) trees were rooted using another member of the Talitroidea, Hyalella azteca (Saussure, 1858). Bayesian phylogenetic trees were constructed for the three sequence sets using MrBayes V3.2 (Ronquist and Huelsenbeck, 2003). For both the COI and 28S genes, the GTR + I + G model was found to be the optimal model using MrModeltest (Nylander, 2004). Analysis was identical for all three datasets, with the exception that the combined dataset was partitioned into genes and the model parameters were unlinked across the analyses. MrBayes analyses were run across four chains for five million generations sampling every 500 generations. Stationarity was determined from an examination of log likelihoods and model parameters. Trees recovered prior to stationarity were discarded and Bayesian posterior probabilities of each bipartition, representing the percentage of times each node was recovered, were calculated from the remaining trees. Multiple runs were performed to ensure that all parameters were not considerably different at stationarity based on alternate prior probabilities. To determine if lineage divergence occurred due to isolation in spring environments or prior the formation of the springs (Guzik et al., 2012; Murphy et al., 2012) molecular clock analyses were undertaken using BEAST version 1.5.2 (Drummond and Rambaut, 2007). The sub-program BEAUTi version 1.4.7 (Drummond and Rambaut, 2007) was used to create input .xml files and Tracer version 1.5 (Rambaut and Drummond, 2007) was used to analyze the parameter distributions estimated from BEAST. An UPGMA starting tree was estimated under the GTR + I + G model with base frequencies estimated, genes partitioned (COI, 28S) and the substitution and clock models unlinked between partitions. For all partitions the substitution model was the GTR + I + G model, and an uncorrelated lognormal clock model was used, employing a specified rate for the COI partition and an estimated rate for both the nuclear partitions. The standard arthropod mtDNA molecular clock of 2.3% divergence per million years (Brower, 1994) was used to calibrate the clock. However, given this is not taxon specific, a normal distribution prior with a mean rate of 0.0115 ± 5% standard deviation was used (Murphy et al., 2012). Three coalescent models (exponential growth, expansion growth and constant size) for tree

Table 1 Distribution and genetic diversity of ESUs identified in this study. Active springs and size are the sum of actively flowing springs and the overall wetland extent (Ha) for each spring group occupied by an ESU (from Fensham et al., 2010), (the number in brackets after the number of active springs is the proportion of springs sampled in this study), N = the number of individuals sampled for each ESU (number in brackets indicates samples taken from previous studies), COI haps = the number of COI haplotypes found, h = haplotype diversity, p = nucleotide diversity, Taj D = Tajima’s D and Fu Fs = Fu’s Fs.

*

ESU

Spring complex

Spring groups

Active springs

Size

N

COI haps

h

p

WG1 WG2 WG3 WG4

Coward Coward Beresford Hermit Hills/Wangianna

28 (0.25) 314 (0.08) 4(0.75) 1016 (0.03)

12 10 1 45

47(16) 130(6) 17(3) 149(all)

32 75 4 106

0.967 0.973 0.596 0.991

0.016 0.040 0.001 0.025

0.625 0.499 0.673 0.918

12.075* 17.980* 1.0551 23.925*

AM FS B1 B2 B3 B4 B5 B6 B7

Strangways/Francis Swamp Francis Swamp Neales Neales Neales Neales Neales Coward Peake

CBC, CBS, CHW, CHS, CCS CEN, CES, CKH, CJS BBH, BWS HBO, HDB, HDD, HWF, HHS, HOF, HOW, WDS, WWS CSS, FFS FFS NHS NTF NBS NTM, NFS NOS CEN EFS TOTAL

13 (0.4) 120 (0.03) 89 (0.03) 1(1) 7 (0.14) 13(0.15) 6 (0.17) 179 (0.03) 49 (0.12) 1839

0.01 10 10 1 1 2 10 0.1 10

25(7) 12(0) 3 10 5 10(5) 3(1) 11 42(2) 464

13 3 3 7 3 7 2 8 30 293

0.903 0.682 1 0.944 0.7 0.911 0.66 0.927 0.979

0.005 0.003 0.012 0.005 0.003 0.006 0.002 0.005 0.009

0.769 2.115 0 1.161 1.092 0.858 0 0.928 1.797*

4.603* 2.475 0.807 2.254 0.276 1.272 1.061 3.056* 20.214*

Significant calculation.

