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With the collaboration of: Research

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Water for a Healthy Country

Spatial and temporal changes in water

quality in Lake Alexandrina and Lake Albert

during a period of rapid water level

drawdown

Kane T. Aldridge, Brian M. Deegan, Sébastien

Lamontagne, Andrew Bissett and Justin D. Brookes

June 2009

Water for a Healthy Country

Spatial and temporal changes in water

quality in Lake Alexandrina and Lake

Albert during a period of rapid water

level drawdown

Kane T. Aldridge1,*, Brian M. Deegan1, Sébastien

Lamontagne2, Andrew Bissett3 and Justin D. Brookes1

1School of Earth and Environmental Science, The University of Adelaide, Adelaide, SA 5005 Australia

2CSIRO Land and Water, PMB 22, Urrbrae, SA 5064 Australia

3Max Planck Institute for Marine Microbiology, Bremen, Germany

*Corresponding author: [email protected]

June 2009

Water for a Healthy Country Flagship Report series ISSN: 1835-095X

ISBN: 978 0 643 09765 0

The Water for a Healthy Country Flagship is a research partnership between CSIRO, state and Australian governments, private and public industry and other research providers. The Flagship aims to achieve a tenfold increase in the economic, social and environmental benefits from water by 2025.

The Australian Government, through the Collaboration Fund, provides $97M over seven years to the National Research Flagships to further enhance collaboration between CSIRO, Australian universities and other publicly funded research agencies, enabling the skills of the wider research community to be applied to the major national challenges targeted by the Flagships initiative.

© Commonwealth of Australia 2009 All rights reserved. This work is copyright. Apart from any use as permitted under the Copyright Act 1968, no part may be reproduced by any process without prior written permission from the Commonwealth.

Citation: Aldridge, K.T., Deegan, B.M., Lamontagne, S., Bissett, A. and Brookes, J.D. (2009). Spatial and temporal changes in water quality and sediment character in Lake Alexandrina and Lake Albert during a period of rapid water level drawdown. CSIRO: Water for a Healthy Country National Research Flagship, Canberra.

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Foreword

The environmental assets of the Coorong, Lower Lakes and Murray Mouth (CLLAMM) region in South Australia are currently under threat as a result of ongoing changes in the hydrological regime of the River Murray, at the end of the Murray-Darling Basin. While a number of initiatives are underway to halt or reverse this environmental decline, rehabilitation efforts are hampered by the lack of knowledge about the links between flows and ecological responses in the system.

The CLLAMM program is a collaborative research effort that aims to produce a decision-support framework for environmental flow management for the CLLAMM region. The framework aims to evaluate the environmental trade-offs for different scenarios of manipulation of management levers, as well as different future climate scenarios for the Murray-Darling Basin.

The ecology of the Lower Lakes (Lake Albert and Lake Alexandrina) and their role in controlling nutrient and organic matter inputs to the Coorong and Murray Mouth region are not well understood. The future status of the Lower Lakes is also unclear due to ongoing low inflows from the River Murray.

Land & Water Australia is co-sponsoring a series of projects on the Lower Lakes in partnership with CSIRO’s Coorong, Lower Lakes and Murray Mouth program and the CLLAMMecology Research Cluster.

CLLAMMecology is a partnership between the University of Adelaide, Flinders University and SARDI Aquatic Sciences to study ecological responses to environmental change in the CLLAMM region, supported by CSIRO’s Flagship Collaboration Fund.

Other partner research and funding agencies involved in the Lower Lakes project include SA Water and the WA Centre for Water Research.

Additional reports relevant from the CLLAMM program and the CLLAMMecology Research Cluster can be found can be found at http://www.csiro.au/partnerships/CLLAMMecologyCluster.html

Water quality in the Lower Lakes during a water level drawdown iv

Table of Contents

Table of Contents .................................................................................................... iv Acknowledgements ................................................................................................. v Executive Summary ................................................................................................ vi 1. The Lower Lakes ............................................................................................... 1

1.1. Background ............................................................................................................... 1 1.2. Limnology of the Lower Lakes .................................................................................. 2 1.3. Management of the Lower Lakes and the current situation ..................................... 3

2. The influence of water level drawdown and salinisation on water quality in the Lower Lakes ................................................................................................... 5

2.1. Introduction ............................................................................................................... 5 2.2. Methods .................................................................................................................... 6

2.2.1. Continuous water temperature measurements ................................................ 6 2.2.2. Sampling ........................................................................................................... 7 2.2.3. Analyses ........................................................................................................... 8

2.3. Results ...................................................................................................................... 9 2.3.1. Salinity and density stratification....................................................................... 9 2.3.2. Dissolved oxygen ............................................................................................ 18 2.3.3. Suspended particles ....................................................................................... 18 2.3.4. pH and alkalinity ............................................................................................. 18 2.3.5. Nitrogen concentrations .................................................................................. 21 2.3.6. Phosphorus concentrations ............................................................................ 21 2.3.7. Other elements ............................................................................................... 22 2.3.8. Summary of results ......................................................................................... 22

2.4. Discussion............................................................................................................... 32 2.4.1. Salinisation of the Lower Lakes ...................................................................... 32 2.4.2. Changes in physical, chemical and biological conditions ............................... 33 2.4.3. Changes in nutrient concentrations ................................................................ 34 2.4.4. Conclusion ...................................................................................................... 35

3. Relationship of sediment character with water depth in the Lower Lakes, as evidence of sediment focussing ......................................................... 36

3.1. Introduction ............................................................................................................. 36 3.2. Methods .................................................................................................................. 37

3.2.1. Sediment collection ......................................................................................... 37 3.2.2. Sediment character ......................................................................................... 38 3.2.3. Water column character .................................................................................. 39 3.2.4. Statistical analyses ......................................................................................... 40

3.3. Results .................................................................................................................... 41 3.4. Discussion............................................................................................................... 47

4. Nutrient flux from permanently inundated and dried-reflooded sediments of Lake Alexandrina undergoing rapid water level drawdown ....... 49

4.1. Introduction ............................................................................................................. 49 4.2. Methods .................................................................................................................. 50

4.2.1. Core collection ................................................................................................ 50 4.2.2. Drying .............................................................................................................. 51 4.2.3. Nutrient flux experiments ................................................................................ 53 4.2.4. Sediment character ......................................................................................... 53 4.2.5. Statistical analyses ......................................................................................... 54

4.3. Results .................................................................................................................... 54 4.4. Discussion............................................................................................................... 58

5. Conclusion ...................................................................................................... 61 6. References ...................................................................................................... 62

Water quality in the Lower Lakes during a water level drawdown v

Acknowledgements

We thank the support of Land & Water Australia, SA Water and the CSIRO Flagship Collaboration Fund.

We also acknowledge the contribution of several other funding agencies to other aspects of the CLLAMM program and the CLLAMMecology Research Cluster, including the Fisheries Research and Development Corporation, the Murray-Darling Basin Authority (formerly the Murray-Darling Basin Commission) and the SA Murray-Darling Basin NRM Board. Other research partners include Geoscience Australia and the Flinders Research Centre for Coastal and Catchment Environments. The objectives of this program have been endorsed by the SA Department of Environment and Heritage, SA Department of Water, Land and Biodiversity Conservation, SA Murray-Darling Basin NRM Board and Murray-Darling Basin Commission.

Water quality in the Lower Lakes during a water level drawdown vi

Executive Summary

Lake Alexandrina and Lake Albert (the Lower Lakes) are a set of large, shallow, fluvial lakes at the downstream end of the Murray-Darling Basin. The lakes host a number of threatened freshwater fish species; are an important nesting habitat for waterbirds; the source of water for a local irrigated agriculture economy; support a substantial fishery; and provide substantial recreational and aesthetic value to South Australia. Together with the Coorong, an estuarine-hypersaline coastal lagoon, the Lower Lakes were declared a Wetland of International Importance under the Ramsar Convention due to the abundant and diverse ecological communities within the region.

From September 2001 until the end of this study (April 2008) the Murray-Darling Basin experienced a severe rainfall deficiency, the second driest seven-year period in its recorded history (MDBC 2008). This combined with the over-allocation of water within the Murray-Darling Basin and presence of the barrages separating the Lower Lakes from the Coorong, has resulted in a dramatic reduction in inflows to the Lower Lakes and rapid water level drawdown. From August 2006 to August 2008, water levels fell from 0.75 m AHD to -0.5 m AHD. This is well below the previous historical low of 0.1 m AHD in April 1968 (MDBC 2008).

This report summarises three limnological studies that were conducted during the drawdown period. The overall purpose of the project was to gain a better understanding of nutrient cycling in the lakes and how it is impacted by changes in water level. The three studies were:

A spatially extensive water quality monitoring program, undertaken between January 2006 and April 2008;

An assessment of the spatial variability in sediment characteristics across the lakes;

A nutrient regeneration experiment simulating a drying and wetting of sediments of Lake Alexandrina.

In addition to documenting ongoing changes during the drawdown, these studies are the most detailed assessment undertaken to date on the limnological properties of this system.

The influence of water level drawdown and salinisation on water quality in the Lower Lakes

Changes in water quality during the drawdown were investigated through an intensive monitoring program. Between January 2006 and April 2008, nineteen sites distributed throughout the Lower Lakes were visited at approximately two-month intervals. Measurements were taken at each site at 0.25 m intervals through the water column for dissolved oxygen, electrical conductivity, pH, turbidity, temperature and light intensity. In addition, integrated water samples were collected and analysed for total suspended solids, particulate organic matter, chlorophyll a, alkalinity, nutrients, organic carbon and sulfur.

Spatial and temporal variations in salinity

Large horizontal and vertical gradients in salinity were found across the lakes during the monitoring program. As the water level fell, salinity within the lakes increased dramatically with average salinity of all sites of 0.9 g/L in January 2007 and 10.6 g/L in April 2008. This was not attributable to evapoconcentration alone, suggesting an external input of salt, the most likely source of which was the barrages. While salinity increased steadily in the main bodies of Lake Alexandrina and Lake Albert, in the area between Goolwa and Clayton (Lake Alexandrina Arm), salinities increased from 1.1 ± 0.5 g/L to 18.4 ± 7.6 g/L. The large variation around this mean at the end of the study period was attributed to longitudinal changes in salinity, with average salinity at Goolwa and Clayton of 22.4 and 9.1 g/L, respectively.

Water quality in the Lower Lakes during a water level drawdown vii

In addition, strong vertical salinity gradients were detected at some locations, in particular in Lake Alexandrina Arm. At Goolwa, salinity stratification was always present, but there was an increase in the extent of stratification during the study period. In January 2007, salinity in the epilimnion (that is, the surface waters when stratification was present) was 0.9 g/L and 5.3 g/L in the hypolimnion (that is, the bottom waters when stratification was present). In comparison, in April 2008 salinity was 20.2 and 28.2 g/L in the epilimnion and hypolimnion, respectively. Stratification was also present at other sheltered sites, particularly in Lake Alexandrian Arm, but not at open water sites such as the middle of Lake Alexandrina. However, at times, some salinity stratification was evident at the more sheltered sites at the upstream end of Lake Alexandrina, probably due to the intrusion of more saline water from the main body of Lake Alexandrina.

Changes in suspended material and physicochemical conditions

While it was hypothesised that suspended solids would decrease due to flocculation of suspended material, this was not the case. Suspended material, including organic matter, inorganic matter and chlorophyll a increased during the study period. This was most likely due to the movement of the shoreline towards the centre of the lake, which contains finer sediments (and higher organic matter and chlorophyll a content) that are resuspended more readily. Although increasing salinity did not have a significant impact upon dissolved oxygen concentration or saturation in surface water, dissolved oxygen concentrations declined towards zero in the hypolimnion, when present. The magnitude of hypolimnetic dissolved oxygen depletion decreased with distance from the barrages.

pH within the lakes remained alkaline throughout the monitoring program. The average pH in the Lower Lakes in January 2007 was 8.9 ± 0.2 and had declined slightly in April 2008, when average pH was 8.3 ± 0.3. However, filtered alkalinity had a positive relationship with salinity within the Lower Lakes and increased during the monitoring program.

Changes in nutrient concentrations

There was a general increase in average total nitrogen (TN) concentrations in the lakes through the study period. However, TN concentrations were particularly high in July 2007, with differences in TN most closely related to light availability, suspended material and alkalinity. Dissolved inorganic nitrogen (NOx-N plus NH4-N) only consisted of approximately 6.7% of TN. However this proportion increased through the study period from 3.1 to 14.8%. Average lake concentrations of NH4-N increased during the study period, particularly in Lake Alexandrina Arm. However, there was considerable variation in concentrations observed in this region with concentrations varying from 0.8 mg/L in Goolwa to 0.3 mg/L at Clayton in April 2008. The increase in NH4-N concentrations was correlated with salinity. However, a number of processes associated with salinity could be involved in this increase in NH4 concentration, including density stratification and displacement of the sediment-bound NH4 by ion exchange. In general, NOx-N concentrations also appeared to increase during the study period, but significant variability was present with particularly high concentrations in May 2007.

Unlike nitrogen, average total phosphorus (TP) concentrations in the Lower Lakes did not change during the study period. Phosphate (PO4-P) consisted of between 1.4 (January 2007) and 14.6% (July 2007) of TP with variation in TP concentrations associated with changes in PO4-P concentration. Although, there was no significant change in average PO4-P concentrations during the study period, PO4-P concentrations did increase in Lake Alexandrina Arm.

Water level drawdown and salinisation in the Lower Lakes have resulted in a shift from aerobic to anaerobic decomposition of organic matter in sediments. Seawater leakage from the barrages resulted in the formation of permanent density stratification in areas where it did not occur or only occurred temporarily under purely freshwater conditions. In turn, anoxia rapidly

Water quality in the Lower Lakes during a water level drawdown viii

developed in the hypolimnion when present due to the consumption of oxygen during organic matter decomposition in the sediments. This process also tends to increase the recycling of nutrients from the sediment pool, as observed from higher concentrations of nitrogen and phosphorus in the hypolimnion than the epilimnion at Goolwa during periods of stratification.

Relationship of sediment character with water depth in the Lower Lakes, as evidence of sediment focussing

The storage and recycling of nutrients in sediments is one of the most important aspects of the biogeochemistry of lakes. Sediment characteristics play an integral role in determining nutrient fluxes between the sediment and water column. However, the distribution of sediments within lakes is typically heterogeneous, owing to the large number of physical, chemical and biological processes that interact and influence their character. In order to gain an understanding of spatial variability in nutrient cycling in the Lower Lakes, this project aimed to evaluate the spatial distribution of various sediment types.

Intact sediment cores were collected from 40 sites in 10 transects across Lake Alexandrina and Lake Albert. Transects were selected to account for all habitat types, including sheltered and open water sites. The top 1 cm of each sediment core was removed and analysed for particle size distribution, moisture content, chlorophyll a, particulate organic matter (POM), total carbon (TC), total nitrogen (TN) and total phosphorus (TP). Comparisons were made between transects and water depth classes.

Sediments were found to be relatively heterogeneous in their character with transitions between sediment types rather than the formation of distinct sediment groups. Sediment characteristics were primarily a function of depth. As depth increased, sediments had a higher proportion of fine particles, increased water content and an increased POM. While TP showed little variation across the whole study area, there was a general decrease in TN with water depth, except for sites with a depth of 0.61-0.9 m, which had relatively low nitrogen concentrations. These characteristics in sediment distribution are consistent with sediment focussing, the tendency in lakes for finer sediments to accumulate in deeper areas, where they are less prone to resuspension. The spatial distribution of sediments has important implications for biogeochemical cycles in the Lower Lakes.

Nutrient flux from permanently inundated and dried-reflooded sediments of Lake Alexandrina undergoing rapid water level drawdown

Water level drawdown results in the drying of previously inundated sediments, causing changes in the physical, chemical and biological character of the sediments to which nutrient cycling is intrinsically linked. Consequently, the cycling of nutrients following drying and reflooding cycles is likely to be different than under permanently inundated conditions. The aim of this project was to gain an understanding of the impacts of reflooding dried sediments from Lake Alexandrina on microbial activity and nutrient fluxes to the water column.

Intact sediment cores were collected from two sites in Lake Alexandrina, a deep site with fine sediments located close to the inlet of the River Murray (Site 1) and a shallow site with coarse sediments located near Tauwitchere Barrage (Site 2). Overlying water was siphoned off of half of the sediment cores of both sediment types and the cores were placed at 40°C until a constant weight was obtained. Following this, permanently inundated cores and dried cores were overlain with artificial River Murray water, the likely source of water for re-inundation of the Lower Lakes. Water samples were extracted from cores over a 10.0 hr period and analysed for phosphate (PO4-P), ammonium (NH4-N), nitrite (NO2-N) and nitrate (NO3-N) concentrations. Flux rates were calculated as the slope of the linear regression between nutrient mass and time. Denitrification rates were measured using the isotope-pairing method.

