Hydrobiologia (2009) 626:67–77 DOI 10.1007/s10750-009-9737-5
SALT LAKE RESEARCH
Response of fringing vegetation to flooding and discharge of hypersaline water at Lake Austin, Western Australia Eddie J. B. van Etten Æ Simone E. Vellekoop
Published online: 26 February 2009 Ó Springer Science+Business Media B.V. 2009
E. J. B. van Etten (&) S. E. Vellekoop Centre for Ecosystem Management, School of Natural Sciences, Edith Cowan University, 270 Joondalup Drive, Joondalup 6027, WA, Australia e-mail:
[email protected]
of the vegetation immediately fringing the lake bed and major inlet channels resulted in dramatic changes to the species composition of annual and short-lived species and growth of perennial species. Flooding resulted in substantial death and damage to perennial shrubs (particularly Halosarcia fimbriata) due most likely to a combination of several weeks/months of inundation and smothering by macroalgae and Ruppia, with smaller plants and those closer to the lake bed impacted upon to a greater degree. Seed germination and recruitment of new Halosarcia plants was substantial as floodwaters receded with the majority of these seedlings surviving some 2 years after flooding despite the severe drought that followed the flood. Growth rates of seedlings differed substantially and were linked to subtle differences in microtopography. Recruitment following flooding was also demonstrated in in vitro experiments involving inundated soil cores, provided water was relatively non-saline (conductivity \30 mS cm-1). A conceptual model is proposed to explain changes in fringing vegetation in response to frequency, depth, period and salinity of flooding in relation to micro-topography. Despite the profound effect of flooding, impacts of discharge were identified, with changes in topsoil pH and salinity greater in areas closer to the discharge than those further away. Impacts on vegetation characteristics were not detected.
Present Address: S. E. Vellekoop OceanaGold Ltd, 1 Hattie Street, Reefton, New Zealand
Keywords Salt lake Fringing vegetation Salt marsh Flooding Mine water discharge
Abstract Patterns and dynamics of the salt marsh vegetation that surrounds many of the salt lake systems of arid/semi-arid Australia are poorly known. Lake Austin is a very large salt lake with extensive areas of fringing salt marsh; it is located in the arid Yilgarn Region of Western Australia. In this study, the changes in this vegetation over a 4-year period (1998–2002), during which both a major flooding event and addition of hypersaline groundwater from a nearby mining operation occurred, are reported. The monitoring program, based on Before-After-ControlImpact (BACI) principles, was designed to detect impacts of discharging hypersaline water into the lake; however, the flooding event, the result of above average rainfall in early 2000, complicated the results. The rains of 2000 and subsequent inundation
Guest Editors: J. John & B. Timms Salt Lake Research: Biodiversity and Conservation—Selected papers from the 9th Conference of the International Society for Salt Lake Research
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Lake Austin is a very large salt lake in this region, but is believed to be an endorheic basin with the lake bed being the ultimate source of surface drainage and groundwater discharge within the catchment (Curry et al., 1994). The lake bed is mostly dry, remarkably flat, variably salt-crusted and devoid of vegetation. Little is known about the biota of the large salt lakes of arid, inland Australia, with most studies concentrating on temporal changes in aquatic invertebrates and water chemistry (e.g. Geddes et al., 1981; Halse et al., 1998; Williams et al., 1998). Such lakes typically are surrounded by salt marsh vegetation dominated by Halosarcia spp. and other succulent and salt tolerant members of the chenopod tribe Salicornieae, commonly known as ‘samphires’ (Wilson, 1980; Datson, 2002). Recently all these taxa have been placed in the genus Tecticornia (Shepherd & Wilson, 2007). In places, this fringing salt marsh vegetation comprises a narrow strip between the lake bed and surrounding dune systems. Elsewhere,
Introduction Large salt lakes (or playas) are a common feature of arid areas, and are particularly widespread in Australia (Shaw & Thomas, 1999). They form a large proportion of inland aquatic ecosystems at a global scale and face a range of threats, not the least being enhanced salinisation and pollution (Williams, 1993, 2002). Some of Australia’s largest and most wellknown salt lakes (such as Lake Eyre and Lake Torrens) represent structural depressions in extensive sedimentary basins. In concordance with the low and highly variable rainfall of inland Australia, most are ephemeral, filling only after abundant rainfall in the catchment, although many have at least some surface water in most years (Geddes et al., 1981). The Yilgarn Plateau, an ancient (Archaean) granitic shield, covers most of central and southern Western Australia; here, the many salt lakes generally form chains along vast paleo-drainage channels (Fig. 1).