Taj D

Fu Fs

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priors were estimated and compared. Analyses were run five times for 10 million generations with a burnin of 25% of the total saved trees. A hypothetical clock constraint (Murphy et al., 2012) was also tested to simulate divergence within GAB spring amphipods occurring post-GAB spring formation. The maximum age of divergence was constrained within each of the major clades to 1 million years (the estimated time of widespread spring formation (Prescott and Habermehl, 2008) and the posterior probabilities for the estimated COI substitution rates were compared with the standard mtDNA substation rates. 2.3. Delimiting taxa As we were interested in delimiting independently evolving amphipod lineages (or evolutionarily significant units; ESUs, (Moritz, 1994)), ‘‘species’’ boundaries were explored using the general mixed Yule–coalescent (GMYC) model, which examines changes from interspecific to intraspecific in branching rates (Pons et al., 2006). Ultrametric trees were estimated for each gene separately using BEAST under the appropriate gene-specific model parameters, with a relaxed uncorrelated log-normal clock and branch lengths estimated using a Coalescent (Constant Size) prior. Analyses were run for 50 million generations. The Maximum Clade Credibility ultrametric trees were analyzed using a log-likelihood ratio test to assess the significance of the estimated shift in branching rates (i.e. how well they fir the GMYC model). Distinct entities (or species) were delineated based on the lineages that are split by interspecific branching rates. These tests were implemented in the SPLITS package (available from http://r-forge.r-project.org/projects/splits) in the R statistical environment (R Development Core Team, 2011). 2.4. Phylogeography, population structure and genetic diversity To test the hypotheses related to the Death Valley Model (i.e. highly geographically structured and the genetic impact of reduced habitat), median joining networks (Bandelt et al., 1999) for each ESU were constructed using NETWORK version 4.5 (http:// www.fluxus-engineering.com) and genetic distances within and between lineages were calculated using Kimura 2 parameter distances. DnaSP v5 (Librado and Rozas, 2009) was used to determine genetic diversity in each ESU by calculating haplotype (h) and nucleotide (p) diversity; and demographic characteristics (e.g. signatures of population change) were estimated with Tajima’s D (Tajima, 1989) and Fu’s Fs (Fu, 1997). Significance of each demographic calculation tested using 1000 random coalescent simulations. Regression analyses were undertaken using the R statistical environment to examine the relationships between genetic diversity or demographic characters and measures of habitat area for each lineage. Habitat area for each lineage was determined from the number of flowing springs or the extent of the wetland area within a spring group (Fensham et al., 2010) occupied by that lineage. As the complexity of number of springs, wetland size and the geographic spread between springs may interact in their impact on genetic diversity an ‘‘interaction factor’’ (calculated using Log (no. springs  wetland size  max dist)) was also calculated. Relationships between genetic and geographic distance were also examined using a regression analyses. 3. Results

