The use of macroinvertebrates in the assessment of wetland ecosystem integrity is an increasingly common tool used for management and conservation. The sensitivity of macroinvertebrates to ecological fluctuation makes them reliable and appealing indicators of ecosystem integrity. However, there is little or no published data available for assessment of wetland ecosystem integrity on the basis of macroinvertebrate species diversity in constructed wetlands of metropolitan Melbourne. The aim of the following study was to assess significant differences in macroinvertebrate diversity in three constructed wetlands in South East metropolitan Melbourne and consequently, suitability as a universal measure of wetland ecosystem integrity.
Three wetlands were chosen randomly, with the requirement that they were entirely man-made, from a list of constructed wetlands in South East metropolitan Melbourne. Between 481 and 629 organisms were found in each wetland comprising 16 different taxa. The one-way ANOVA for species richness (P > 0.05, F = 0.19) and Shannon-Weiner diversity index (P > 0.05, F = 2.54) indicate no significant differences between the wetlands in both of these measures. The data collected in the present study compared with published species richness and Shannon-Weiner diversity index data suggests macroinvertebrate species diversity can be used as a universal measure of wetland ecosystem integrity in constructed wetlands in South East metropolitan Melbourne. This is important where there is need for a rapid and streamlined tool for assessment of ecosystem integrity and consequently, the management and conservation of constructed wetlands.
The importance of wetlands lies in their environmental, economic and social values (Bennett and Whitten, 2001) and is equally applicable to constructed wetlands in developed and urbanized areas (Azous and Horner, 2001). Of the 117 wetlands in Melbourne, Australia, managed by Melbourne Water, a large number are constructed for the purpose of receiving and treating storm-water and stream inputs (Melbourne Water, 2003).
In view of Melbourne Water's treatment aims of reducing pollution from nitrogen and other contaminants in local waterways and the receiving waters of Port Phillip Bay, appropriate strategies for management and conservation of wetlands must be employed. One of these strategies involves assessment of wetland ecosystem integrity, which is defined by De Leo and Levin, (1997) in the context of conservation and economic value, as desirable stability domains that guarantee optimal exploitation rates. It is often referred to as bioassessment, biomonitoring or biological monitoring (USEPA, 1997).
The assessment of ecosystem integrity involves sampling, identification and analysis of some biotic elements in a water body, such as phytoplankton, bacteria, larval and adult aquatic insects, and other macroinvertebrates. Vertebrates such as fishes, amphibians, bird and mammal species may also be used in assessment of ecosystem integrity, but in some cases are more laborious and time consuming to sample (Adamus and Brandt, 1990). Macroinvertebrate community centrality and interrelationship with other organisms in food webs (Anamaet et al., 2005) as well as their sensitivity to changes in the trophic state of a water body, makes them highly reliable and appealing for use as indicators of ecosystem integrity and in bio-monitoring programs.
However, there is very little published data available for assessment of wetland ecosystem integrity on the basis of macroinvertebrate species diversity in constructed wetlands of South East metropolitan Melbourne (Bryant and Papas, 2007). Consequently, little is known of the macro-invertebrate community assemblages, diversity and richness in these constructed wetlands and their suitability as a measure of wetland ecosystem integrity, for the purpose of conservation or bio-monitoring programs. In response to this, the following study set out to assess the relative dynamics of macroinvertebrate species diversity in three constructed wetlands. The specific aims of the study were to collect baseline data for species richness and Shannon-Weiner species diversity index and establish any significant differences in these data between the three wetlands. Further, this data would be used to determine if macro-invertebrate species diversity could be applied to these constructed wetlands for the assessment of ecosystem integrity, and as a rapid, universal measure for comparison of constructed wetlands in other, similarly urbanized and developed geographical regions.
Materials and Methods
Description of study sites
The three representative sites chosen for the study shown in Figure 1 were selected randomly from a list of locations in the South East metropolitan Melbourne region as described in Melbourne Water's Water Sensitive Urban Design web page (Melbourne Water, 2009). The three wetlands all constructed between 2003 and 2004 treat similar discharge types from a combination of residential, commercial, industrial and roadside areas in the region. Macrophyte coverage is described in Table 1.