Western Australia 400 km
Cue
Lake Austin 40 km
(b) North
(a) Fig. 1 Extensive paleo-drainage systems of Western Australia showing location and detail of Lake Austin and the township of Cue
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expansive salt marshes are found, particularly around inlet channels (Curry et al., 1994; Datson, 2002). The fringing vegetation immediately adjacent to the lake bed and inlet channels is subject to inundation but only following substantial rainfall episodes. At Lake Austin, almost twice the mean annual rainfall (of 232 mm) was recorded over a 5-month period in early 2000. This resulted in flooding of the lake and inundation of the fringing vegetation for several months. This was followed by a period of well below average rainfall during which the lake slowly dried out. The impact of this flooding and subsequent drought on fringing vegetation at Lake Austin is the primary subject of this paper. Salt marsh responses to flooding and drying are well known for tidal/coastal salt marshes (e.g. Huckle et al., 2000); however, inland playa systems have received little attention, with most research concentrated on the Great Plains playas of USA (e.g. Haukos & Smith, 2001). The Yilgarn Plateau is a major centre for mining and minerals processing. Shallow and saline groundwater systems often represent a challenge for mining in the region. In recent years, it has become common practice for mining companies to dispose of excess groundwater into salt lakes (Ward, 2002). Although a licence is required for such discharge into a salt lake, impacts are poorly understood. From 1999 to 2002, Fig. 2 Monthly rainfall for Cue from January 1996 to August 2002 (solid line). Monthly means are also shown (broken line). Arrows indicate monitoring dates (Bureau of Meteorology, 1998)
hypersaline groundwater (110–180 g l-1 total dissolved solids) intercepted from a nearby mine was pumped into Lake Austin. Monitoring before and during this period occurred both adjacent and distant from the discharge point in an attempt to detect impacts of hypersaline water discharge on fringing vegetation; this is the second subject of the paper.
Materials and methods Study area The study area is located on the north-west side of Lake Austin, a large salt lake located in the Murchison region of Western Australia, some 650 km NNW of Perth (27°290 S, 117°390 E; Fig. 1). Mean annual rainfall is 232 mm (1894–2004) with a bimodal monthly distribution (Fig. 2). Inter-annual rainfall variability is high (Fig. 2; Bureau of Meteorology, 1998). Summers are characteristically hot (mean min. 21.9°C, mean max. 36.9°C), and winters are mild (mean min. 7.6°C, mean max. 19.3°C). Landforms are characterised by salt lakes with extensive fringing saline plains, dunes and sandy banks, supporting low halophytic shrublands and scattered tall acacia shrublands (Curry et al., 1994).
160
RAIN AVERAGE
140
120
RAIN (mm)
100
80
60
40
20 0 02 20 AY 02 M 20 N JA 001 2 P 1 SE 00 2 AY 1 M 0 20 N JA 000 2 P 0 SE 00 2 AY 0 M 0 20 N JA 999 1 P 9 SE 199 AY M 999 1 N JA 998 1 P 8 9 SE 19 AY 8 M 9 19 N 7 9 JA 19 P 7 SE 99 1 AY 7 M 99 1 N 6 9 JA 19 P 6 9 SE 19 AY 6 9 19 M
N JA
MONTH
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The salt lake bed is flat, often salt encrusted and unvegetated. Vegetation patterns surrounding the lake are related to salinity-topographic gradients from lake edge to higher sand dunes. Halosarcia fimbriata dominated salt marsh typically occurs on flats at the lake edge giving way to Halosarcia halocnemoides salt marsh on crests and upper slopes of fringing banks. Inundation during episodic flood events seems limited to the H. fimbriata community. When not flooded, soils of this community are highly saline (EC of 10–80 mS cm-1) loams at the surface, but clayey and saturated just below the surface, suggesting active groundwater discharge (Mabbutt, 1977). Rainfall and hydrology Monthly rainfall for Cue varied from 0 to 150 mm between January 1996 and August 2002 (Fig. 2). High rainfall variability is clearly a feature, with nil or negligible rain possible in any month. In contrast, monthly rainfall several times the average is also a regular feature. This occurred in: June and July of 1996; February, April and August of 1997; May, July, August and December of 1998; March and December of 1999; and January, March and April of 2000. Overall, 1,700 mm of rain was received between December 1995 and June 2000, which is 37% above the average expected for this period. The period between December 1999 and April 2000, during which a number of cyclonic, low pressure systems moved inland from the north-west coast, was clearly the wettest 5 months of recent times. This resulted in the lake being full during the first half of 2000. Rainfall was mostly below average following the flood (Fig. 2). In summary, monitoring covered both a pluvial and a drought period. Little is known of the hydrology of the lake and no detailed measurements of lake levels are reported. The lake is usually dry, but fills in response to large rainfall episodes in the surrounding catchment. The degree to which water entering the lake is derived from surface run-off via drainage lines, as opposed to surface expression of rising groundwater, is unknown. At the time of the initial (baseline) survey in September 1998, the lake contained a reasonable amount of water, contributed by above-average rainfall during winter of that year, but was not nearly full. Lake levels remained below fringing salt marsh vegetation. Substantial summer-autumn rains of
