J Insect Conserv (2013) 17:1209–1219 DOI 10.1007/s10841-013-9602-8
ORIGINAL PAPER
Grassland area determines beetle assemblage dissimilarity from surrounding floodplain forest Philip S. Barton • Matthew J. Colloff • Kimberi R. Pullen • Saul A. Cunningham
Received: 20 May 2013 / Accepted: 4 October 2013 / Published online: 11 October 2013 Ó Springer Science+Business Media Dordrecht 2013
Abstract Patch size is known to affect biodiversity in fragmented landscapes, but is usually examined in systems where the surrounding matrix habitat is unfavourable. We examined beetle diversity in a floodplain ecosystem that is characterised by naturally occurring grassland patches within a dominant matrix of contrasting yet habitable forest. We asked whether differences in the beetle assemblage between grassland and forest vegetation depended on the area of the grassland patch, which is a function of its flooding frequency and duration: smaller grasslands tend to be higher on the floodplain and are flooded less often and for shorter periods than larger grasslands. We found a negative relationship between grassland area and beetle abundance and species richness, and a positive relationship between grassland area and compositional dissimilarity from the surrounding forest. As expected, we found an overall difference in composition between forest and grassland assemblages, with five beetle species more common in the grasslands. Our study indicates that floodplain grasslands not only support beetle assemblages that are distinct from the surrounding forest, but that assemblages from the larger grasslands are compositionally more distinct than those from smaller grasslands. A likely cause of this pattern is the reduced edge effects and greater Electronic supplementary material The online version of this article (doi:10.1007/s10841-013-9602-8) contains supplementary material, which is available to authorized users. P. S. Barton (&) Fenner School of Environment and Society, Australian National University, Canberra, ACT 0200, Australia e-mail:
[email protected] M. J. Colloff K. R. Pullen S. A. Cunningham CSIRO Ecosystem Sciences, GPO Box 1700, Canberra, ACT, Australia
environmental contrast between forest and large grasslands that may be exposed to greater variation in local climate. Ongoing changes to flood regimes and potential encroachment of forest plants may decrease grassland area in the future, which may reduce spatial heterogeneity in the insect community in this unique floodplain ecosystem. Keywords Assemblage Barmah forest Composition Edge Hydrology Murray river Spatial heterogeneity
Introduction Ecologists have long been interested in understanding how patchy or fragmented landscapes affect patterns of biodiversity (Driscoll et al. 2013; Haila 2002). A major outcome of this research has been the recognition of the importance of patch size and contrast with the surrounding habitat for species abundance and composition. Although many studies have examined the effects of patch attributes on species assemblages (Ewers and Didham 2006; Peyras et al. 2013; Ries et al. 2004), many of these studies are conducted in modified landscapes where the patch is remnant vegetation, and the focus of the study is on the change in biota within ¨ ckinger et al. 2012). the patch (Didham et al. 1998; O Further, the patches of remnant vegetation are often forest or woodland, and the surrounding matrix has been converted to less habitable cropland or pasture (Campbell et al. 2011). Far less common are studies set in ecosystems where the reverse is true: the patches are grassland, the surrounding matrix is forest, and both are natural and habitable vegetation. Insects make up a substantial component of biodiversity and animal biomass in terrestrial ecosystems (Basset et al.
