Potential impacts of salinity and
turbidity in riverine ecosystems
Characterisation of impacts and a discussion of
regional target setting for riverine ecosystems in
Queensland
Jason Dunlop, Glenn McGregor, Nelli Horrigan
The National Action Plan for Salinity and Water Quality (NAPSWQ) is a joint Australian and Queensland
Government initiative that encourages governments and regional communities to work together to address
salinity and water quality issues in priority catchments throughout Queensland. This document has been
produced under the NAPSWQ using Australian and Queensland Government financial support.
Acknowledgments
The authors wish to thank the assistance of Vivienne McNeil and Roger Clarke. We thank Patrick Burke for
assistance with map production. Thanks also to a number of reviewers for their valuable contributions.
© The State of Queensland 2005
ISBN 1 74172 078 8
QNRM 05523
Project undertaken by:
Aquatic Ecosystem Health Unit
Water Quality and Monitoring
Natural Resource Sciences
Queensland Department of Natural Resources and Mines
Email: aeh@nrm.qld.gov.au
For copies of this publication contact:
The National Action Plan for Salinity and Water Quality
Water Quality State-level Investment project
www.regionalnrm.qld.gov.au
While all care has been taken in the preparation of this document, the views and conclusions expressed in
this document may not represent the Queensland or Australian Government views or policy. The Queensland
and Australian Governments therefore accept no liability for any decisions or actions taken on the basis of this
document.
Readers should be aware that some information might be superseded with further scientific studies and
evolving technology and industry practices.
II
Potential impacts of salinity and turbidity in riverine ecosystems
Table of contents
Executive Summary _____________________________________________________________iii
List of Tables and Figures _ ________________________________________________________ v
Background ___________________________________________________________________ vii
Chapter 1 Salinity Impacts ________________________________________________________ 1
1.1
Measurement _ ________________________________________________________ 1
1.2
Salination process_ _____________________________________________________ 1
1.3
Tolerance 1.4
1.5
1.6
1.7
1.8
__________________________________________________________ 1
1.3.1
Pre-exposure and duration of exposure ____________________________ 3
1.3.2
Osmotic regulation ____________________________________________ 4
1.3.3
Avoidance _ __________________________________________________ 5
Physical and chemical aspects ____________________________________________ 5
1.4.1
Solubility ____________________________________________________ 5
1.4.2
Composition __________________________________________________ 6
1.4.3
Contaminant interactions _______________________________________ 7
Effect on stream biota ___________________________________________________ 8
1.5.1
Bacterial communities __________________________________________ 9
1.5.2
Algae ________________________________________________________ 9
1.5.3
Macroinvertebrates ___________________________________________ 10
1.5.4
Vertebrates __________________________________________________ 11
1.5.5
Plants _ _____________________________________________________ 13
Impacts on aquatic ecosystems __________________________________________ 15
1.6.1
Ecosystem processes __________________________________________ 17
1.6.2
Stream habitat _______________________________________________ 18
Salinity in Queensland surface waters _____________________________________ 18
1.7.1
Queensland salinity zones ______________________________________ 18
1.7.2
Queensland ionic composition __________________________________ 20
Salinity sensitivity index for macroinvertebrates _ ___________________________ 29
Chapter 2 Turbidity Impacts _ ____________________________________________________ 38
2.1
Measurement _ _______________________________________________________ 38
2.2
Sediment sources _____________________________________________________ 39
2.3
Physical and chemical aspects ___________________________________________ 39
2.3.1
Fine sediment ________________________________________________ 40
2.3.2
Erosivity _ ___________________________________________________ 40
2.3.3
Contaminant interactions with fine sediment ______________________ 42
2.3.4
Light penetration and temperature _ _____________________________ 42
2.3.5
Gill flushing _ ________________________________________________ 43
III
Potential impacts of salinity and turbidity in riverine ecosystems
2.4
2.5
Effect on in-stream biota _ ______________________________________________ 43
2.4.1
Tolerance of individuals ________________________________________ 43
2.4.2
Invertebrates _ _______________________________________________ 44
2.4.3
Vertebrates __________________________________________________ 45
2.4.4
Change in species composition _ ________________________________ 45
Impacts to aquatic ecosystems _ _________________________________________ 46
2.5.1
Light limitation and primary productivity _ ________________________ 46
2.5.2
Stream habitat _______________________________________________ 47
2.5.3
Avoidance ___________________________________________________ 47
2.5.4
Food web interactions _________________________________________ 48
Chapter 3 Determining acceptable concentrations for Salinity and Turbidity ______________ 49
3.1
Existing salinity and turbidity guidelines ___________________________________ 49
3.2
Regional target setting _________________________________________________ 49
3.2.1
Applying a risk assessment approach _____________________________ 50
Discussion ____________________________________________________________________ 52
References ____________________________________________________________________ 53
IV
Potential impacts of salinity and turbidity in riverine ecosystems
List of Tables and Figures
Tables
Table 1
General salinity thresholds for freshwater biota ______________________________ 3
Table 2
Summary of acute 72-hour salinity tolerance of selected macroinvertebrates
to marine salts ________________________________________________________ 11
Table 3
Salinity tolerance of selected freshwater fish _ ______________________________ 12
Table 4
Salinity tolerance of selected aquatic plants ________________________________ 14
Table 5
Electrical conductivity percentiles for Queensland salinity zones _______________ 19
Table 6
Summary of stream water chemistry in Queensland _ ________________________ 20
Table 7
Ranges of default trigger values for conductivity (EC, salinity), turbidity and
suspended particulate matter of slightly disturbed ecosystems in south-west
Australia. _ ___________________________________________________________ 64
Table 8
Ranges of default trigger values for conductivity (EC, salinity), turbidity and
suspended particulate matter of slightly disturbed ecosystems in tropical
Australia. _ ___________________________________________________________ 64
Figures
Figure 1 Conceptual model of salinity impacts on a freshwater ecosystem ______________ 16
Figure 2 Water composition provinces for Queensland _ _____________________________ 22
Figure 3 National Action Plan for salinity and water quality, priority catchments __________ 23
Figure 4 Water types of the Mary and Burnett catchments ____________________________ 24
Figure 5 Water types of the Burdekin catchment _ __________________________________ 25
Figure 6 Water types of the Brisbane and Western catchments ________________________ 26
Figure 7 Water types of the Queensland Murray Darling catchments _ __________________ 27
Figure 8 Water types of the Fitzroy catchments _____________________________________ 28
Figure 9 Salinity index in 12 equal data groupings along increasing conductivity gradient
for edge (a) and riffle (b) habitats. Median values with boxes corresponding to
80th and 20th percentiles and horizontal bars to maximum and minimum. _ _____ 30
Figure 10 Percentage of sensitive and very tolerant taxa in 12 equal data groupings with
increasing conductivity for (a) edge habitat and (b) riffle habitat. Median values
with boxes corresponding to 80th and 20th percentiles and horizontal bars to
maximum and minimum. _______________________________________________ 32
Figure 11 Salinity index for the Mary and Burnett Catchments _ ________________________ 33
Figure 12 Salinity Index for the Burdekin Catchments _ _______________________________ 34
Figure 13 Salinity Index for the Brisbane and Western catchments ______________________ 35
Figure 14 Salinity index for the Queensland Murray Darling catchments _ ________________ 36
Figure 15 Salinity index for the Fitzroy catchments ___________________________________ 37
Figure 16 Conceptual model of turbidity impacts on aquatic ecosystems _ _______________ 41
Figure 17 Risk assessment model _________________________________________________ 51
Potential impacts of salinity and turbidity in riverine ecosystems
Executive Summary
This report reviews the current state of knowledge on the effects of salinity and turbidity
as stressors on riverine ecosystems in Queensland. The report explores their physical
and chemical properties and characterises their potential ecological impacts in aquatic
environments. The current scientific understanding of the mechanisms by which salinity and
turbidity result in impacts are discussed in the context of target setting under the National
Action Plan for Salinity and Water Quality.
Salinity and turbidity are naturally occurring and ubiquitous components of freshwater
ecosystems. However, given elevated concentrations and durations of exposure, they may
result in profound ecological impacts. The prediction of the effects of salinity and turbidity
is complicated by the fact that each can consist of markedly different components that
all contribute to a single measure of salinity or turbidity. The composition of individual
components contributing to salinity and turbidity measures can vary temporally and spatially
according to broad-scale geomorphic, geological and geographic variability. The toxicity of
salinity and turbidity to freshwater biota is likely to be affected by the different components
that contribute to measurement as well as their total concentrations. There are many other
factors that can either directly or indirectly compound their ecological impacts including the
life stage of organisms exposed, ecological interactions including predator/prey interactions,
prevailing environmental stressors including rates and frequency of flows, interactions with
other contaminants, and the effect of genetic variation over generations of exposure.
When considering the biological effects of salinity in Queensland, it is important to appreciate
the extent and type of salinity found in Queensland’s surface waters. With a focus on National
Action Plan for Salinity and Water Quality priority catchments we provide an analysis of the
surface salinity and biological patterns in Queensland. To this end we include spatial and
tabular information describing the broad salinity zones in Queensland and their percentiles,
a characterisation of surface waters according to their ionic composition, and indicate the
general sensitivity of water bugs (macroinvertebrates) at sites across Queensland derived
using a mathematical modelling (Artifical Neural Networks) approach.
The use of a risk assessment methodology to assess the impacts of salinity and turbidity
impacts on aquatic ecosystems is discussed. The risk assessment method is based on
a nationally accepted risk assessment framework. The requirements for the practical
application of such a model for assessing the risk of salinity and turbidity are discussed.
The risk assessment model requires substantial information about the likelihood of impacts
and the consequence of their effects. Although there is substantial information available
that quantifies the effect that each of these factors, particularly that of salinity, may exert on
species sensitivity, there remains gaps in the existing information. We present an approach
to the assessment of risk that utilises available information to determine acceptable
concentrations for salinity and turbidity for target setting and provide recommendations
for the information requirements for improvement of the model.
VI
Potential impacts of salinity and turbidity in riverine ecosystems
Background
It is well recognised that as Australia is the driest continent on earth, water in Australia
is a valuable and scarce resource. Therefore maintaining the health and integrity of those
finite aquatic resources is important. The National Action Plan for Salinity and Water Quality
(NAPSWQ) is developing partnerships between government and communities to enhance
the management of natural resources. A critical part of the program activities is the setting
of resource condition targets for natural resource management. Under Schedule 4 of the
NAPSWQ agreement on the framework for target setting, there is a requirement that the
integrity and diversity of aquatic and terrestrial biodiversity and ecosystems be maintained
or enhanced.
Impacts from salinity and turbidity are known to result in significant impacts to
the integrity and diversity of aquatic ecosystems. The Australian Dryland Salinity
Assessment (National Land and Water Resources Audit 2000a) estimates that in
Queensland 48 000 ha of land are currently affected by salt. This is predicted to
rise to 3.1 million ha by the year 2050. In Queensland, increases in salinity have
been recorded in parts of the Condamine catchment area, Lockyer Creek, the lower
Mary catchment area, the South Burnett catchment area, Three Moon Creek and
some tributaries in the Fitzroy catchment area (National Land and Water Resources
Audit 2000b). Whilst salinity impacts are regarded as the greatest priority for
management, sediment (a major contributor to measures of turbidity) impacts are
also a significant proplem. The National River Contaminants Program rated sediments as the
third greatest priority for waterway management in Australia (Land and Water Australia 2002).
High in-stream turbidity concentrations are a natural phenomenon in Queensland’s streams
and by world standards many are considered to be highly turbid.
Despite some views that salinity and turbidity are natural and hence not contaminants, it is
now well recognised in the scientific literature that impacts from increased concentrations
of salinity and turbidity can have profound and measurable effects on riverine ecosystems.
As they are natural components of aquatic ecosystems, and their impacts are relative to
background concentrations, it is difficult to establish the concentrations at which impacts
are likely to occur when considering the many different aquatic ecosystems in Queensland.
This is reflected in the fact that there are no concentrations of salinity and turbidity that
are acceptable or safe for all aquatic ecosystems. Although state agencies are continually
reviewing water quality guidelines in relation to new and developing approaches and
techniques, the existing guidelines for these stressors are indicative only and generally
related to geographically large areas containing different types of ecosystems. Hence there is
a need to better understand the impacts of salinity and turbidity, and to develop the capacity
to determine safe concentrations in the environment. The National Water Quality Guidelines
(ANZECC/ARMCANZ 2000) outline a general framework to determine locally relevant
guidelines for stressors. However, this framework provides only a general approach and
provides no specific guidance with respect to salinity and turbidity.