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1999/2000 contributed to extremely high lake levels, which inundated much of the lower parts of the fringing vegetation around the lake and inlet channels. At the June 2000 survey, the edges of fringing vegetation were still flooded in many places, although floodwaters had receded from their peak of April that year by several centimetres (T. Wilkeis, pers. comm.). Water levels gradually receded after the 2000 flood through evaporation. During March 2001, there were some discrete ponds of water on the lake bed, but at the time of the final assessment (April 2002), water remained only in the deeper drainage lines entering the lake and in the area immediately around the discharge point. Analysis of rainfall records over 110 years revealed that rainfall to produce flooding similar in extent to that of 2000 occurs approximately every 10 years, with the lake filling (but not flooding fringing vegetation) some three times per decade. Monitoring program and experimental design During September 1998 (before commencement of discharge), 14 permanent monitoring sites were established in the fringing salt marsh vegetation immediately adjacent to the lake edge, all of which were inundated to some extent during the floods of 2000. Of these, seven sites were close (within 1 km) to the discharge point (hereafter called ‘discharge’ sites), with the other seven (called ‘non-discharge’ sites) located distant ([5 km) from the discharge point. These non-discharge sites were originally selected to act as a control as it was anticipated that they would not be impacted by discharge of hypersaline water. At each site, a 10 9 5 m plot was randomly placed (with long side parallel to lake edge). Percentage cover and abundance (i.e. number of individual plants) of species were estimated. Three specimens of each dominant shrub species within each plot were randomly selected, tagged and then measured for height, shrub width (along the direction of transect) and % health (percentage of the total plant volume living and deemed healthy). Topsoil (surface to 1 cm depth) and subsoil (1–4 cm depth) samples were taken at random points across the plots. Samples were analysed later for electrical conductivity (EC) and pH using appropriate probes placed into a 1:5 soil to deionised water solution. The particle size
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distribution of these soils was also determined following sieving. Discharge commenced in April 1999 at around 50,000 kL per month, rising to 180,000 kL per month in late 2000, and then gradually declining to zero by October 2002 as mining activities were scaled down. Regular water monitoring revealed that the salinity of discharge water varied from 110 to 180 g l-1 (EC 100–130 mS cm-1), mainly due to NaCl. Sites were remeasured during June 2000 (some 3 months after the peak flooding event) and in April 2002. Inundation of some low lying sites prevented soil collection at all sites in 2000. In May 2002, 20 additional sites were established in fringing H. fimbriata community and measured as before. In addition, three 1-m2 quadrats were placed randomly inside each plot and the height, width and % health of individuals of each perennial species were recorded to examine patterns of recruitment and persistence. Each of these additional sites was located at a similar distance from the lake bed in the same vegetation and soil type, with equal numbers of sites close to and distant from the discharge point. Sixteen intact surface soil cores (to 4–5 cm depth) were collected for a soil seed bank experiment from 11 of the additional sites (six discharge zone and five non-discharge zone) and placed in plastic punnets. The punnets were transported to Perth, placed in large flooding trays within a glasshouse and subjected to inundation with discharge water of one of four levels of dilution for a period of 64 days after which they were allowed to slowly dry out. The four levels of dilution were 100% (undiluted), 50%, 25% and 0% of discharge water (collected from the discharge pipe) with deionised water used for the dilution. The 0% was all deionised water. The 25% had an electrical conductivity similar to that found at the lake during the flooding event of June 2000. In addition, 16 extra soil samples from one site were not inundated but placed in the irrigated section of the glasshouse as a control to gauge the effect of flooding on germination and recruitment. The pH, electrical conductivity and temperature of the treatment water were measured and numbers of emergent seedlings were counted weekly. Replicate units in this case were flooding trays that each contained single punnets from each site, giving four replicates for each flooding treatment. Impact of flooding was tested using repeated measures analysis of variance to compare mean site
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values across three monitoring dates, with zone used as a between-subject factor to test for impacts of discharge (or, in other words, differences in response across time between sites close to discharge compared with those distant from discharge). Mauchly’s test of sphericity was applied to test homogeneity of variance assumptions, with Huynh-Feldt correction used where variances were not shown to be equal. Post-hoc power tests were performed to determine probability of detecting differences.