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Fig. 1 Top an aerial photograph of a floodplain grassland surrounded by the dominant river red gum forest. The grassland is partially inundated with water, representing an intermittent disturbance process. Bottom a clear boundary is obvious between grassland and forest. Our study was conducted at a time when the grassland lakes were dry
2012), and contribute to many important ecological services (Losey and Vaughan 2006). Further, insects can be very sensitive to environmental change due to often high levels of resource specialisation and fine-scale perception and response to landscapes (Andersen et al. 2004; Pik et al. 2002).This makes insects an ideal group to examine the effects of patch size on differences between vegetation types (Dangerfield et al. 2003; Didham et al. 1996). In this study, we examined the effect of grassland patch size on beetle diversity in the Barmah Forest on the River Murray in south-eastern Australia (see Fig. 1). This unique floodplain ecosystem is characterised by river red gum
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(Eucalyptus camaldulensis) forest interspersed with open grass plains dominated by the aquatic Moira grass (Pseudoraphis spinescens). These grasslands are subject to more frequent flooding than the surrounding forest (Bren 1992; Stokes et al. 2010). The topography and hydrology of the area affects the distribution of the floodplain plants. The abrupt distinction between forest and grassland is based in part on their tolerance to intermittent soil waterlogging as well as their partial or complete inundation during flood events (Casanova and Brock 2000; Colloff and Baldwin 2010). This ecosystem is of considerable conservation significance, and is the largest continuous extent of river red gum
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forest (71,575 ha). It is a Ramsar listed wetland (DSE 2008) and is recognised as a Wetland of National Significance (Environment Australia 2001). Over recent decades, there have been significant changes to the flooding regime due to river regulation and water resource development (Bren 2005; Bren et al. 1988; Dexter et al. 1986). Changes in flood regimes, combined with extended, severe drought between 1997–2010, have led to major changes in the vegetation communities (Catford et al. 2011; Chesterfield 1986; Mayence et al. 2010; Stokes and Cunningham 2006), including the encroachment of river red gum saplings into the now drier grasslands (Bren 1992). Although there is an understanding of the negative impacts of changed flood regimes on vegetation and vertebrates in the Barmah Forest (Ballinger and Mac Nally 2006; Mayence et al. 2010; Stokes et al. 2010; Ward and Colloff 2010), far less is understood about the grasslands and their associated invertebrate community. The additional threats to the grasslands have given further importance to understanding how both grasslands and forest contribute to the biodiversity in this ecosystem. The aim of this study was to examine whether grassland size affected the contrast between forest and grassland vegetation. To date, research on invertebrates in river red gum floodplain forest has identified a positive relationship between flooding frequency and duration and the diversity and abundance of invertebrates in river red gum floodplain forests (Ballinger et al. 2005, 2007). However, no studies have explicitly compared insect communities between forest and grassland vegetation. Although a difference in beetle assemblages between forest and grassland might be obvious, it is not clear whether the size of the grassland is a factor influencing its difference from the surrounding forest. Identifying such a pattern may have implications for the conservation of biodiversity in this important floodplain forest ecosystem.
Methods Study area and sampling design Our study was carried out at Barmah Forest in southeastern Australia (approximately 35°55 South, 145°00 East). We selected eight grasslands dominated by Moira grass (P. spinescens), an emergent aquatic species. The grassland areas and perimeters were: Bucks Lake (1.32 ha, 591 m), Top Lake (5.43 ha, 1,392 m), Little Rushy Swamp (9.78 ha, 1,952 m), Rabbit Lagoon (10.19 ha, 2,058 m), Harbour Lake (27.39 ha, 3,061 m), Hut Lake (33.96 ha, 3,423 m), Keyes Point (111.24 ha, 10,698 m), and Steamer Plain (361.92 ha, 10,410 m). Each of the grasslands differs in size due to their unique local topography. This leads to different flooding frequency
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and duration, which in turn determines the boundary between each grassland and the surrounding forest. The daily volume of river flow at which each grassland location commences to flood shows a negative relationship with area (Reid and Quinn 2004), with larger grasslands flooding at lower flows (Electronic Supplementary Material, Appendix 1). In the 10 years prior to this study (Sep 1997– Sep 2007), which corresponded with the Millennium Drought of 1997–2010 (Potter and Chiew 2011), the lowest commence-to-flood volume for a grassland (12 GL/day) was exceeded on 17 occasions totalling 598 days, whereas the highest (18 GL/day) was exceeded on seven occasions totalling 193 days, calculated from daily flow data downstream of Yarrawonga Weir (Gauge No. 409025). This indicates that smaller grasslands were flooded less frequently than larger grasslands. At each of the eight grasslands, we established a pair of transects: one in the grassland and a one in the adjacent forest. These two transects were each 50 m into their respective vegetation type, and ran parallel to the grassland-forest boundary. We placed three plots (5 m 9 5 m) spaced approximately 100 m apart along each transect (each transect