VII
Potential impacts of salinity and turbidity in riverine ecosystems
1.0
Salinity Impacts
Impacts from salinity have been identified as one of Australia’s most serious environmental
issues. In areas already affected salinity has devastated ecosystems resulting in massive loss
of habitat, biodiversity, native vegetation and water resource value (Land and Water Australia
2002). An estimated 48 000 hectares of land in Queensland is seriously affected by induced
salinity and an assessment undertaken for the Land and Water Resources Audit found that
3.1 million hectares of land could be affected by salinity by 2050 (Department of Natural
Resources and Mines 2002).
Soluble salts occur naturally in aquatic ecosystems and are a vital component of the normal
functioning of freshwater biota. They are ubiquitous in Australia’s soils and are a remnant
of geological history. Salts are also an integral part of the biochemistry of life in terrestrial
and aquatic environments though for many freshwater aquatic animals exposure to high
concentrations of salt can have toxic effects. Similarly, a lack of salt can also act as a toxicant in
saline and estuarine environments for freshwater species.
1.1 Measurement
Salinity is an indicative measure of the total concentration of cations that include
sodium, calcium, magnesium, and potassium (Na+, Ca2+, Mg2+, K+), and anions that
include sulphate, carbonate, bicarbonate, and chloride (i.e. SO42-, CO32-, HCO32-, Cl-) in
solution (ANZECC/ARMCANZ 2000). Salinity may also be expressed as Total Dissolved
Solids (TDS) or Total Soluble Salts (TSS), which refer to the residual weight of salts
after drying and filtration. Measures of TDS closely resemble those of TSS (ANZECC/
ARMCANZ 2000). Conductivity is often used as a surrogate for TSS and TDS and is
a measure of the ability of a solution to conduct an electrical current between two
points. It accurately reflects measures of TDS and TSS except at very high salinities
where the relationship between TSS and conductivity diminishes and can vary
depending on which ions are dominant (Williams and Sherwood 1994). This is due to the different
ionic conductivities of various salts. For example, if the dominant ions in solution are not sodium
chloride (Na+Cl-) then conductivity may not be a suitable surrogate measure of salinity at high
salinities (Bailey et al, 2002). The most accurate measure of salinity requires a full ionic laboratory
analysis to be performed. However, it may not be practical or feasible to perform a full ionic
analysis and in its place a measure of Electrical Conductivity (EC) is often used.
1.2 Salination process
Stream salinity is closely linked with the geology of and water movement within a catchment.
Much of Australia’s soils have a high concentration of salts with about 5% of continental
Australian soils described as saline (McTainsh and Boughton 1994). Different soil types create
differences in potential salinity hazard. Alluvium valleys often have low hydraulic gradients
that can act as an impediment to groundwater movement, hence potentially increasing upward
movement of groundwater. Salt loads can accumulate in alluvial plains in the unsaturated
zone between a shallow watertable and tree roots and is common in lithology consisting
Potential impacts of salinity and turbidity in riverine ecosystems
of weathered basalt overlaying less permeable sandstones and mudstone (McTainsh and
Boughton 1994).
Salinisation is the process whereby concentrations of dissolved salts in soil
and freshwater become “unnaturally” elevated (Williams 2001, Hart et al, 1989).
Dryland or secondary salinity is the result of changes in the water balance of
landscapes following the removal of native vegetation and its replacement
by annual crops and pastures that results in increased groundwater recharge
(National Dryland Salinity Program 2001). Runoff from soils that have undergone
secondary salinisation can be highly saline and can result in impacts to the
freshwater environment. Secondary salinisation and hence the salinity of inland
waters tends to increase gradually over time making its short-term impact subtle
and difficult to detect.
Flow regimes can influence in-stream salinity concentrations altering the exposure regimes and
hence the biological impacts. Under base flow or low flow conditions where the evaporation
potential exceeds the rainfall, in-stream salinity concentrations are likely to increase naturally.
Likewise, in-stream salinity concentrations decrease in scenarios of high rainfall and high flow
conditions. Therefore rainfall/run-off dynamics have an appreciable effect on the duration and
frequency of exposure to salinity and thus alter the biological effects of salinity increases.
The disposal of saline water into streams from salt drainage schemes designed to alleviate
the effects of large-scale irrigation and mine drainage schemes are also sources of stream
salinisation. In these circumstances, the biological impacts from stream salinisation can be
profound as disposal of saline water in freshwater systems increases stream salinity rapidly
and can sometimes occur over long time periods. Disposal of saline water can contribute
significantly to the volume of stream base flow particularly in ephemeral streams or in streams
having limited base flow. A study by Kefford and Robley (1996) found that saline water disposal
affected the following water quality parameters, total phosphate, total Kjeldahl nitrogen,
suspended solids, electrical conductivity and total discharge, and that these changes were
associated with impacts to the abundance of several macroinvertebrate taxa and the structure
of macroinvertebrate communities. There is also a level of uncertainty associated with the
prediction of likely impacts from saline water disposal as most established biological effects
are based on an ionic composition of seawater. In some instances groundwater pumped
directly into streams may have an ionic composition markedly different to that of seawater and
the presence or absence of certain ions in solution may increase or decrease the biological
impact of saline water. This issue of ionic composition is discussed further in section 1.4.2
and the potential for contaminant interactions that may co-occur with saline water disposal
schemes is discussed in section 1.4.3.
1.3 Tolerance
Salinity tolerance refers to the ability of an animal to withstand exposure to salinity for an
indefinitely long period without dying. There are many factors that can affect the ability of
Potential impacts of salinity and turbidity in riverine ecosystems
aquatic organisms to survive increases or decreases in salinity and these factors
all contribute to their tolerance. For all environmental toxicants, including salinity
and turbidity, the dosage, duration of exposure and frequency of occurrence
all contribute to their toxicity and their ecological impacts. Tolerance to salinity
is in part due to the physiological mechanisms and morphological adaptations
that act to balance concentrations of salts in the cells and tissue of an organism
against the external environment. In this way, salinity tolerance can vary between
species, populations and to some extent can vary between individuals of the same
population over time. Salinity tolerance may also be due to environmental factors
that affect the duration of exposure and rate of increase in salinity concentrations.
Some indicative tolerance ranges are shown in Table 1. This table highlights the range of
different salinity tolerances observed between taxonomic groups.
Table 1 General salinity thresholds for freshwater biota
Taxa
Threshold (µS cm-1)
Effect
Small, multicellular organisms (e.g. hydra,
leeches, flatworms)
Not tolerant to elevated salinity
levels
Lethal effects
Macroinvertebrates without impermeable
exoskeletons (e.g. pulmonate gastropods)
Not tolerant to elevated salinity
levels
Lethal effects
Microinvertebrates
3000
Lethal effects
Majority of macroinvertebrates
3000
Adverse effects
Adult fish
13000
Most are tolerant up to this
level
Juvenile fish: pre-hardened eggs
3000–6600
Adverse effects
Juvenile fish: growth rate, survivorship, sperm
motility
6600–7300
Optimal between these
values
Most submerged macrophytes
1500–3000
Sublethal effects, lethal
effects for some species
Source: (James et al, 2003)
1.3.1 Pre-exposure and duration of exposure
If a population has been exposed to elevated salinity for extended periods it is possible that
they may have evolved greater tolerance than would another population of the same species
that had been exposed to a much lower concentration of salinity. Little is known about
the effect of pre-exposure to salinity on freshwater biota and its potential to affect salinity
tolerance. In addition there is little information available about the effect of pulsed increases in
salinity. The effect of pulsed exposure to salinity is particularly pertinent to saline water disposal
schemes and to first flow events in ephemeral streams that may carry concentrated loads of
soluble salts.
Despite these limitations, it is well established from laboratory studies that, when salinity
increases slowly, some organisms are able to tolerate incremental increases in salt
concentrations (between 10% and 50% of the original concentration) (James et al, 2003). In
these types of laboratory studies, the slow increase of salinity over time increases the chances
of survival of the adult stages of freshwater fish when compared with direct transfer studies
Potential impacts of salinity and turbidity in riverine ecosystems
(Kefford et al, 2004, Hart et al, 1991). A greater correlation was observed between the slow
acclimation LC50 (Lethal Concentration for 50% test population) results for Australian freshwater
fish and their maximum field distribution than with results from direct transfer studies and
their maximum field distribution (Kefford et al, 2004). Having said that, in the same study the
LC50 results for early life stage aquatic animals were observed to be lower than the maximum
field distribution for those species and the direct transfer LC50 results for adult life stage aquatic
animals.
There is also a spatial and temporal dimension to species tolerance that has been observed
for freshwater fish whose tolerance has been observed to be variable within a species and
between catchments. Williams (1987) observed geographically isolated pockets of Flyspeckled
Hardyhead (Craterocephalus stercusmuscarum fulvus), Smelt (Retropinna semoni ), Crimson
Spotted Rainbowfish (Melanotaenia fluviatilis) and Western Carp Gudgeon (Hypseleotris
klunzingeri ) to have significantly different salinity tolerances. Williams and Williams (1991)
argue that the variability in the salt tolerance of different fish species can be attributed to
their long-term and short-term ancestral and/or life history. Some freshwater fish such as
Barramundi (Lates calcarifer) have evolved a diadromous lifecycle, meaning that they migrate
between freshwater and saltwater to complete their lifecycle. Thus, Barramundi have well
developed physiological mechanisms for the regulation of salt to allow them to survive in salt
and freshwater. Other fish species remain in freshwater environments throughout their lifecycle
and cannot survive in saltwater. Some groups of fish that have had a relatively recent marine
ancestral background and are generally more tolerant to saline conditions include, Atherinidae
(Hardyheads), Eleotids (Gudgeons), Gobiidae (Gobies) and Ariidae and Plotosidae (Catfishes)
(Hart et al, 1991).
Genetic diversity may also be reduced with the onset of increased salinity due to selection
pressure thus affecting tolerance. It is possible that a loss in genetic diversity occurring over
a long period of time may reduce the resilience of the species and may mean they are more
susceptible to environmental stressors including diseases and contaminants. When exposure
to salinity occurs slowly, it is possible for some organisms to tolerate salinity increases higher
than their 72 hour lethal tolerance values. In an acclimation experiment on a freshwater bivalve
(Veneroidea corbicula), the 72 hour LC50 concentration was observed to be lower than their
336 hour LC50 when exposure to salinity was incrementally increased over a two week period
(NR&M unpublished data). However, acclimation to salinity may come at a cost, requiring an
additional energy input and potentially reducing long-term viability.
1.3.2 Osmotic regulation
Another aspect that can affect the tolerance of an organism is its ability to maintain the optimal
internal osmotic concentration required for survival. The biological complexity of freshwater
organisms influences their capacity to balance internal ionic composition against a salinity
gradient in water. Simple structured organisms including single-celled algae and bacteria
have limited detoxification mechanisms to regulate salt. These are named osmoconformers as
they are subject to the passive flow of molecules from the lower concentration to the higher
Potential impacts of salinity and turbidity in riverine ecosystems
concentration across a selectively permeable membrane (i.e. a cell wall) (Hart et al, 1991). This
passive flow of molecules is due to diffusion and in this case, a difference in osmotic potential
between an organism’s internal cells and the external environment induces a loss of water
and results in the eventual loss of functioning of the cells. More complex organisms such as
fish or invertebrates generally have a greater capacity to maintain optimal internal osmotic
concentration. These are named osmoregulators as they actively regulate their internal and
external ionic concentrations via active or mediated transport mechanisms (Bently 2002).
Although some species are able to osmoregulate, because it is an active mechanism, the
requirement to osmoregulate extracts a metabolic cost that may affect the organism’s longterm viability or resilience. The ability of an organism to regulate salt and their determination
as either an osmoconformer or an osmoregulator in part determines the salinity tolerance of
an aquatic organism (Hart et al, 1991).