Results Temporal change Fringing vegetation was found to be floristically simple with generally only 1–3 perennial Halosarcia shrub species and nil to several species of annuals (depending on rainfall) per site. H. fimbriata was dominant and present at 12 of the 14 fringing sites, whilst two morphologically different forms of H. halocnemoides were found at three sites each, and H. pergranulata at two sites. Major changes in richness and composition of annual plant species occurred over the 4 years of monitoring, while composition of woody perennials was highly consistent. Daisies (family Asteraceae) dominated the short-lived flora during September 1998, whereas grasses (family Poaceae) were the most common component in June 2000. No annuals were found in fringing vegetation in 2002. Although composition of perennial species, in terms of presence/absence, was consistent, significant differences in cover and abundance of the shrub species occurred (Table 1). Total abundance of all shrubs (Fig. 3c) and specifically that of H. halocnemoides form ‘b’, increased at each monitoring date, whereas cover, particularly of H. fimbriata (Fig. 3e), declined (Table 1). Concomitant with declining cover, the mean % health of shrubs declined between monitoring periods, with H. fimbriata (Fig. 3d) and H. halocnemoides form ‘a’ (Table 1) showing significant declines. No growth of shrubs was demonstrated over the monitoring period except for H. halocnemoides form ‘a’. This taxon occurs on higher ground than the others and grew significantly in width (Table 1). Mean surface soil salinity (EC) and pH were both significantly lower after
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Table 1 Mean values (with standard deviation, SD, in parenthesis) of plant and soil variables at each monitoring date and results of mean comparison using repeated measures ANOVA showing F-value and significance levels (*** P \ 0.001; Species
Variable
** P \ 0.01; *P \ 0.05; ns: not significant) for both time and time 9 zone (discharge vs. non-discharge area) interaction (nm = not measured)
Mean values (SD) 1998 (Pre-discharge)
All shrubs
H. fimbriata
H. halocnemoides form ‘a’
H. halocnemoides form ‘b’
H. pergranulata
Top soil (0–1 cm)
Sub soil (1–4 cm)
2000 (Post-discharge; flood)
2002 (Post-discharge & flood)
Factor (Time)
Time 9 Zone
Height (m)
0.36 (0.13)
0.35 (0.14)
0.34 (0.15)
0.93 ns
0.64 ns
Width (m)
0.34 (0.18)
0.38 (0.24)
0.37 (0.23)
1.9 ns
0.45 ns
Health (%)
60.4 (15.4)
53.4 (20.9)
40.1 (21.6)
30.4***
0.24 ns
Cover (%)
20.2 (20.4)
19.3 (20.1)
16.6 (17.2)
3.0 ns
1.1 ns
Abundance
67.2 (61.2)
72.4 (81.8)
103 (113)
5.1*
0.98 ns
Height (m)
0.41 (0.14)
0.36 (0.15)
0.36 (0.15
2.6 ns
0.22 ns
Width (m)
0.36 (0.20)
0.39 (0.28)
0.37 (0.25)
0.33 ns
0.66 ns
Health (%)
66.2 (15.0)
52.4 (23.0)
43.4 (24.3)
17.3***
0.57 ns
Cover (%)
15.6 (11.9)
14.2 (10.3)
10.6 (9.3)
3.7*
0.47 ns
Abundance Height (m)
47.2 (22.4) 0.37 (0.17)
50.6 (34.4) 0.42 (0.16)
77.3 (86.3) 0.44 (0.16)
1.0 ns 2.8 ns
0.92 ns 0.72 ns
Width (m)
0.41 (0.17)
0.44 (0.23)
0.51 (0.26)
6.2*
4.4*
Health (%)
56.9 (14.4)
65.9 (18.6)
37.5 (14.6)
20.0***
3.7 ns
Cover (%)