about 200 m long). Because each transect was placed the same distance from the grassland-forest boundary in all our grassland sites, distance from edge is not confounded with grassland area, whereas distance to the middle of the grassland would have been. In each plot, we dug five pitfall traps into the soil, equally spaced along their diagonals, giving three plots of five pitfall traps in each of the 16 grassland and forest sites (total = 240 traps). We used pitfall traps (80 mm diameter) with 70 % aqueous ethanol and 5 % glycerol as a killing agent and preservative. Traps were open for 1 week during September 2007 (southern hemisphere spring), which avoided the extremes of cold in winter and heat in summer. Pitfall trap catches are sensitive to insect activity levels and factors that might influence activity, such as vegetation density (Samways et al. 2010). However, experimental research has indicated that such bias is most notable at extreme levels of ground-layer vegetation density (Melbourne 1999). Due to the prolonged drought in our study region, vegetation was relatively sparse in the grasslands, with considerable areas of bare ground. We visually assessed ground-layer vegetation density (but not composition) to be similar between the grassland and forest. This reduced the potential for bias in sampling efficacy of the pitfall traps, and we considered potential for bias in our measures of abundance to be small. A further important consideration is the potential detrimental impacts of our sampling on the beetle fauna or vegetation across our study sites. We disturbed a combined 3 m2 of soil to a depth of 10 cm to dig in our pitfall traps, relative to an area of over 600 m2 across all out plots. Thus,
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we consider the impact of our research to be negligible compared with the prolonged and significant effects of the Millennium Drought (1997–2010) that was occurring during the time of our study. All beetles were identified to family and subfamily, by experienced entomologists with reference to appropriate taxonomic keys (Lawrence and Britton 1994; Lawrence et al. 1999), and then sorted to morphospecies (sensu Oliver and Beattie 1996). Common species were identified to genus and species by taxonomists at the Australian National Insect Collection. Each morphospecies was placed in a generalised feeding group, based on its family or subfamily-level phylogeny (Hunt et al. 2007; Lawrence and Britton 1994). Voucher specimens were deposited in the Australian National Insect Collection, Canberra, Australia.
Statistical analyses We first examined differences in species richness between grassland and forest with regard to sampling effort and the number of individuals sampled (Gotelli and Colwell 2001). We used EstimateS 8.0 (Colwell 2013) to generate accumulation curves using the analytical Mao Tau method (Colwell et al. 2004), and calculated means and standard deviations from 1,000 randomisations of the samples. We used generalised linear mixed models in GenStat 14 (VSNI 2013) to test for effects of vegetation type on beetle abundance or species richness per plot. For this analysis we pooled the five pitfall traps within each plot to give 48 ‘samples’. We fitted vegetation type as a fixed effect, and fitted each grassland-forest pair as a random effect. For these analyses we assumed a Poisson error distribution and used a log-link function. We were also interested in examining the relative contribution of different families and feeding groups to the beetle assemblages in the grassland and forest vegetation types. We used a G test of independence, similar to the Chi square test (Sokal and Rohlf 1995), to see whether the relative proportions of the observed number of individuals and species of beetles in the eight common families and the four feeding groups was independent of vegetation type. To examine differences in species composition between grasslands and forest, and the influence of grassland area, we used a two different multivariate methods. First, we used a two-way Permutational Analysis of Variance (PERMANOVA, Anderson (2001)) in PC-ORD 5.1 (McCune and Mefford 2006) to test for interactive effects of vegetation type and grassland-forest pair on the composition of beetle species. For this analysis we pooled the five pitfall traps in each plot to give 48 plot-level samples, thus giving three replicates within each transect, producing a split-plot design. Second, we used non-metric multi-dimensional
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scaling ordination (NMS) in PC-ORD 5.1 to graphically depict differences in composition between the eight grassland and eight forest sites. For this analysis, we pooled the trap data to transect level, and used 500 iterations of the data to find a two-dimensional solution. For both these multivariate analyses we used the Bray-Curtis dissimilarity index calculated from square-root transformed species abundance data. We used indicator species analysis (Dufrene and Legendre 1997) in PC-ORD 5.1 to identify any species that were associated with either grassland or forest vegetation, more than expected by chance. To account for the greater overall abundance of beetles in the forest vegetation, we first relativised species abundances by total abundance of beetles at each site (McCune and Grace 2002). Significance was determined using 10,000 permutations of the data. We accompanied the indicator species analysis with plots of mean abundances of species that had a significant association with either vegetation type. Finally, we used linear regression to examine the relationship between log grassland area (hectares) and beetle abundance, species richness, and Bray-Curtis dissimilarity from each paired forest transect.