1.3.3 Avoidance
To survive periods of high salinity some animals will attempt to avoid it. Some species can
remain in diapause during which the eggs or cysts can tolerate higher salinities and periods
of low flow and drought and emerge again when environmental conditions are favourable
(Skinner et al, 2001, Bailey et al, 2004). This extended diapause is a natural part of the lifecycle
of many freshwater taxa that include for example: Diptera, Nematoda, Turberllaria, Cladocera,
Harpacticoida, Cyclopoida, Anostraca, Conchostraca, Notostraca, Rotifera, Cyanophyta,
Bacillariophyta, Sarcodina and Ciliophora (Bailey et al, 2004). Mobile species such as fish may
migrate to areas with lower conductivity to avoid areas of high conductivity. Smaller species
with limited mobility can move to a shallower depth that may allow survival. In some cases
highly mobile and semi-aquatic animals are able to obtain resources from saline systems while
using fresher systems nearby for drinking and breeding (James et al, 2003). This may place
greater demands on freshwater resources and where this type of change is long term it may
change community dynamics and alter the structure of food webs resulting in further impacts.
1.4 Physical and chemical aspects
Increases in salinity can alter the physical and chemical properties of a solution
including the solubility of ions, pH, reduction and oxygenation potential, and
can lead to stratification in still waters (refer to Figure 1). The ions present in
solution that contribute to measures of salinity in rivers can be highly variable
and dependent on the geochemical characteristics of the catchment.
1.4.1 Solubility
Increased concentrations of suspended particles and salts often lead to their
precipitation out of solution in the presence of high concentrations of salinity.
Settling of matter on stream substrates can affect periphyton growth that in some
systems can result in the loss of an important functional group (Hart et al, 1991). This may
contribute to physical effects of blanketing substrate and habitat. It may also result in increased
light infiltration into the water column, that if combined with high concentrations of nutrients
may result in algal blooms (Murray Regional Algal Coordinating Committee 2002).
Potential impacts of salinity and turbidity in riverine ecosystems
Some alterations in pH can occur with increased salinity. Dominance of either hydrogen
(H+) or hydroxide (OH-) will influence pH. The resulting pH of a solution after an increase in
conductivity is proportional to the concentration of hydrogen (H+) or hydroxide (OH-) ions in
solution when at equilibrium. In a highly alkaline solution, its buffering capacity or its ability to
accept hydrogen ions is increased therefore increasing the solution’s pH. Therefore the pH of
surface waters that have naturally high alkalinity is less likely to be affected by increases in salts.
1.4.2 Composition
As previously discussed, salinity is a mixture of anions and cations in solution and given
that some organisms (osmoregulators) possess varied mechanisms for regulating different
ions, the ionic composition is likely to be a significant factor in determining the toxicity of
the salts to freshwater organisms. For this reason the use of integrative measures of salinity
(TSS and EC) may not be reliable predictors of toxicity for all water types (Mount et al, 1997).
For example Burnham and Peterka (1975) noted that Fathead Minnows (Pimephales promelas)
could tolerate TDS concentrations up to 15 000 mg L-1 (22 000 µS cm-1) in Saskatchewan
lakes dominated by sodium (Na+) and sulphate (SO42-), but did not persist above 2000 mg L-1
(3000 µS cm-1) in sodium (Na+), potassium (K+), bicarbonate (HCO3-) dominated lakes of
Nebraska. Similarly, a study by Dwyer et al (1992) demonstrated that the toxicity of high TDS
waters to the Water Flea (Daphnia magna) and Striped Bass (Morone sexatilis) was dependent
on the specific ionic composition of those waters. In a review of the effects of increasing
salinity on freshwater ecosystems in Australia Nielsen et al (2003) suggested that the ratio of
sodium and potassium (Na+ and K+) to magnesium and calcium (Mg2+ and Ca2+) is an important
determinant of the toxicity of salts on freshwater organisms. This notion is supported by Bayly
(1969) who found that the monovalent ions sodium and potassium (Na+ and K+) are more
toxic than the divalent ions calcium and magnesium (Ca2+ and Mg2+). This means that higher
proportions of sensitive taxa could be found in calcium bicarbonate-dominated water than in
sodium chloride-dominated water under equal conductivities.
Ionic composition is known to affect the distribution of copepods in saline lakes (Clunie et al,
2002). Also, in a review of available toxicity data for salts, Warne et al (2004) found that the
toxicity of calcium chloride (CaCl2), magnesium sulphate (MgSO4), sodium chloride (NaCl),
magnesium chloride (MgCl2), and calcium sulphate (CaSO4) was greater than that of artificial
sea salt which is dominated by NaCl but consists of many salts. This highlights the potential for
inadequacies in assessing the potential impacts of salts using only an integrative measure of
total salts.
Much of the existing species sensitivity information available has been derived using a
standardised salinity solution with ionic proportions similar to that of seawater (Bailey et al,
2002). These ionic proportions are common in Australian inland waters (Bayly and Williams
1973, in Kefford et al, 2004). However, in Queensland, the ionic proportions of surface waters
are known to vary according to their geographical location (McNeil et al, 2005).
Potential impacts of salinity and turbidity in riverine ecosystems
1.4.3 Contaminant interactions
In determining safe or acceptable concentrations of salinity in impacted systems, it is
important to consider the potential for interactions between contaminants and salinity as
degraded systems suffering from impacts from salinity may often be exposed to impacts from
other contaminants. Increases in aqueous salinity concentrations can directly alter the toxicity
of contaminants. The toxicity of contaminants may increase, be additive, remain constant or
even reduce in toxicity with increased salinity. In this way the presence of contaminants may
increase the susceptibility of an animal or population of animals to salinity. Previous studies of
the interactive effects of salinity and contaminants have demonstrated exponential increases
in toxicity for some contaminants in the presence of increased salinity (Dassanayake et al,
2003, Hall and Anderson 1995), whilst for other contaminants no observable increases in
toxicity have been reported. Thus it is useful to consider the established information available
on the combined effects of salinity with various classes of contaminants.
Organophosphate compounds are one such group that display altered toxicity in the presence
of salinity. Atrazine is a triazine herbicide known to bioaccumulate at low concentrations, and
is slightly to moderately toxic to aquatic animals (EXTOXNET 1996). Dassanayake et al (2003)
found that the toxicity of Atrazine to the Water Flea (Daphnia carinata) increased synergistically
with increasing concentrations of salinity; at low concentrations the effects were found to
be additive, and at higher concentrations the effects were found to be synergistic. Molinate
is a thiocarbamate pesticide that is highly toxic to some aquatic organisms (EXTOXNET
1996). The toxicity of Molinate to Daphnia carinata was found to be additive with increasing
concentrations of salinity. Chlorpyrifos is an organophosphate insecticide that under normal
conditions is very highly toxic to aquatic organisms (EXTOXNET 1996). Dassanayake et al
(2003) found that the toxicity of chlorpyrifos to Daphnia carinata was antagonistic with
increasing salinity. However, at higher concentrations the effect was solely due to salt toxicity.
Hall and Anderson (1995) reported a similar trend for organic compounds finding that the
toxicity of some organic compounds increased with increasing salinity and the toxicity of
others decreased with increasing salinity. Given these limited yet mixed results, and the fact
that many herbicides and pesticides are found in the environment, it is difficult to predict what
the likely interactive effects of organic herbicides and pesticides with salinity will be, though
serious consideration should be given to the effect of salinity-contaminant interactions where
these contaminants are found.
In a review of the influence of salinity on the toxicity of various classes of chemicals to
aquatic biota, Hall and Anderson (1995) found that the toxicity of metals (including cadmium,
chromium, copper, mercury, nickel and zinc) decreases with increasing salinity. Conversely, the
increased toxicity of metals is due to the greater bioavailability of the free metal ion (which is
the more toxic form) at lower salinities (Hall and Anderson 1995).
The effects of Petroleum Aromatic Hydrocarbons (PAHs) on marine species of crabs were
mixed (Hall and Anderson 1995). However, a study by Dange (1986) on the mortality of Tilapia
Potential impacts of salinity and turbidity in riverine ecosystems
(Oreochromis mossambicus) (a freshwater euryhaline species) was found to increase at higher
salinities after exposure to Naphthalene (a PAH).
There is little established information about the interactive effects of nutrients and salinity.
However, in isolation, excessive concentrations of each are often associated with biological
impacts. In a study of the biological effects of saline lake water disposal in the Lough Calvert
drainage scheme in Southwest Victoria Kefford (1998a) found that the operation of the scheme
resulted in changes to abundance of macroinvertebrate community structure. In this study
Kefford (1998a) notes that increases in salinity were also associated with increases in nutrients
and suspended solids. The effects of each could not be isolated due to the correlative nature of
the study, though it was likely that, when combined, salinity and nutrients were responsible for
the observed effects on the macroinvertebrate community.
Endocrine Disrupting Compounds (EDCs) have also been reported to display altered toxicity in
the presence of salinity. Each receptor responds to a hormone compound that is secreted by
the endocrine glands to trigger a specific response. EDCs occur naturally in the environment
but can also be due to anthropogenic sources. Disruption of the normal function of the
endocrine system can be caused by hormones and by compounds that can act as hormone
mimics. Interactive effects of salinity with EDCs are likely to be important as osmoregulation
processes in vertebrates are in many instances controlled by hormones that, according to
Bently (2002) allow the absorption and secretion of water and electrolytes, especially sodium
(Na+), potassium (K+), and chloride (Cl-) across epithelial membranes. The steroid hormone
cortisol is known to be responsible for inducing changes when animals move from saltwater to
freshwater, and prolactin concentrations increase with decreasing salinity (Knox et al, 1995).
Highly specialised chloride cells are responsible for the active transport of ions across the gills,
and there is good evidence that their numbers and functions change in the presence of these
two hormones (Knox et al, 1995). The mimicking or disruption of cortisol and/or prolactin may
result in increased salt sensitivity in vertebrates due to a reduced ability to osmoregulate.
Given that the endocrine system controls the mechanisms used by vertebrates to regulate salt,
it seems logical that interruption of the normal functioning of these systems may alter their
sensitivity to salt.
1.5 Effect on stream biota
There are seven major groups of organisms known to inhabit freshwater ecosystems. These
comprise vertebrates (e.g., fish, amphibians, reptiles, birds, and mammals), invertebrates
(e.g., protozoa, myxozoans, rotifers, worms, molluscs), plants, algae, fungi, bacteria, and
viruses (USEPA 2003). There are many ways that salinity can have a direct physiological effect
on these aquatic animals though many animals have evolved mechanisms for regulating
salinity. Salinity therefore has varied impacts on different animals and extent and type of
impact are dependent on their biological and physiological characteristics and of the duration
and exposure to salinity. Although salinity has the potential to affect all stream biota, only the
major groupings for which there are sufficient data available to make valid conclusions are
discussed here.
Potential impacts of salinity and turbidity in riverine ecosystems
1.5.1 Bacterial communities
Bacterial communities play a key role in the functioning of ecosystems through
nutrient and carbon cycling processes. Current scientific understanding suggests
that the presence of individual microbes within microbial communities is to some
extent determined by the physical and chemical characteristics of the ecosystem.
Some bacteria are known to survive in a wide range of environmental conditions
and sometimes exist in hostile environments where most other life is unable to
survive.
Freshwater bacteria appear to have some ability to adapt to slight changes
in salinity within a specified range after which there may be a change in the
community composition and associated changes in ecosystem function (Hart et al, 1991, Bailey
and James 2000). However, there is generally a lack of salinity tolerance information available
for bacterial communities so it is difficult to predict what effect that changes in salinity will have
on bacterial communities. Much of the information available regarding the effect of salinity on
microbial communities has been inferred from changes in microbial communities over salinity
gradients in estuaries. The relevance of such findings to freshwater ecosystems is limited,
though in the absence of information regarding the effect of salinity on freshwater bacterial
communities they can at least be used as a guide to likely effects. Despite these learnings it
remains unclear what effect changes in microbial community composition are likely to have on
ecosystem function.
1.5.2 Algae
Microalgae are known to drive primary production and play a key role in food web interactions
and inorganic nutrient cycling processes in many ecosystems, making them important
components of those aquatic ecosystems. Microalgae provide structural stability to substrates
and create mats that form habitat for invertebrates and fish (Fore and Grafe 2002). Despite
their importance to ecological systems, there is generally a lack of information about the
sensitivity of microalgae to salinity (Hart et al, 1991, Bailey and James 2000).
A review by (Bailey and James 2000) concluded that, as salinity increases, the number and
diversity of diatoms species is expected to fall. Previous studies have found that the richness
of diatom communities is correlated with ionic composition including the proportion of
sodium (Na+), potassium (K+), magnesium (Mg2+), and chloride (Cl-) (Bailey and James 2000).