13.3 (7.6)
12.7 (6.8)
18.7 (11.1)
0.52 ns
0.09 ns
Abundance
52.3 (32.1)
46.0 (24.2)
88.3 (48.1)
2.2 ns
0.40 ns
Height (m)
0.28 (0.06)
0.30 (0.09)
0.33 (0.11)
0.69 ns
0.52 ns
Width (m)
0.28 (0.11)
0.31 (0.12)
0.32 (0.14)
0.43 ns
1.8 ns
Health (%)
47.5 (10.0)
57.5 (10.5)
45.0 (10.4)
3.7 ns
3.7 ns
Cover (%)
40.0 (30.4)
44.0 (31.4)
35.0 (21.7)
1.4 ns
3.6 ns
Abundance
166 (95.6)
203 (145)
308 (102)
16.3*
7.2 ns
Height (m)
0.40 (0.06)
0.39 (0.08)
0.34 (0.21)
0.52 ns
n.a.
Width (m)
0.40 (0.08)
0.48 (0.15)
0.36 (0.29)
1.09 ns
n.a.
Health (%)
51.0 (8.9)
49.6 (15.8)
34.0 (27.2)
2.8 ns
n.a.
Cover (%)
36.5 (47.4)
31.5 (40.3)
31.5 (40.5)
1.1 ns
n.a.
Abundance Salinity (mS cm-1)
47.0 (22.4)
50.6 (34.4)
77.3 (86.3)
0.9 ns
n.a
15.3 (15.8)
7.6 (4.6)
26.7 (22.5)
8.2*
2.7 ns
pH
8.7 (0.7)
8.2 (0.4)
8.6 (0.4)
4.4*
5.5*
% [2 mm
2.4 (2.2)
3.6 (7.4)
1.50 (0.87)
1.5 ns
0.75 ns
Salinity (mS cm-1)
6.8 (5.2)
nm
16.0 (13.1)
9.2*
0.35 ns
pH
8.7 (0.7)
nm
8.4 (0.3)
2.1 ns
1.5 ns
floodwaters receded in 2000, but then increased during the drying period of 2000 to 2002 when they were greater than the baseline (1998) levels (Table 1). Flooding stimulated prolific growth of Ruppia sp. (a fine-leaved aquatic angiosperm) and macroalgae in shallow waters. This growth completely covered most submerged parts of shrubs in the flooded fringing
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F-values
vegetation and remained as dead plant matter (‘wrack’) on plants and soil surface after floodwaters receded. Fully submerged plants (due to smaller size and/or lower elevation) were often completely covered in wrack for several months. The impact of flooding was greater on smaller sized shrubs, with a modest but significant correlation (r = 0.42; P = 0.008) between the size of H. fimbriata prior to flooding and
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(a)60000
Discharge Non-discharge
Discharge Non-discharge
9
40000 8.5
pH
EC (uS/cm)
50000
(b) 9.5
30000
8 20000 7.5
10000
7
0 Sep-98
Discharge Non-discharge
140
Apr-02
Discharge Non-discharge
60
120 100 80
50 40 30
60
20
40
0
Jun-00 Monitoring period
70
10
20
(e)
Sep-98
(d) 80
180 160
Abundance (#individuals)
Apr-02
% Health
(c)
Jun-00 Monitoring period
Sep-98
Jun-00 Monitoring period
Apr-02
0
Sep-98
Jun-00 Monitoring period
Apr-02
25 Discharge Non-discharge
Cover (%)
20
15
10
5
0
Sep-98
Jun-00 Monitoring period
Apr-02
Fig. 3 Trends in mean values (with standard error bars) over monitoring period comparing discharge and non-discharge sites for: (a) topsoil EC; (b) topsoil pH; (c) total abundance (all shrubs); (d) health status of H. fimbriata and e) cover of H. fimbriata
decline in health across the monitoring period. Additionally, the health status of fringing shrubs in 2002 was significantly correlated with plant height (r = 0.46; P = 0.005). Large (but highly variable) amounts of seedling recruitment were present in 2002 following the recession of floodwaters with densities of seedlings varying from\1 to 113 m-2 (mean = 13; SD = 28). Seedling heights also varied considerably with patches of seedlings of similar height observed seemingly in response to surface soil patterns.