analyses of individual and the combined data sets resulted in identical, well-supported evolutionary lineages from the Lake Eyre GAB spring amphipods. These were the two named species, W. guzikae from the southern and central springs and A. murphyi from Strangways springs, as well as a new lineage from Francis Swamp (closely related to the southern Australian riverine A. subtenuis), and large diverse clade (Lineage B) encapsulating all of the northern springs. Relationships amongst these lineages and with the non-Lake Eyre species vary between analyses, although the Lake Eyre amphipods never form a single monophyletic group. All analyses (Figs. 2 and 3 and Supp. Fig. 1) support at least four reciprocally monophyletic clades within W. guzikae; 3 from the central springs (WG1 = CBC/CHW/CBS/CCS, WG2 = CES/CJS/CEN/ CKH, and WG3 = BBH/BWS) and one (WG4) encompassing all of the southern-most springs. Within Lineage B all of the northern spring groups formed well-supported reciprocally monophyletic clades, with the exception of a close relationship between NTM/ NFS. Within Lineage B there is also a well-supported lineage from the central CEN spring group, which is clearly unrelated to all other central spring lineages. The COI dataset (Fig. 3) identified further well-supported clades within W. guzikae; including the separate spring groups within WG1. There is clearly a large degree of genetic divergence within WG2, including the genetically distinct CEN springs, and some evidence of geographically-based clades within WG4. There are also a number of well-supported distinct clades within WG2 and WG4 that are not reciprocally monophyletic with respect to spring group. There were large differences between the results of GMYC analyses for the two gene trees. For the COI tree (Supp. Fig. 2), 32 distinct entities (CI = 21–55) were identified from the Lake Eyre GAB springs, most of these reflected the lineages identified above, however this also included many lineages that were not well-supported by Bayesian posterior probabilities, and some which represented a single slightly divergent haplotype. For example, 10 distinct entities identified within WG4 and 7 within WG2 often divide single springs into multiple ‘‘species’’. The results of the 28S analyses are more conservative, finding 13 distinct entities (CI = 6– 17) within the Lake Eyre springs (Supp. Fig. 3) all of which are wellsupported by Bayesian posterior probabilities for the three phylogenetic analyses, and are hereafter refer to as ESUs. This includes A. murphyi, the Francis Swamp lineage, all seven of the reciprocally monophyletic clades identified from Lineage B, and the four main W. guzikae groups. Interestingly, GMYC analysis of the COI data set suggested that the change from inter- to intraspecific branching patterns occurs approximately 1–2 million years ago; the hypothesized time of spring origins. During this time population structure may have changed from a free ranging riverine environment into the isolated spring-living that exists today, which could explain the changes in branching patterns. Molecular clock dating (Table 2) suggests that most of the lineages identified above evolved prior to the hypothesized origin of the Lake Eyre GAB springs1 million years ago. Even some of the more recent divergence events evident from the COI tree appear to have a ‘‘pre-spring’’ origin, for example the common ancestry of amphipods from the 4 spring groups that make up WG2 is suggested at 1.84–5.26 million years ago. Constraining the diversification within the two major Lake Eyre lineages (W. guzikae and Lineage B) to occur within the last million years (after spring origin) required a COI divergence rate of 12.6–26.4% per million years, clearly supporting the idea that most of the divergence occurred prior to the origin of the springs.

3.1. Phylogenetic analyses and ESUs 3.2. ESU distribution and diversity A total of 293 COI and 27 28S haplotypes were generated from 492 individuals sampled from 89 springs, from 29 spring groups and eight spring complexes (Table 1, Supp. Table 1). Phylogenetic

Most of the ESUs identified have incredibly narrow distributions and are present in only a small number of geographically

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Fig. 2. Phylogenetic tree resulting from MrBayes analyses of combined COI mtDNA and 28S rDNA data. Numbers on branches = Bayesian posterior probabilities.

restricted springs (Supp. Fig. 4, Table 1). The only exception is WG4, which occurs across separate spring complexes and Lake Eyre catchments. Only 3 of the 6 ESUs found in multiple spring groups exhibit any evidence of population structure (Supp. Fig. 4), while 7 out of 13 ESUs are actually located within a single spring group, the most extreme being B2, which exists only in the single spring from the Fountain spring group. There is a significant correlation between COI genetic and geographic distance (Fig. 4). Close examination of these data, however reveals three clusters of genetic distance measures; (1) between springs within ESUs, (2) between ESUs from the same evolutionary lineage (i.e. between W. guzikae or lineage B ESUs), and (3) between ESUs from separate evolutionary lineages. When these three data sets are examined separately there is no relationship between genetic and geographic distances. The extreme geographic restriction of some ESUs meant that limited samples were available for calculating genetic diversity statistics (Table 1), therefore the discussion of diversity statistics does not include B1, B3 and B5, all of which had <10 individuals sequenced. The majority of ESUs had high (>0.90) haplotype diversity; the exceptions were FS (0.68) and WG3 (0.60). Nucleotide diversity values differed by an order of magnitude amongst ESUs; ranging from >0.015 amongst the three diverse W. guzikae ESUs (WG1, WG2, WG4) to 0.001 for WG3. Tests for Tajima’s D or Fu’s Fs found most ESUs with non-significant values for at least one of the statistics (indicating demographic stability). Only B7 had significantly negative values for both statistics, suggesting long-term