Results We collected 5,179 beetles comprising 109 species from 24 families, and we sampled more beetle individuals and species from the forest than the grassland sites in absolute terms (Appendix 1). Species accumulation curves showed there was no notable difference in the expected number of beetle species sampled from either vegetation type due to sampling effort or number of individual beetles sampled (Appendix 2). Generalised linear mixed model analysis indicated there was a significant difference between vegetation types for beetle abundance (F1,32 = 4.65, P = 0.037), but not for species richness (F1,39 = 0.78, P = 0.381). We also found that the relative abundance of beetles in the major families (G = 448.9, P \ 0.001, Fig. 2b) and feeding groups (G = 333.72, P \ 0.001, Fig. 2d) differed significantly from expected between forest and grassland vegetation. However, we did not find a difference in relative proportion of beetle species in the major families (G = 4.28, P = 0.831, Fig. 2a), or feeding groups (G = 1.29, P = 0.729, Fig. 2c). PERMANOVA revealed a significant interaction between site and vegetation type (F = 1.582, P = 0.004), indicating that there was a difference in species composition between grassland and forest vegetation, but this depended on the grassland location. This was illustrated by the NMS ordination, which showed an overall separation of grassland and forest sites (Fig. 3). Further, each paired
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(b)100%
90%
90%
Percentage of individuals
(a)100% Percentage of species
Fig. 2 Relative proportion of beetle species and individuals from dominant families (a, b) and four generalised feeding groups (c, d) across forest and grassland vegetation types
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80% 70% 60% 50% 40% 30% 20% 10%
Forest
70%
Staphylinidae
60%
Leiodidae
Chrysomelidae
30%
Carabidae
20%
Anthicidae
Forest
90%
Percentage of individuals
(d)100%
90%
Percentage of species
Curculionidae
40%
(c)100% 80% 70% 60% 50% 40% 30% 20%
Grassland
80%
Predator
70%
Herbivore
60%
Fungivore
50%
Detritivore
40% 30% 20% 10% 0%
0%
Forest
Stress = 10.7% Steamer
Grassland sites
Bucks
Elateridae
50%
0%
Grassland
10%
Axis 2
Other
10%
0%
Harbour
80%
Top
Little Rushy Rabbit Keyes Hut Forest sites
Axis 1
Fig. 3 Ordination of beetle samples from forest (black circles) and grassland (grey circles) in Barmah National Park. Dashed lines connect forest-grassland transect pairs. Grassland areas were: Bucks Lake (1.32 ha), Top Lake (5.43 ha), Little Rushy Swamp (9.78 ha), Rabbit Lagoon (10.19 ha), Harbour Lake (27.39 ha), Hut Lake (33.96 ha), Keyes Point (111.24 ha), and Steamer Plain (361.92 ha)
grassland and forest site were separated from each other in approximately the same direction in ordination space. The largest separation in ordination space was apparent for the largest grassland (Steamer Plain), and the smallest
Grassland
Forest
Grassland