The flux of nutrients began immediately upon inundation, with a large immediate flux of NH4-N to the water column. At both sites and for both treatments, PO4-P and NO3-N were consistently lost from the water column over the 10 hour period, but the flux from the water column to the

Water quality in the Lower Lakes during a water level drawdown ix

sediments was greater in reflooded treatments than permanently wet treatments. In addition, the flux of PO4-P from the water column in Site 2 was greater than that of Site 1.The greatest flux rates of the studied nutrient forms were observed for NH4-N. In permanently wet sediments, NH4-N concentrations in the water column increased at a rate of 237.9 ± 94.4 mg/m2/hr for Site 1, while for Site 2 NH4-N concentration decreased at a rate of 191.8 ± 91.1 mg/m2/hr. For both sites, the flux of NH4-N to the water column was greater following drying-reflooding than for permanently wet sediments. Permanently wet sediments were found to have denitrification rates of 0.2 ± 0.1 mg/m2/hr at Site 1 and 0.3 ± 0.1 mg/m2/hr at Site 2. However, following drying-reflooding, denitrification was reduced to 0.00 ± 0.00 mg/m2/hr and 0.00 ± 0.01 mg/m2/hr for Sites 1 and 2, respectively. Thus, upon refilling of the Lower Lakes, a large flux of nutrients, particularly nitrogen can be expected. This may result in an increase in algal productivity and even algal blooms or increased bacterial productivity, causing the development of anoxic conditions.

Summary and recommendations

The key finding of the three studies is that water level drawdown in the Lower Lakes has dramatically altered the biogeochemistry of the system. Water level drawdown has resulted in salinisation of the lakes, resulting in strong horizontal and vertical salinity gradients. While salinity can directly influence nutrient cycling, it is the generation of density stratification that could have the most important impacts on water quality. In areas close to the barrages, it appeared that the settling of dense saline water to the bottom of the water column resulted in the development of anoxic conditions and release of nutrients from sediments. This may lead to increased productivity of autotrophic or heterotrophic microbial communities.

The impact of increased nutrient concentrations on productivity within the Lower Lakes is further complicated by the increased suspended solid concentrations that were observed. This is likely to reduce primary productivity, including aquatic plants within the system, thus reducing habitat and organic matter resources for higher organisms. The increased suspended solid concentrations were thought to be a result of the accumulation of finer sediments towards the middle of the lakes as a result of sediment focussing. It was found that there is a strong relationship between sediment character and water depth in the Lower Lakes in recently deposited sediments. This predictable distribution was thought to reflect river regulation, ultimately resulting in reduced habitat diversity within the Lower Lakes.

This study has highlighted that the water level and salinity regimes are key ecosystem drivers in the Lower Lakes and that they impact on ecological processes in complex ways. Planning management interventions in this ecosystem will require that the effects on water level and salinity are properly understood. Currently there exists a lack of knowledge on the likely ecological responses to various scenarios and effort must be placed to reduce uncertainty around these responses. This will allow appropriate management decisions to be made.

Water quality in the Lower Lakes during a water level drawdown 1

1. The Lower Lakes

1.1. Background

Lake Alexandrina and Lake Albert, collectively known as the Lower Lakes, are set of large (over 750 km2), shallow (maximum depth of 4.1 m), connected, terminal lakes of the Murray-Darling Basin (Figure 1), Australia’s largest drainage basin (1,063,000 km2). Together with the Coorong, an estuarine-hypersaline coastal lagoon, the Lower Lakes were declared a Wetland of International Importance in 1985 under the Ramsar Convention. This recognition was due to the abundant and diverse ecological communities within the region. The Lower Lakes are an important refuge for a number of threatened freshwater fish species and are an important nesting habitat for waterbirds. The lakes are also an important source of water and resources to the Coorong and near shore environment, thus supporting these downstream ecosystems (Cook et al. 2008). In addition, the Lower Lakes are the source of water for a local irrigated agriculture economy; support a substantial fishery; and provide substantial recreational and aesthetic value to South Australia.

The River Murray, one of the world’s longest rivers, carries the largest and most constant flow of water to the Lower Lakes. The Darling River also contributes significant flow, although this is more variable flow and carries high loads of fine particles, resulting in extremely high turbidity. Although several local streams discharge into Lake Alexandrina, their overall contribution to total annual flow are only considered to be significant during periods of low River Murray inputs (Anon 2007). The original flow regime of the lower River Murray at the Lower Lakes was characterised by spring floods and summer low flows, but the magnitude of both was variable from year-to-year (Jolly 1996). During periods of extreme low inflow, water levels would have fallen, exposing areas of sediment to the atmosphere and allowing the intrusion of seawater into the Lower Lakes (Von Der Borch and Altman 1979; Geddes 1984a; Sims and Muller 2004). Following European settlement and the development of the irrigation industry upstream, more frequent intrusions of salt water occurred into the Lower Lakes (Sims and Muller 2004). To prevent these intrusions, five barrages were constructed between 1935 and 1940 to maintain the lakes as a permanent freshwater supply for irrigation and human use. These barrages separate the Lower Lakes from the Coorong, an estuarine-hypersaline coastal lagoon (Geddes 1987) that is dependent upon inflows from the Lower Lakes to maintain the estuarine component of the system. The recognition of the region as Wetland of International Importance is due to the abundant and diverse ecological communities within the region, which are a result of high diversity of habitats created largely by the salinity gradient that exists.


Figure 1. Map of the Lower Lakes, Lake Alexandrina and Lake Albert.

1.2. Limnology of the Lower Lakes

The natural salinity balance of the Lower Lakes prior to European settlement in Australia remains contentious. Anecdotal evidence of the occurrence of salinity intrusions into the lakes following European settlement has been used as evidence of system that was influenced extensively by marine water prior to European settlement. However, an in depth investigation of the anecdotal evidence (Sims and Muller 2004), as well as paleolimnological studies (Barnett 1993; Fluin et al. 2007) have demonstrated that the Lower Lakes were predominately fresh. As with all estuaries, the position of the estuary would have been dynamic, dependent upon the balance between river and oceanic inputs. Only during periods of extremely low River Murray inflow would significant marine water have entered the main bodies of the lakes (Sims and Muller 2004; Fluin et al. 2007)

Given the ecological and socio-economic importance of the region, it is surprising that very little is known about the ecological functioning of the Lower Lakes. Early studies in the Lower Lakes region were dominated by geological and geomorphic investigations. These investigations are summarised by Bourman and Barnett (1995), who suggested that since European settlement river regulation has resulted in reduced inputs of coarse sediment due to sedimentation upstream of weirs (Thoms and Walker 1993). However, there has been an apparent increase in sedimentation rates within the Lower Lakes (Barnett 1993) as a result of increased erosion of the lake shore due to elevated water levels (Coulter 1992). This has resulted in the retreat of lake perimeter at an average rate of 1 m/yr (Coulter 1992).

Although our limnological understanding of the Lower Lakes is generally poor it is a historically important site, as it was site of the first detailed scientific account of toxic cyanobacteria (Francis 1978; Codd et al. 1994). More recently, paleolimnological studies have demonstrated that following European settlement the diatom community has changed, reflecting the altered salinity


regime and increased nutrient inputs from upstream agriculture (Fluin et al. 2007). Indeed, changes in the source of organic matter and nutrients have occurred following European settlement, with an apparent increase in nitrogen from agriculture (Herzeg et al. 2001). The Lower Lakes are now considered to be eutrophic (Geddes 1984a), which is supported by the regular occurrence of algal blooms (Baker 2000). However, Geddes (1984a) considered the Lower Lakes to be a marginal environment for phytoplankton growth as it is highly turbid and most nutrients are not in bio-available forms. Indeed, Cook et al. (2008) found that the lakes assimilated inorganic nutrients rapidly and exported them to the Coorong in organic forms.

Geddes (1984b) found that the turbidity also appears to play an important role in structure the community composition of higher trophic levels in the Lower Lakes. The zooplankton community of the Lower Lakes is relatively dense and contains a high proportion of large zooplankters in comparison to lakes elsewhere (Geddes 1984b). The high turbidity is thought to result in the absence of significant vertebrate predation. While planktivores, including smelt, galaxias and other juvenile fish do occur in the littoral region, no open-water planktivores exist and so predation, particularly on large zooplankton, is thought to be low (Geddes 1984b).

This disconnection of the Lower Lakes and Coorong due to the construction of the barrages is thought to reduce the abundance of organisms in the region. Species that normally have high densities in estuaries, such as nematodes, are low due to an absence of a permanent estuary (Nicholas et al. 1992). However, as a result of the conversion of inorganic nutrients into organic forms within the lakes, including incorporation into zooplankton, flow provisions across the barrages could provide significant benefits for the foodwebs of the Coorong (Geddes 1984b; Cook et al. 2008). Indeed, studies of two of the most recent flow provisions to the Coorong have revealed such benefits, with the aggregation and spawning of estuarine fish near the barrages (Geddes 2005b; Geddes 2005a). However, the benefits of small inputs are short-lived with estuarine conditions not persisting within the region for a majority of the time.

1.3. Management of the Lower Lakes and the current situation

The flow regime of the River Murray has been gradually modified following European settlement by a complex system of reservoirs, weirs and barrages. Extraction of water upstream for irrigation and human use has severely reduced the amount of water passing into and through the Lower Lakes. Consumptive water use within the Murray-Darling Basin has reduced average streamflow through the Murray Mouth from 12,233 GL/yr to 4733 GL/yr, a 61% reduction (CSIRO 2008).

However, the operation of the barrages for irrigation and recreational purposes has had further impacts on the hydrology of the system. Current management of the barrages maintains an average water level of 0.75 m AHD (Anon 2007). As river flow increases, barrage gates are opened to maintain this level. The lakes are surcharged to 0.85 m AHD at the beginning of summer to ensure that evaporation and irrigation demand during summer does not lower the water level below 0.6 m AHD by autumn. This, along with extractions, has resulted in a reduced the size (magnitude and duration) of peak flow; brought forward the timing of peak flow; reduced the frequency and duration of small and medium floods; increased the average lake water residence time; increased average water levels; and reduced the range of water level fluctuations (Jolly 1996; CSIRO 2008).

From September 2001 until the end of this study (April 2008) the Murray-Darling Basin experienced severe rainfall deficiencies, the second driest seven-year period in its recorded history (MDBC 2008). This combined with the over-allocation of water within the Murray-Darling Basin and presence of the barrages separating the Lower Lakes from the Coorong, has resulted in a dramatic reduction in inflows to the Lower Lakes and rapid water level drawdown. From August 2006 to August 2008, water levels fell from average levels of 0.75 m AHD to -0.5 m AHD (Figure 2). This is well below the previous historical low of 0.1 m AHD in April 1968 (MDBC 2008) and has resulted in the intrusion of saline water into the lakes, presumably from leakage from the barrages. Early during the draw down period, significant leakage of marine water


through the barrages was evident and attempts were made to seal barrage gates between winter and spring 2007 by sealing stop logs with PVC pipe, flat pipe and foam rubber (Jarrod Eaton, DWLBC, personal communication). For similar reasons, boat access through the barrage lock at Goolwa was curtailed in winter 2007.

During periods of low-flow, water quality of lakes is strongly influenced internal biogeochemical processes. The aim of this project was to provide an understanding of internal processes influencing water quality within the Lower Lakes under a period of low inflow and water level drawdown. This study represents the most comprehensive assessment of the limnology of the system to date.

Date

01-06 04-06 07-06 10-06 01-07 04-07 07-07 10-07 01-08 04-08

Wate

r le

vel (m

AH

D)

-1.0

-0.6

-0.2

0.2

0.6

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cha

rge

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L/d

ay)

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2

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12Coorong water level

Lake Alexandrina water level

Murray River Discharge

Figure 2. Lake Alexandrina and Coorong water levels and River Murray flow at Lock 1 from January 2006 to May 2008. Flow data provided by the Murray-Darling Basin Authority (Bigmod). Water level data provided by the Department of Water, Land and Biodiversity Conservation.


2. The influence of water level drawdown and salinisation on water quality in the Lower Lakes

Kane T. Aldridge

Brian M. Deegan

Sébastien Lamontagne

Justin D. Brookes

2.1. Introduction

From September 2001 until April 2008 the Murray-Darling Basin experienced a severe rainfall deficit, the second driest seven-year period in its recorded history (MDBC 2008). This combined with the over-extraction of water and presence of the barrages separating the Lower Lakes from the Coorong, resulted in rapid water level drawdown in the Lower Lakes. Average water levels fell from of 0.75 m AHD to -0.5 m AHD, well below the previous historical low of 0.1 m AHD in April 1968 (MDBC 2008) and also below sea level (see Chapter 1 for a more detailed description).

Terminal water bodies, such as the Lower Lakes, are particularly vulnerable to the over extraction of water resources because these impacts are cumulative and can result in extended drawdown periods. One example is the Aral Sea, where water level dropped nearly 13 meters and its area decreased by 40 percent between 1960 and 1987 as result of over extraction of water for irrigation (Micklin 1988). As the Aral sea shallowed and shrunk, biological productivity declined rapidly and native flora and fauna communities became degraded or completely disappeared (Micklin 1988). Significant changes in water quality have also been observed as a result of drawdown in the man-made reservoirs (Naselli-Flores 2003; Sanchez-Carrillo et al. 2007; Baldwin et al. 2008).

Many of the observed changes in response to water level drawdown are associated with salinisation, the process whereby the concentration of dissolved salts increase (Williams 1987). Secondary salinisation of inland waters accompanying water level drawdown results from evapoconcentration; rising saline groundwater as a result a reduced pressure head from surface water; or in low-lying, end of system water bodies, such as the Lower Lakes, the encroachment of ocean water. However there is little information available on the combined impact of water level drawdown and salinisation on the biogeochemistry of inland waters.

The combination of water level drawdown and salinisation will impact physical, chemical and biological processes within the water body that control water quality. However, predicting the response is complex. For example, water level drawdown and salinisation may have contrasting impacts on concentrations of phosphorus, a limiting nutrient of algal growth. As water levels fall, the penetration of oxygen into sediments is likely to increase (Baldwin and Mitchell 2000). This may result in increased affinity of sediments for phosphorus due to the oxidation of ferrous sulfides into amorphous ferric oxyhydroxides, which have a high affinity for phosphorus (De Groot and Fabre 1993; De Groot and Van Wijck 1993; Baldwin 1996). In contrast, salinity stratification that often accompanies intrusions of saline water may lead to oxygen depletion within the sediment, resulting in the release of phosphorus into the water column (Davis and Koop 2006; Baldwin et al. 2008). Examples of the contrasting impacts of water level drawdown


and salinisation can also be found for the cycling of nitrogen (Knowles 1982; Rysgaard et al. 1999; Baldwin and Mitchell 2000; Baldwin et al. 2006; Laverman et al. 2007) and other elements.

These biogeochemical processes are further complicated by changes in physical conditions as a result of water level drawdown and salinisation. For example, increasing salinity can alter light penetration through salt induced aggregation and flocculation of suspended material (Grace et al. 1997). This may increase pelagic and benthic productivity and demand for nutrients. Furthermore, increased ionic strength due to increased salinity can lead to reduced oxygen solubility and salinity stratification (Nielsen et al. 2003). This may reduce oxygen penetration into sediments, shifting biogeochemical cycling of nutrients from aerobic to anaerobic processes (Donnelly et al. 1997) and increasing the flux of nutrients from sediments to the water column.

The impact that water level drawdown and salinisation will have on the Lower Lakes is unknown. This report summarises an extensive 18 month water quality monitoring program for the Lower Lakes during a period of water level drawdown. This sampling program represented the most comprehensive assessment of the limnology of this system since the work of Geddes (1984a; 1984b). A particularity of the monitoring program was that it was designed to represent all the different sub-basins of the system, including offshore locations. In addition, vertical temperature and salinity profiles were monitored at each site (including continuous temperature profiles monitoring at 4 locations) to evaluate if drawdown and salinisation increased the likelihood of density stratification. The report first summarises the trends in water level and salinity over the sampling period, with an emphasis on the complex spatial patterns in salinity change. Major trends in other physical (temperature, suspended particles, etc), chemical (oxygen concentration, nutrients and major ions) and biological properties (chlorophyll a etc.) are presented. Due to the recently discovered presence of Potential and Actual Acid Sulfate Soils in exposed areas of the Lower Lakes (Fitzpatrick et al. 2008), trends in lake water pH and alkalinity are also summarised.