population growth. Conversely, only FS had positive values for both statistics, suggesting a recent population bottleneck. None of the diversity statistics were correlated with the proportion of springs sampled from a spring group (Table 3). Neither haplotype diversity nor Tajima’s D were correlated with the habitat characteristics. Nucleotide diversity and Fu’s Fs were not correlated with either the number of springs or wetland size on their own, however both of these statistics were significantly correlated with the habitat interaction factor (although only Fu’s Fs remained significant after correction for multiple tests), suggesting that the combination of number of springs, wetland size and distance between springs plays some role in amphipod genetic diversity.

4. Discussion This study has revealed even higher levels of genetic, species and generic diversity than recently proposed within the Lake Eyre GAB spring amphipods (King, 2009; Murphy et al., 2009). Strikingly, this study has identified a number of very geographically restricted and highly allopatric ESUs that show genetic divergences reflective of species-level differences or beyond and have origins older than their spring habitat (King and Leys, 2011). Comparative morphological studies are currently being undertaken on both lineage B and the Francis Swamp ESU, and it is likely that these two lineages will ultimately be recognized as separate genera, bringing the total of amphipod genera within Lake Eyre springs

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Fig. 3. Phylogenetic tree resulting from MrBayes analyses of entire COI mtDNA data. Numbers on branches = Bayesian posterior probabilities. Clades and lineages are colored based on their location in the Lake Eyre GAB springs, as shown on inset map.

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N.P. Murphy et al. / Molecular Phylogenetics and Evolution 66 (2013) 645–653 Table 2 BEAST generated molecular clock results for either the divergence of distinct lineages or the common ancestry of ESUs. Common ancestry/divergence

Clock dates (million years)

Com. anc. of WG1 (CBC/CHW/CBS/CCS) Com. anc. of WG2 (CES/CJS/CEN/CKH) Com. anc. of WG3 (Hermit Hills/Wangianna) Com. anc. of W. guzikae(WG1/WG2/WG3/WG4) Com. anc. of Francis Swamp/A. subtenuis Div of A. murphyi Div of NBS Div of NTF Com. anc. of main Neales clade (NBS/NTM/NTF/ NFS/NHS) Div of NOS Div of Elizabeth North in Lineage B Div of Lineage B

0.69–2.27 1.84–5.26 1.20–3.24 5.61–13.02 3.28–11.70 16.65–36.60 1.44–4.36 2.38–6.08 3.04–7.32 6.36–14.86 5.54–12.934 6.87–15.97

to four (R.A. King pers comm). Unfortunately though, thorough examination of W. guzikae across its range failed to find any morphological characters to delineate the four well-defined genetic lineages (King, 2009; Murphy et al., 2009), and it is unclear whether lineage B ESUs can be delineated morphologically. Morphological uniformity among species that display extreme levels of genetic divergence is apparent in other GAB spring invertebrates (Murphy et al., 2012), and groundwater invertebrates in general (Witt et al., 2006), and traditional taxonomic appraisals have not diagnosed candidate species that diverged millions of years ago (Murphy et al., 2009, 2012). As a result, these species languish in the now increasingly common ‘‘twilight zone’’ occupied by well-defined, genetically distinguishable taxa that cannot be