separation was for the smallest grassland (Bucks Lake), corresponding with the negative relationship between area of the grasslands and their commence-to flood volumes (Appendix 3). Of the 109 beetle species identified in this study, 23 species were represented by 20 or more individuals. Indicator species analysis showed that eight of these species had an association with either grassland or forest vegetation significantly greater than that expected by chance (Table 1). The strongest association with grassland vegetation (high relative abundance and highly statistically significant) was for two species of anthicid (Floydwernerius australis and Omonadus. hesperi), a curculionid (Steriphus parvus), an elaterid (Agrypnus variabilis), and a staphylinid (Tmesiphorus sp.). The strongest association with forest vegetation was apparent for a species of anthicid (Pseudotomoderus sp.), a chrysomelid (Monolepta sp. near arida) and a nitidulid (Thalycrodes sp.; Table 1). Plots of mean abundances of these species (Fig. 4) indicated that Pseudotomoderus sp. (Anthicidae) was more abundant in forest (Fig. 4a), whereas O. hesperi (Anthicidae) was more abundant in grassland (Fig. 4e). We found a negative linear relationship between absolute abundance (F1 = 7.60, P = 0.033, r2 = 0.49, Fig. 5a) of beetles and grassland area, and a similar pattern for species richness (F1 = 5.69, P = 0.054, r2 = 0.40, Fig. 5b). In contrast, we found a positive relationship between grassland area, and assemblage dissimilarity
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Table 1 Summary of indicator species analysis for 20 common beetle species showing number of individuals sampled, relative abundance, number of sites the species was recorded from, its associated vegetation type Family
Species or morphospecies
No. individuals
Relative abundance
Number of sites
Forest
Forest
Grassland
Grassland
Vegetation association
P value
Aderidae
Aderidae sp.
275
48
52
7
7
Anthicidae
Pseudotomoderus sp.
114
86
14
8
5
Forest
0.001
0.852
Anthicidae Anthicidae
Pseudocyclodinus glaber (King) Floydwernerius australis (King)
161 56
56 3
44 97
6 2
5 8
Grassland
0.665 0.001
Anthicidae
Omonadus hesperi (King)
103
1
99
1
8
Grassland
0.001
Anthicidae
Omonadus sp.
77
55
45
4
6
0.872
Scarabaeidae
Ataenius sp.
23
2
98
1
2
0.469
Carabidae
Platycoelus prolixus (Erichson)
98
27
73
6
6
0.371
Carabidae
Pseudoceneus iridescens (Castelnau)
39
74
26
5
3
0.219
Chrysomelidae
Monolepta sp. nr. arida Lea
Corylophidae
Sericoderus sp.
29
84
16
8
3
221
25
75
7
8
Curculionidae
Steriphus parvus (Blackburn)
135
11
89
7
8
Curculionidae
245
44
56
8
6
Elateridae
Xyleborinus saxesenii (Ratzeburg) Agrypnus guttatus (Cande`ze)
Elateridae
Agrypnus variabilis (Cande`ze)
Leiodidae
Nargomorphus sp.
182
73
27
8
6
73
13
87
4
8
2,209
66
34
8
6
Forest
0.008 0.074
Grassland
0.010 0.961 0.058
Grassland
0.001 0.066
Nitidulidae
Thalycrodes sp.
51
93
7
6
1
Forest
0.015
Staphylinidae Staphylinidae
Tmesiphorus sp. Aleocharinae sp. 1
65 161
15 75
85 25
4 6
7 4
Grassland
0.035 0.142
Staphylinidae
Aleocharinae sp. 2
337
74
26
8
8
0.058
Staphylinidae
Sepedophilus sp.
49
20
80
6
7
0.162
Staphylinidae
Aleocharinae sp. 3
52
73
27
3
3
0.640
Staphylinidae
Staphylinidae sp.