Several hypotheses were tested with the monitoring program, including that:

Salinisation will increase water clarity due to flocculation of suspended material;

Salinisation will increase the prevalence of density stratification due to the input of more dense saline water;

Ammonia and phosphate concentrations will increase due to the development of anaerobic conditions associated with the increased prevalence of density stratification;

Particulate nutrient concentrations will decrease due to flocculation of suspended material

The implications of this study for the management of the Lower Lakes are discussed.

2.2. Methods

2.2.1. Continuous water temperature measurements

In December 2006, Stow-Away Tidbit Temperature Loggers were deployed at four sites (Figure 3). At three sites in Lake Alexandrina (Murray-Lake Alexandrina Intersection, Lake Alexandrina Opening and Poltalloch) five sensors were distributed through the water column between the water surface and a water depth of 3.0 m. At Lake Albert Entrance, one sensor was placed immediately beneath the surface and another was placed at 3.0 m. During the study period water levels at each site were less than 3.0 m and so measured temperature at 3.0 m reflects that at sediment/water interface. Measurements of water temperature were recorded every 15 minutes and sensors were downloaded, cleaned and redeployed every 2 months.


Figure 3. Location of study sites in Lake Alexandrina and Lake Albert. Filled red circles show locations of water quality monitoring sites, unfilled red circles show locations of temperature loggers. Large circles show groupings of sites used for analyses: Lake Alexandrina Body (brown), Lake Albert (green) and Lake Alexandrina Arm (blue).

2.2.2. Sampling

Between January 2007 and April 2008, nineteen sites within the Lower Lakes (Figure 3) were visited at approximately 2 month intervals (9-11 January 2007, 19-21 March 2007, 21-23 May 2007, 9-11 July 2007, 2-4 October 2007, 26-28 November 2007, 22-24 January 2008 and 31 March-2 April 2008). To assist discussion of spatial patterns the sites are grouped into regions (Figure 3). On occasions during the course of the study, sites were not inundated with water and so samples were not collected. In addition, rough weather occasionally prevented sampling at some of the offshore locations.

At each site, sensors of a calibrated TPS 90-FLT logger were lowered into the water column. Upon equilibration measurements of dissolved oxygen, electrical conductivity, pH, turbidity and temperature were recorded at 0.25 m intervals. Dissolved oxygen measurements were automatically corrected for salinity since the sensitivity of electrodes is influenced by salinity. Dissolved oxygen saturation (DOSAT [%]) was calculated through the equations of Eaton et al. (2005):

P

SATC

DODO

100 (1)


where CP is equilibrium concentration at nonsaturated pressure (mg/L) and DO is salinity corrected dissolved oxygen concentration (mg/L).

Electrical conductivity was converted to salinity (S, g/L) using the equation of Williams and Sherwood (1994) for Australian lakes:

0878.1

25)(466.0 ECS (2)

where EC25 is the electrical conductivity corrected to 25°C (mS/cm). However, when S is less than 3 g/L the relationship between S and electrical conductivity is linear and so the following equation was used (Tucker and Beatty 1974):

25640 ECS (3)

Three instantaneous measurements of light intensity were also recorded at 0.25 m intervals using a LI-COR underwater spherical quantum sensor (LI-193SA). For each depth interval, the light attenuation coefficient (k) was calculated using the Beer-Lambert equation:

Z

II

kO

Zln

(7)

where IZ is light intensity at water depth Z, and IO is the light intensity at the water surface. The euphotic depth (ZEU) was estimated using:

kZEU

6052.4 (8)

Average k and ZEU values were calculated as the average of each depth interval of a single profile.

Integrated water samples were collected using a polyvinyl chloride tube with an internal diameter of 5.4 cm. Tubing was lowered through the water column, sealed and retrieved. Approximately 2.5 L of unfiltered water was collected and all samples were immediately stored in the dark below 3°C. These samples were analysed for TP, TN, TSS, OM, chlorophyll a from January 2007. Alkalinity (as “unfiltered alkalinity”) and TOC were also analysed from October 2007. Approximately 100 mL of mixed sample was filtered through a Millex® AP 20 GF prefilter followed by a Millex ® 0.22 μm PES Membrane filter. Filters were not pre-rinsed as they were found not to leach detectable levels of nutrients; however, the first 5 mL of filtered sample was not dispensed into the sample bottle. Filtered samples were analysed for ammonia (NH4-N), oxidised nitrogen (NOx-N), nitrite (NO2-N), phosphate (PO4-P), S, and dissolved organic carbon (DOC) from January 2007. Also, alkalinity (as “filtered alkalinity”), pH and total organic carbon (TOC) were analysed from October 2007; and sulfate (SO4-S) from January 2008. In November 2007 and April 2008 discrete depth samples were also collected from the surface and bottom of the water column at Goolwa and analysed for the parameters described above.

2.2.3. Analyses

Unfiltered alkalinity (as CaCO3) was measured in the field using a Hach AL-DT test kit. Samples were titrated to pH 4.5 with 1.6N H2SO4 with Bromcresol Green-Methyl Red indicator. On return to the laboratory, filtered alkalinity was measured using APHA method 2320 (Eaton et al. 2005). A known volume of sample was titrated to pH 4.5 using standardised 0.02M HCl to determine total alkalinity. Sample pH was measured using a pH meter (Orion 940) and combined glass electrode calibrated at pH 4.0, 7.0 and 10.0 (Rayment and Higginson 1992).


All analyses of nutrients, organic carbon and other elements were analysed by the CSIRO Analytical Laboratory, South Australia following methods outlined in Lamontagne et al. (2004). Particulate organic carbon was calculated as the difference between TOC and DOC.

Total suspended solids were determined by APHA method 2540 (Eaton et al. 2005), whereby suspended particulate material is concentrated onto pre-combusted Whatman International GF-C filters. The increase in weight (dried to constant weight at 105°C) represented total suspended solids. Filters and suspended material were then ignited to constant weight at 550°C to determine fixed and volatile solids following APHA method 2540-E (Eaton et al. 2005), reported as particulate organic matter (POM). Material concentration on to additional GF-C filters were used to measure suspended chlorophyll a concentration following Golterman et al. (1978), using 99.8% methanol and a Hitachi U-2000 spectrophotometer (Hitachi Ltd., Tokyo, Japan), with a path length of 10 mm. The particle size distribution of samples was determined using a Laser In-Situ Scattering and Transmissometry instrument (LISST-100X, Sequoia scientific). This instrument measures scattering of light at 32 angles and this information is used to calculate the size distribution of 32 size classes log-spaced between 2.5 and 500 μm.

For the purposes of providing an overview of trends in water quality parameters through the study period, average concentrations of analytes at all of the sites were calculated for each sampling period. This average is not volume weighted since the volumes of different lake regions were not available. Therefore these values not represent average values within the Lower Lakes, but rather the average values of the sampled sites. Consequently, sites that represent smaller volumes have a disproportionate influence on the calculated average values. These averages were used to calculate total mass within the Lower Lakes in January 2007 and April 2008. This was done by multiplying the average concentration by the estimated lake volume. The lake volume was calculated with the water level-volume relationship calculations of the Department of Water, Land and Biodiversity Conservation (DWLBC, unpublished data). The average water level data was provided by Bureau of Meteorology (BOM, unpublished data). In January 2007 average water levels were 0.42 m AHD with a calculated volume of 1760 GL. In April 2008 average water levels were -0.50 m AHD with a calculated volume of 1128 GL.

The association between parameters was tested through regression analysis using JMP-IN®. All variability is reported as standard deviations.

2.3. Results

2.3.1. Salinity and density stratification

In January 2007, average salinity at the sampled sites within the Lower Lakes was 0.9 ± 0.4 g/L (Table 1). As the water depth fell, salinity within the lakes increased rapidly (Table 2) and by May 2007 average salinity had increased to 6.8 ± 8.1 g/L (Table 1). Following May 2007, there was a small decrease in average salinities to October 2007, followed by another increase, particularly between January 2008 and April 2008, with average salinities reaching 10.6 ± 9.6 g/L (Table 1). However, as well as an increase in average salinities, there was increase in the variability in salinity within the lakes, associated with an increasing salinity gradient between the barrages and the River Murray (Figure 4 and Figure 5). In January 2007 there was a gradual increase in salinity from the outlet from the River Murray to the barrages, with salinities ranging from 0.4 g/L at Lake Alexandrina Opening to 2.1 g/L at Goolwa. However, by April 2008, salinities ranged from 1.0 g/L at Lake Alexandrina Opening to 30.0 g/L at Ewe Island (Figure 4). Within Lake Alexandrina Arm there existed considerable spatial variation, with a longitudinal gradient of decreasing salinity away from the barrages (Figure 5). In April 2008, average salinity at Goolwa was 22.4 g/L, while salinity at Clayton was 9.1 g/L (Figure 4). The salinity gradient within Lake Alexandrina Body and Lake Albert was more gradual, although salinities tended to be higher in Lake Albert than Lake Alexandrina Body (Figure 4). In Lake Alexandrina, salinities decreased with distance from the barrages, but in Lake Albert, salinities tended to increase from North to South (Figure 4).


Table 1. Average physical, chemical and biological conditions in the Lower Lakes, January 2007-April 2008.

Parameter Jan-07 Mar-07 May-07 Jul-07 Oct-07 Nov-07 Jan-08 Apr-08

Wind speed (km/h) 26.7 ± 11.5 22.4 ± 5.7 31.3 ± 6. 7 12.6 ± 4.4 31.3 ± 6.2 19.1 ± 6.8 24.5 ± 8.0 24.3 ± 10.1

Water depth (m) 2.2 ± 1.0 1.8 ± 0.9 1.6 ± 1.0 1.6 ± 1.0 1.4 ± 0.8 1.5 ± 1.1 1.3 ± 1.0 1.0 ± 0.9

Salinity (g/L) 0.9 ± 0.4 2.35 ± 3.0 6.8 ± 8.1 5.2 ± 5.9 4.3 ± 4.2 6.0 ± 6.1 6.0 ± 6.4 10.6 ± 9.6

Temperature (°C) 22.1 ± 1.0 20.4 ± 1.0 13.2 ± 0.5 10.8 ± 0.5 16.0 ± 1.9 21.4 ± 1.9 21.7 ± 2.7 18.3 ± 1.9

DO (mg/L) 8.2 ± 1.3 8.5 ± 0.8 10.2 ± 0.8 10.3 ± 1.1 9.4 ± 0.8 7.8 ± 0.6 8.6 ± 1.1 8.2 ± 0.9

DO saturation (%) 93.7 ± 15.6 94.8 ± 8.5 97.3 ± 7.9 93.1 ± 10.9 94.7 ± 6.9 88.6 ± 7.8 98.3 ± 15.0 87.4 ± 11.6

pH 8.9 ± 0.2 8.9 ± 0.2 8.7 ± 0.3 8.7 ± 0.2 8.9 ± 0.2 8.4 ± 0.1 8.7 ± 0.4 8.3 ± 0.3

Total alkalinity (CaCO3 meq/L)

--- --- --- --- 3.4 ± 0.4 3.2 ± 0.5 --- 4.0 ± 1.0

Filtered alkalinity (CaCO3 meq/L)

--- --- --- --- 2.9 ± 0.3 3.1 ± 0.5 3.1 ± 0.5 3.3 ± 0.8

Turbidity (NTU) 25.7 ± 17.2 20.4 ± 11.7 142.1 ± 151.0 31.4 ± 21.5 160.7 ± 97.3 35.1 ± 27.1 37.0 ± 38.9 147.2 ± 383.1

TSS (mg/L) 69.8 ± 42.7 54.5 ± 21.9 183.1 ± 152.0 66.0 ± 24.5 196.3 ± 100.3 83.5 ± 37.9 80.5 ± 39.1 409.9 ± 1200.0

SPOM (mg/L) 28.3 ± 9.2 24.7 ± 8.0 59.6 ± 38.0 27.5 ± 7.5 65.2 ± 22.5 34.5 ± 15.1 37.1 ± 14.1 91.2 ± 208.4

TISS (mg/L) 41. 5 ± 34.7 29.7 ± 14. 7 123.4 ± 116.2 38. 5 ± 18.6 131.2 ± 78.7 49.1 ± 25.5 43.4 ± 26.6 318.7 ± 991.9

K 3.8 ± 1.9 3.4 ± 1.0 4.2 ± 3.1 4.4 ± 2.5 9.4 ± 5.6 4.6 ± 2.2 4.3 ± 3.4 4.8 ± 4.12

ZEU (m) 1.5 ± 0.8 1.5 ± 0.6 2.0 ± 1.7 1.4 ± 0.7 0.8 ± 0.6 1.3 ± 0.8 1.8 ± 1.3 1.7 ± 1.3

Chlorophyll a (μg/L) 12.4 ± 6.8 36.4 ± 9.8 56.3 ± 41.3 43.0 ± 25.5 76.2 ± 28.9 41.8 ± 26.4 47.3 ± 16.5 84.1 ± 176.7


0

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Ale

xandri

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iddle

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(g

/L)

Jan-07

Mar-07

May-07

Jul-07

Oct-07

Nov-07

Jan-08

Apr-08

Figure 4. Changes in salinity in the Lower Lakes from January 2007 to April 2008.


Figure 5. Interpolated salinity in the Lower Lakes from January 2007 to April 2008. Note that figure has been rotated for modelling purposes and units are PSU. Provided by Matthew Hipsey, the University of Western Australia.


Table 2. Relationship between salinity and water level with various parameters. Only statistically significant regressions are presented, where p < 0.05.

Parameter Water depth Salinity

p r2 Relationship p r

2 Relationship

Salinity 0.002 0.81 Inverse NA NA NA

TSS 0.03 0.56 Inverse 0.03 0.59 Positive

SPOM 0.03 0.59 Inverse 0.03 0.57 Positive

TISS 0.04 0.55 Inverse 0.03 0.59 Positive

Chlorophyll a 0.005 0.76 Inverse 0.02 0.60 Positive

pH --- --- --- 0.01 0.70 Inverse

Filtered alkalinity --- --- --- 0.04 0.92 Positive

NH4-N 0.004 0.77 Inverse 0.009 0.70 Positive

PO4-P 0.02 0.64 Inverse 0.02 0.65 Positive

TS 0.003 0.80 Inverse <0.0001 0.94 Positive

DOC 0.02 0.61 Inverse 0.02 0.60 Positive

Within Lake Alexandrina Arm there also existed strong vertical salinity gradients with the development of salinity stratification (Figure 6). These were more apparent near the barrages but were still evident at Clayton for a majority of the study period (Figure 6). Salinity stratification was also evident at Points in May 2007, but was not evident in Lake Alexandrina Middle or in Lake Albert (Figure 6). However, at times, vertical salinity gradients were also evident at Lake Alexandrina Opening (Figure 6). This may indicate the occasional movement of cooler, more saline water from the main body of Lake Alexandrina to the hypolimnion of Lake Alexandrina Opening (Figure 6 and Figure 7).

The continuous temperature monitoring showed that while persistent thermal stratification was not evident initially in Lake Alexandrina Opening, it developed in October 2007 and persisted for the remainder of the study period (Figure 8). In comparison, at the intersection of the River Murray and Lake Alexandrina, thermal stratification was evident throughout the study period, suggesting that earlier in the study stratification broke down moving from the River Murray to Lake Alexandrina Middle (Figure 8), where stratification was only evident during warmer periods but was not persistent (Figure 9). Overall, there was no evidence of temperature stratification in open water sites, but stratification was evident at sheltered sites, such as Clayton (Figure 7). However, in comparison to other sites, hypolimnetic waters around Goolwa were often found to have higher water temperatures (Figure 7), suggesting the source of saline water was also warmer. Surface water temperature variations were greater at the end of the study period, especially during summer months (Figure 8).


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-9

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Figure 6. Salinity profiles in the Lower Lakes in January 2007 (full dark line), May 2007 (large dashed line or triangle), October 2007 (small dashed line) and April 2008 (full grey line or square).

B

Tauwitchere

Clayton Points

Lake Alexandrina Middle Lake Albert Entrance

Lake Alexandrina Opening River Murray

Goolwa


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Figure 7. Temperature profiles in the Lower Lakes in January 2007 (full dark line), May 2007 (large dashed line), October 2007 (small dashed line) and April 2008 (full grey line).

Goolwa

Lake Alexandrina Middle

Lake Alexandrina Opening

Clayton


Date

01/01/07 01/05/07 01/09/07 01/01/08

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°C)

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ter

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0.0 m

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Figure 8. Water temperature profiles in the Lower Lakes, January 2007 to March 2008.