assigned a morphological diagnosis. Thus, the use of ESUs to identify units for conservation is perhaps the most practical for the GAB spring system. Alternatively, it is now increasingly common for taxonomists to formally diagnose new species based solely on molecular genetic characters (e.g. Egge and Simon, 2006). Assigning species names to GAB spring amphipod ESUs will ensure them greater recognition and more likely greater protection (Isaac et al., 2004; Mace, 2004). However, as most of these species will almost automatically qualify for endangered species lists, the heightened political sensitivity surrounding the use of water resources in the GAB demands that any newly-named species are robustly defensible (Harvey et al., 2011). Pragmatically, this will require concordant taxonomic support from a range of additional nuclear markers. Clearly, the geographic isolation of the springs has been a major influence on amphipod diversity, leading to highly restricted distributions. Whilst we cannot rule out the possibility that the identified ESUs have wider ranges than reported here, this is very unlikely given the highly allopatric distributions found and our thorough coverage of all Lake Eyre spring groups and complexes where amphipods are found. The GAB spring amphipods therefore easily fit within the 10,000 km2 criteria suggested for short-range endemics (Harvey, 2002). Whilst ESU B2 is the extreme example, found in only a single springs with a 0.01 km2 wetland extent, even the most widespread, WG4, is restricted <1 km2 of wetland across 50 km of desert. Indeed, with all 13 ESUs inhabiting <1 km2 of wetland, they are best described as micro-endemics. Our genetic data, in particular the very large divergences between neighboring springs, strongly support the first of the two hypotheses implicit in the Death Valley Model for GAB spring

Fig. 4. Scatterplot of relationships between genetic and geographic distances. Overall regression slope and R2 value is shown by the dotted line. The shaded circles and individual regression slopes indicate the relationships amongst spring groups within ESUs (bottom), ESUs within major lineages (middle) and ESUs among major lineages (top). Shown in the inset is the histogram of K2P genetic distances.

Table 3 Results of regression analyses of genetic diversity statistics vs. habitat statistics. = significant at 0.05. h = haplotype diversity, p = nucleotide diversity, Taj D = Tajima’s D and Fu Fs = Fu’s Fs. b = Standardized regression coefficients, R2 = R-squared values, P = P-values. Statistically significant relationships (P < 0.05) are shown in boldface. No. springs

h

p D Fs

Wetland size R2

b 0.045 0.298 0.157 2.591

0.025 0.286 0.020 0.255

P 0.300 0.064 0.391 0.078

Interaction factor R2

b 0.025 0.199 0.094 2.00

0.056 0.159 0.070 0.209

P 0.492 0.147 0.540 0.103

R2

b 0.070 0.492 0.164 4.921

0.025 0.345 0.077 0.455

Prop. springs sampled P 0.300 0.042 0.567 0.019

R2

b 0.034 0.263 0.315 1.738

0.089 0.014 0.057 0.051

P 0.619 0.318 0.50 0.474

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amphipods, namely very localized population structure and no evidence for isolation by distance within ESUs. There is less evidence, however, for a reduction in genetic diversity in the ESUs inhabiting smaller habitats, a second key assumption of the DVM. Whilst it appears that there is a relationship between the complexity of the habitat and nucleotide diversity, no relationships with haplotype diversity were evident. The relationship between nucleotide diversity and habitat size, whilst significant, is not strong and could also be explained by population structure within the three most widespread ESUs driving the extensive nucleotide divergence. Although contrary to the DVM expectations for highly fragmented species, the lack of a clear relationship between habitat size and genetic diversity, together with evidence of long term demographic stability or population growth in most populations, is not unexpected. To persist in such harsh environments, these amphipods must have evolved mechanisms to cope with the deleterious effects of limited distributions. A metapopulation structure, with individual spring vents operating as distinct populations and a constant cycle of extinction and recolonisation within spring groups, as described previously in GAB spring invertebrates (Wilmer et al., 2011), could explain the high levels of genetic diversity, as would a higher mutation rate. 4.1. Conservation implications Given that these taxa appear to have evolved to cope with small population size, they may be able to cope with a small level of spring extinction, particularly since spring extinction is a natural part of the geomorphic evolution of this environment (Prescott and Habermehl, 2008). Clearly though, no taxon is immune to major habitat loss, thus ESUs with the smallest ranges are under threat simply because they are restricted to tiny patches of habitat. Even a genetically diverse ESU such as WG2, found in over 300 springs, would be threatened by localized aquifer drawdown, given it only occupies a 2 km2 region. Only three of the ESUs (WG1, WG2, B6) are currently protected by national parks. The remaining are located on pastoral land, which has suffered in varying degrees of degradation due to differing management practices across the region (Fensham et al., 2010). The micro-endemism evident within these amphipods implies that local habitat stresses and degradation could lead to the extinction of entire species. A number of springs within the Lake Eyre region are either completely degraded or have dried up due to reduced water pressure and therefore no longer support endemic invertebrates (Fensham et al., 2010). As a result, it is quite possible that unknown ESUs and species have already gone extinct, particularly from more isolated springs. The results of this study can therefore help assess whether existing ESUs are on the verge of extinction, vulnerable to any further habitat reduction, or seemingly quite robust to habitat loss. The two ESUs with the lowest genetic diversity are WG3 and FS, therefore these two are seemingly at risk of negative genetic impacts. Interestingly, these two ESUs occupy one of the smallest and one of the larger habitats respectively, however both the Beresford (occupied by WG3) and Francis Swamp (occupied by FS) appear degraded, with discolored water and damage from feral and domestic animal access (Murphy, pers. obs.). Possibly the degree of habitat degradation rather than the size of the habitat has more influence on this reduction in genetic diversity. Other springs, in particular in the genetically diverse Neales complex (B1–B5), also appear on observation to be significantly degraded. Thus further work is required to determine whether habitat quality or quantity is the better predictor of within-ESU genetic diversity. Clearly though, immediate attention is required for spring groups on pastoral land that show obvious signs of spring degradation and have highly restricted ESUs with low genetic diversity. The more widespread and genetically diverse ESUs