During inundation of soil samples in the glasshouse with discharge water of various dilutions, Ruppia grew in samples of the 0% and 25% treatments only. Dense macroalgae also developed on the water surface in 0% and 25% treatments, but only in soil samples collected from two particular sites. Once water levels receded through evaporation, seedlings of H. fimbriata were only found in the 0% treatment (mean of 28.3 plants with std. error of 12.3), apart from one seedling emerging in the
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irrigated (i.e. non-flooded) control. Water salinity (EC) and pH increased over the period of flooding as water evaporated: 100% treatment increased, on average, from 117 mS cm-1 to 156 mS cm-1; 50% increased from 79 mS cm-1 to 120 mS cm-1; 25% from 54 mS cm-1 to 81 mS cm-1; and 0% from 20 mS cm-1 to 27 mS cm-1. Two of the four 0% replicates had significantly (P = 0.024) lower conductivity as refilling was necessary early in the experiment to replace water that leaked from cracks in the flooding trays. One of these replicates had by far the highest seedling number (48).
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was found between discharge and non-discharge zones, both in the field and flooding experiments. Seedling survivorship in the field as at May 2002 however showed that, where recruitment of H. fimbriata on the shoreline had been substantial, more seedlings were dead than alive at sites of enhanced soil salinity near the discharge point, whereas in areas of lower salinity further away from the discharge point, the opposite trend occurred. The difference in ratio of dead to live seedling abundance was significant (Mann-Whitney U-test; z = -2.2; P = 0.03).
Discharge impacts Discussion With one exception, there were no statistically significant interactions between time and zone (Table 1), demonstrating that almost none of the changes in vegetation and plants measured over time were different between sites near to discharge and those distant from discharge. The exception was width of H. halocnemoides form ‘a’, which grew more in the non-discharge zone than the discharge zone. Surface soil pH increased to a significantly greater degree during the period 2000 to 2002 in the discharge zone compared with the non-discharge zone (Table 1; Fig. 3b). The trend in surface soil salinity was similar (Fig. 3a), but the difference between zones was not significant (Table 1). pH and salinity of surface soils were highly correlated (r = 0.76; P \ 0.001). Post-hoc power tests showed that there was a high probability that significant differences could not be detected, assuming of course they exist. This was generally due to sampling intensity not being great enough to counter the high spatial variability in the level of change over time. Power tests suggest that the sampling effort should have been in the order of 30 to 150 monitoring sites (depending on the parameter), far more than the 14 established in this study. Extra sampling in the fringing H. fimbriata vegetation during May 2002 demonstrated that the EC of topsoil in the non-discharge zone (mean = 17.7 mS cm-1) was lower than that in the discharge zone (mean = 49.8 mS cm-1; P = 2.4 9 10-5). EC values varied widely in the fringing vegetation of the discharge zone (SD = 19.2), but were clearly higher at sampling points immediately around the discharge point. Despite being extremely patchy, no difference in seedling recruitment
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Fringing salt marsh dynamics Assessment was primarily designed to detect impacts of mine water disposal and the marked pluvialdrought transition was fortuitous as it facilitated measurement of impacts and recovery of fringing vegetation in response to flooding and subsequent drying, a one in 10-year event at Lake Austin. This study represents a preliminary account of vegetation dynamics in that only one such transition in one section of a single salt lake has been studied. As there is an obvious dearth of published information on such phenomena, this study is valuable. Specifically, flooding of fringing vegetation has resulted in substantial death and damage to perennial shrubs (particularly Halosarcia fimbriata) due most likely to a combination of several weeks to months of inundation and smothering by Ruppia and macroalgae. Smaller plants and those closer to the lake bed suffered to a greater degree due to longer and deeper inundation. Seed germination and recruitment of new Halosarcia individuals was substantial, although patchy, as floodwaters receded, a phenomenon not observed in the fringing vegetation at other times. Inundation experiments on intact soil cores confirmed the necessity of flooding with relatively fresh water to stimulate germination and early growth with only one germinant recorded in non-flooded (but irrigated) samples and no germination where water salinity exceeded 30–40 g l-1. Elsewhere, germination of halophytes of arid, inland salt lakes has been linked to flooding with fresh to brackish water (e.g. Khan & Rizvi, 1994; Ungar, 1995; Katembe et al., 1998;