River Murray- Lake Alexandrina intersection


Poltalloch Lake Albert Entrance


Date

01/03/08 06/03/08 11/03/08 16/03/08 21/03/08

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mp

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)

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0.0 m

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Figure 9. Water temperature profiles in the Lower Lakes during March 2008.

River Murray- Lake Alexandrina intersection


Poltalloch

Lake Albert Entrance


2.3.2. Dissolved oxygen

Although increasing salinity did not have a significant impact upon dissolved oxygen concentration or saturation (Table 2) in surface water, dissolved oxygen concentrations declined towards zero in the hypolimnion when present (that is, the bottom waters when stratification was present) (Figure 10). As for salinity stratification, the magnitude of hypolimnetic dissolved oxygen depletion decreased with distance from the barrages. However, at sites upstream from the barrages oxygen consumption within the bottom waters was apparent with a decrease in dissolved oxygen concentrations with water depth (Figure 10). Within Lake Alexandrina Arm there also appeared to be a general decrease in dissolved oxygen saturation between January 2007 and April 2008, but this was less apparent in upstream sites (Figure 11).

2.3.3. Suspended particles

Suspended material (TSS, SPOM, TISS, chlorophyll a) increased with falling water depth and increasing salinity during the study period (Table 2). However, there was no relationship between wind velocity and suspended material. The large variation in average values of suspended material in April 2008 (Table 1) was a result of extremely high concentrations of suspended material in Lake Albert. There was no apparent change in the composition of suspended material during the study period, with TSS consisting of between 30 and 40% SPOM, except for April 2008 when it was 22%.

2.3.4. pH and alkalinity

Average pH within the lakes fell with increasing salinity during the study period (Table 2). Average pH in January 2007 was 8.9 ± 0.2, but in April 2008 was 8.3 ± 0.3 (Table 1). However, this was not accompanied by a drop in alkalinity. In fact, filtered alkalinity had a positive relationship with salinity (Table 2), increasing from 2.9 ± 0.3 CaCO3 meq/L in October 2007 to 3.3 ± 0.8 CaCO3 meq/L in April 2008.


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Figure 10. Dissolved oxygen concentration profiles in the Lower Lakes in January 2007 (full dark line), May 2007 (large dashed line), October 2007 (small dashed line) and April 2008 (full grey line).

Goolwa


Clayton



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Figure 11. Dissolved oxygen saturation profiles in the Lower Lakes in January 2007 (full dark line), May 2007 (large dashed line), November 2007 (small dashed line) and April 2008 (full grey line).

Goolwa Clayton




2.3.5. Nitrogen concentrations

Average NH4-N concentrations increased at all sites during the study period (Figure 12). In Lake Albert there appeared to be decreasing NH4-N concentrations from North to South, but this was more evident in Lake Alexandrina, particularly in Lake Alexandrina Arm (Figure 12). Within Lake Alexandrina Arm considerable variation was observed, with concentrations varying from 0.8 mg/L at Goolwa to 0.3 mg/L at Clayton in April 2008. At Goolwa, hypolimnetic samples had NH4-N concentrations 2.1 and 3.3 times greater than epilimnetic samples in November 2007 and April 2008, respectively (Table 3). Concentrations of NH4-N were most closely related to salinity (Table 2) and euphotic depth (p <0.0001, r2 = 0.24). Thus, density stratification clearly resulted in increased NH4-N concentrations in lake water.

Table 3. Nutrient and organic carbon concentrations in epilimnion and hypolimnion at Goolwa in November 2007 and April 2008.

Month Sample NH4-N (mg/L)

NOx-N (mg/L)

TN (mg/L)

PO4-P (mg/L)

TP (mg/L)

SO4-S (mg/L)

DOC (mg/L)

TOC (mg/L)

Nov-07 Epilimnion 0.1 <0.005 1.5 0.004 0.03 --- 12.3 14.1

Hypolimnion 0.5 0.02 1.6 0.006 0.05 --- 11.6 12.6

Apr-08 Epilimnion 0.5 <0.005 1.5 0.003 0.08 494.1 13.1 12.0

Hypolimnion 1.0 0.02 1.6 0.005 0.1 607.6 8.7 9.1

In contrast NOx-N concentrations were highly variable and displayed no consistent temporal or spatial patterns (Figure 13). There was a particularly high NOx-N concentration in Finniss in July, during inflows from the Finniss River. High NOx concentrations were also observed in Milang, Narrung and Mundoo in April 2008 (Figure 13).

Total nitrogen consisted of only 6.7% dissolved inorganic nitrogen (NOx-N plus NH4-N). Thus, a majority of the nitrogen was in organic forms. However the proportion of inorganic nitrogen increased through the study period from 3.1 to 14.8%. Total nitrogen concentrations were highest in Lake Albert and Boundary Creek (Figure 14). There was a general increase in average TN concentrations during the study period, although there were relatively high concentrations in October 2007, particularly in Lake Alexandrina Body (Figure 14). This was most likely associated with the higher amount of suspended material, with TN closely related to suspended material (p <0.0001 r2 = 0.30 for linear regression with <4.48 μm particle size class and TN) and light availability (p <0.0001 r2 = 0.40 for linear regression with k and TN).

2.3.6. Phosphorus concentrations

There were only small changes in phosphorus concentrations during the study period. It appeared that PO4-P concentrations tended to be higher in sites in Lake Albert and close to the barrages, with particularly high concentration observed in Boundary Creek in April 2008 (Figure 15). In many sites concentrations appeared to increase initially during the study period to May-October 2007 before falling again to April 2008 (Figure 15). Although TP only comprised of between 1.4 (January 2007) and 14.6% (July 2007) PO4-P, temporal variation in TP was associated with changes in PO4-P, with concentrations highest in May-October 2007 (Figure 16). Average TP concentrations tended to be higher in Lake Albert, followed by Lake Alexandrina Body and Lake Alexandrina Arm (Figure 16). An exception to this was Boundary Creek which had particularly high TP concentrations. These spatial patterns were most likely associated with the higher amount of suspended material that was observed in the Lower Lakes, with TP closely related to suspended material (p = <0.0001 r2 = 0.26 for POM) and light availability (p = <0.0001 r2 = 0.30 for relationship with k).


2.3.7. Other elements

Total sulfur concentrations increased during the study period at all sites (Figure 17) as water levels fell and salinity increased (Table 2). Concentrations increased gradually downstream through Lake Alexandrina Body, but increased dramatically through Lake Alexandrina Arm with closeness to the barrages (Figure 17). The increase in TS concentrations during the study period was also much greater in those sites located near the barrages (Figure 17). TS was also related to salinity ZEU (p = <0.0001 r2 = 0.94). These patterns were also observed for SO4-S (Figure 18), which constituted an average of 87.8% ± 4.9 and 96.6% ± 2.4 of TS in January 2008 and April 2008, respectively.

There was a general increase in DOC concentrations during the study period (Figure 19) as water levels fell and salinity increased (Table 2). Concentrations were lower in Lake Alexandrina Body and sites near the barrages, with exception of Boundary Creek which had particularly high concentrations (Figure 19). TOC consisted of between 44.1 and 78.4% DOC and POC displayed different spatial and temporal patterns. Unlike DOC, there was a general decrease in POC concentrations from October 2007 (Figure 20). However, in Lake Alexandrina Body and Lake Albert, POC increased between January and April 2008 (Figure 20). In general, concentrations of POC were lower in Lake Alexandrina Arm (Figure 20).

2.3.8. Summary of results

As water levels fell in the Lower Lakes during the study period there was a concurrent increase in salinity (Table 1 and Table 2). Salinities in the main bodies of Lake Alexandrina and Lake Albert increased gradually during the study period, whereas salinity and the rate of salinity change increased with proximity to the barrages. In addition, there was increased prevalence of salinity stratification and oxygen depletion of the hypolimnion observed in the Lake Alexandrina Arm. Ammonia, PO4, TS, SO4-S and DOC concentrations increased as water levels and salinity increased during the study period. As for salinity, for NH4, PO4, TS and SO4-S this increase was strongly associated to proximity to the barrages. TN and TP concentrations displayed less temporal or spatial variation than NH4-N and PO4-P, with variation in TN and TP associated with suspended material. There was an increase in suspended material (TSS, SPOM, TISS and chlorophyll a) during the study period. While these changes were related to water level and salinity (Table 1), they were not related to wind velocity.


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Figure 12. Changes in ammonia concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 13. Changes in nitrate-nitrite concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 14. Changes in total nitrogen concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 15. Changes in phosphate concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 16. Changes in total phosphorus concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 17. Changes in total sulfur concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 18. Changes in sulfate concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 19. Changes in dissolved organic carbon concentrations in the Lower Lakes from January 2007 to April 2008.


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Figure 20. Changes in particulate organic carbon concentrations in the Lower Lakes from January 2007 to April 2008.


2.4. Discussion

2.4.1. Salinisation of the Lower Lakes

In arid climates, water level drawdown and salinisation often coincide due to evapoconcentration of dissolved salts (Williams 1987; Micklin 1988). However, in the case of the Lower Lakes salinisation did not appear to be due to evapoconcentration alone. Given the average water volume and salinities of the Lower Lakes between January 2007 and April 2008, salt present in the water column of the Lower Lakes increased by approximately 650% (Table 4), whereas the salt present would have remained constant if only evapoconcentration was involved. Thus, a large external load of salt occurred during the study period. However, the calculated increase in salt mass is considered to be an overestimate, due to the smaller volume of Lake Alexandrina Arm than Lake Alexandrina Body and Lake Albert. Consequently, it is likely that the high salinity values of Lake Alexandrina Arm have a disproportionate influence on the calculated mass. This is also the case for other discussed analytes.

Table 4. Change in average concentrations of sampling sites and resultant mass of elements in the Lower Lakes between January 2007 and April 2008. Average concentrations are not volume-weighted. Average water level data provided by Bureau of Meteorology (BOM, unpublished data) and water level-volume relationship based on data from Department of Water, Land and Biodiversity Conservation (DWLBC, unpublished data). In January 2007 average water levels were 0.42 m AHD with a calculated volume of 1760 GL. In April 2008 average water levels were -0.50 m AHD with a calculated volume of 1128 GL.

Parameter Unit Jan-2007 Apr-2008 % Increase

in mass

Salinity g/L 0.9 10.6

646 Tonne 1596200 11913514

NH4-N mg/L 0.04 0.4

495 Tonne 69 408

NOx-N mg/L 0.007 0.02

27 Tonne 13 17

TN mg/L 1.5 2.6

9 Tonne 2636 2879

PO4-P mg/L 0.002 0.01

404 Tonne 3 14

TP mg/L 0.1 0.1

-27 Tonne 197 144

DOC mg/L 8.5 13.9

4 Tonne 14995 15651

TS mg/L 28.2 278.7

534 Tonne 49580 314328


Cook et al. (2008) also found a large unknown source of salts into the Lower Lakes between 1979 and 1997. Potential sources include saline groundwater discharge (Lamontagne et al. 2005), atmospheric deposition and seawater leakage through the barrages (Cook et al. 2008). Cook et al. (2008) hypothesised that the dominant unknown supply of salts to the Lower Lakes was likely to be groundwater rather than from seawater leakage as the ratio of the major ions to chloride in the Murray-Darling River is slightly more elevated than in seawater. However, in this study there was a strong relationship between salinity and proximity to the barrages, suggesting that seawater leakage was the dominant source of salts during the period of low water levels. The slowing of salinity incursions between winter and spring 2007 may have been associated with the works undertaken to reduce barrage leakage during this period (Jarrod Eaton, DWLBC, personal communication). However, during summer and autumn 2008 there was another rapid increase in salinity. This may have been associated with further leakage of the barrages or increased groundwater discharge, since saline groundwater intrusions tend to increase with falling water levels. However, groundwater–surface water interactions in the Lower Lakes region are not well known and require further investigation.

As hypothesised, the input of saline water into the Lower Lakes resulted in the development of salinity stratification. This is a common phenomenon associated with salinisation of inland waters (Anderson and Morison 1989; Naselli-Flores 2003; Nielsen et al. 2003; Baldwin et al. 2006; Davis and Koop 2006), whereby current or wind energy is not sufficient to prevent the formation of discrete water column layers of different salinities: dense saline water in the hypolimnion and less dense, less saline water in the epilimnion. The occurrence of density stratification in the lakes means that only monitoring surface water quality may not appropriately represent the overall lake conditions. Temperature-dependent density stratification was also apparent in Lake Alexandrina Arm. However, at Goolwa water temperatures were higher in the hypolimnion, suggesting an input of warm saline water. While persistent thermal stratification appeared to be limited to more sheltered sites, there was a dramatic increase in the temperature variability in warmer months toward the end the study compared to warmer months at the beginning of the study. This is most likely due to the lower water levels, resulting in a greater surface area to volume ratio for heating and cooling.

The occurrence of salinity-driven density stratification is not unique to the Lower Lakes. Anderson and Morison (1989) found that in the Wimmera River saline groundwater intrusion resulted in the development of salinity-driven density stratification in an extensive area of the river. This saline water tended to travel downstream and stratification was only being broken down during periods of major flows. In this study, the prevalence of salinity stratification increased through the study period with a salt wedge of denser saline water moving from the barrages through Lake Alexandrina Arm. Salinity stratification appeared to be broken down upon entry to open water sites of Lake Alexandrina on most sampling occasions through wind-induced mixing. However, at times, vertical salinity gradients were also evident at Lake Alexandrina Opening, perhaps indicating the intrusion of cooler, more saline (more dense) water from the main body of Lake Alexandrina to the hypolimnion of Lake Alexandrina Opening. This is a likely response to the temperature difference between the main body of Lake Alexandrina and Lake Alexandrina Opening, resulting in the movement of saline water upstream from Lake Alexandrina. This is a likely mechanism of the observed saline intrusions into the hypolimnion of the River Murray (Rob Daly, SA Water, unpublished data).

2.4.2. Changes in physical, chemical and biological conditions

The large increase in salinity would be expected to have a large impact on physical conditions within the Lower Lakes. It was hypothesised that as salinity increased within the lakes, suspended solids would decrease due to flocculation of suspended material (Grace et al. 1997; Nielsen et al. 2003) and reduced suspended solid concentrations from the River Murray under low flow (Cook et al. 2008). In fact, this was not the case with suspended material increasing during the period of water level drawdown and salinisation. The most likely explanation for this is the movement of the shoreline towards the centre of the lake, which contains finer sediments


(see Chapter 3) due to sediment focussing. Since finer particles have a lower critical shear stress (Bloesch 1995), the movement of the shoreline towards these sediments will result in increased rates of resuspension. In addition, water level drawdown results in the concentration of suspended material into a smaller volume. The increased suspended solid concentrations are likely to reduce primary productivity, especially for submerged macrophytes, and thus reducing habitat and organic matter resources for higher organisms. Indeed, in the Lower Lakes the inherent high wind velocities and wave generated turbulence results in high suspended solids that influences primary and secondary productivity of the lakes (Geddes 1984a; Geddes 1984b).

2.4.3. Changes in nutrient concentrations

The study highlighted that sediment resuspension and the establishment of permanent density stratification are key drivers for nutrient concentrations in the water column of the Lower Lakes. It was hypothesised that TN and TP concentrations would decrease during the study period because of the tendency for increased flocculation and deposition of suspended sediments at higher salinities and low inputs under low flow (Cook et al. 2008). However, this was not the case. There was a small increase in TN concentrations during the study period, consistent with an increase in suspended sediment concentration. As suspended sediment inputs from the River Murray were low during the study, sediment resuspension within the lakes played a key role in maintaining relatively elevated nutrient concentrations.

The generation of permanent density stratification was a determinant factor in increasing inorganic nutrient concentrations in the Lower Lakes during the monitoring program. The principal source of oxygen to lakes is exchange with the atmosphere. This process is curtailed for the hypolimnion following the onset of stratification. Hypolimnetic oxygen can be rapidly consumed by the decomposition of sedimentary organic matter (Davis and Koop 2006). This may cause a shift from aerobic to anaerobic decomposition of organic matter in sediments (Naselli-Flores 2003; Nielsen et al. 2003; Baldwin et al. 2008), resulting in the release of inorganic nutrients, such as ammonium and phosphorus, into the water column (Baldwin et al. 2008). In addition, the input of saline water over sediments in equilibrium with less saline water will result in convective transport of the less saline water up through the water column. This will result in greater rates of nutrient transport from the sediments to the water column than what would happen through diffusion alone.