such as WG1, WG2 and WG4, although still geographically restricted, should be the least vulnerable to any localized reduction in spring habitat, or degradation. However, their high levels of genetic structure also argues the need for knowledge of if and how dispersal and gene flow occur in this highly structured environment (Fahrig and Merriam, 1994). There are a number of springs (either dry or degraded) which once had an endemic invertebrate fauna. The highly restricted distributions and evidence of deep population structure revealed herein suggests it is doubtful that amphipods would ever naturally repopulate restored springs (Hughes, 2007). Assisted colonization is therefore likely to be required to restore amphipods to rehabilitated springs and could also be used to improve the evolutionary potential of genetically impoverished populations. Of course this option is not without risk, as the large degree of genetic divergence over such a fine scale may mean that local adaptation is playing a major role in population structure (Bohonak and Jenkins, 2003), for example to differences in water chemistry among springs. Thus assisted recolonisation may actually breakdown local adaptations, leading to greater extinction potential (Kreyling et al., 2011). In fact, one of the main management recommendations in highly structured species, and one that appears to be most appropriate in the case of the GAB spring amphipods is to avoid creating any artificial dispersal and gene flow between divergent populations (Meffe and Vrijenhoek, 1988). Acknowledgements This research was undertaken with Australian Research Council Discovery (DP0770979) and Linkage (LP0669062) grant programs with research partners; The Department of Environment and Heritage (SA), BHP Billiton, Nature Foundation SA and the South Australian Museum. Thanks to Steve Delean, Travis Gotch and Rachael King for fieldwork assistance. We appreciate the access given to us by the traditional owners of the GAB spring country, particularly Reg Dodd (Arabunna people), which enabled us to undertake our field collection through permission to access culturally sensitive land. We also thank S. Kidman & Co Ltd for permission to access springs and collect specimens from pastoral land (Anna Creek, Stuart Creek and The Peake Stations). Thanks to Rachael King for discussions of all things Chiltoniidae, and Rachael and three anonymous reviewers for improving the manuscript. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.ympev.2012. 10.013. References Bandelt, H., Forster, P., Rohl, A., 1999. Median-joining networks for inferring intraspecific phylogenies. Mol. Biol. Evol. 16, 37–48. Bohonak, A.J., Jenkins, D.G., 2003. Ecological and evolutionary significance of dispersal by freshwater invertebrates. Ecol. Lett. 6, 783–796. Brower, A.V.Z., 1994. Rapid morphological radiation and covergence among races of the butterfly Heliconius erato inferred from patterns of mitochondrial DNA evolution. Proc. Natl. Acad. Sci. USA 91, 6491–6495. Drummond, A., Rambaut, A., 2007. BEAST: Bayesian evolutionary analysis by sampling trees. BMC Evol. Biol. 7, 214. Drummond, A., Ashton, B., Buxton, S., Cheung, M., Cooper, A., Duran, C., Field, M., Heled, J., Kearse, M., Markowitz, S., Moir, R., Stones-Havas, S., Sturrock, S., Thierer, T., Wilson, A., 2010. Geneious v5.3. . Echelle, A.F., Echelle, A.A., Edds, D.R., 1989. Conservation genetics of a springdwelling desert fish, the Pecos Gambusia (Gambusia nobilis, Poeciliidae). Conserv. Biol. 3, 159–169. Egge, J.J.D., Simon, A.M., 2006. The challenge of truly cryptic diversity: diagnosis and description of a new madtom catfish (Ictaluridae: Noturus). Zool. Scr. 35, 581– 595.