54
93
7
3
2
0.321
between each grassland and forest pair (F1 = 9.48, P = 0.022, r2 = 0.55, Fig. 4c).
Discussion We have shown that floodplain grasslands support beetle assemblages that are distinct from the surrounding floodplain forest, and that these assemblages become more compositionally distinct with increasing area of the grassland. Below we discuss the potential causes of these differences, and the conservation implications of changed flooding regimes for beetle assemblages in this ecosystem. Compositional differences We found differences in the composition of the beetle fauna between grassland and forest vegetation, but not overall beetle species richness or abundance. For example, we found differences in the relative proportion of feeding groups as well as dominant families present in the forest and grassland vegetation. The primary driver of these
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compositional differences is likely to be the differences in microclimate and resources in each vegetation type (D’Odorico et al. 2012). It is widely understood that many families of beetle prefer particular habitat types that provide suitable micro-environmental conditions and food resources (Hunt et al. 2007; Lawrence and Britton 1994). Thus, the dominance of particular families in each of the grassland and forest vegetation types matches other studies that have demonstrated differences in abundances of particular families in environments with contrasting microenvironments (Barton et al. 2009; Lindsay and Cunningham 2009). Underlying the vegetation differences between grassland and forests is the topography of the area, which leads to spatial variation in flooding disturbance, and pooling of water in the grasslands. Periodic flooding of the grasslands temporarily removes the availability of this habitat for terrestrial species. The species that persist in floodplain grasslands may have submersion-tolerant eggs, which has previously been identified as a mechanism for some groups of arthropods in floodplain ecosystems (Rothenbucher and Schaefer 2006). However, differential immigration into grasslands by some species after the water has receded is also
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Pseudotomoderus sp. (Anthicidae) -Forest Detritivore
(a)
Steriphus parvus (Curculonidae) - Grassland Herbivore
(b) 25
16
Abundance
Forest
12
Grassland
10 8 6 4
2
Omonadus hesperi (Anthicidae) - Grassland Detritivore
Steamer
Keyes
Hut
Harbour
Rabbit
Rushy
Thalycrodes sp. (Nitidulidae) - Forest Detritivore
(f)
12
4
Bucks
Steamer
Keyes
Hut
Harbour
Rabbit
0 Rushy
0 Top
2 Bucks
2
Monolepta sp. nr. arida (Chrysomelidae) - Forest Herbivore
Tmesiphorus sp. (Staphylinidae) - Grassland Predator
(h)
5
Steamer
4
6
Keyes
6
Forest Grassland
8
Hut
8
10
Harbour
Forest Grassland
Rabbit
10
Abundance
12
Rushy
(e)
Bucks
Steamer
Keyes
Hut
Harbour
Rabbit
Rushy
Top
Bucks
0
Steamer
4
Keyes
6
Hut
Grassland
8
Forest Grassland
Harbour
Forest
10
10 9 8 7 6 5 4 3 2 1 0 Rabbit
12
Abundance
14
Agrypnus variabilis (Elateridae) - Grassland Herbivore
Rushy
16
Top
Bucks
(d)
18
Abundance
10
Top
Floydwernius australis (Anthicidae) - Grassland Detritivore
Abundance
15
0
Steamer
Keyes
Hut
Harbour
Rabbit
Rushy
Top
Bucks
0
(g)
Forest Grassland
5
2
(c)
20
Top
Abundance
14
8
Abundance
Abundance
7 4 Forest Grassland
3 2
6
Forest Grassland
5 4 3 2
1
1 Steamer
Keyes
Hut
Harbour
Rabbit
Rushy
Top
Steamer
Keyes
Hut
Harbour
Rabbit
Rushy
Top
Bucks
Bucks
0
0
Fig. 4 Abundance of eight different species (a–h), each with a significant association with grassland or forest vegetation (see Table 1). Locations are ranked left to right in order of increasing area of the floodplain grassland. Columns show mean ± SE
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(a)
600 R 2 = 0.49
Abundance
500 400 300 200 100 0 1.0
10.0
100.0
1000.0
Area (Hectares)
(b)
50 R 2 = 0.40
Species richness
40 30 20 10 0 1.0
10.0
100.0
1000.0
conclusions about temporal variability in forest-grassland differences. Nevertheless, we hypothesise that abundances would be higher following a recent flood across both vegetation types, with time since flooding previously identified as a driver of invertebrate abundance in this ecosystem (Ballinger et al. 2005). We also suspect that a measurable contrast between forest and grassland would remain due to the persistent differences in vegetation composition and structure. We found that two of the eight species that displayed a significant association with a particular vegetation type were herbivores (one chrysomelid found in forest, and one curculionid found in grassland), suggesting plant specialisation may be a factor affecting compositional differences. Although we did not measure detailed floristic variables in our grassland or forest sites. It is possible that these herbivorous beetles feed on particular plant species restricted to each vegetation type. Our indicator species analysis also revealed four other species of beetle (detritivores and predators) that were associated with the grasslands. This suggests there are several beetles species strongly associated with the grassland vegetation per se, and could therefore be at particular risk of decline if the grasslands are invaded by forest plants or become reduced in area.