Potapova and Charles (2002) and Pan et al (1996) correlated salinity concentrations and major
ions distributions with diatom taxa in rivers in the United States of America. In a study of the
distribution of diatoms in the Northern Kimberley region in Western Australia Tudor, Blinn
and Churchill (1991) found that 92 species (78% of the total number of species) were found
to be distributed in discontinuous groups at sites ranked along a TDS gradient. In this study
the diatom fauna was found to be clearly divided between freshwater sites of between
0–660 mg L-1 TDS (0–970 µS cm-1) and between 460–42 000 mg L-1 TDS (676–61 764 µS cm-1).
Further to this, a study conducted by Blinn and Bailey (2001) showed that salinity and
phosphorus interacted to determine stream diatom structure in drainages with high secondary
Potential impacts of salinity and turbidity in riverine ecosystems
salinisation. A study by Pilkaitytë et al (2004) used a mesocosm approach to demonstrate that
as waters become increasingly saline, benthic algal communities are dominated by microbial
mats composed almost entirely of filamentous cyanobacteria. It can be concluded then that
some species of microalgae are sensitive to salinity changes and that community level changes
can be observed with increasing salinity. It may be useful then to use algal communities as a
sensitive biological indicator of the effect of changes in salinity.
1.5.3 Macroinvertebrates
Macroinvertebrates form an important component of aquatic food webs making up herbivores,
detritivores and predators. There is a substantial body of knowledge on the acute tolerance
of macroinvertebrates to marine salts gained from laboratory observations (Bailey et al,
2002, Kefford et al, 2003). Their sensitivity as a group is varied and they consist of species
that are highly sensitive to salt and species that are highly tolerant to salt (Clunie et al, 2002,
Bailey and James 2000). Macroinvertebrate salinity sensitivity does not correspond well
with taxonomic groups. Rather, different genera and species within the same family in some
instances may have a markedly different sensitivity to salinity.
For macroinvertebrates, the primary driver for salinity regulation is related to osmosis
(Hart et al, 1991). As mentioned earlier in section 1.3.2, this is likely to be primarily driven
by the lack of complexity in these organisms, and a lack of mechanisms to assist in the
regulation of internal ionic composition against an external gradient. The variation in
tolerance among species of macroinvertebrates has been attributed to the internal ionic
concentrations of invertebrates (Hart et al, 1991). For example, the higher the internal ionic
concentration of a species the higher the salinity tolerance is likely to be for that species.
Many freshwater macroinvertebrates have internal ionic concentrations of 1000 to 15 000 mg L-1
(1470–22 058 µS cm-1) (Hart et al, 1991). Shrimp are known to have a relatively high internal
ionic concentration compared with their surrounds (Hart et al, 1991) and it follows also that
they have a relatively high tolerance to salinity (refer to Table 2).
The invertebrate species that are often found to be relatively tolerant to salinity include
beetles and dipteran flies. Other groups such as stoneflies, mayflies, caddisflies and
dragonflies are generally sensitive to even minor increases in salinity (Hart et al, 1991). From
the limited sensitivity information available for molluscs and snails, pulmonate snails are
often found to be particularly sensitive to increasing salinity (Hart et al, 1991, Clunie et al,
2002, Bailey and James 2000). Crustaceans are generally thought to be relatively tolerant of
conductivity, though there are some that are quite salt-sensitive (Hart et al, 1991, Clunie et al,
2002, Bailey and James 2000).
Despite the complexities associated with the interpretation of macroinvertebrate sensitivity to
salt, some generalisations can be made about their salinity tolerance. A study by Hart et al
(1991) indicated that salinities in excess of 1000 mg L-1 (1470 µS cm-1) are likely to be the point
at which adverse effects are likely to be observed in invertebrate communities. Subsequent
work by various authors has indicated that adverse individual and therefore community
effects may be occurring below this concentration. In the case study presented in section 1.8,
10
Potential impacts of salinity and turbidity in riverine ecosystems
a distinctive shift from communities with a high proportion of salinity-sensitive taxa to
communities of more tolerant individuals has been observed to occur between 544 and
680 g L-1 (800–1000 µS cm-1) for edge habitats in Queensland (Horrigan et al, 2005).
This figure is lower for the riffle habitat where this effect has been observed to occur at
around 440 mg L-1 (300 µS cm-1) (Horrigan et al, 2005). In section 1.7 the percentiles of EC for
Queensland streams are shown. EC values for Queensland are generally below 1500 µS cm-1
but there are several zones as classified in Table 5, that have the 90th percentiles of EC data
exceeding conductivities of 800 to 1000 µS cm-1. The data given in section 1.7 are of
percentiles grouped by catchments and therefore may obscure outliers, so within these
and other catchments there is likely to be sites having greater EC values.
Kefford (1998b) investigated potential linkages between EC and macroinvertebrate
communities in four river systems of southwest Victoria, Australia. The results of this study
showed that macroinvertebrate community structure was associated with EC for the river
systems investigated. The results of this study were not confounded by geographical scale
parameters due to sampling of paired sites upstream of the confluences of the two streams
having different stream salinities.
Table 2 Summary of acute 72-hour salinity tolerance of selected
macroinvertebrates to marine salts
Order
Family
Genus
Species
LC50
95% CI
Ephemeroptera
Baetidae
Cloeon
centroptilum
5 500
0.76 9.8
Ephemeroptera
Baetidae
Genus 1
NA
6 200
3.7–3.9
Diptera
Chironomidae
NA
NA
10 000
6.8–15.0
Gastropoda
Physidae
Physa
acuta
14 000
13–15
Trichoptera
Ecnomidae
Ecnomus
NA
16 000
9–28
Hemiptera
Corixidae
Micronecta
annae
17 000
16–29
Plecoptera
Gripopterygidae
Dinotoperla
twaitesi
18 000
15–24
Trichoptera
Leptoceridae
Triplectides
australicus
22 000
19–24
Trichoptera
Calamoceratidae
Anisocentropus
NA
23 000
19–26
Trichoptera
Leptoceridae
Notilina
spira
25 000
22–29
Decapoda
Atyidae
Paratya
australiaiensis
38 000
34–42
Amphipoda
Ceinidae
Austrochiltonia
NA
52 000
47–59
(Modified from Kefford et al (2003), all figures are in µS cm-1, NA = Not Available, CI = Confidence Interval)
1.5.4 Vertebrates
Vertebrate species are at the top of aquatic food webs and the group contains many iconic fish,
bird, amphibian, and reptile species. While data are lacking for many vertebrate species there
is some salinity tolerance information available for Australian freshwater fish that mainly stem
from descriptions of optimum conditions for aquaculture (Clunie et al, 2002). Hart et al (1991)
suggest that freshwater fish appear to be quite tolerant up to salinities around 10 000 mg L-1
(14 705 µS cm-1). Bacher and Garnham (1992) suggest that most freshwater fish can tolerate
salinities up to 13 000 mg L-1 (19 117 µS cm-1) as teleost fish maintain their internal ionic
concentration in the same range. Once internal concentrations are exceeded, osmoregulatory
11
Potential impacts of salinity and turbidity in riverine ecosystems
mechanisms break down rapidly (Clunie et al, 2002). There is evidence of freshwater fish being
more sensitive to salt in the early life stages of development, with non-hardened eggs being
particularly vulnerable to increased salinity (Table 3). The results of slow acclimation of tests
and chronic tests as indicated in Table 3 suggest that many species of fish are able to tolerate
higher salinities if they are introduced to them incrementally over extended periods of time
than if they were directly exposed to the same maximum concentration.
Table 3 Salinity tolerance of selected freshwater fish
Species
Common Name
Gadopsis marmoratus
River Blackfish
Hephaestus fuliginosus
Sooty Grunter
Perca fluviatilis
Redfin
Maccullochella macquariensis
Trout Cod
Cyprinus carpio
European Carp
Atherinosoma microstoma
Small Mouthed Hardyhead
Slow
(chronic)
LC50
Early life
stage
LC50
8 800 1
2
11 800
10 300f 2
11 760 3
d4
5 000
6 600e 5
12 000 5
10 700 6
18 800 7
13 200 h 8
158 800 9
d4
2 900
3 030 e 5
19 800 5
Macquaria australasica
Macquarie Perch
Maccullochella peelii peelii
Murray Cod
19 400 10
23 100 10
13 800 5
Bidyanus bidyanus
Silver Perch
20 100
10
23 500
10
2 200 11
26 500c 11
26 500i 12
8 800d 13
Pseudaphritus urvilli
Tupong/Congoli
25 000 1
26 200
10
Tandanus tandanus
Freshwater Catfish
21 250
10
10
Mogurnda adspersa
Purple-Spotted Gudgeon
21 800
Melanotaenia splendida splendida
East Queensland Rainbowfish
13 200 14
4 400 14
26 200 6
25 000 15
13 200b 14
Carassius auratus
Goldfish
10 700 6
18 800 16
19 200 17
28 200 17
Macquaria novemaculeata
Australian Bass
29 400a 18
Melanotaenia duboulayi
Duboulay’s Rainbowfish
32 400 d
30 900b 15
Gambusia holbrooki
Mosquito Fish
28 700 19
36 800 f 4
Prototroctes maraena
Australian Grayling
44 100 20
7 400c 20
15
15
25 000 d
17 600b 15
Melanotaenia fluviatilis
Crimson Spotted Rainbowfish
44 100
31 000 15
43 800
Macquaria ambigua
Golden Perch
21 200 10
45 600 10
12 200 5
Kuhlia rupestris
Jungle Perch
51 500 g 9
Salmo gairdneri
Rainbow Trout
51 500 21
4 400 21
Salmo trutta
Brown Trout
51 500 21
4 400 21
Leiopotherapon unicolour
Philypnodan grandiceps
12
Direct
(acute)
LC50
Spangled Perch
Flat-headed Gudgeon
32 400
10
34 900
10
52 200
10
58 800
10
Potential impacts of salinity and turbidity in riverine ecosystems
Species
Common Name
Direct
(acute)
LC50
Craterocephalus stercusmuscarum
fulvus
Fly-specked Hardyhead
64 300 22
Hypseleotris klunzingeri
Western carp Gudgeon
55 900 15
Retropinna semoni
Galaxias maculatus
Australian Smelt
Common Galaxias
Slow
(chronic)
LC50
Early life
stage
LC50
73 500 15
86 800
22
66 200
23
23
91 200
8 800 1
(modified from Clunie et al (2002), values have been standardised to µS cm-1)
1 (Bayly 1993)
2 (Bisson and Bartholomew 1984)
3 (CF&L 1988)
4 (Clucas and Ladiges 1980) 5 (Allen and Cross 1982)
6 (Bunn and Davies 1992)
7 (Chessman and Williams 1974) 8 (Chessman and Robinson 1987)
9 (De Decker and Geddes 1980) 10 (Allen 1982)
11 (Bailey and James 2000) 12 (Denne 1968)
13 (Brock 1981)
14 (Brock and Lane 1983)
15 (Beumer 1979)
16 (Campbell 1995)
17 (Clemen et al, 1983)
18 (Blake 1981)
19 (Chessman and Williams 1975)
20 (Brock and Shiel 1983)
21 (Bird 1978)
22 (Bayly 1969)
23 (Ackrill et al, 1969)
Comments: a Spawning requirement, b Fry LC50 tolerance, c Limit of egg development, d Egg LC50 tolerance prior to cleavage,
e Egg LC50 tolerance prior to hardening, f Limit to sperm motility hatching from larvae, g Limit to osmotic ability
Amphibians are particularly sensitive to salt as they are generally poor osmoregulators,
and most species are completely absent from brackish and saline environments. These
have a unique physiology that can make them vulnerable to many toxicants including salinity
(Mann and Bidwell 1999). The skin of an adult amphibian is a permeable organ used for
respiration and water balance, whereas the larval stage relies predominantly on gills for
respiration (Mann and Bidwell 1999). During the aquatic larval stage, amphibians have
highly exposed eggs that are vulnerable to salinity. A wide variation in salt tolerance and
physiological responses to increased salinity has been observed in species of amphibians
(Mann and Bidwell 1999). There are few tolerance data available for amphibians; however,
the cane toad (Bufo marinus) has been reported to be found at a salinity of up to 14 000
mg L-1 (20 588 µS cm-1) (Liggins and Grigg 1985) has been observed to increase plasma
concentrations to moderate the effects of osmotic pressure (Liggins and Grigg 1985).