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Egan & Ungar, 1999; Haukos & Smith, 2001). Keiffer & Ungar (1997) demonstrated that germination decreased with both duration and intensity of salt exposure for five species of halophytes collected from salt marsh in Ohio, USA. The majority of seedlings measured in the field survived (with patchy mortality) some 2 years after flooding. English et al. (2002) found that H. pergranulata seedling density increased substantially after flooding at Hannan Lake (some 400 km southeast), but then slowly declined as soil dried and surface salinity rose. Therefore, continued monitoring is recommended at Lake Austin to gauge the long-term fate of this recruitment event. Growth rates of seedlings varied substantially, associated with micro-topography. On slightly higher ground within the fringing vegetation, groups of seedlings attained 10–20 cm height, whereas on lower parts, seedlings were often less than 5 cm tall (often exhibiting symptoms of water-stress). The height plants attain in the inter-flood period is likely to play an important role in their ability to survive the next inundation. Plants on the slightly lower slopes closer to the lake bed are less likely to survive than those on higher parts of the slopes (or on slightly raised soil mounds) due to slower growth, as well as their exposure to greater frequency, depth and duration of inundation. This would tend to control the position and, possibly, plant density of the edge of fringing vegetation, which would be expected to fluctuate spatially in response to flooding regime. Substantial number of years (say 10?) between flooding events could result in an extension of the fringing vegetation toward the lake bed as plants have a chance to get to a reasonable size between floods and disperse seed at a further distance. By the same principle, a retraction in the edge is possible when flooding frequency is increased. This model requires more research to validate and quantify responses, especially in terms of spatial and edaphic constraints on colonisation toward the lake bed. Badger & Ungar (1989) suggested that shifts in salt marsh vegetation zones could occur over time in response to soil salinity changes driven by rainfall and flooding fluctuations, although this would be unlikely at Lake Austin given the substantial differences in soil and flooding between the shoreline H. fimbriata community and vegetation of adjacent fringing banks.
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Impacts of hypersaline water disposal The monitoring program was designed a priori to detect impacts (if any occurred) arising from disposal of hypersaline mine water into Lake Austin by comparing change near the discharge point to that deemed sufficiently distant to be uninfluenced by such water. Two events occurred during the discharge phase to reduce the likelihood for impacts to occur in the fringing vegetation. The first was the extension of the pipeline in the second year of the discharge some 600 m from the lake edge. The second was the flooding event of 2000, which substantially decreased and homogenised salt levels in the water around the discharge point for at least several months. Despite these events, an impact of discharge was detected in soil properties of the lake edge. Since the recession of floodwaters from June 2000, topsoil pH of fringing vegetation increased by 10% in the area close to the discharge point, significantly greater than at more distant areas (0.2% increase). Topsoil salinity revealed a similar trend (differences not statistically significant), and it is clear that certain sites close to the discharge point increased substantially in salinity. Overall, topsoil salinity is currently much higher in areas close to discharge than in areas remote from it. As pH is strongly related to salinity levels of the soil, it is reasonable to conclude that the hypersaline levels in the discharge water have increased salt and pH levels in the topsoil of the fringing vegetation. How then has this occurred? First, since the recession of floodwaters, it is possible that discharge water has spread to fringing vegetation nearest to the discharge pipe, particularly at times of high discharge volumes. Indeed, observations of salt scalds on the lake surface at April 2002, both on the ground and from the air, support this contention. In addition, the pipe was displaced by a flush of water from a nearby inlet channel in 2001 to a position quite close to fringing vegetation. The second possibility is that accumulation of salt that occurred in the fringing vegetation prior to the pipe extension persisted to some degree. Although it would be expected that salt in the topsoil would have dissolved when flooded, perhaps deeper stored salt has since risen in response to capillary rise. There is little evidence, however, that this enhanced salt loading of fringing soils close to the discharge point has led to an impact on the vegetation