(c)
0.8
Bray-Curtis dissimilarity
Area (Hectares)
0.7
The importance of grassland area
0.6 0.5 R 2 = 0.55
0.4 0.3 1.0
10.0
100.0
1000.0
Area (Hectares)
Fig. 5 The relationship between floodplain grassland area and beetle a abundance, and b species richness of beetles in the grasslands, and c assemblage dissimilarity between each grassland-forest pair
likely to be important (Rothenbucher and Schaefer 2006). These species may recolonise from other grasslands that were not flooded, or perhaps from forest where they persist in lower densities until able to re-enter the grassland environments. Only during extreme and rare flooding events would all grasslands be inundated, with smaller grasslands flooded less frequently. This indicates the smaller grasslands might act as refuges for some grassland specialists to recolonise the larger and more frequently flooded grasslands. Our sampling was conducted for only a single week and during a prolonged drought, therefore limiting any
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Our study demonstrated a positive relationship between grassland area and beetle assemblage dissimilarity from the surrounding floodplain forest. Reasons for the abundance differences include a potential ‘spill-over’ effect from the forest (sensu Tscharntke et al. 2012), where some highly abundant forest species move into or through the smaller grasslands. A likely explanation for this is the stronger influence of edge effects on the smaller grasslands. That is, the greater contact of smaller grasslands with forest relative to their area (perimeter to area ratio) increases the likelihood of forest species crossing into the smaller grasslands. However, fewer beetles may occur in the larger grasslands because of the environmental contrast with the forest. That is, more variable temperatures, wind speed, and sun light may produce a less favourable environment for forest species (Barton et al. 2009; Lindsay and Cunningham 2009; Mazia et al. 2006). Thus, the bigger the grassland, the more different the environmental conditions are from the surrounding forest, leading to a reduced edge effect. Habitat contrast has also been shown to moderate edge effects on dung beetles in the tropical forests of South America (Peyras et al. 2013). This assumes that while we sampled the grasslands at the same distance from the nearest forest edge, the distance to the other edge is also important. In small grasslands the other side (edge) of the grassland is closer, the local climate is closer to that of the
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surrounding forest, and edge effects may extend further into the grassland. Increasing distinctness of beetle assemblages in larger grasslands also may be due to population ‘mass’ effects (Shmida and Wilson 1985). That is, populations of grassland beetles, such as F. australis and S. parvus, was higher in the larger grasslands, increasing their likelihood of being sampled. Flooding frequency and duration, which affects soil moisture and vegetation, are also important factors that may be driving the patterns we report, as indicated by the negative relationship between area of the grasslands and their commence-to flood volumes. Comparatively lower soil moisture in areas subject to less frequent flooding have been shown to decrease the richness or abundance of insect assemblages in several parts of the world, including beetles and springtails in flooded grasslands in Germany (Lessel et al. 2011; Vasconcelos et al. 2010), and ants in flooded forests in Brazil (Vasconcelos et al. 2010). Conversely, a higher frequency of flooding in river grasslands in Germany was associated with a greater diversity of species (Gerisch et al. 2012). Previous research in river red-gum floodplain forests by Ballinger et al. (2005, 2007) has shown that time-since-flooding is also an important variable determining the abundance and biomass of invertebrates. Specifically, abundance was highest soon after the flooding events, as invertebrates take advantage of new resources (Ballinger et al. 2005), indicating soil moisture is an important factor determining arthropod abundance and composition in river red-gum floodplain forests. The present study was conducted after several years of drought conditions, during which time the forest received floods from managed delivery of environmental water (King et al. 2010; Murray Darling Basin Authority 2012). It is not clear how the drought might have reduced or enhanced the effect of grassland area on the contrast between forest and grassland patches. Key knowledge gaps therefore remain about the effect of flooding on beetle assemblages in this important ecosystem. These include the timing of migration of beetles back into grasslands following inundation, the role of forest as refugia during flooding, and what traits of beetle species might influence their temporal response to flooding. Conservation implications The floodplain forests in our study area, which lies within the Murray-Darling Basin, have been subject to changes in flood regimes, with reductions in frequency and duration of flooding events (CSIRO 2008; Mac Nally et al. 2011; Sims et al. 2012). This has led to a number changes to the ecology of the floodplain forest ecosystem, including dieback of river red gum trees within the forest itself (Cunningham et al. 2009), invasion of weeds into drier areas
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(Catford et al. 2011; Stokes et al. 2010), and encroachment of river red gum into the grasslands (Bren 1992). Thus, in future there is a risk that grasslands in the Barmah-Millewa forest are likely to become smaller, with the smallest ones at risk of disappearing. Under natural flood regimes, the grassland plains become temporary wetlands for up to 6 months of the year, with Moira grass exhibiting maximum growth during floods of C45 GL/day that inundated the grasslands to depths of C1.5 m. Flow regulation of the River Murray commenced in 1936 with the completion of the Hume Dam. The frequency of modelled high flow events (C45 GL/day) has halved since then, the period between events has doubled and their mean duration reduced by 26 % (Colloff et al. 2013). Flooding limits the extent of some forest plant species such as the river red gum, as well as some invasive species which cannot withstand prolonged inundation (Catford et al. 2011; Vivian and Godfree 2012). From a longer-term perspective, there is increasing likelihood of the grasslands decreasing in area, and a reduction in the vegetation contrast between floodplain grasslands and the surrounding forest as grasslands become in-filled by forest (Colloff et al. 2013). This would have the effect of reducing the spatial heterogeneity of the beetle fauna (and likely the other insect fauna) as well as reducing the habitat available to grassland specialists. Efforts to return environmental water flows to this ecosystem, as outlined in the Barmah-Millewa Forest environmental water management plan (Murray Darling Basin Authority 2012) are likely to be beneficial for maintaining the floristic composition and structure of the grasslands, the landscape-scale mosaic this produces, and spatial heterogeneity in the broader insect community. Acknowledgments We are grateful to Raphael Didham and John Evans for providing thoughtful comments on an early version of the manuscript. We also thank Rolf Oberprieler, Adam Slipinski, Tom Weir and Hermes Escalona Garcia at the Australian National Insect Collection, CSIRO Ecosystem Sciences, for confirming the identity of some beetle species. We thank our research partners, Keith Ward and Neville Atkinson, Goulburn-Broken Catchment Management Authority; Paul O’Connor, Department of Sustainability and Environment; Lee Joachim, Yorta Yorta Nation Aboriginal Corporation; and Kane Weekes, Parks Victoria. Fieldwork was conducted under permit from Department of Sustainability and Environment and Parks Victoria. This research was funded by CSIRO Water for a Healthy Country National Research Flagship.
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