Although, there was no large change in PO4 concentrations across the system as a whole during the study period, PO4 concentrations increased in Lake Alexandrina Arm where the salinity increase and subsequent oxygen depletion was most evident. Oxygen depletion in the sediments is thought to increase the flux of PO4 from sediments to the water column due to the reduction and solubilisation of iron, to which PO4 is adsorbed to in sediments (Boström et al. 1988). However, Baldwin et al. (2006) found decreased phosphorus concentrations with increasing salinity due to increased solubilisation of iron and precipitation of PO4 with iron. In the study of Baldwin et al. (2006) only sodium chloride was used as a source of salinity. However, in the Lower Lakes a part of the increase in salinity was caused by increased SO4 concentration. Under anaerobic conditions, sulfate reduction can occur, leading to the production of Fe-S compounds and the release of PO4 from the sediment pool (Wetzel 2001). Thus, the salinisation of the Lower Lakes may have favoured the recycling of PO4 by promoting anoxic conditions and increased SO4 reduction rates.

The generation of an anoxic hypolimnia and the increase in SO4 reduction rates following salinisation could also have a strong impact on the nitrogen cycle in the Lower Lakes. Higher NH4 concentrations were observed in the hypolimnion than the epilimnion at Goolwa during periods of stratification, consistent with reduced nitrification rates under anoxic conditions. Displacement of NH4 from sediment exchange sites by cation exchange (Rysgaard et al. 1999; Baldwin et al. 2006; Laverman et al. 2007) could also be caused by rising salinity. Alternatively, the increased NH4 concentration in the Lower Lakes could have been caused by increased decomposition rates in sediment exposed to the atmosphere during the drawdown (Sah et al. 1989; Mitchell and Baldwin 1999; Baldwin and Mitchell 2000; Baldwin et al. 2005). In this case,


occasional reflooding of sediments by rainfall, seasonal increases in water level, or by wind-driven hydrodynamic processes would be required to transport the remineralised NH4 to the water column.

While it also appeared that there was a source of NOx to the Lower Lakes (Table 4), this was a due to high average NOx concentration in April 2008 as a result of high concentrations at Milang, Narrung and Mundoo. These sites were all particularly shallow under drawdown water levels, perhaps resulting in greater nitrification rates with increase delivery of oxygen to the sediment in shallow, well mixed water. Overall, NOx was highly variable with no clear spatial or temporal patterns, suggesting that differences were associated with internal cycling rather than external inputs. Drawdown potentially increases oxygen delivery to the sediment and increases nitrification, but can also increase denitrification rates (Baldwin and Mitchell 2000). In comparison, NH4 loss from sediments induced by salinity (Rysgaard et al. 1999; Baldwin et al. 2006; Laverman et al. 2007) may result in reduced nitrification rates. Furthermore, salinity may reduce nitrification rates through inhibitory effects on nitrifying bacteria (Rysgaard et al. 1999).

Similarly, there are a number of processes that may have caused the observed increase in DOC concentrations in the lakes during the study period. Firstly, the drying-reflooding of fringing sediments may have resulted in an increased flux of organic carbon to the overlying water column (Baldwin and Mitchell 2000). In addition, the increased algal biomass, as indicated by an increase in chlorophyll a during the study period, may have resulted in greater DOC production rates by algae (Nguyen et al. 2005). However, most of the increase in DOC concentrations during the study period can be accounted for by evapoconcentration (Table 4). Due to the long water residence time of the River Murray, riverine DOC inputs are probably be relatively refractory and conservative once in the lakes (Curtis and Schindler 1997).

2.4.4. Conclusion

This study has highlighted that the water level and salinity regimes are key environmental drivers in the Lower Lakes and can interact in complex ways. In particular, the generation of density stratification by seawater leakage and the exposure of sediments to the atmosphere are key processes influencing water quality during drawdown periods. This is also consistent with recent findings showing that water level lowering has exposed Potential Acid Sulfate Soils and in some cases generated localised areas with Actual Acid Sulfate Soils along the margins of the lakes (Fitzpatrick et al., 2008). These can be a significant threat to water quality, especially following re-flooding. To evaluate potential consequences for water quality, management interventions in the Lower Lakes will need to carefully evaluate what impacts these will have on their water level and salinity regimes.


3. Relationship of sediment character with water depth in the Lower Lakes, as evidence of sediment focussing

Kane T Aldridge

Brian M Deegan

Justin D Brookes

3.1. Introduction

Sediment is considered to be all settled material in aquatic environments and consists of various forms of organic matter, particulate mineral matter and inorganic components of biogenic origin (Wetzel 2001). Sediments provide essential habitat for a range of organisms including periphytic algae, macrophytes, macroinvertebrates and heterotrophic microbial organisms. Perhaps the most important function of sediments within aquatic ecosystems is providing habitat and resources for heterotrophic microbial communities, thus controlling biogeochemical cycling of elements. Sediments are a major store of nutrients (Marsden 1989; Martinova 1993; Søndergaard et al. 1993); contain a majority of the heterotrophic microbial organisms that are involved in biogeochemical cycling (Wetzel 2001); and are the major site of decomposition of dead organic matter and its recycling back into foodwebs (Wetzel 2001). Consequently, their role in controlling physicochemical conditions within aquatic ecosystems is fundamental and the distribution of sediments of different character exerts a strong control over the functioning of aquatic ecosystems (Odum 1971).

Within lakes, sediments are typically heterogeneous in their distribution owing to the large number of physical, chemical and biological processes that interact to influence their character (Hilton et al. 1986; Bloesch 1995). Sediments may be of internal or external origin, the latter brought in from inflowing rivers and streams or atmospheric deposition (Wetzel 2001). Upon entering lakes, sediment derived from the drainage basin may be deposited due to lower current velocities. This often results in longitudinal gradients of sediment character, with finer particles carried further before deposition due to their lower settling velocities (Hilton et al. 1986; Kumke et al. 2005). Similarly, elements stored within sediments, such as nutrients and contaminants, may have higher concentrations closer to the source due to greater deposition rates and uptake by the sediments (Xu et al. 2003). However, sediments are continually redistributed throughout lakes by resuspension and deposition, driven by internal physical processes, such as wind driven mixing, seiching and current and wave erosion (Bloesch 1995). The net result of these processes is the deposition of finer sediments in areas of low energy, often resulting in what is known as sediment focussing, the accumulation of finer sediments in deeper and sheltered regions (Hilton et al. 1986; Matty et al. 1987; Lebo and Reuter 1995).

Once settled, sediment character may be altered by chemical and biological processes, further enhancing heterogeneity. For example, the presence of aquatic macrophytes can result in: the deposition of particulate organic matter (Boon and Sorrell 1991; Brix 1997; van der Putten et al. 1997); the deposition of finer sediments due to dampening of wave energy (Sand Jensen 1998; Sand-Jensen and Pedersen 1999); and alteration of the abundance of various elements as a result of oxygen release from roots (Chen and Barko 1988; Moore et al. 1994; Wigand et al. 1997; Aldridge and Ganf 2003). Similarly, macroinvertebrates can result in the redistribution of sediment particles and alter chemical conditions within sediments through bioturbation and bioirrigation (Aller 1982; Vaughn and Hakenkamp 2001).


The transformation of organic material inherently alters the sediment composition and physicochemical conditions. While labile organic matter will be utilised rapidly and re-enter foodwebs, refractory organic matter may remain in sediment for extended or indefinite periods (Wetzel 2001). The rate and quantity of organic matter decomposition is dependent upon the availability of electron acceptors and nutrients (Wetzel 2001), which also interact with sediment character. Oxygen is an important electron acceptor that results in the greatest rates of organic matter decomposition. However, oxygen content of sediments is dependent upon sediment type, with lower oxygen found within finer sediments. This is due to low rates of oxygen transport within fine sediments and increased trapping of fine organic material, large surface area for microbial colonisation and higher oxygen consumption (Eagle 1983). The oxygen content will have further implications for biogeochemical cycling. For example, nitrification requires the presence of oxygen, while denitrification rates are higher in anoxic conditions (Wetzel 2001). Similarly, phosphorus desorption from iron complexes within sediment occurs when these iron complexes are reduced under anoxic conditions (House and Denison 1998; Webster et al. 2001).

It is clear that sediment distribution within lakes is a result of a complex interaction of physical, chemical and biological processes, which contribute to their heterogeneous distribution. The aim of this project was to gain an understanding of the spatial distribution of various sediment types within Lake Alexandrina and Lake Albert, two terminal lakes of the River Murray, collectively known as the Lower Lakes. The natural flow regime of the Lower Lakes has been modified substantially. In the absence of large currents, associated with reduced river flows and barrage construction, the development of longitudinal gradients in sediment character are likely to be less pronounced then under natural conditions. Instead, the dominant source of energy distributing sediments in the lakes are presumably now winds, with resuspended sediments deposited in deep, low energy areas. It was hypothesised that high rates of sediment resuspension would result in a high degree of sediment focussing, with sediment character distributed according to water depth. In comparison, it was not expected that sediment character would differ longitudinally. This information is used to examine the processes most likely to be influencing the distribution of sediments within the Lower Lakes and provide a basis for understanding the spatial heterogeneity in biogeochemical processes within the Lower Lakes.

3.2. Methods

3.2.1. Sediment collection

Sediment samples were collected from 40 sites along 10 transects in Lake Alexandrina and Lake Albert (Figure 21). Transects were selected to account for all habitat types, including sheltered and open water sites. Transects were run from the shallowest point accessible by boat on one shoreline to the opposite side of the lake or channel, approximately perpendicular to the lake-shore. Sampling locations were selected along transects to reflect changes in water depth (Figure 22). In Lake Albert, shallow water prevented a boat from being launched and samples were collected by wading. Consequently, an accurate reflection of the water depth gradient was not achieved for these transects. No samples could be collected from T2 or P3 due to the presence of rock substrate. For the purposes of this study, sediment cores were assigned to a water depth class, divided to give approximately equal number of replicates in each depth class. Depth class 1 represents <0.6 m, depth class 2 represents 0.61-0.9 m, depth class 3 represents 0.91-1.5 m, depth class 4 represents 1.51-2.5 m and depth class 5 represents >2.5 m.


Figure 21. Location of transects and sampling sites of Lake Alexandrina and Lake Albert. Only peripheral sites in each transect are listed, with a sequential increase in site numbers between the two peripheral sites.

Between 12th and 14th February 2007, one intact sediment core was collected from each site in Perspex cores with an internal diameter of 57 mm. Unsealed cores were attached to a pole and gently pushed into the sediment perpendicular to the sediment surface. The cores were pushed 10-15 cm into the sediment, sealed and gently extracted. The core was retrieved and a rubber stopper was pushed into the base to prevent sediment loss. The upper seal was then broken and overlying water was siphoned off, except for approximately 1 cm to maintain the moisture content of the sediment. Cores were immediately stored in the dark, below 3°C.

3.2.2. Sediment character

On return to the laboratory samples were removed gently from the cores and divided into 0-1 cm, 1-5 cm and 5-10 cm depth intervals and fresh-weight was recorded. For the purposes of this study only 0-1 cm sediment was characterised since it is considered the sediment most active in biogeochemical cycling. Sub-samples were collected and placed into centrifuge tubes for determination of chlorophyll a following Golterman et al. (1978), using 99.8% methanol and a Hitachi U-2000 spectrophotometer (Hitachi Ltd., Tokyo, Japan), with a path length of 10 mm. Sequential extractions were conducted, but revealed no significant additional extraction of chlorophyll a.

Sub-samples were also removed for analysis of particle size distribution. Sediments were suspended in deionised water and the particle size distributions were determined using a Laser In-Situ Scattering and Transmissometry instrument (LISST-100X, Sequoia Scientific). This instrument measures scattering of light at 32 angles and this information is used to calculate the size distribution of 32 size classes log-spaced between 2.73 and 462 μm. For the purposes of this study size classes were divided into six categories of different particle size class (PSC): <4.48 μm (PSC1), 4.48-6.24 μm (PSC2), 7.36-19.9 μm (PSC3), 23.5-63.3 μm (PSC4), 74.7-186 μm (PSC5), 219-462 μm (PSC6). Data is presented as percent composition of each PSC by volume. Although this analysis did not include particles greater than 462 um, sediment


descriptions recorded at the time of collection indicated that there were no particles coarser than fine sands (approximately 500 μm).

Figure 22. Bathymetry of Lake Alexandrina, Lake Albert and the Northern Coorong. Legend shows position of sediment surface (m AHD). Data provided by Department of Water, Land and Biodiversity Conservation.

The remaining sediment was dried to a constant weight at 105°C and dry-weight was recorded, with moisture content measured as the difference between fresh-weights and dry-weights (expressed as a percentage of fresh-weight). Sediments were ground with a mortar and pestle and analysed for benthic particulate organic matter (BPOM), total carbon (TC), total nitrogen (TN) and total phosphorus (TP). BPOM was determined by igniting dried sediment to constant weight at 550°C to determine fixed and volatile solids following APHA method 2540-E (Eaton et al. 2005). TC and TN were analysed using a LECO TruSpec Carbon/Hydrogen/Nitrogen Determinator (LECO, St. Joseph, USA). TP was analysed using the standard persulphate digestion method, 4500-N C (Eaton et al. 1995). The molar ratios of phosphorus, carbon and nitrogen were calculated.

3.2.3. Water column character

Immediately after sediments were collected water depth was recorded. In addition, a number of water column characteristics that relate to sediment character were measured. Three instantaneous measurements of light intensity were recorded at 0.25 m intervals using a LI-COR underwater spherical quantum sensor (LI-193SA). For each depth interval, the light attenuation coefficient (k) was calculated using the Beer-Lambert equation:

1.0 m

0.0 m

-1.0 m

-2.0 m

-3.0 m

-4.0 m

-5.0 m

-6.0 m

-7.0 m

-8.0 m

20 km


Z

II

kO

Zln

(1)

where IZ is light intensity at water depth Z, and IO is the light intensity at the water surface. For each depth interval, the euphotic depth (ZEU) was calculated as:

kZEU

61.4 (2)

Average k and ZEU values were calculated as the average of each depth interval measured for a single profile. The ratio of ZEU and Z was calculated to reflect light availability at the sediment surface.

Integrated water samples were collected using polyvinyl chloride tubing with an internal diameter of 54 mm. The tubing was gently lowered through the water column and sealed. The sample was retrieved and unfiltered water was collected and immediately stored in the dark, below 3°C. These samples were analysed for total suspended solids (TSS), suspended particulate organic matter (SPOM), total inorganic suspended solids (TISS), suspended chlorophyll a and particle size distribution.

Total suspended solids were determined by APHA method 2540 (Eaton et al. 2005), whereby suspended particulate material is concentrated onto pre-combusted and pre-weighed Whatman International GF-C filters. The increase in weight of the filters (dried to constant weight at 105°C) represented total suspended solids. SPOM was analysed by combusting these filters, following methods described for sediment (Eaton et al. 2005). Total inorganic suspended solids (TISS) were calculated as the difference between TSS and SPOM. Material concentrated onto additional GF-C filters was used to measure suspended chlorophyll a (1978), as described previously. The suspended particle size distribution of samples was determined using a Laser In-Situ Scattering and Transmissometry instrument (LISST-100X, Sequoia Scientific), as described previously.

3.2.4. Statistical analyses

For the purposes of this study sediment character was considered to encompass the composition of particles size classes, moisture content, BPOM, benthic chlorophyll a, TC, TN and TP. To examine the overall distribution of sediment character throughout the lakes a NMS Ordination of sediment character was carried out on these parameters using PC-ORD® with a Euclidean distance measure. To determine which measured parameters were most strongly correlated with the distribution of sites within the ordination, a second matrix was overlain, which included these parameters as well as water depth, k, ZEU, TSS, SPOM, TISS and suspended chlorophyll a. A successful two-dimensional ordination was found with a stress level of 8.8. However, for this CC-3 had to be removed from the analysis as it was found to have a disproportionate influence over the analysis due to its extremely high BPOM. Successful ordinations of more than two dimensions were not found.

Spatial patterns of physical characteristics (moisture content and particle size classes) and biogeochemical characteristics (BPOM, benthic chlorophyll a, moisture content, TC, TN and TP) were examined in more detail. In addition, ratios of BPOM, nitrogen and phosphorus were examined to reflect nutrient availability. For these analyses one-way analysis of variances were conducted using JMP-IN® to determine whether the observed patterns of individual characteristics were associated with different transects or water depth class. Where significant differences were detected, Tukey’s HSD All Pairs Test was carried out to determine where the differences occurred.