Rana esculenta and Rana temporaria have been reported to have a salinity range of up to
7000 mg L-1 (approximately 10 294 µS cm-1) (Ackrill et al, 1969) and Rana pipiens have been
reported to die at 35 000 mg L-1 (approximately 51 470 µS cm-1) (Bentley and SchmidtNielsen 1971) and 7000 mg L-1 (Ackrill et al, 1969). In general, amphibians must maintain
hyperosmoticity to their environment and salinities greater than 25% of that of seawater
are likely to be problematic for survival (Mann and Bidwell 1999).
1.5.5 Plants
Aquatic plants are regarded as producers as their growth converts light and nutrient energy
into oxygen. Aquatic macrophytes provide important habitat for many animals including fish.
To some extent, plant species composition, distribution and percentage cover of aquatic plants
may determine the fish species composition, and individual fish species production. Aquatic
plants have highly variable salinity tolerance ranges (Table 4). Many aquatic plants have
been observed to be tolerant as adults but are known to be sensitive in their early life stages
13
Potential impacts of salinity and turbidity in riverine ecosystems
(Bailey et al, 2002, Hart et al, 1991). A large proportion of aquatic macrophytes are sensitive to
salinity at concentrations between 1000 and 2000 mg L-1 (1470–2941 µS cm-1) above which the
growth and reproductive success of aquatic macrophytes are likely to be significantly reduced
(Hart et al, 1991). A reduction in the biomass of the water lettuce Pistia stratiotes has been
observed to occur at a salinity of 830 mg L-1 (1220 µS cm-1) whilst mortality was observed to
occur at a salinity of 2500 mg L-1 (3676 µS cm-1) (Haller et al, 1974).
Table 4 Salinity tolerance of selected aquatic plants
Species
Common Name
Reduction in Biomass
Mortality
Pistia stratiotes
Water Lettuce
1220 1
3600 1
Typha domingensis
Cumbungi
4300 = slight,
8600 = severe 2
Eichornia crassipes
Water Hyacinth
2500 1
Cyperus involucratus
NA
6000 biomass reduced,
12 800 necrosis evident 3
Baumea arthrophylla
NA
8500 = 43% 9
Amphibromus fluitans
Graceful swamp
wallaby-grass
8800 no change 5
Myriophyllum crispatum
Watermilfoil
1500 6
10 300 = 48% mortality 6
Eleocharis acuta
Common Spike Rush
1500 and above growth
reduced 6
10 300
Potamogeton tricarinatus
Floating Pondweed
1500 6 8800 severe 5 8800 12
10 300 44% mortality 6
Triglochin procera
Water Ribbons
1500 6 8800 5
10 300 6
Bolboschoenus medianus
NA
13 000 = 37 – 58% 7
Typha domingensis
Cumbungi
5100 = 50% 8
4900 1
22 000 = 75% mortality
9
Typha domingensis
Cumbungi
8500 = 33% – 53%
Typha latifolia
Cumbungi
8500 9
Salvinia rotundifolia
Salvinia
9800 1
Hydrilla verticullata
Hydrilla
9700 1 12 000 biomass
reduced by ~80% 10
Lemna minor
Duckweed
14 700 1
Vallisneria americana
Ribbonweed
9800 1 17 600 10
19 600 mortality 1
Najas quadalupensis
NA
14 700 1
19 600 mortality 1
Myriophyllum spicatum
Eurasian Watermilfoil
14 700 1
24 500 mortality 1
Phragmites australis
Common Reed
14 700
11
14 700 1
33 100 = 88% mortality
(for seedlings) 11
(values have been standardised to µS cm-1)
1 (Haller et al, 1974)
2 (Hocking 1981a)
3 (Hocking 1981b)
4 (Morris 1998)
5 (Warrick and Bailey 1997)
6 (James and Hart 1993)
7 (Morris and Ganf 2001)
8 (Glenn et al, 1995)
9 (Anderson 1977)
10 (Twilley and Barkon 1990)
11 (Lissner and Schierup 1997)
12 (Warrick and Bailey 1998)
NA = Not Available
The available sensitivity information as reviewed by Hart et al (1991) indicates that many of
the higher plants associated with lowland rivers are also salt-sensitive with upper tolerance
14
Potential impacts of salinity and turbidity in riverine ecosystems
ranges around 2000 mg L-1 (2941 µS cm-1). Waterlogging of plant roots associated with raised
watertable levels may result in the impairment of growth and death. The combined effect of
waterlogging was reported to have greater detrimental impact on the growth and survival
of Melaleuca and Eucalyptus seedlings than either salinity or waterlogging did alone (Macar
(1993) in Clunie et al (2002)). Loss of riparian vegetation is associated with a loss of stream
shading that can result in increased water temperatures and stream metabolism and may also
reduce bank stability (Figure 1). The health of riparian vegetation is an important component of
stream habitat as discussed further in section 1.6.2.
1.6 Impacts on aquatic ecosystems
Whilst some animals have a capacity for survival in hyper-saline conditions and
some are highly tolerant of dissolved salts, most freshwater ecosystems will be
impacted by increased concentrations of dissolved salts. There is a wealth of
literature that supports the view that increased stream salinity is associated with
ecological impacts including a loss in aquatic biodiversity (Williams 2001, Hart et al,
1991, Clunie et al, 2002, Bailey and James 2000).
When considering how ecosystems will be impacted and what we might expect
to observe in affected ecosystems, we must consider that salinity can impact
upon riverine ecosystems in many different ways. These can include changes
in community structure (i.e. loss of aquatic biodiversity), breakdown in food webs, shift in
community composition to more tolerant species, alterations to normal ecosystem function
through reduction in nutrient cycling and metabolism resulting in changes to the physical and
chemical parameters of waters. These processes are a complex set of interactions that occur
at the individual organism and the ecosystem level. These interactions have been depicted in
a simplified illustration in Figure 1.
None of the impacts depicted in Figure 1 act in isolation from each other. Rather, each of the
individual impacts is interlinked and has follow on effects. The linkages between cause and
effect impacts from salinity on aquatic ecosystems can be seen in the following example of a
stepwise impact process. A raised watertable in riparian zones due to salinity impacts can lead
to increased salinity in the root zone and result in die back of riparian vegetation. Die back of
riparian vegetation reduces stream shading that in turn results in increased stream metabolism
and changes to food web structure from a heterotrophic to an autotrophic system (Boulton
and Brock 1999). This can alter ecosystem processes that include the cycling of nutrients
and organic contaminants. Stream riparian zones are known to reduce the influx of nutrients
of overland flows and their loss is likely to result in increased nutrient and sediment input
into streams. Vegetation in riparian zones also provides stability to stream banks preventing
morphological alterations to stream channels such as bank slumping and stream braiding that
can also have negative impacts on aquatic ecosystems. The ecological effect of sedimentation
on riverine biota is discussed further in section 1.5.
15
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 1 Conceptual model of salinity impacts on a freshwater ecosystem
16
Potential impacts of salinity and turbidity in riverine ecosystems
Studies of the ecological impacts of saline water disposal provide some indication of the
changes expected to occur in community structure with increased in salinity. A study by Kefford
(1998a) found that saline water disposal from the Lough Calvert Drainage Scheme in Victoria
significantly affected the macroinvertebrate community structure. Piscart et al (2005) observed
macroinvertebrate richness to decrease by 30% downstream of a 1.4 g L-1 (2060 µS cm-1) salinity
input and also observed a slight change in realtive abundances of invertebrate feeding groups
to follow the salinity gradient. At high conductivity sites Piscart et al (2005) found a greater
number of exotic species to be present than were in lower conductivity sites. Skinner et al (2001)
looked at species richness recovered from propagules in dry wetland sediments along a salinity
gradient in temporary saline lakes. In this study, a reduction in species richness and an increase in
abundance of emergent organisms correlated with an increasing salinity gradient with algae and
protists found to dominate at high salinity. The authors suggest that in nature this could result
in insufficient food for animals higher in the food chain, including fish and waterfowl. Metzeling
(1993) used multivariate analysis to show that a clear distinction between salinity regimes could
be seen with rare taxa compared with that of common taxa. This suggests that an alternative
stable state community is likely to occur at high salinities. However, what is not clear is whether
or not there is likely to be a return from this alternative state if salinity concentrations subside,
returning to their previous condition.
Shifts in community structure may be dominated not only by species tolerance but also other
factors co-occurring in the environment including chemical interactions. For example, at high
salinity concentrations, suspended particulates precipitate out of solution, improving the clarity of
waters. Species having a high tolerance to salinity and those that prefer clear waters for predatory
behaviour are likely to be more successful under these conditions. Other examples of factors
co-occurring with salinity increases and that result in shifts in community structure and their
resulting impacts are discussed further in the following sections 1.6.1 and 1.6.2.
1.6.1 Ecosystem processes
Changes in community structure may result in alterations to ecosystem processes. At this
stage there is little quantitative information available that describes the changes in ecosystem
processes that may be expected to occur as a result of increased salinity. An example of
expected changes in ecosystem processing due to salinity increases can be seen in bacterial
communities. As bacteria grow they undergo cellular respiration that results in the release
of energy and conversion of compounds. In this way, bacteria process minerals, nutrients
and carbon, converting these compounds into food and assisting the natural treatment and
cycling of raw compounds that may otherwise pollute aquatic ecosystems (Hart et al, 1991).
So the loss of bacteria responsible for nutrient conversions may cause an increase in nutrients
or other organic compounds, potentially altering the functioning of the ecosystem. This
process occurring in a system already degraded by salinity impacts is likely to result in further
degradation. The existing information suggests that increased salinity does have an appreciable
effect on primary productivity.
In a study by Davies (2004) changes in the metabolic processes of gross primary production
(GPP) and community respiration (CR24) were measured in two streams in adjacent catchments
17
Potential impacts of salinity and turbidity in riverine ecosystems
with varying salinity. The results of the study showed that even low levels of salinity
substantially suppressed GPP and CR24. Food web structure (fish and macroinvertebrate
community structure) showed an associated shift to a decreased incorporation of algal
carbon towards increased reliance on detrital carbon. At high salinities, a differential loss of
algal-grazing species was recorded. Consequently benthic metabolism may be describing
a fundamental underlying mechanism by which salinity influences changes observed
in community structure (e.g. reducing GPP to a level where algal consumers cannot be
supported) with increasing salinity.
1.6.2 Stream habitat
Stream habitat is essential for the healthy functioning of aquatic ecosystems as many
organisms require appropriate habitat at critical times in their lifecycle. Aquatic macrophytes
comprise a significant component of stream habitat and their tolerance is an important
consideration when determining the ecological effects of salinity. In some cases habitat
alteration due to salinity may be a greater threat to aquatic ecosystems than the direct
toxic effect of some species. For this reason it is important that plant species be included
in assessments of salinity risk. For example, freshwater fish can have much higher salinity
tolerance than the aquatic plants that they may depend on for habitat. Since many fish are
dependent on habitat for feeding and reproductive success, in some cases measuring the
loss of habitat may be a more obvious predictor of impacts to those species than measures of
direct tolerance. However, after clear and measurable impacts have occurred to alter stream
habitat, it may be too late to prevent irreparable damage to aquatic ecosystems.
1.7 Salinity in Queensland surface waters
When considering the biological effects of salinity in Queensland, it is important
to appreciate the extent and type of salinity found in Queensland’s surface
waters. Studies by McNeil and Clarke (2004) and McNeil and Cox (2000)
have been compiled here to provide a useful indication of salinity and ionic
composition zones for Queensland. Datasets used in these studies are based
on monitoring programs conducted by Queensland government departments
including Natural Resources and Mines and the Environmental Protection
Agency and other organisations. The original data from these studies have
been used to prepare maps characterising the ionic composition of the surface
waters in Queensland NAP priority regions. This section of the report provides a
valuable insight into the broad salinity and ionic composition patterns of Queensland streams.
1.7.1 Queensland salinity zones
Individual sites from across Queensland were categorised into high, moderate or low salinity
groups. Membership to a group was determined by fitting the 50th and 80th percentile of all
EC values into the categories (McNeil and Clarke 2004). Site membership was mapped and
then grouped into zones according to like salinity descriptions. Table 5 shows the proposed
Queensland salinity zones and their percentiles.
To assist in creating the boundaries between salinity zones, other water chemistry
characteristics were also considered. Table 5 shows proposed salinity zones for Queensland.