Sampling point selection
Random selection of sampling points in each wetland, as per Spieles and Mitsch (2000), was used, with 4 sites being selected so as to be representative of the whole wetland (Sharma and Rawat, 2009). Davis et al. (1999) similarly suggests water bodies be divided into 4 sectors labeled, north, east, south and west, or however is deemed appropriate, allowing one site to be randomly selected by the numbered grid method as per Sutherland (2006) in each sector. Random selection eliminates influence of confounding variables such as substrate, macrophyte growth and detritus (Barber and Kevern, 1973), on macro-invertebrate distribution through the wetland.
Sampling at fortnightly intervals is recommended (NRM, 2008) to overcome confounding effects of seasonal changes in hydrology, adult macroinvertebrate migrations and local climate (USEPA, 2002; Balla and Davis, 1995). Samples from each sector were taken in each wetland and pooled once a fortnight for 3 months, between 29/4/2009 and 30/9/2009.
Barlow et al. (1982) describe sweep and dredge nets as being most satisfactory in capturing motile macroinvertebrates, with dredge nets allowing greater areas to be sampled. Sweep netting is also described by Sutherland (2006) as the most suitable sampling method as it captures the lower water column and upper sediment layers and was used in conjunction with benthic grabs to allow sampling of deeper sediment layers (Davis et al., 1993). Benthic grabs were performed with a hand trowel and sweeps with a 500μm mesh sweep net as per Sharma and Rawat (2009). Samples were preserved in 70 per cent ethanol. Samples were spread on a Petri dish for viewing and identification under a dissecting microscope, using identification keys from Ingram et al. (1997), Miller (1983), and Gooderham and Tsyrlin (2002). Samples were identified to the furthest level of classification possible.
The Shannon-Weiner diversity index, a function of species richness and abundance (Wilhm and Dorris, 1968; Nzengy’a and Wishitemi, 2000) was used to determine species diversity. The equation for Shannon-Weiner diversity index, where H is diversity, p is proportion or number of individuals of each taxon or species in sample and ln is log110, is shown below.
The value calculated is negative; however, this was written as a positive figure in the results for practical reasons. Summary statistics for macroinvertebrate species richness and Shannon-Weiner diversity index were calculated to determine normality of data as per Boix et al. (2007). Analysis of variance as employed by Muli (2005) and Nzengy’a and Wishitemi (2000) was the most accessible and affordable data analysis method for assessing significant differences in Shannon-Weiner diversity index and species richness, using MS Office Excel. Temporal variation of Shannon-Weiner diversity index and species richness was plotted on a time graph (Anamaet et al., 2005). Summary statistics on raw data show they were not normal; however the degree of normality present was sufficient for the use of one-way ANOVA.
A total of 16 taxa across the 3 wetlands were identified with 481 organisms collected from Police Road Retarding Basin (PRRB), 629 from Troups Creek Wetland (TCW) and 530 from the Dandenong Retarding Basin (DRB) wetlands. The taxa found in the samples and the 5 dominant taxa in each wetland are shown in Table 2.
|.||PRRB .||TCW .||DRB .|
|FAMILY||General taxa||5 dominant taxa||General taxa||5 dominant taxa||General taxa||5 dominant taxa|
|.||PRRB .||TCW .||DRB .|
|FAMILY||General taxa||5 dominant taxa||General taxa||5 dominant taxa||General taxa||5 dominant taxa|
The Shannon-Weiner diversity index (SWDI) was calculated for PRRB, TCW and DRB wetlands every fortnight after sampling and identification. One way ANOVA at the 0.05 significance level was performed on species richness, or number of species per sample, recorded across the study period. The range and means were 3–9 and 6.17; 3–9 and 6.33; 4–8 and 5.92 for PRRB, TCW and DRB, respectively. The three wetlands did not significantly differ (P > 0.05, F = 0.19) in terms of species richness.
The same one way ANOVA at the 0.05 significance level was performed on SWDI calculated across the study period. The range and means were 0.84–1.48 and 1.19; 0.47–1.50, and 1.03; 0.74–1.62, and 1.31 for PRRB, TCW and DRB wetlands, respectively. The three wetlands, though not significantly different (P > 0.05, F = 2.54) were more likely to show significant differences in species diversity than species richness.
Temporal variation in SWDI was found to have an increasing trend in all wetlands. Temporal variation in SWDI and species richness can be seen in Figures 2 and 3, respectively, between April and September 2009. The SWDI range documented for constructed wetlands is between 2.1 (Anamaet et al., 2005) and 0.4 (Webb and Mitsch, 2001), in the United States. SWDI values of 1.9 to 0.46 observed in the 3 wetlands may be representative of the true range in constructed wetlands in metropolitan Melbourne, since they fall within this range.