To determine the effect of physical characteristics on biogeochemical characteristics (BPOM, benthic chlorophyll a, moisture content, TC, TN and TP and nutrient ratios) multiple regression analysis was conducted. The BPOM value of CC-3 was removed from these analyses due to its extremely high value. To limit the number of explanatory variables of the analyses, parameters describing light availability were ignored, since they were considered to be a function of suspended solids, which remained in the analysis. Since TSS, TISS and SPOM were highly correlated, only TSS was included in the analysis. However, for chlorophyll a, ZEU:Z was included in the analysis to determine the influence of benthic light availability. Particle size was the only measured factor considered to influence moisture content and so this relationship was examined through single regression analysis. For all analyses α = 0.05 and variability is reported as standard errors.

3.3. Results

Sediments were found to be relatively heterogeneous in their character with gradual transitions between sediment types rather than the formation of distinct groups, as shown in the ordination (Figure 23). Benthic chlorophyll a had the strongest relationship to sediment character in the ordination, most likely due to the much higher values at CC-4. However, greatest separation of sites occurred parallel with Axis 2, which was not associated with different transects, but instead corresponded with water depth. Shallow, shoreline sites tended towards the base of Axis 2, with a transition to deep, open water sites along Axis 2. This transition was strongly associated with increasing moisture, BPOM, PSC1-4 and decreasing PSC6 (Figure 23).

These results were supported by the analysis of variance, with differences in sediment character related to water depth rather than site (Table 5). This was indeed the case for all particle size classes, with no differences observed between transects for any size classes due to the large variation that was observed in sediment particle distribution between sampling sites within each transect. Instead particle size class was closely related to changes in water depth, with finer sediments associated with sediments collected from deeper water and coarser sediments associated with sediment collected from shallower sites (Table 5). Each particle size varied between the depth classes except for PSC5, which was consistently between 35 and 40% of the sediment (Table 5). PSC6 accounted for approximately 50% of the sediment in depth classes 1-3, but was much less abundant in depth class 5, where it was 14.7 ± 1.3% of the sediment (Table 5). The proportion of PSC 1-4 increased with increasing depth class (Table 5).


Figure 23. NMS Ordination of sediment character in the Lower Lakes. Stress level of 8.8 and r2 of

0.1 with the length of red vectors representing strength of parameters in controlling distribution of sites within the ordination. Parameters include benthic chlorophyll a (BChl), euphotic depth (ZEU), sediment moisture content (Bmoist), sediment particulate organic matter (BPOM), sediment total nitrogen (BTN), sediment total carbon (BTC), water depth (depth) and particle size classes of sediment (numbered ranges in µm).

Moisture content was also higher in sediments collected from deeper water (Table 5). Consequently, moisture content was also related to particle size, with moisture content increasing within increasing composition of smaller particle size classes (Figure 24, p = <0.0001 for PSC 1-4) and decreasing composition of larger particle size classes (p = 0.02 and <0.0001 for PSC5 and 6, respectively).

Despite BPOM also appearing to be higher in sediments collected from deeper areas (Table 5), the difference was not significant. However, CC-3 had BPOM 3.5 times higher than all other sites and with removal of this value the difference was significant (p = 0.007), with greater BPOM in depth class 5 than depth class 1. Despite this relationship and the relationship between particle size classes and water depth (Table 5), BPOM was not found to be related to particle size (Table 6). BPOM did have positive relationships with moisture content and TC (Table 6). This was also the case for benthic chlorophyll a (Table 6), although benthic chlorophyll a did not differ between depth classes (Table 5).

TC was found to be greater in the shallower depth classes, with highest TC observed in depth class 2 (Table 5). Consequently, TC was also found to decrease with increasing moisture content (Table 6). However, the relationship of TC with other parameters was strongly affected by the occurrence of two distinct groups: one with high moisture content and high TC and one with low moisture content and low TC. The observed inverse relationship between TC and moisture was largely driven by the separation of two distinct groups: <50% moisture-high TC and >50% moisture-low TC (Figure 25). Separate analysis of these relationships for the individual groups revealed that the relationship was not significant (p = 0.1 for < 50% moisture and 0.08 for >50% moisture). In contrast, although TC was not found to be related to BPOM, TC had a positive relationship with BPOM for the different sediment types of low and high moisture


content (p = 0.02 and p = 0.0003) (Figure 26). The separation of distinct groups was not associated with sediment particle size classes.

Table 5. Differences in sediment character between transects and depth classes. Average values shown for different depth classes. Letters shared (

WXYZ) denote values not significantly different

between depth classes according to Tukey’s HSD All Pairs Test. NS

denotes analyses where no significant differences were detected by analysis of variance (p > 0.05).

Parameter Transect Depth class

p p 1 2 3 4 5

PSC1 (% volume) 0.2 0.0004 0.2 ± 0.1

X

0.3 ± 0.3

X

2.0 ± 1.2

X

3.2 ± 1.3

X

8.2 ± 1.8Y

PSC 2 (% volume) 0.4 <0.0001 0.2 ± 0.1

X

0.4 ± 0.2

XY

0.8 ± 0.4

XY

1.3 ± 0.3

Y

2.3 ± 0.2Z

PSC 3 (% volume) 0.3 <0.0001 1.2 ± 0.5

X

1.9 ± 1.0

X

4.9 ± 2.5

X

7.4 ± 2.0

X

15.8 ± 1.2Y

PSC 4 (% volume) 0.3 <0.0001 3.7 ± 1.3

X

4.8 ± 2.4

XY

9.7 ± 4.6

XY

16.5 ± 3.7

Y

32.7 ± 1.8

Z

PSC 5 (% volume) 0.2 0.24 38.3 ± 5.6

NS

39.1 ± 6.7

NS

33.6 ± 4.4

NS

38.6 ± 4.6

NS

26.3 ± 1.9

NS

PSC 6 (% volume) 0.3 <0.0001 56.5 ± 6.8

X

53.6 ± 8.4

X

49.0 ± 8.4

X

33.0 ± 4.5

XY

14.7 ± 1.3Y

Moisture (% weight)

0.4 <0.0001 19.2 ± 0.8

Z

29.4 ± 6.9

YZ 42.4 ± 11.8

XY 58.2 ± 9.3

WX

75.6 ± 2.6W

BPOM (g/m2) 0.4 0.12

38.0 ± 6.5

W

69.4 ± 21.7

W

64.7 ± 15.9

W

101.6 ± 18.3

W

125.0 ± 14.8

W

Benthic chlorophyll a (mg/m

2)

0.3 0.55 82.5 ± 16.8

NS

120.8 ± 25.6

NS

77.3 ± 20.5

NS

85.2 ± 19.0

NS

96.9 ± 10.0

NS

TC (g/m2) 0.7 0.0001

102.5 ± 7.8

WX 122.9 ± 10.9

W 81.4 ± 11.3

XY 74.9 ± 9.8

XY 54.5 ± 4.8Y

TN (g/m2) 0.001 0.04

10.9 ± 1.6

W 6.7 ± 1.0

W

10.0 ± 1.6

W

7.9 ± 1.0

W

6.5 ± 0.6W

TP (g/m2) 0.3 0.70

0.6 ± 0.1

NS

0.8 ± 0.1

NS

0.6 ± 0.1

NS

0.8 ± 0.2

NS

0.7 ± 0.1NS

BPOM to TN ratio 0.1 0.0005 4.2 ± 1.1

W

11.0 ± 2.7

WXY

8.5 ± 2.8

WX

12.3 ± 2.8

XY

19.2 ± 1.2Y

BPOM to TP ratio 0.3 0.03 66.6 ± 16.8

W

90.9 ± 29.8

WX

118.4 ± 31.6

WX

153.8 ± 40.2

WX

191.2 ± 17.8

X

TN to TP ratio 0.0004 0.001 19.4 ± 3.6

X

8.9 ± 1.7

W

16.8 ± 2.0

WX

11.9 ± 0.9

W

10.0 ± 0.7W


There was also a significant difference in TN concentrations between depth classes, however it was not clear statistically where the differences occurred (Table 5). There was a general decrease in TN with water depth, except for depth class 2 which had relatively low nitrogen concentrations. Unlike most other sediment characteristics, there were some observed differences in TN between transects (Table 5). This difference was associated with high TN in Lake Albert transects with LAlbT having higher TN than LAM, T, M and CC. TN was not significantly related any other measured parameters.

Since no differences in phosphorus were observed between transects (Table 5), TN to TP molar ratios also differed between transects, with LAT and LAB having higher values than several other sites (p = 0.0004). TN to TP ratios were also found to be different amongst depth classes (p = 0.001) with depth class 1 having higher values than all other depth classes except depth class 3. Differences in TN to TP ratios were most closely related to suspended solids (Table 6). Sediments collected from deeper areas appeared to have a high inorganic carbon component since TC was greater than BPOM (Table 5). Consequently, availability of nitrogen and phosphorus to organic carbon were made by comparing to BPOM. It was found that both BPOM to TN and BPOM to TP ratios increased with water depth, with deeper depth classes having greater ratios than shallower depth classes (Table 5).

0

20

40

60

80

100

0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5

PSC2 (% composition)

Mo

istu

re c

on

en

t (%

weig

ht)

Figure 24. Relationship between Particle Size Class 2 composition and sediment moisture content (p = <0.0001, r

2 = 0.8 for regression with natural log transformation of PSC2).


Table 6. Relationship of sediment characteristics to measured parameters. Multiple regression analyses, with p values presented and a description of the relationship between parameters in parentheses for significant relationships (p < 0.05).

Parameter Statistic TC TN TP BPOM Benthic

chlorophyll a BPOM to TN

ratio BPOM to TP

ratio TN to TP

ratio

Model p <0.0001 0.007 0.0006 <0.0001 0.02 <0.0001 <0.0001 0.003

r2 0.88 0.62 0.70 0.91 0.61 0.93 0.89 0.62

Moisture content

p (relationship)

0.0001

(inverse) 0.9222 0.7580

0.0003

(positive)

0.0149

(positive)

<0.0001

(positive)

0.0003

(positive) 0.8813

PSC1 0.8 0.2 0.3 0.8 0.3 0.2 0.6 0.3

PSC2 0.8 0.2 0.3 0.9 0.2728 0.2 0.5 0.3

PSC3 0.9 0.2 0.3 0.8 0.3 0.2 0.6 0.3

PSC4 0.8 0.2 0.3 0.8 0.3 0.2 0.6 0.3

PSC5 0.8 0.2 0.3 0.8 0.3 0.2 0.6 0.3

PSC6 0.8 0.2 0.3 0.8 0.3 0.2 0.6 0.3

BPOM 0.0006 (positive) 0.4 0.7 NA 0.6 NA NA 0.6

Benthic chlorophyll a

0.0069 (positive) 0.4 0.05 0.6 NA 0.4 0.1 0.8

TC NA 1.0 0.008

(positive)

0.0006

(positive) 0.06

0.0006

(positive) 0.6 0.1

TN 0.9604 NA 0.2 0.4 0.6 NA 0.3 NA

TP 0.0076

(positive) 0.2 NA 0.7 0.08

0.01

(inverse) NA NA

TSS 0.6977 0.06 0.2 0.1 0.9 0.3 0.2 0.04

(positive)

Suspended chlorophyll a

0.6 0.6 0.7 0.05 0.9 0.2 0.9 0.7

ZEU:Z NA NA NA NA 0.7170 NA NA NA


0

40

80

120

160

0 20 40 60 80 100

Moisture content (% weight)

TC

(g

/m2)

Figure 25. Relationship between moisture content of sediment and total carbon (TC). Note that the relationship was not significant for sediments with greater or less than 50% moisture content.

0

40

80

120

160

0 60 120 180 240

BPOM (g/m2)

TC

(g

/m2)

>50%Moisture

<50% Moist

Log. (>50%Moisture)

Expon.(<50%Moist)

Figure 26. Relationship between benthic particulate organic matter (BPOM) and total carbon (TC) for sediments with greater than and less than 50% moisture content. For regression of sediments less than 50% moisture content p = 0.02, r

2 = 0.26 with natural log transformation of BPOM. For

regression of sediments greater than 50% moisture content p = 0.0003, r2 = 0.57 with natural log

transformation of TC. For both analyses values of CC-3 were removed.


3.4. Discussion

River regulation and the extraction of water for human-use in the River Murray have had a clear impact on the distribution of sediments within the Lower Lakes. The maintenance of elevated water levels within the Lower Lakes has caused an increase in sedimentation rates within the Lower Lakes (Barnett 1993) since European settlement as a result of increased erosion of the lake-shore (Coulter 1992). This has resulted in the retreat of lake perimeter at an average rate of 1 m/year (Coulter 1992). Based on deposition rates of 3 mm/year (Herzeg et al. 2001), the top 10 mm of sediment studied in this investigation is likely to represent sediments deposited in the 3-4 years prior to sampling. These sediments were dominated by particle sizes in the range of 23.5-462 µm, making them silt-fine sands. The dominance of intermediate sized particles deposited during this period is likely to be a result of a number of processes associated with river regulation. Firstly, the presence of over-flow weirs on the River Murray has resulted in increased sedimentation of coarser sediment upstream of weirs (Thoms and Walker 1993). Similarly, reduced peak flows due to water extractions and regulation in the River Murray may increase deposition rates of larger particles along the River Murray. Furthermore, the construction of barrages at the downstream end of the Lower Lakes would prevent the deposition of coarser marine sediments in the Lower Lakes during periods of marine intrusions.

The Lower Lakes are prone to resuspension due the their shallow water depth and large fetch (Geddes 1984a). Consequently, in the absence of large currents in the Lower Lakes, associated with reduced river flows and barrage construction, the dominant source of energy in the lakes are winds: wind-driven waves will be the dominant source of resuspension and the wind-driven currents will move suspended material around the lake basin rather than current flow. The relatively constant water levels maintained in the Lower Lakes means that large wind driven events continuously resuspend fine sediments from shallow regions. Subsequently, these resuspended sediments are deposited in areas of low energy, including areas of deeper water (Hilton et al. 1986; Matty et al. 1987; Lebo and Reuter 1995). This is a common phenomenon in lakes known as sediment focussing, resulting in sediments with coarser particles tending to shallow regions and sediments with finer particles tending to deeper regions, as was observed in this study. Since the Lower Lakes have a gently sloping lake bed, there existed a transition in sediment character according to water depth rather than the formation of distinct groups of different sediment types.

The transition in particle size with water depth was coupled a transition of increasing moisture content and BPOM with water depth and decreasing particle size. The correlation between parameters describing their character is common in such studies of sediment distribution in lakes (Kumke et al. 2005), making it difficult to determine the basis of observed patterns. For example, the greater BPOM in deeper areas than in shallow areas is likely to be function of both depth and particle size: enhanced rates of organic matter decomposition in shallow regions due to enhanced sediment resuspension; and focussing of fine (recalcitrant) organic material in deeper low energy regions due to low specific gravity (Jones and Simon 1981; Bloesch 1995). However, as with several sediment characteristics, BPOM was surprisingly not related to particle size. Instead, BPOM was strongly related to moisture content, reflecting differences in sediment porosity, which is a function of both particle size and particle sorting. Overall, TC did not follow the same pattern as BPOM, which appeared to be due to sediments with >50% moisture content containing a large component of inorganic carbon. For sediment with <50% moisture content, TC was approximately 50% of BPOM, suggesting that organic carbon was a major component of TC with the carbon content of aquatic plants being approximately 40% of dry-weight (Demars and Edwards 2007).

Unlike other parameters, TN was more closely related to properties of suspended solids than benthic sediments with a general increase in TN with decreasing water depth and increasing water turbidity. These parameters are likely to be auto-correlated due to shallow sites being more prone to bank erosion and resuspension of benthic sediments (Hilton et al. 1986; Bloesch 1995). Although TN was found to be highest in Lake Albert, this is likely to be a result of all sampling locations being in shallow water as a result of access problems. A possible explanation for low TN in deeper sediments is greater rates of denitrification in anoxic


sediments, thus lowering nitrogen concentrations. This would also explain the low TN to TP molar ratios in deep water, with TP showing little variation across the whole study area. However, TN in depth class 2 was found to be relatively low and was not related to differences in BPOM, perhaps indicating depths of tightly coupled nitrification-denitrification within the sediment profile, thus efficient loss of nitrogen. Others have found low C:N in shallow, high energy areas due to low organic content, rapid degradation of labile organic material and reduced productivity (McLusky 1989). Indeed in this study BPOM:TN increased with water depth.