18
Potential impacts of salinity and turbidity in riverine ecosystems
The zones tend to follow catchment boundaries, though in some instances consideration was
given to existing management boundaries of the NAP regions when delineating boundaries for
salinity zones. For example, the Northern Murray Darling zone and the Fitzroy zone had similar
characteristics in terms of their representation in salinity zones but were separated according
to their catchment.
Table 5 Electrical conductivity percentiles for Queensland salinity zones
Percentiles of EC μS cm-1
Data used
Zone
Relative salinity
Sites
Number
of
samples
90
75
50
25
10
Wet Tropics
Generally very low
49
6199
130
92
71
50
36
Cape York
Mainly low, quite variable
92
3166
198
125
82
57
42
Belyando Suttor
Low
5
271
225
168
135
109
80
Western Murray
Darling Basin
Appears to be low
36
253
312
169
118
88
70
Lake Eyre
Low
4
383
410
200
128
90
71
Fitzroy Central
Low to moderate
42
4376
510
340
242
175
130
Central Coast North
Low to moderate, variable
17
1916
560
375
200
120
88
Burdekin Bowen
Moderately low but some
high outliers
18
1944
470
271
176
129
98
Maranoa Balonne
Border Rivers
Moderately low
28
2872
471
325
234
165
123
Gulf
Moderate
12
565
630
500
245
157
100
Southern Coastal
Moderate but variable
45
6717
732
520
340
212
121
Fitzroy North
Moderately high and
variable
11
755
1250
720
355
209
130
Sandy Coastal
Moderate to high, very
variable
11
1195
1310
626
368
216
90
Condamine
Macintyre
Moderate to high
33
4003
755
500
355
255
189
Callide Upper Burnett
High, very variable
28
2501
1450
760
500
339
240
Central Coast South
High and variable
6
653
1500
970
640
444
230
Don
High
10
372
1058
680
346
214
170
Southern Divide
Generally very high
59
5935
1570
1120
760
481
289
The salinity zones proposed here do not represent natural reference ranges, as reference
and test sites were used together to provide an indication of ambient salinity ranges. Rather
they do provide a broad scale categorisation of areas of related water chemistry based on
ionic composition and their conductivity. Hence, the percentiles presented do not constitute
salinity targets. It should also be recognised that a certain level of salinity at the catchment
scale does not necessarily indicate that for specific sites within the broader salinity zones
of the catchment. It is also important to mention that although a large dataset was used to
determine salinity zones (approximately 63,000 independent EC measures), in future it would
be beneficial to include more data for the study. This would allow greater spatial resolution and
refinement of the zones.
19
Potential impacts of salinity and turbidity in riverine ecosystems
The salinity zones identified here can be used to identify sites or sub-catchments where
the EC is unusually high or low compared with the rest of the region and provide a valuable
overview of regional salinity zones for Queensland.
1.7.2 Queensland ionic composition
In considering the biological effects of salinity at the local or regional scale it is important to
consider the composition of the total ions that make up measures of salinity locally as the
composition of ions is likely to have a bearing on its potential biological impacts. The effects
of the ionic composition of salts on biological impacts are discussed further in section 1.4.2.
Sites were grouped according to the proportion of ions typical at sites in Queensland using
a Principle Components Analysis (PCA) (McNeil et al, 2005). In this study ionic ratios were
shown to follow geographic trends. An investigation into the effect of the stage of flow on
total salinity and ionic proportions showed that the stage of flow did have an appreciable
effect on total salinity, but not to have a significant effect on ionic ratios. The geographical
locality of sites had a greater impact on ionic ratios (McNeil et al, 2005) than did flow alone.
Figure 2 shows the broad composition patterns in Queensland as water composition provinces.
Figure 4 to 8 show the trends in ionic composition in the Queensland NAP priority catchments.
Figure 3 shows the locality of the NAP catchments in Queensland and indicates the relative
scale of each of the catchments against that of Queensland. The water types and the codes
used to represent them in subsequent maps are described in detail in Table 6. Maps of ionic
composition shown here indicate the expected composition under base flow conditions.
Table 6 Summary of stream water chemistry in Queensland
Water Types
(Ionic Composition)
Number of
samples
Salinity Range
(90-10%)
1601
High Salinity Na Cl
Major Ions (%)
45–440
60–80
10–25
3–17
60–90
8–36
0–9
339
735–36 800
55–90
5–25
2–20
70–95
0.6–25
0.3–10
Low Salinity
Ca+HCO3-
12 586
66–590
25–55
15–40
18–50
10–50
40–85
0–10
High Salinity
Na+ Mg+ Cl-
5073
230–5900
40–55
20–40
10–30
50–90
5–43
0–10
Moderate Salinity
Mg+ HCO3-
1155
300–1470
20–55
40–70
12–40
3–50
40–95
0–5
Low Salinity
Na+Cl- HCO3-
2038
40–600
50–80
10–25
10–25
36–65
25–55
0–10
Low Salinity Na Cl
+
-
Cl-
SO4-
Mg2+
-
Ca2+
HCO3-
Na+
+
(Reclassification of water chemistry zones from table in McNeil et al (2005), all figures standardised to µS cm-1)
McNeil et al (2005) found the regional chemical trends in ionic composition to be consistent
with trends in geology and climate. Streams from northeast Queensland, with short, steep
catchments and high rainfall, generally have low salinity and are dominated by sodium
chloride. This pattern is consistent with the sandy southern catchments (refer to Table 5 for
examples). The proportion of sodium chloride was generally found to decrease westward
from the east coast. Streams draining the western side of the Great Dividing Range or flowing
into the southern Gulf of Carpentaria were found to contain relatively hard water given the
low concentrations of salinity. Streams in Western Queensland are higher in salinity and
bicarbonate. In large catchments draining southwest from Queensland into central Australia,
20
Potential impacts of salinity and turbidity in riverine ecosystems
the composition was found to be extremely variable and commonly high in sulphate. Also,
high proportions of magnesium in low salinity waters were found in areas within the vicinity
of basalts.
Sites in the southwest of the Mary Burnett catchments (Figure 4) were dominated by sodium
magnesium chloride. In the northern and southern parts of the Mary Burnett all anions
and cations were found to be present with no specific ions found to dominate the ionic
composition. The eastern sector of the Mary Burnett is generally dominated by sodium and
chloride.
The ionic composition of sites in the Burdekin catchment is (Figure 5) generally dominated
by high bicarbonate and calcium with some areas having elevated sodium sulphate. Sodium,
chloride, bicarbonate dominance was observed in the northeast of the Burdekin and some
high magnesium bicarbonate in the northwest part of the catchment.
The composition of ions in the northwest of the Brisbane catchments (Figure 6) is generally
dominated by sodium magnesium chloride. Sites in the southern part of the catchment were
not dominated by any specific ions. Sites in the northeast of the Brisbane catchment near the
coast were generally dominated by sodium chloride.
The Queensland Murray Darling catchments have some isolated sites with elevated sodium
sulphate (Figure 7). No dominant ions were observed in central sites. Southwest sites in the
Queensland Murray Darling were generally dominated by sodium sulphate ions and sites in
the northeast are dominated by sodium magnesium chloride.
The Fitzroy catchment generally has no dominant ions in the north part of the catchment
(Figure 8). However, one site is dominated by sodium chloride in that part of the catchment.
Sites in the southeast are dominated by high bicarbonate and calcium. Two sites in the
southwest exhibit moderate salinity with high magnesium content. Pockets of sodium
magnesium chloride also occur in parts of the catchment.
21
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 2 Water composition provinces for Queensland
22
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 3 National Action Plan for salinity and water quality priority catchments
23
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 4 Water types of the Mary and Burnett catchments
24
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 5 Water types of the Burdekin catchment
25
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 6 Water types of the Brisbane and Western catchments
26
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 7 Water types of the Queensland Murray Darling catchments
27
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 8 Water types of the Fitzroy catchment
28
Potential impacts of salinity and turbidity in riverine ecosystems
1.8 Salinity sensitivity index for macroinvertebrates
The relationship between stream macroinvertebrates and measures of conductivity
in Queensland river systems was examined to assess if there were any broad
patterns in community composition that were attributable to salinity. Family level
presence/absence stream macroinvertebrate data from edge (2580 samples) and
riffle (1367 samples) habitats collected throughout Queensland in spring and
autumn from 1994 to 2002 was used in this analysis. Salinity Sensitivity Scores (SSS)
were derived for individual macroinvertebrate families in Queensland. SSS were
derived from the results of a sensitivity analysis using predictive Artificial Neural
Network (ANN) models (for a more detailed account of these methods see Horrigan
et al, 2005). An SSS was assigned to each taxon (1 – very tolerant, 5 – tolerant,
10 – sensitive) based on the resulting frequency and sensitivity plots and the taxon-specific
mean conductivity values.
After establishing the SSS for individual macroinvertebrates, a Salinity Index (SI) was proposed
to reflect changes in macroinvertebrate communities caused by changes in conductivity. The SI
was calculated using the following formula:
ΣX  T
SI ______
​ Ni i​ Where:
(SI)  Salinity Index
Xi  1 if taxon i was present, Xi  0 if absent
Ti  Salt Sensitivity Score of taxon i
N  Total number of taxa in the sample
The SI can theoretically vary from a value of 1 when all the taxa in a sample are highly tolerant
to a value of 10 when all the taxa are sensitive. In practice opportunistic taxa are expected to
be present in unimpacted and impacted sites. These species will contribute to maintaining the
total score less then 10 and higher than 1. An SI based on the cumulative SSS was proposed to
reflect changes in macroinvertebrate communities caused by changes in conductivity.
The results show that as conductivity increases, sensitive taxa are being replaced by tolerant
taxa, and this is reflected in decreasing values of SI with increasing conductivity (Figure 9). This
trend is obvious in both habitats but appears to be more prominent in riffles. Figure 10 shows
changes in the percentage of sensitive and very tolerant taxa with increasing conductivity
(12 equal intervals). With reference to riffle data, sites having an EC in the range of 800 and 1500
µS cm-1 were observed to have a decrease in the mean percent of sensitive taxa from 33 to
16.7 relative to the low conductivity category (22-99 µS cm-1) and percent of very tolerant taxa
increased accordingly from 9.4% to 32%.
29
Potential impacts of salinity and turbidity in riverine ecosystems
(a)
(b)
Figure 9 Salinity index in 12 equal data groupings along increasing conductivity gradient
for edge (a) and riffle (b) habitats. Median values with boxes corresponding to 80th and 20th
percentiles and horizontal bars to maximum and minimum.
30
Potential impacts of salinity and turbidity in riverine ecosystems
SI values generally decreased along the salinity gradient. However, low SI values are also
observed in many cases with low conductivities, especially in the edge habitats. One possible
explanation for this is that several taxa such as Leptophlebiidae and Helicopsychidae are known
to be sensitive to many water quality parameters other than salinity. The SI may therefore be
influenced by other water quality parameters. To address this possibility, additional analyses
were applied to test whether or not the SI indeed reflects changes in macroinvertebrate
communities due to salinity alone and are not affected by natural variability or the effect of
other widespread stressors such as concentration of nutrients or turbidity. To achieve this,
sites judged to have good water quality (i.e. turbidity  5 NTU, total nitrogen  0.37 mg L-1,
total phosphorus  0.05 mg L-1, pH between 6.5 and 9 and dissolved oxygen  5 mg L-1) were
subjected to a partial Canonical Correspondence Analysis (CCA). This analysis allows for the
identification of water quality factors that may account for natural variability (Horrigan et al,
2005) and also allows for these factors to be removed from the analysis. From this analysis,
it was concluded that the change in macroinvertebrate communities as reflected in the SI
was most likely attributable to conductivity. This process identified conductivity as the major
determinant of the SI. However, it is impossible to rule out all the possible interactions with
other stressors.
Despite the outcomes of this analysis we know that other factors potentially influence the
SI, including natural taxa distribution patterns, lag effects related to previous conductivity
exposure, and localised ionic composition. Secondary salinisation is also associated with
a range of indirect environmental impacts including waterlogging, increases in nutrients,
sedimentation etc. (James et al, 2003). Therefoe it is essential that consideration be given
to the full range of potential multiple confounding influences apart from conductivity when
assessing sites using the SI.