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of this area. Although H. halocnemoides (form ‘a’) grew more slowly closer to the discharge point, this species typically dominates the small levee banks above H. fimbriata and was only marginally flooded at a small number of sites. As it is unlikely that this species has been in direct contact with discharge water, it is more likely that other differences between zones (such as in grazing intensity or groundwater depth) may explain such differences. There are a number of possible explanations as to why impact on fringing vegetation was not detected despite the increases in soil salinity around the discharge point. The first is that, as salt tolerant plants, Halosarcia spp. may be able to withstand the effects of increased soil salinity and pH. Short & Colmer (1999) showed that H. pergranulata, another fringing species, could grow in soil containing 800 mol m-3 of NaCl (*45 mg l-1) for several weeks. H. fimbriata was typically found in soils with EC1:5 of 10–20 mS cm-1, with values of up to 50 mS cm-1 recorded before commencement of discharge. These levels of soil salinity are typically 2 to 3 orders of magnitude higher than other salt marsh communities on slightly higher ground. This species is therefore in a league of its own in terms of salinity tolerance and it is not unexpected that it could persist and grow at elevated levels of salt (the highest recorded EC in April 2002 was 75 mS cm-1). Although no impacts on adult plants were detected, it is possible that the effects of increased salt levels may arise sometime in the future, particularly on other stages of the life-cycle. It is known that Halosarcia spp. are more sensitive to salt at the seed germination and seedling stages (English et al., 2002). The glasshouse flooding experiments showed that germination of H. fimbriata is curtailed where floodwaters are saline, with flooding by discharge water, even when diluted to 25%, resulting in no recruitment. Therefore, direct flooding of fringing vegetation by discharge waters should be avoided (e.g. through extended pipelines). The phenomenon of extensive lake flooding, albeit infrequent, effectively dilutes discharge water. The recruitment response following inundation with fresh/brackish water was shown in the field and flooding experiments to be unaffected by enhanced soil salinity closer to the discharge point. However, seedling survivorship was lower closer to the discharge point.
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Continued monitoring is warranted despite discharge ceasing in October 2002. The half a million tonnes of extra salt deposited through discharge should be regarded as a long-term addition to this endoreic lake system (Curry et al., 1994). As in 2000, much of this extra salt would become dissolved during flood events and, given the huge volumes of water at these times, would be diluted to a level where it would make only a marginal contribution to total water salinity (Williams et al., 1998). However, in contrast to what was anticipated, the deposition of salt as the floodwaters evaporated and receded was not even across the lake bed. Water drains to the lowest points in the system, which were observed to be, somewhat counter-intuitively, inlet channels. The fact that inlet channels now contain evaporative salt deposits up to 1 m in depth means that the first large flows into the lake following the end of drought periods are likely to be very high in salinity. The buildup of salt in drainage lines is likely to challenge both aquatic biota and fringing vegetation in these areas.
Conclusion The inundation of the fringing shoreline vegetation at Lake Austin following abundant rainfall resulted in death and decline of perennial halophytes (mainly H. fimbriata); however, this was compensated by substantial recruitment as floodwaters receded. Disposal of hypersaline water from a nearby mine pit resulted in raising pH and salinity of topsoil, presumably because discharge waters have at times been in contact with or close to fringing vegetation near the discharge point. This enhanced salt level seems to be having a detrimental impact on the survivorship of seedlings, but otherwise, no impacts were detected on soil seed store, recruitment potential, health and size of the main fringing species, H. fimbriata. Impacts may occur some time in the future, however, as Lake Austin is widely believed to be an enclosed drainage system. Acknowledgements We would like to thank Harmony Gold Australia, Wirralie Gold Mines P/L and Newhampton Goldfields P/L for funding the research and Bea Sommer, Robyn Loomes, Joanne Bowyer, Mike Griffiths and Charelle Harkins for helping with field work. Jo Ward, Toni Wilkeis and Mark Hewitt are also thanked for logistical support at Big Bell and Lake Austin. Paul Wilson helped with species identifications.
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