N.P. Murphy et al. / Molecular Phylogenetics and Evolution 66 (2013) 645–653 Fahrig, L., Merriam, G., 1994. Conservation of fragmented populations. Conserv. Biol. 8, 50–59. Fensham, R.J., Price, R.J., 2004. Ranking spring wetlands in the great artesian basin of Australia using endemicity and isolation of plant species. Biol. Conserv. 119, 41– 50. Fensham, R., Ponder, W., Fairfax, R., 2010. National recovery plan for the community of native species dependent on natural discharge of groundwater from the Great Artesian Basin. In: Department of the Environment, W., Heritage and the Arts, Canberra. Queensland Department of Environment and Resource Management, Brisbane. (Ed.). Fu, X.-Y., 1997. Statistical tests of neutrality of mutations against population growth, hitchhiking and background selection. Genetics 147, 915–925. Guzik, M.T., Adams, M.A., Murphy, N.P., Cooper, S.J.B., Austin, A.D., 2012. Desert springs: deep phylogeographic structure in an ancient endemic crustacean (Phreatomerus latipes). PLoS ONE 7, e37642. Harvey, M.S., 2002. Short-range endemism among the Australian fuana: some examples from non-marine environments. Inver. Syst. 16, 555–570. Harvey, M.S., Rix, M.G., Framenau, V.W., Hamilton, Z.R., Johnson, M.S., Teale, R.J., Humphreys, G., Humphreys, W.F., 2011. Protecting the innocent: studying short-range endemic taxa enhances conservation outcomes. Inver. Syst. 25, 1– 10. Hershler, R., Mulvey, M., Liu, H.-P., 1999. Biogeography in the death valley region: evidence from springsnails (Hydrobiidae: Tryonia). Zool. J. Linn. Soc. 126, 335– 354. Hess, G.R., 1996. Linking extinction to connectivity and habitat destruction in metapopulation models. Am. Nat. 148, 226–236. Hughes, J.M., 2007. Constraints on recovery: using molecular methods to study connectivity of aquatic biota in rivers and streams. Freshw. Biol. 52, 616–631. Isaac, N.J.B., Mallet, J., Mace, G.M., 2004. Taxonomic inflation: its influence on macroecology and conservation. Trends Ecol. Evol. 19, 464–469. Juan, C., Guzik, M.T., Jaume, D., Cooper, S.J.B., 2010. Evolution in caves: Darwin’s ‘wrecks of ancient life’ in the molecular era. Mol. Ecol. 19, 3865–3880. King, R.A., 2009. Two new genera and species of chiltoniid amphipods (Crustacea: Amphipoda: Talitroidea) from freshwater mound springs in South Australia. Zootaxa 2293, 35–59. King, R.A., Leys R., 2011. The Australian freshwater amphipods Austrochiltonia australis and Austrochiltonia subtenuis (Amphipoda: Talitroidea: Chiltoniidae) confirmed and two new cryptic Tasmanian species revealed using a combined molecular and morphological approach. Inver. Syst 25, 171–196. Kodric-Brown, A., Brown, J.H., 2007. Native fishes, exotic mammals, and the conservation of desert springs. Front. Ecol. Environ. 5, 549–553. Kreyling, J., Bittner, T., Jaeschke, A., Jentsch, A., Jonas Steinbauer, M., Thiel, D., Beierkuhnlein, C., 2011. Assisted colonization: a question of focal units and recipient localities. Restor. Ecol. 19, 433–440. Librado, P., Rozas, J., 2009. DnaSP v5: a software for comprehensive analysis of DNA polymorphism data. Bioinformatics 25, 1451–1452. Liu, H.-P., Hershler, R., 2007. A test of the vicariance hypothesis of western North American freshwater biogeography. J. Biogeogr. 34, 534–548.