These preliminary findings provide an indication that community changes are occurring at
concentrations lower than previously thought to be suitable for macroinvertebrates. However,
as the data are of coarse taxonomic resolution care should be taken when applying proposed
sensitivity scores to the management of risk posed by salinity on specific species. Increased
taxonomic and geographical resolution combined with a greater proportion of data from higher
conductivity (brackish and saline categories) sites and the use of abundance data instead of
presence/absence data are likely to improve the accuracy of the SI. The SI has been mapped
for the NAP priority catchments where data existed for the period 1996 to 2002 (Figures 10–14).
31
Potential impacts of salinity and turbidity in riverine ecosystems
(a)
(b)
Figure 10 Percentage of sensitive (green) and very tolerant taxa (orange) in 12 equal data
groupings with increasing conductivity for (a) edge habitat and (b) riffle habitat. Median values
with boxes corresponding to 80th and 20th percentiles and horizontal bars to maximum and
minimum.
32
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 11 Salinity index for the Mary and Burnett Catchments
33
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 12 Salinity Index for the Burdekin Catchment
34
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 13 Salinity Index for the Brisbane and Western catchments
35
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 14 Salinity index for the Queensland Murray Darling catchments
36
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 15 Salinity index for the Fitzroy catchment
37
Potential impacts of salinity and turbidity in riverine ecosystems
2.0
Turbidity Impacts
Much of Queensland’s landscape is arid and prone to extended periods of drought and periodic
flooding. Turbid streams occur naturally in Queensland, though when natural erosional
processes are accelerated streams can become highly turbid (Dodds 2001) and can result in
negative impacts on stream biota (Waters 1995).
2.1 Measurement
Turbidity is a measure of water clarity or cloudiness and is defined as the optical
property of a liquid that causes light to be scattered and absorbed rather than
transmitted in straight lines (Bruton 1985). It is thus an integrated measure of
the suspended and dissolved load and the properties of particulates held in
solution (Bruton 1985). Turbidity is most commonly measured in Nephelometric
Turbidity Units (NTU) using a turbidity meter. A Secchi disc may also be used
to measure water clarity though this method is less commonly used than a
turbidity meter in riverine environments due to a lack of depth. Secchi discs are
more commonly used in lakes and reservoirs, but are often used in conjunction
with turbidity meters. Secchi discs are mostly used in the field situation, as they
provide a low-cost and rapid measure of visual clarity and photic zone depth.
Solids held in solution that contribute to the measurement of turbidity consist of Total
Dissolved Solids (TDS) and Total Suspended Solids (TSS). In some cases the term Total
Particulate Matter (TPM) is used interchangeably with (TSS) or Non Filterable Residue (NFR). TSS
refer to the portion of a water sample able to be filtered out of solution using a predefined filter
size (typically 0.45 µm for soluble metals and typically 1.2 µm for TSS) (American Public Health
Association 1998). Dissolved substances are considered to be the residue other than water able
to pass through a predefined filter size. In some instances TSS is used to infer the turbidity of a
sample using a site-specific calibration based on a regression curve relating TSS to turbidity. The
relationship between the two can vary according to the characteristics of the suspended solids.
Fine colloids and other material can have little effect on the concentration of TSS but have a
major effect on measures of turbidity. To some extent flow rates can also affect particle size
distribution and hence the relationship between turbidity and TSS.
Particulates can include Organic Matter (OM), phytoplankton, colour, mineral content and
suspended sediment. OM can be found dissolved in solution as Dissolved Organic Matter
(DOM) or suspended in particulate form. DOM refers to the complex molecules including humic
and fulvic acids that are the products of the breakdown of OM. The colour of a sample can also
contribute to a measure of its turbidity. Despite the contributions of colour organic matter and
biological matter to measures of turbidity, sediment is often a major contributor to measures of
turbidity in Queensland streams.
38
Potential impacts of salinity and turbidity in riverine ecosystems
2.2 Sediment sources
Stream sediments can originate from outside the river channel from colluvial
processes or from alluvial processes within the channel itself. The supply of
sediment from outside the channel is highly variable and is dependent on the
erodability of the soil, land use practices, and geographical features (Wood 1997).
Out of channel sources of sediment can be reduced or exacerbated by different
land uses and land management practices (Wood 1997). Different sediment sources
can result in varying degrees of sedimentation and can have different particle size
distribution and geochemical characteristics that are related to their source.
Remobilisation of deposited sediments is the primary source of sediments from
within streams. The frequency and extent of sediment remobilisation is dependent on the
erosivity of flows and the stability of the channel bed (Wood 1997). Disturbance of stream
sediments is to some extent a natural process though it may be exacerbated in the presence
of exotic species. Carp (Cyprinus carpio) is an example of an exotic fish species that is known
to cause excessive sediment disturbance. They have a specialist feeding technique that allows
them to sieve through the sediment and source what may otherwise be under-utilised food
sources. This adaptation results in their competitive ecological advantage over many Australian
native fish species (Koehn 2004) that do not possess these capabilities. This feeding activity
results in increased suspended sediment and therefore increased turbidity (King et al, 1997).
Carp are also known to graze on the soft leaves of some aquatic plants. Loss of submerged
aquatic vegetation may reduce streambed stability allowing further re-suspension of sediments
from within the stream channel. Other exotic species such as the goldfish (Carassius auratus)
are also known to generate high levels of turbidity (Richardson et al, 1995).
Suspended solids are highly temporally variable. Very high concentrations can occur naturally
and the bulk movement of suspended solids is often associated with peak flow events.
Many aquatic organisms are able to survive these short-term exposures to the very high
concentrations although many may be impacted by subtle increases in the duration and
magnitude of exposure that may be due to accelerated erosion processes.
2.3 Physical and chemical aspects
Fine particulates can remain in suspension when there is enough water turbulence to keep
them in suspension otherwise they tend to settle out on the streambed. The rate of
re-suspension is proportional to the size of the particles and the volume and velocity of water
passing across the streambed (Waters 1995). Suspended matter may be dissolved or in a
particulate phase. Suspended matter can precipitate out or remain in suspension indefinitely
as a colloid. A ‘colloid’ is characterised as being a substance having particles in solution
approximately 100 to 10000 nm diameter (Thain and Hickman 2000). An increase in water
salinity can cause a colloidal substance to precipitate out of solution through a process of
coagulation. In this case, the addition of an electrolyte causes neutralisation of the charge of
surface particles thereby removing the electrostatic repulsion forces between particles causing
them to collide and coagulate and thus precipitate out of solution. There are many naturally
39
Potential impacts of salinity and turbidity in riverine ecosystems
and perennially turbid waters in Queensland, some of which include the Fitzroy River, Cooper
Creek and the Warrego River.
The size, shape, and chemical characteristics of particulates in rivers are highly variable
(Wood 1997). Sediments can have different ecological and biological impacts depending
on their geochemical and physical characteristics. The nature and extent of impacts from
suspended particulates can also vary according to physical dynamics of the system within
which they are transported. The following sections describe some common sediment
characteristics and the relevant biological implications.
2.3.1 Fine sediment
Fine materials suspended in the water column are responsible for reduced light penetration,
gill clogging and reduced visibility (refer to Figure 16). Fine suspended particles have a high
capacity for ion exchange that causes them to bind directly with biological membranes such
as fish and invertebrate gills resulting in clogging (Bond and Downes 2003).
2.3.2 Erosivity
Increased velocity of suspended particles enhances their erosive potential to
scour streambeds and banks, resulting in sediment re-suspension and increased
stream sediment loads. Fast flows remobilise large particles from in-stream
deposits that would otherwise not be mobilised under normal flow regimes
(Wood 1997). This scouring effect can result in the loss of habitat suitable for
reproduction.
As flow and turbidity can have independent and cumulative effects on aquatic
biota it is difficult to separate the effects of increased flow velocity from that of
turbidity. Bond and Downes (2003) found that when flow rates and suspended
sediment concentrations were increased in artificial streams simultaneously and separately,
flow was a greater determinant than turbidity in reducing the numbers and diversity of
macroinvertebrates. In this study the numbers and diversity of macroinvertebrates were not
observed to be significantly affected by the addition of fine sediment in concentrations similar
to those observed in upland streams in southeast Australia. However, gross sedimentation
clearly has an impact on macroinvertebrate communities and the limited biotic response
to sediment reported in this study is different to those reported in other studies where the
response to sediment was distinctive (Doeg et al, 1987, Doeg and Milledge 1991). A possible
explanation for the difference in response may be due to the particular size distribution of
fine sediments used in the study by (Bond and Downes 2003) and the time scale over which
the response of the organisms was measured. The particles used by (Bond and Downes 2003)
were larger (500-1000 µm) than by Doeg and Milledge (1991). The evidence from these studies
suggest that the most detrimental effects on macroinvertebrates are in the first instance most
likely to be due to the finer silt and clay particles.
40
Potential impacts of salinity and turbidity in riverine ecosystems
Figure 16 Conceptual model of in-stream processes and the potential effects of increased turbidity on aquatic
ecosystems
41
Potential impacts of salinity and turbidity in riverine ecosystems
2.3.3 Contaminant interactions with fine sediment
Contaminants can interact with suspended particulates through adsorption, desorption
and transport processes. Fine colloidal particles can include particulate OM and silt and clay
particles. Silt and clay have a very high ion-exchange capacity allowing them to bind with
contaminants such as nutrients and metals whilst particulate OM has an affinity to adsorb
organic contaminants.
The tendency of a particulate to bind with a contaminant is dependent on the chemical
affinity of the contaminant for the particulate. For example, hydrophobic (mid-polar to nonpolar) contaminants have a stronger affinity to bind with organic particles than with silts
and clays. Given that hydrophobic organic contaminants readily sorb to carbon sources, and
organic carbon can exist either as a solid or a liquid, dissolved contaminants sorbed to carbon
molecules can also exist in either phase.
Adsorption and desorption processes are related to the bioavailability of contaminants, and
hence their toxicity to aquatic life. Bioavailability is dependent on the tendency of a compound
to be associated with the particulate or aqueous phase. Contaminants in the aqueous phase are
more bioavailable and subsequently more toxic to aquatic life. Contaminants in the solid phase
are generally less bioavailable and less toxic to aquatic life. The bioavailability of a contaminant
associated with particulate matter is to some extent dependent on whether or not the
particulate they are bound with is in a dissolved or in a particulate form.
Individually, fine particles have a very large surface area relative to their volume. Collectively
this large surface area allows large volumes of organic and inorganic contaminants to be
adsorbed onto their surface. Contaminants sorbed to the surface of particles are subject to the
same transport processes as the sediments they are associated with. Contaminants bound in
the particulate phase in sediments can leach into sediment pore waters and become toxic to
benthic fauna. They may also enter overlying waters through diffusion where they may be toxic
to aquatic life (Batley in Boulton and Brock 1999).
2.3.4 Light penetration and temperature
The primary source of energy in virtually all aquatic ecosystems is light
(Boulton and Brock 1999). Light is required by photosynthetic organisms to
convert inorganic compounds (carbon dioxide CO2) into organic compounds
(carbohydrates CH2O) via photosynthesis (ANZECC/ARMCANZ 2000).
Submerged macrophytes rely upon light penetrating the water column to reach
them for photosynthesis to occur and where turbidity is high, light cannot
reach them and they cannot survive. A Natural Resources and Mines study
(unpublished data) of the distribution of macrophytes within the Condamine
Balonne River in the Murray Darling catchments, found that submerged species
were not present where turbidity was greater than 20–30 NTU. In another
study by Blanch et al (1998) highly turbid water was found to prevent the growth of Vallisneria
americana in the Murray Darling. Decreased water temperatures due to decreased light
42
Potential impacts of salinity and turbidity in riverine ecosystems
penetration in the upper surface of the water column may also affect temperature-sensitive
species by altering breeding cues or in some cases through direct physiological effects. High
concentrations of turbidity in standing waters can also result in temperature stratification. This
phenomenon is due to increased light scattering at the surface of water resulting in a reduction
in the penetration of light and a subsequent reduction in solar heating of water (Ryan 1991).
A reduction in light penetration will result in a net decrease in photochemical processes
including the breakdown of contaminants by photolysis. This may reduce the capacity for
waters to naturally degrade contaminants including many herbicides and pesticides to be
broken down in natural systems.