It is evident that the water depth plays an important role in determining sediment character in the Lower Lakes. However, the interaction between the sediment and the water column is also likely to have implications for higher trophic organisms, by influencing light availability. While Geddes (1984a) proposed that resuspension in the Lower Lakes is an important process in the productivity of phytoplankton, it is also likely to control the productivity of periphytic algal and macrophyte communities. Across the Lower Lakes there were no differences detected in chlorophyll a concentrations between transects and depth classes. This may have been a result of increased deposition of pelagic chlorophyll in deeper sties (Ostrovsky and Yacobi 1999) despite shallow sites containing higher abundances of benthic algal populations. Indeed, benthic chlorophyll a was not related to ZEU:Z, despite average values of >1, indicating light availability at the sediment surface. However, it is likely that the high level disturbance to the sediment surface in the Lower Lakes due to sediment resuspension prevents the establishment of an extensive benthic algal community.

Overall, this study has found that there existed a transition in sediment character according to water depth in the Lower Lakes. This ultimately results in reduced habitat complexity as a result of the deficiency of current energy from river flows; and the maintenance of relatively constant water levels (Kumke et al. 2005). It was interesting to note that sediments that displayed the greatest variance in character were those at CC, a site located adjacent to a small inflowing stream, Currency Creek. This site would experience seasonal water level fluctuations and inputs of current energy, demonstrating the importance of inflows in creating heterogeneity in sediment distribution.

The strong association of sediment character with water depth as a result of sediment focussing means that falling water levels are likely to have implications for the behaviour of sediment particles, and subsequently biogeochemical processes. As water levels fall, finer sediments with lower critical sheer stress will become more exposed to wave energy, presumably resulting in increased suspended solids loads, as was observed in Chapter 2. This will have implications for primary productivity and biogeochemistry. It is likely that increased suspended solids will reduce primary productivity, including aquatic plants within the system, thus reducing habitat and organic matter resources for higher organisms. However, these may be counteracted somewhat by the salinity intrusions into the lakes, inducing flocculation and settling of very fine particles (<200 μm) (Grace et al. 1997).


4. Nutrient flux from permanently inundated and dried-reflooded sediments of Lake Alexandrina undergoing rapid water level drawdown

Kane T Aldridge

Andrew Bissett

Justin D Brookes

4.1. Introduction

Nutrients are fundamental components of river and lake ecosystems that undergo continuous transformation as they pass downstream (Newbold et al. 1981). Although cycling of nutrients has been studied in great detail, much of this information has been developed in temperate regions, which are not subject to extreme changes in water level. Inland water ecosystems in regions of mediterranean and arid climates are subject to changes in water level on a range of spatial and temporal scales, owing largely to variability in rainfall (Gasith and Resh 1999). These changes in water level can be compounded by over extraction of water for human purposes.

Water level drawdown results in the drying of previously inundated sediments, which contain a major portion of nutrients stored within inland water ecosystems (Marsden 1989; Martinova 1993; Søndergaard et al. 1993). Drying of sediments will result in changes to the physical, chemical and biological character of the sediments to which nutrient cycling is intrinsically linked (Baldwin and Mitchell 2000). Consequently, the cycling of nutrients following drying and reflooding cycles is likely to be different than under permanently inundated conditions. Despite this, only few have investigated the role of drying-reflooding cycles on nutrient cycling in aquatic ecosystems.

Baldwin and Mitchell (2000) have provided a comprehensive review of the likely effects of drying-reflooding on nutrient dynamics of inland water ecosystems. The major effects of drying of previously inundated sediments are to the mineralogy and microbial ecology, both of which respond to increasing oxygen concentrations within the sediment (Baldwin and Mitchell 2000). As water levels fall, the oxygen penetration depth within sediments will increase. This is initially due to increased delivery of oxygen rich water to the sediments. As sediments begin to dry oxygen penetration will continue to increase due to contact with the atmosphere and finally through the development of cracks accompanying the loss of moisture. The increasing oxygen concentrations will influence the mineralogy by causing the oxidation of reduced mineral phases (Baldwin and Mitchell 2000). The most notable change that influences nutrient cycling is the oxidation of ferrous sulfides into amorphous ferric oxyhydroxides which have a high affinity for phosphorus (De Groot and Van Wijck 1993). This is thought to be responsible for the increase in affinity for phosphorus of aerated sediments (De Groot and Fabre 1993; Baldwin 1996). However, as drying continues, the affinity of sediments for phosphorus has been shown to be reduce due to oxyhydroxides becoming crystalline (Lijklema 1980), reducing the number of phosphorus binding sites (Sah et al. 1989; Qiu and McComb 1994; Baldwin 1996).

The penetration of oxygen into inundated sediments during drying will also influence nutrient cycling by altering the microbial composition and activity (Baldwin and Mitchell 2000). Initially, the expansion of the oxygen penetration depth may allow aerobic and anaerobic nutrient cycling


processes to occur concurrently (Baldwin and Mitchell 2000). These process rates may be enhanced due to the coupling of processes such as nitrification and denitrification, with denitrification rates enhanced by the presence of nitrification through the provision of nitrate (Knowles 1982). As oxidation continues however, obligate anaerobic heterotrophs will be killed or form resting stages (Lynch and Hobbie 1988), thus reducing the rate of the processes that they carry out. Furthermore, the loss of moisture from the soil as sediments continue to dry will result in a further decrease in bacterial biomass and activity (De Groot and Van Wijck 1993) and extreme drying will result in high bacterial mortality and cell lysis (West et al. 1988; Qiu and McComb 1994).

Upon re-inundation lysed cells may leach nutrients resulting in the flux of large amounts of nutrients into the porewater and overlying water column (Sah et al. 1989; Mitchell and Baldwin 1999; Baldwin and Mitchell 2000; Baldwin et al. 2005). This is thought to result in increased rates of microbial activity, including rates of nitrification and denitrification (Baldwin and Mitchell 2000) if microbial communities have survived the desiccation processes. Denitrifying bacteria are predominately facultative anaerobes (Knowles 1982) and so drying-reflooding cycles have been shown not to reduce denitrification rates (Kern et al. 1996). However, if obligate anaerobes are dominant then rates of microbial processes will presumably be lower upon reinundation than prior to drying.

Previous studies of the influence of drying-reflooding cycles on nutrient fluxes from sediments have focussed on aquatic ecosystems that undergo regular periods of drying-reflooding. However, the accumulation of nutrients and the development of microbial communities are likely to be different under permanently inundated conditions than in response to drying-reflooding. Consequently, nutrient fluxes are also likely to differ. The Lower Lakes of the River Murray (Lake Alexandrina and Lake Albert), South Australia are permanently inundated, but have undergone rapid water level drawdown to extreme drought conditions (MDBC 2008). This has exposed vast areas of sediment to the atmosphere, resulting in drying of the sediments. The aim of this project was to investigate the impacts of reflooding dried sediments from Lake Alexandrina on microbial activity and nutrient fluxes. This is used to gain an understanding of the impacts of the re-inundation of exposed sediments on nutrient cycling within Lake Alexandrina. In doing so the response of permanently inundated sediments to water bodies that undergo regular drying-reflooding cycles is made.

4.2. Methods

4.2.1. Core collection

On 19th March 2007 sediment cores were collected from two sites in Lake Alexandrina (Figure 27 and Figure 28). The two sites chosen for this study represent a deep site located close to the inlet of the River Murray (Site 1) and a shallow site located near Tauwitchere Barrage, the largest of the five barrages (Site 2, Figure 27). Site 1 sediments were overlain with a water depth of 3.3 m at the time of sampling and Site 2 with 0.9 m. In Lake Alexandrina, deep areas are characterised by fine sediments, high moisture content and high particulate organic matter, while shallow areas are characterised by coarser sediments, low moisture content and high particulate organic matter (see chapter 3).

The sediment was collected in Perspex cores with an internal diameter of 5.7 cm. Unsealed cores were attached to a pole and gently pushed into the sediment perpendicular to the sediment surface. The cores were pushed in 10-15 cm into the sediment, sealed and gently extracted from the sediment. The core was retrieved and bottom sediment was disposed of leaving the top 5 cm. A rubber stopper was pushed into the base of the core to prevent sediment from falling from the core. The seal was broken and overlying water was siphoned off, except for approximately 1 cm to maintain the moisture content of the sediment. Cores were immediately stored in the dark, below 3°C.


Figure 27. Location of sampling sites within Lake Alexandrina.

On return to the laboratory sediment cores were pooled according to site and mixed to reduce the level of heterogeneity that is observed with sediment character (Figure 28) (Whitmore et al. 1996). Thirty sediment cores (5 cm of sediment) of both sediment types were then collected in Perspex cores with internal diameters of 3.5 cm. The bases of the cores were sealed and cores were placed in an aquarium and overlain carefully with artificial lake water (Figure 28). The composition of this water was based on historical water quality within the River Murray (Williams and Buckney 1976; Cook et al. 2008), the likely source of water for re-inundation of the Lake Alexandrina (Table 7). Aquarium water was continually mixed and aerated and the aquarium was stored in the dark at 20°C for a 3 week equilibration period, during which artificial lake water was changed regularly.

4.2.2. Drying

Fifteen cores of both sediment types were removed from the aquarium. Overlying water was siphoned off and sediments were then placed at 40°C until a constant weight was obtained (Figure 28). The aim of the drying processes was to reflect the loss of moisture that would occur under an extended drought period in southern Australia, which would result in total loss of moisture content. Under in situ conditions it is likely that the moisture content of the sediments would periodically increase by rainfall prior to reinundation from increased water levels. However, for the purposes of this experiment the assumption was made that this periodic wetting process would not significantly influence the nutrient fluxes upon permanent re-inundation.


Figure 28. Flow diagram of sediment treatment prior to experimental incubation.

Table 7. Composition of experimental water used for nutrient flux experiments.

Element Concentration

(mg/L) Source

NO3 1.2 Twice values of average River Murray concentrations, 1997-

2007 (Cook et al. 2008)*

PO4 0.2 Average River Murray concentrations, 1997-2007 (Cook et al.

2008)

SO4 10

HCO3+CO3 89 Williams and Buckney (1976)

Na 62 Williams and Buckney (1976)

Ca 18 Williams and Buckney (1976)

Mg 16 Williams and Buckney (1976)

K 6 Williams and Buckney (1976)

Fe 0.07 Average River Murray concentrations, 1997-2007 (Cook et al.

2008)

Mn 0.004 Williams and Buckney (1976)

Zn 0.004 Williams and Buckney (1976)

Cl 121 Williams and Buckney (1976)

*Concentrations of River Murray water were found to be low and so concentrations were doubled to ensure denitrification rates could be detected.

Sediments pooled and homogenised (sites separate)

30 sediment cores collected for both sites and overlain with artificial lake water

Collection of sediments from Site 1 (deep) and Site 2 (shallow)

15 sediment cores from both sites dried at 40°C to a

constant weight (“Reflooded”)

15 sediments cores from both sites permanently inundated

with artificial lake water (“Wet”)

Incubation with experimental water (nutrient flux experiment)


4.2.3. Nutrient flux experiments

For the purposes of calculating denitrification rates experimental water was made up as previously described, with NO3 consisting of 15NO3. Immediately prior to the experiment three samples were removed from experimental water for analysis of phosphate (PO4-P), ammonium (NH4-N), nitrite (NO2-N) and nitrate (NO3-N) concentrations. Samples were filtered through 0.22 μm syringe filters and immediately frozen prior to analysis by the Max Planck Institute Laboratory, Bremen, Germany.

Overlying water on permanently wet sediment cores was siphoned off and both wet and dried sediments were overlain with experimental water. In addition, three cores containing no sediment (blanks) were filled with experimental water. Sediments were purged with experimental water by forcing water from a syringe connected to the top of the core and out of a valve on the base of the core. A membrane on the base prevented the loss of sediment during the purging process. Floatable magnetic stirrers were placed inside the cores, which were then sealed and placed around rotating magnets within an aquarium. The rotation of the magnetic stirrers effectively mixed the water columns of individual cores.

At 30 s, 0.5 hr, 2.0 hr, 5.0 hr, 10.0 hr after inundation with experimental water, 3 cores of each sediment type removed from the aquarium for extraction of water samples. For blanks, samples were only collected at 0 and 10 hr. The caps were removed and 1 ml of ZnCl2 was added to the overlying water (Dalsgaard et al. 2000). The cap was placed back on the core was turned to ensure that labelled N2 produced was homogenously distributed (Dalsgaard et al. 2000). Care was taken to prevent the exchange with the atmosphere. Sediments were allowed to settle and water samples were collected with a syringe for analysis of PO4-P, NH4-N, NO2-N and NO3-N concentrations, as described earlier. Flux rates were calculated as the slope of the linear regression between nutrient mass and time (Dalsgaard et al. 2000). All rates were corrected for the volume of overlying water within the core during the experimental period and the surface area of the sediment. Since the cores were mixed prior to collecting the water samples flux rates represent total flux from sediments to both the porewater and overlying water. However, calculated flux rates are reported as flux to and from the water column and assume equilibration of nutrient concentrations between the two. The immediate flux of nutrients from the sediment to the water upon addition of experimental water was also calculated since there is likely to be large immediate flux to or from the water column. This was done by calculating the difference in nutrient concentrations of experimental water and in cores at 30 s.

Additional samples were collected from cores for calculation of denitrification rates. These samples were extracted with a syringe and carefully injected into a vial containing HgCl2. Sample vials were filled with no head space and stored in the dark at room temperature for later analysis of 29N2 and 30N2 concentrations. The isotopic composition of N2 of each sample was analysed using a mass spectrometer following the methods of Dalsgaard et al. (2000). In this procedure, samples were extracted from the water by introducing a helium headspace. The vial was shaken and samples extracted for analysis of 29N2 and 30N2 concentrations. Denitrification rates were calculated through methods of Dalsgaard et al. (2000), whereby the production of 29N2 and 30N2 per unit time was calculated as the slope of the linear regression between of the mass of labelled N2 and time. All rates were corrected for the volume of overlying water within the core and the surface area of the sediment.

4.2.4. Sediment character

At the time of sediment collection, three wet sub-samples of each sediment type were collected and their particle size distribution was determined by suspending sediment in deionised water and measuring particle size of the samples using a Laser In-Situ Scattering and Transmissometry instrument (LISST-100X, Sequoia scientific). This instrument measures scattering of light at 32 angles and this information is used to calculate the size distribution of 32 size classes log-spaced between 2.73 and 462 μm. For the purposes of this study size classes were divided into six categories of different particle size class (PSC): <4.48 μm (PSC1), 4.48-


6.24 μm (PSC2), 7.36-19.9 μm (PSC3), 23.5-63.3 μm (PSC4), 74.7-186 μm (PSC5), 219-462 μm (PSC6). Data is presented as percent composition of each particle size class by volume.

Following the extraction of water samples in the flux experiment remaining water was siphoned and sediments samples were collected at 0 and 10 hr. The sediment was dried to a constant weight at 105°C and dry-weight was recorded, with moisture content measured as the difference between fresh-weights and dry-weights (expressed as a percentage of fresh-weight). Sediments were ground with a mortar and pestle and analysed for particulate organic matter (POM), total carbon (TC), total nitrogen (TN) and total phosphorus (TP). POM was determined by igniting dried sediment to constant weight at 550°C to determine fixed and volatile solids following APHA method 2540-E (Eaton et al. 2005). TC and TN were analysed using LECO TruSpec Carbon/Hydrogen/Nitrogen Determinator (LECO, St. Joseph, USA). TP was analysed using the persulfate digestion method of standard method 4500-N C (Eaton et al. 1995).

4.2.5. Statistical analyses

All statistical analyses were performed using JMP-IN®. Differences in nutrient flux and denitrification rates were tested using a two-way analysis of variance with sediment and treatment (wet/reflooded) as fixed effects. Differences in sediment character (nutrients, POM) between times were tested using t-tests.

4.3. Results

The sediment character of the two sites was found to be different, with sediments of Site 1 having a greater portion of finer sediments and greater moisture content than Site 2 (Table 8). Site 1 had greater POM, TN, TC and TP concentrations than Site 2 (Table 9, p = <0.0001 for all analyses with comparison at 0 hr). There was no significant effect of the experimental treatment on POM, TN, TC or TP, with no differences found between both treatments at the two sites (p = >0.05 for all analyses).

Table 8. Moisture content and sediment particle size distribution of sampling locations in Lake Alexandrina.