One way ANOVA for species richness with a P-value of 0.83, (P > 0.05, F = 0.19), indicates that the macro-invertebrate species richness in the 3 wetlands is similar. This may appear to be contradicted by the fluctuation in temporal variation of SWDI between April and October, seen in Figure 3. However, if considered in conjunction with the dominant taxa listed in Table 2, it could be suggested that the 5 dominant taxa in each wetland may be ecological generalists, capable of exploiting a wide range of conditions (Kratzer and Batzer, 2007). As a result, the consistent occurrence and relatively high frequency of the same 5 taxa in samples from their respective wetland is reflected in the average species richness of the 3 wetlands: being 6.17 for PRRB, 6.33 for TCW and 5.92 for DRB.
Of the 16 taxa found in the three wetlands, the majority of these may be considered cosmopolitan with the exception of Trichoptera, Atyidae and Amphipoda in PRRB, Glossiphonidae, Stratyomidae and Parastacidae from TCW and Trichoptera, Parastacidae and Atyidae from DRB wetlands. As Bryant and Papas (2007) suggest, it is likely these taxa are common to most constructed and naturally occurring wetlands and other water bodies in the Eastern, Northern and Western metropolitan regions of Melbourne. The community composition of macroinvertebrates in each wetland was dominated by 5 main taxa listed in Table 2, in order of highest to lowest frequency.
It is known that increased nutrient load and thus reduced dissolved oxygen affect a reduction in overall macroinvertebrate diversity (Cyr and Downing, 1988). One way ANOVA for SWDI with P-value of 0.09 (P > 0.05, F = 2.54) indicated a comparatively higher probability for significant difference, than for species richness. This may have resulted from higher nutrient loads in one or two of the three wetlands due to long or short term enrichment through run-off and other nutrient inputs; however, the source and magnitude of the variation in enrichment and water quality could not be quantified or confirmed, as this could not be completed in the course of the study. The variation in macrophyte coverage of the three wetlands may have also been a factor in the increased probability for significant difference in SWDI; however, the percentages shown in Table 1 are only estimates and cannot be seen as an influencing factor without further investigation.
The remaining, less frequent and abundant taxa found in each wetland may be what are termed ‘weaker’ or less dominant taxa, and may be more sensitive to adverse conditions (USEPA, 2002) such as pollution or excessive nutrient concentration. They may simply be less competitive within the ecosystem and this may explain the fluctuation in SWDI values over the study period. Their relatively inconsistent occurrence between fortnightly samples explains their lack of influence on average species richness and the similarity of species richness between the 3 wetlands.
Collection of physico-chemical data, including dissolved oxygen, total dissolved solids, salinity, pH and temperature for each wetland was required to validate any differences in macroinvertebrate diversity between the wetlands (Bryant and Papas, 2007). However, the volume of work required for the limited fortnightly time frame for sampling meant this part of the study could not be carried out and only macroinvertebrate diversity was assessed.
The application of macroinvertebrate diversity as a measure of wetland ecosystem integrity was confirmed as being a reliable tool for the purpose of conservation and management of these constructed wetlands. The similarities in the species richness and Shannon Weiner diversity index suggested such results may be common, with some regional or local variation, and so macro-invertebrate diversity may be considered a universal measure of wetland ecosystem integrity.
The similarity of collected data for species richness and Shannon Weiner diversity index to published data (Anamaet et al., 2005; Webb and Mitsch, 2001) indicated that these measures of macroinvertebrate diversity are universally suitable for use in constructed wetlands. The accuracy and validity of this data would be strengthened with further similar studies in alternative climates, other constructed wetlands and other geographical locations.
The temporal variation in SWDI and species richness showed a definite trend, and while this agrees with known climatic and ecosystem variation (Adamus and Brandt, 1990; Kratzer and Batzer, 2007), further studies may assist in determining if local variation in wetland ecosystems in different geographical locations and climates is significantly different, or similar enough to one another so as to be negligible.
The overall implications from the present study were that, there was high potential for its universal application in measuring ecosystem integrity in most if not all constructed wetlands despite location, climate or ecosystem characteristics.
Further studies to determine if these conclusions apply during spring and summer would be desirable to strengthen the results of the present study, since the study was completed over autumn and winter. Whether pollution and macrophyte coverage have any confounding effects on the reliability of these measures should also be observed.