653

Mace, G.M., 2004. The role of taxonomy in species conservation. Philos. Trans. R. Soc. London Ser. B: Biol. Sci. 359, 711–719. Meffe, G.K., Vrijenhoek, R.C., 1988. Conservation genetics in the management of desert fishes. Conserv. Biol. 2, 157–169. Miller, R.R., Williams, J.D., Williams, J.E., 1989. Extinctions of North American fishes during the past century. Fisheries 14, 22–38. Mitchell, B.D., 1985. Limnology of mound springs and temporary pools, south and west of Lake Eyre. In: Greenslade, J., Joseph, L., Reeves, A. (Eds.), South Australia’s Mound Springs. Nature Conservation Soc, SA Inc., Adelaide, pp. 51– 63. Moritz, C., 1994. Defining ‘evolutionarily significant units’ for conservation. TRENDS in Ecol. Evol. 9, 373–374. Murphy, N.P., Adams, M., Austin, A.D., 2009. Independent colonization and extensive cryptic speciation of freshwater amphipods in the isolated groundwater springs of Australia’s Great Artesian Basin. Mol. Ecol. 18, 109–122. Murphy, N.P., Guzik, M.T., Worthington Wilmer, J., 2010. Understanding the influence of landscape and dispersal on species distributions in fragmented groundwater dependent springs. Freshw. Biol. 55, 2499–2509. Murphy, N.P., Breed, M.F., Guzik, M.T., Cooper, S.J.B., Austin, A.D., 2012. Trapped in desert springs: phylogeography of Australian desert spring snails. J. Biogeogr.. Nevill, J.C., Hancock, P.J., Murray, B.R., Ponder, W.F., Humphreys, W.F., Phillips, M.L., Groom, P.K., 2010. Groundwater-dependent ecosystems and the dangers of groundwater overdraft: a review and an Australian perspective. Pac. Conserv. Biol. 16, 187–208. Nylander, J.A.A., 2004. MrModeltest v2. Program distributed by the author. Ponder, W.F., Eggler, P., Coglan, D.J., 1995. Genetic differentiation of aquatic snails (Gastropoda: Hydrobiidae) from artesian springs in arid Australia. Biol. J. Linn. Soc. 56, 553–596. Pons, J., Barraclough, T.G., Gomez-Zurita, J., Cardoso, A., Duran, D.P., Hazell, S., Kamoun, S., Sumlin, W.D., Vogler, A.P., 2006. Sequence-based species delimitation for the DNA taxonomy of undescribed insects. Syst. Biol. 55, 595–609. Prescott, J.R., Habermehl, M.A., 2008. Luminescence dating of spring mound deposits in the southwestern Great Artesian Basin, northern South Australia. Aust. J. Earth Sci. 55, 167–181. Purvis, A., Hector, A., 2000. Getting the measure of biodiversity. Nature 405, 212– 219. Rambaut, A., Drummond, A., 2007. Tracer v1.4. . Ronquist, F., Huelsenbeck, J.P., 2003. MrBayes 3: Bayesian phylogenetic inference under mixed models. Bioinformatics 19, 1572–1574. Tajima, F., 1989. Statistical method for testing the neutral mutation hypothesis by DNA polymorphism. Genetics 123, 585–595. Wilmer, J.W., Murray, L., Elkin, C., Wilcox, C., Niejalke, D., Possingham, H., 2011. Catastrophic floods may pave the way for increased genetic diversity in endemic artesian spring snail populations. PLoS ONE 6, e28645. Witt, J.D.S., Threloff, D.L., Hebert, P.D.N., 2006. DNA barcoding reveals extraordinary cryptic diversity in an amphipod genus: implications for desert spring conservation. Mol. Ecol. 15, 3073–3082.