Time

(hr)

Moisture content

(%)

PSC1

(%)

PSC2

(%)

PSC3

(%)

PSC4

(%)

PSC5

(%)

PSC5

(%)

Site 1 67.7 ± 7.91 14.6 ± 7.4 3.7 ± 0.4 15.3 ± 0.8 18.0 ± 1.2 21.4 ± 2.5 27.1 ± 4.3

Site 2 39.6 ± 3.4 6.6 ± 2.9 2.0 ± 0.24 7.6 ± 1.1 6.7 ± 0.6 40.4 ± 3.3 36.7 ± 2.2

However, despite no changes in nutrient concentrations of the sediments before and after the treatment, there were a number of differences in the flux rate of nutrients from and to the water column. This flux began immediately upon inundation with experimental water, with a large immediate flux of NH4-N to the water column (Table 10). There were no differences in the immediate flux between sites for any nutrient forms. However, reflooded treatments had a greater immediate flux rate for NH4-N and permanently wet treatments had greater immediate flux rates for NO2-N and NO3-N (Table 10 and Table 11). For NO2-N there was a net flux to the water column in the permanently wet treatments, but a net flux from the water column in reflooded treatments (Table 10). For NO3-N, there was also a net flux from the water column in reflooded treatments, whereas there was not apparent net change in nutrient concentrations in the permanently wet treatments (Table 10).


Table 9. Sediment character at 0 and 10 hr in permanently wet and dried-reflooded treatments of Site 1 and 2.

Treatment Time (hr) POM (mg/g) TN (mg/g) TC (mg/g) TP (mg/g)

Site 1 wet 0 44.6 ± 1.7 1.8 ± 0.2 29.9 ± 1.2 0.27 ± 0.04

10 47.6 ± 2.9 2.0 ± 0.2 31.4 ± 0.9 0.26 ± 0.01

Site 1 reflooded

0 44.6 ± 5.9 2.3 ± 0.1 30.3 ± 0.7 0.27 ± 0.05

10 47.8 ± 2.9 2.3 ± 0.1 30.8 ± 0.7 0.31 ± 0.03

Site 2 wet 0 6.5 ± 0.2 0.9 ± 0.1 21.9 ± 0.1 0.04 ± 0.00

10 6.4 ± 0.9 0.7 ± 0.2 21.9± 0.1 0.06 ± 0.01

Site 2 reflooded

0 6.0 ± 0.7 0.8 ± 0.2 21.9 ± 0.1 0.10 ± 0.04

10 6.3 ± 0.8 0.7 ± 0.2 21.8 ± 0.1 0.06 ± 0.01

Table 10. Immediate flux of nutrients in permanently wet and dried-reflooded treatments of Site 1 and 2. Positive values denote flux from sediment to water column, negative values denote flux from water column to sediments.

Site Treatment PO4-P (g/m2/hr) NH4-N (g/m

2/hr) NO2-N (g/m

2/hr) NO3-N (g/m

2/hr)

Site 1 Wet -0.5 ± 0.0 425.7 ± 120.0 0.3 ± 0.0 0.1 ± 0.4

Reflooded -0.5 ± 0.4 262.4 ± 190.9 -0.2 ± 0.1 -3.1 ± 1.6

Site 2 Wet -0.5 ± 1.4 73.3 ± 110.9 -0.2 ± 0.5 -3.2 ± 3.9

Reflooded 0.7 ± 2.1 442.0 ± 51.5 0.2 ± 0.1 -0.5 ± 0.4

In comparison, flux rates of nutrients during the subsequent 10 hour experimental period were found to differ considerably between sites and treatments. At both sites and for both treatments, PO4-P was consistently lost from the water column over the 10 hour period except for the reflooded treatment of Site 1, where there was no apparent flux (Figure 29A). The flux of PO4-P from the water column was inconsistent between sites and treatments and so there was no overall effect of site or treatment (Table 11). For Site 1, permanently wet treatments appeared to have higher flux from the water column, while for Site 2 reflooded treatments had higher flux rates from the water column (Figure 29A). Reflooded treatments of Site 2 had greatest flux rates of 1.4 ± 0.8 mg/m2/hr from the water column.

Differences in NO3-N flux between sites and treatments were similar to that of PO4-P, with a positive linear relationship between NO3-N and PO4-P flux (Figure 29, p = 0.01, r2 = 0.51). However, for NO3-N, both the permanently wet treatment of Site 2 and reflooded treatment of Site 1 had no apparent net flux (Figure 29B). Overall, there was greater flux from the water column in reflooded than permanently wet treatments and greater flux in Site 2 than Site 1 (Figure 29B, Table 11). Much of this difference was due to the large flux rates of the reflooded treatments of Site 2, which had flux rates of 6.5 ± 3.1 mg/m2/hr. The flux of NO2-N from and to the water column was much smaller than that of NO3-N (Figure 29B-C). In comparison to NO3-N, NO2-N concentrations in the water column generally increased during the experimental period for both sites in permanently wet treatments. However, there was not net flux of NO2-N to the water column of both sites in reflooded sediments (Figure 29C). While the permanently wet treatment had greater flux of NO2-N at Site 1, there appeared to be little difference between treatments at Site 2 (Figure 29C) and so the effect of treatment was inconsistent between the two sites (Table 11).


Table 11. P-values for the effect of site and treatment on nutrient flux rates.

Flux period Nutrient flux

rate Site Treatment Site*Treatment

Immediate

PO4-P 0.5 0.5 0.5

NH4-N 0.3 0.007 0.2

NO2-N 0.7 0.02 0.5

NO3-N 0.8 0.04 0.9

10 hour

PO4-P 0.05 0.3 0.04

NH4-N 0.0005 0.008 0.9

NO2-N 0.6 0.02 0.3

NO3-N 0.1 0.34 0.002

Denitrification 0.3 0.0002 0.3

As for the immediate flux of nutrients, greatest flux rates of the studied nutrient forms were observed for NH4-N. In permanently wet sediments, NH4-N concentrations in the water column increased at a rate of 247.1 ± 94.4 mg/m2/hr for Site 1, while for Site 2 NH4-N concentration decreased at a rate of 139.0 ± 91.1 mg/m2/hr (Figure 29D). For both sites, the flux of NH4-N to the water column was greater following drying-reflooding than for permanently wet sediments, with greatest rates observed in the reflooded sediments of Site 1 (Figure 29D). Consequently, greater NH4-N flux to the water column was observed in Site 1 than Site 2 and in reflooded treatments than permanently wet sediments (Table 11).

Permanently wet sediments were found to have denitrification rates of 0.2 ± 0.1 mg/m2/hr at Site 1 and 0.3 ± 0.1 mg/m2/hr at Site 2 (Figure 30). However, following drying-reflooding denitrification was reduced to 0.00 ± 0.00 mg/m2/hr and 0.00 ± 0.01 mg/m2/hr for Sites 1 and 2, respectively. Consequently, there was an effect of treatment and this effect was consistent across sites (Table 11).


-4

-3

-2

-1

0

1

Wet Reflood Wet Reflood

Site 1 Site 2

PO

4-P

Flu

x (

mg

/m2/h

r)

-12

-8

-4

0

4


Site 1 Site 2

NO

3-N

Flu

x (

mg

/m2/h

r)

-0.4

0.4

1.2

2.0


Site 1 Site 2

NO

2-N

Flu

x (

mg

/m2/h

r)

-250

0

250

500

750


Site 1 Site 2

NH

4-N

Flu

x (

mg

/m2/h

r)

Figure 29. Flux of PO4-P (A), NO3-N (B), NO2-N (C) and NH4-N (D) in permanently wet and dried-reflooded treatments of Site 1 and 2. Positive values denote flux from sediment to water column, negative values denote flux from water column to sediments.

A) B)

C) D)


-0.10

0.10

0.30

0.50


Site 1 Site 2

De

nit

rifi

cati

on

rate

(m

g N

/m2/h

r)

Figure 30. Denitrification rate in permanently wet and dried-reflooded treatments of Site 1 and 2.

4.4. Discussion

The observed flux of NH4 from sediments of Lake Alexandrina was far greater than those observed elsewhere (Baldwin et al. 2005). This flux was enhanced by a drying-reflooding cycle, most likely due to cell lysis, with nutrients leached from cells upon re-inundation (Baldwin and Mitchell 2000). A possible explanation of the extremely high NH4 flux is that since flow regulation, Lake Alexandrina has essentially acted as a terminal system to the River Murray. Based on sediment deposition rates of 3 mm/year in Lake Alexandrina (Herzeg et al. 2001), the top 5 cm of sediment collected represents sediments deposited in the previous 15-20 years, a period of permanent inundation, small water level fluctuation and few outputs downstream. Those flows that have occurred have been managed releases of water rather than flow-through events generated from episodic inputs from the River Murray. Consequently, during this period, Lake Alexandrina is likely to have acted as a deposition site of nutrients and sediments of the River Murray.

This is supported for the larger flux of NH4 to the water column from Site 1, which is located close to the inlet of the River Murray and contained greater nutrient and POM contents than Site 2. Upon entering Lake Alexandrina, flow velocity from the River Murray would dissipate owing to the wide nature of Lake Alexandrina, thus resulting in the deposition or suspended material (Hilton et al. 1986; Kumke et al. 2005), including organic material. Another explanation is the differences in sediment types, with the finer sediments of Site 1 developing large cracks following the drying process owing to the high moisture content. Consequently upon reinundation there was unrestricted exchange between porewater and the overlying water column.

Site 1 is also a deeper site of Lake Alexandrina and it is likely that oxygen penetration into sediments is limited, thus limiting the development of a large denitrifying community and resulting in the accumulation of organic material and NH4. Indeed, even under permanently wet conditions, flux rates of NH4 from Site 1 to the water column were high. In comparison, NH4 concentrations in the water column of the permanently wet Site 2 treatment decreased. Since


NO3 concentrations in the water column were also found to decrease, it suggests that the NH4 was lost to coupled nitrification-denitrification in Site 2.

Significant denitrification was found to be present in both sediments under permanent inundation. The measured rates of approximately 0.3 mgN/m2/hr are on the higher end of values measured in lakes across the world (Seitzinger et al. 2006). However, it is likely that under ambient NO3 concentrations, denitrification rates would be lower due to NO3 limitation, as was observed for a reservoir on the River Murray (Mitchell and Baldwin 1999). It was clear that the drying-reflooding process of this study resulted in a complete reduction of denitrification. The likely explanation is death of the microbial community, which is possibly an obligate anaerobic community due to the permanent inundation of these sediments. Elsewhere, denitrification rates of floodplain soils and reservoir sediments have been shown not the be affected by drying-reflooding (Kern et al. 1996), since denitrifying bacteria are predominately facultative anaerobes (Knowles 1982). Given this, an alternative explanation is the reduction in NO3 concentrations in the water column limiting denitrification. Although nitrate concentrations decreased there still remained significant concentrations after the 10 hr incubation. Alternatively, the increased oxygen penetration into the sediments, due to the exposure to the atmosphere during the drying process, may have prevented denitrification during the 10 hour incubation period.

With the apparent cessation of denitrification following drying-reflooding and desiccation of organic material, NO3 concentrations would be expected to accumulate within the water column, as observed elsewhere (Qiu and McComb 1994; Qiu and McComb 1996; Baldwin and Mitchell 2000). However, this was not the case in this study, with either no net change or enhanced loss of NO3 from water column observed. Since there were high NH4 concentrations, this suggests that negligible nitrification was occurring following drying-reflooding, which has also been observed for a River Murray Reservoir (Mitchell and Baldwin 1999). In fact, on the addition of NO3 into the overlying water column, Mitchell and Baldwin (1999) observed all NO3 was lost from the water column, but only 50-75% was lost through denitrification. Other possible pathways identified included absorption to sediment, assimilation into the sedimentary micro-biota, conversion to NH4 by dissimilarity nitrate reduction or abiotic reduction of NO3 facilitated by manganese or iron-sulfides (Mitchell and Baldwin 1999). It is possible that the mechanism responsible for this is associated with the flux of PO4 from the water column since the two were related. The loss of PO4 from the water column following drying-reflooding is likely to be due to the increased affinity of sediments for phosphorus due to the oxidation of ferrous sulfides into amorphous ferric oxyhydroxides which have a high affinity for phosphorus (De Groot and Fabre 1993; De Groot and Van Wijck 1993; Baldwin 1996). As drying continues through time, the affinity of sediments for phosphorus has been shown to be reduced due to aging of oxyhydroxides, becoming more crystalline (Lijklema 1980) and reducing the number of binding sites and affinity for phosphorus (Sah et al. 1989; Qiu and McComb 1994; Baldwin 1996).

The contrasting findings of this and other studies demonstrate that the impact of drying-reflooding on biogeochemical cycling of nutrients varies considerably. It is likely that the extent of drying, the microbial community present and the mineralogy of the sediment play important roles in determining the different responses. The 10 hour incubations conducted in this study reflect the initial flux of nutrients upon re-inundation. It was found that drying-reflooding enhanced the loss of NO3 and PO4 from the water column, particularly in coarser sediments with low nutrient content. However, in the site with finer sediments and high nutrient and POM contents, there was an extremely large flux of NH4, particularly following a drying-reflooding cycle. It is likely that the observed results would be quite different if studied over longer periods. For example, although it appeared that there was negligible nitrification following drying-reflooding in this study, Qui and McComb (1996) found that there was a four day lag in the recovery of nitrifying organisms. Many of the observed results therefore appear to be a result of the flux of NH4 from lysed cells and abiotic adsorption of PO4 and NO3 to sediments. Through time a recovery of the microbial community would be expected, which would have implications for the net flux of nutrients between the sediment and water column.

This study also reflects the flux of nutrients that would occur if sediments remained completely dry prior to reinundation. In reality, dried sediments will undergo partial wetting following rainfall and the movement of water across the dried lake bed due to wind induced seiching. The


impacts of these intermediate wetting processes are largely unknown (Baldwin and Mitchell 2000) and will depend on the extent of drying and inundation. It is however clear, that as levels in Lake Alexandrina are increased the subsequent reinundation of exposed sediments will have major implications for the flux of nutrients between the sediment and water column.

In this study the assumption was made that there is equilibration between porewater and water column nutrient concentrations. In reality, this is not the case, with nutrient concentrations higher in pore-water than the water column. The level of equilibration between the two will be dependent upon diffusion rates across the sediment-water interface. Diffusion rates are affected by the concentration gradient, sediment characteristics, bioturbation and the level of mixing within the water column. Shallow, turbulent lakes, such as Lake Alexandrina, will tend to be well mixed resulting in high rates of diffusion and maintenance of concentration gradients. This is a result of the rapid dispersion of released nutrients due to mixing; and the exposure of anoxic sediments with high nutrient concentrations due to resuspension of surface sediments. In addition, mixing will result in mechanical mixing of pore-water and the water column. It is likely that the measured flux rates of this project are higher than what would occur in situ upon reinundation of Lake Alexandrina. However, the results of this study do reflect the total flux of nutrients from the sediment that will increase the potential pool for autotrophic and microbial organisms.

The large initial flux of NH4 would be expected to result in an increase in autotrophic or heterotrophic productivity. However, this may be offset by the increased affinity of sediments for phosphorus unless significant external inputs of phosphorus are supplied. Increased assimilation of nutrients by the heterotrophic microbial community may result in the development of anoxic conditions (Baldwin and Mitchell 2000). Similarly, an increase in the available nutrient pool may increase the standing biomass of phytoplankton since the maximum carrying capacity of phytoplankton will ultimately be limited by nutrient availability (Dillon and Rigler 1974).


5. Conclusion

The key finding of the three studies is that water level drawdown in the Lower Lakes has dramatically altered the biogeochemistry of the system. Water level drawdown has resulted in salinisation of the lakes, resulting in strong horizontal and vertical salinity gradients. While salinity can directly influence nutrient cycling, it is the generation of density stratification that could have the most important impacts on water quality. In areas close to the barrages, it appeared that the settling of dense saline water to the bottom of the water column resulted in the development of anoxic conditions and release of nutrients from sediments. This may lead to increased productivity of autotrophic or heterotrophic microbial communities.

The impact of increased nutrient concentrations on productivity within the Lower Lakes is further complicated by the increased suspended solid concentrations that were observed. This is likely to reduce primary productivity, including aquatic plants within the system, thus reducing habitat and organic matter resources for higher organisms. The increased suspended solid concentrations were thought to be a result of the accumulation of finer sediments towards the middle of the lakes as a result of sediment focussing. It was found that there is a strong relationship between sediment character and water depth in the Lower Lakes in recently deposited sediments. This predictable distribution was thought to reflect river regulation, ultimately resulting in reduced habitat diversity within the Lower Lakes.

This study has highlighted that the water level and salinity regimes are key ecosystem drivers in the Lower Lakes and that they impact on ecological processes in complex ways. Planning management interventions in this ecosystem will require that the effects on water level and salinity are properly understood. Currently there exists a lack of knowledge on the likely ecological responses to various scenarios and effort must be placed to reduce uncertainty around these responses. This will allow appropriate management decisions to be made.


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