2.3.5 Gill flushing
Fine sediments can inhibit respiration by clogging the gills of fish by blocking invertebrate
feeding membranes. Some fish species and some invertebrates can expel water in the reverse
direction across their gills to flush sediments from their gills. However, this action requires
energy expenditure that when prolonged can result in depletion of energy reserves and
ultimately death. Species such as the burrowing thalassinidean shrimps Nihonotrypea japonica
and Upogebia major employ a passive gill cleaning mechanism. By frequently moving limbs,
they are able to remove particulates from their gills allowing them to tolerate suspended
particulate matter (Batang and Suzuki 2003).
2.4 Effect on in-stream biota
There are many ways in which excess stream sediments can have direct
physiological impacts on aquatic organisms. Direct impacts are those that act at the
organism level to reduce its chances of survival and/or its viability and can include
mortality, reduced physiological function, avoidance, depressed rates of growth,
reproduction and recruitment (Henley et al, 2000). Indirect effects are those that
can act at the ecosystem level to reduce the chances of survival and can include the
interruption of food webs through, for example, the loss of a food source or the loss
of habitat. The following sections describe the current scientific understanding of
how turbidity-related impacts affect aquatic ecosystems, and provide a summary
of this conceptual understanding against which changes to aquatic ecosystems
exposed to elevated turbidity may be assessed.
2.4.1 Tolerance of individuals
There is very little tolerance information available for turbidity. Part of the reason for this is
that it is difficult to determine a standard for turbidity due to the highly variable properties
of suspended particulates that contribute to measures of turbidity. Without a clear definition
of what turbidity is, it is not possible to undertake comparative toxicity studies. Another
complication is that suspended particulates are known to cause physical effects through
abrasion and smothering. Therefore traditional dose-response relationships do not provide
generic response models applicable all the effects due to measures of turbidity. For these
reasons, turbidity tolerance ranges determined in laboratory experiments provide only a partial
explanation of the potential impacts.
43
Potential impacts of salinity and turbidity in riverine ecosystems
Some studies have derived tolerance ranges using ‘standardised’ concentrations of suspended
particulates (Herbrandson et al, 1999). While these studies are relevant to those conditions used
in the test, they may not be applicable to turbidity conditions different from that
re-created in the test. Standardised tests do provide a relative indication of species tolerance
that is useful for comparative purposes though care should be exercised in applying turbidity
effects data to conduct localised assessments of turbidity risk. The thresholds for change and
the intensity of stress on stream ecosystems as a result of impacts from sediment transport
have not been fully quantified and local targets for rehabilitation remain undefined. Although
turbidity has the potential to affect virtually all stream biota, only the major groupings, for
which there are sufficient data available to make valid conclusions, are discussed here.
2.4.2 Invertebrates
The literature indicates that invertebrates are generally affected by turbidity. However, there are
few determinations of the likely impacts on abundance from which to base predictions of likely
effects. A field study by Campbell and Doeg (1989) found that macroinvertebrate community
structure was affected by forestry practices in general. However, in this case a direct linkage
between the presence/absence of field caught organisms and increased turbidity attributable
to forestry practices was only loosely defined. This was because there were many other water
quality impacts associated with forestry practices other than just an increase in turbidity.
These included changes in temperature, sediment loads, nutrients and flow all of which may
result in changes to macroinvertebrate community structure. Although this study could not
separate out the effects that turbidity and sediments may have had on the macroinvertebrate
community structure from the other water quality parameters, it did show that changes in
turbidity have the potential to affect macroinvertebrate community structure.
Another study by Quinn et al (1992) conclusively demonstrated that fine inorganic suspended
particulates downstream of a gold mine in New Zealand did affect macroinvertebrate
community structure. In this case, the gold mine discharged only suspended sediments into
a stream in a catchment that was otherwise not impacted by other land uses. Samples were
taken upstream (mean turbidity of 7 NTU) and downstream of sedimentation impacts (mean
turbidity of 145 NTU). Invertebrate abundance was significantly lower at all downstream sites
ranging from 9% to 45% of the abundance observed at matched upstream sites. Taxonomic
richness was also significantly lower at downstream sites. The authors suggested that higher
turbidity at the downstream sites was associated with lower biomass and productivity, and
hence degraded food sources. In a review by Chutter (1969) it was suggested that complete
smothering of streambeds is required to cause large reductions in macroinvertebrate numbers,
but that smaller streambed changes could bring about a shift in species.
One group of macroinvertebrates that may be particularly vulnerable to increased turbidity
concentrations are filter feeders (e.g. freshwater bivalves). The rationale for this is that increased
particulates may inhibit feeding. A study by Aldridge et al (1987) found bivalve feeding was
impaired in lab experiments that simulated increased suspended sediment loads and water
turbulence.
44
Potential impacts of salinity and turbidity in riverine ecosystems
2.4.3 Vertebrates
There is generally limited information about the effects of turbidity on vertebrates though
there is some information available for fish. Impacts on fish are often difficult to detect and are
of a chronic nature. In sensitive species clogging of the gills often results in death (Ryan 1991).
A study by Schulz (1996) found that the number of fish species was negatively correlated with
the transport of bed loads. The major effect on fish populations was suspected to be due to
a reduction in food sources and the ability to feed. In another study by Russell et al (2003) of
spatial variation in fish assemblage structure of the Barron and the Mitchell river systems in
Northern Queensland there was some evidence that suspended solids influenced the diversity
of fish species present in the Barron. Other authors including Blaber and Blaber (1980) and
Cyrus and Blaber (1987) have also observed turbidity gradients to be important factors in
structuring fish communities especially in estuarine environments.
2.4.4 Change in species composition
Despite the limitations of high-turbidity environments, some animals have developed the
ability to survive in low-light conditions making them very successful in waters that are highly
turbid. For example some species of fish possess highly developed olfactory systems and
barbels to facilitate prey detection and foraging success allowing them to thrive in highly
turbid environments (Pusey et al, 2004). For animals with highly developed olfactory systems
(senses of smell) it may only take very limited stimuli for the detection of food or the avoidance
of predators. An example of such a species is Hyrtl’s Tandan (Neosilurus hyrtlii) that is known
to inhabit turbid and relatively clear streams. Burrows et al (1999) in Pusey et al (2004) found
Neosilurus hyrtlii to occur in turbid environments (as high as 581 NTU in the Belyando River).
In a study by Natural Resources and Mines this species was found to be highly abundant in
Cooper Creek where turbidity can reach approximately between 1500 and 2000 NTU (Natural
Resources and Mines unpublished data).
Turbidity is also likely to affect the species composition of submerged macrophytes.
A point touched on earlier in section 2.3.4 referring to the effect of light penetration, was that
submerged macrophytes are affected by high concentrations of turbidity. A study by Natural
Resources and Mines (unpublished data) of the prevalence of submerged macrophytes in
the upper parts of the Condamine River found that macrophyte species with floating leaves
or emergent growth forms were found throughout the catchment whereas submerged
macrophytes were found only in the upper parts of the catchment where turbidity is
comparatively less than in the lower parts of the catchment.
In addition to the changes in fish and aquatic macrophyte species, increased turbidity and
sediment deposition has been shown to negatively affect the composition of species found
in the benthic zone. The ability of a species to survive in highly turbid environments is often
species-specific and may be related to their ability to avoid starvation due to reduced filtration
and an increased requirement to clear sediment from their gills (Waters 1995).
45
Potential impacts of salinity and turbidity in riverine ecosystems
2.5 Impacts on aquatic ecosystems
There are many ecosystem-level impacts that can indirectly impact individuals
within an ecosystem affected by stream sedimentation. Reduction in the visual
clarity of water and sedimentation can alter the structure and functioning of
aquatic ecosystems. In some circumstances where turbidity is high light can
become a limiting factor in the functioning of an ecosystem. Stream habitat can
be completely lost by the smothering of benthos by sedimentation. Predator/
prey interactions may also be altered for those animals that rely on sight for the
detection and avoidance of prey.
Many of Queensland’s river systems have naturally high concentrations of
turbidity making it difficult to determine how much of a change in visual clarity and stream
sedimentation is acceptable. It is when turbidity rises above normal threshold concentrations
that it may become a problem. Benthic organisms may be smothered at concentrations that are
much lower than is likely to have a direct effect on fish thus reducing the quantity of available
food sources, but also making the food source harder to detect (Ryan 1991). In this example,
food web interactions are of greater impact to the ecosystem than species tolerance.
Recovery from the effects of short pulses of suspended sediments can be rapid, once the
source of contamination is removed and provided that the pulse was not of a magnitude that
may be catastrophic to the ecosystem (Ryan 1991). However, prolonged inputs of suspended
particulates are likely to have a long-term impact on ecological integrity and in some cases
may be irreversible in the short to medium term. Especially given the long time periods that
the in-stream sediment slugs may take to pass through a river system. It is essential that excess
turbidity in aquatic ecosystems be recognised at an early stage to allow the implementation of
strategies to reduce the inputs of sediment in streams.
2.5.1 Light limitation and primary productivity
Primary production forms the basis of most aquatic food webs and as light is a major factor
governing rates of in-stream primary production, decreased light penetration due to high
turbidity can limit aquatic primary production. This scenario can result in an ecosystem that
is light-limited. A complication to this conceptual model of the effect of light limitation on
primary productivity is that some ecosystems have food webs driven by carbon consumers
not by photosynthesis and also in some circumstances highly productive bands of algae can
occur in the photic zones of highly turbid waters allowing in-stream productivity to remain
high (Bunn et al, 2003). However, in the absence of narrow bands of productive algae in the
upper part of the water column, reduced light penetration can in some cases eliminate primary
productivity altogether (Dodds 2001, Henley et al, 2000) resulting in follow-on effects within
food webs (Wood 1997).
Lloyd et al (1987) developed a model that related stream turbidity to measures of Gross Primary
Production (GPP). In this study turbidity levels as low as 5 NTU were shown to decrease primary
productivity by 3% to 13% in streams with low background concentrations of turbidity. In
46
Potential impacts of salinity and turbidity in riverine ecosystems
addition to reduced primary productivity, the study also showed that turbidity was linked to
negative impacts on benthic macroinvertebrate and fish communities. It is not clear in this
case whether turbidity acted directly at the organism level to cause these effects or whether
turbidity acted indirectly at the ecosystem level to impact on benthic macroinvertebrate and
fish communities.
2.5.2 Stream habitat
The loss of aquatic habitat from sedimentation is known to have deleterious and sometimes
irreversible effects on aquatic ecosystems and the food webs that they support (Figure 16).
Sediment deposition can modify the characteristics of stream substrate and can smother the
substrate removing food sources and preventing the movement of solutes in the hyporheic
zone. This limited solute transport results in a net reduction in nutrient and carbon cycling
processes (Waters 1995).
Loss of habitat due to increased sedimentation can also result in a loss of biodiversity and a
decrease in the aesthetic value of aquatic resources. For example, reproductive cycles of fish
can be interrupted by the loss of habitat as many egg-laying fish rely on suitable habitat for
successful breeding (Pusey et al, 2004, Pusey et al, 1993). Suspended and deposited sediment
may also alter fish community composition by interfering with riffle-run-pool sequences and
preventing migration into preferred habitats for spawning. Smothering of stream substrate
is also known to alter the distribution of macroinvertebrates. Typically, streams subjected
to increased sediment loads have a less diverse macroinvertebrate fauna (Water and Rivers
Commission 2000). Macroinvertebrates such as caddisflies, stoneflies and mayflies have
preference for clean gravel riffles and become less abundant where these types of habitat are
lost through sedimentation. Conversely, worms and midge larvae, which prefer fine sediment,
can become more abundant where there is high sedimentation (Waters 1995).
Loss of stream habitat through streambed blanketing is most likely to occur where sediment
loads are high and are periodically transported by large volumes of water then deposited in
the stream channel. Streambed blanketing may not be the primary impact for regulated rivers
where sediment loads are held in impoundments with water released gradually over time.
The most pronounced changes to predator/prey interactions are likely to occur in waters that
normally have low turbidity that, on some occasions, receive large quantities of sediment.
Given a constant exposure to turbidity, it is expected that pre-exposed populations are likely to
experience much less impact than a stream community that resides in relatively clear streams.
2.5.3 Avoidance
A study by Richardson et al (2000) found that the Banded Kokopu (Galaxias fasciatus) from New
Zealand avoided turbidity levels of 25 NTU during migration. A study by Russell et al (2003)
provided evidence that fish distributions in the Barron catchment were affected by suspended
solids. There is also some evidence that increased sediment loads increase macroinvertebrate
drift by inducing night like darkness and triggering dispersal. Ryan (1991) found that an
increase in suspended solids can increase macroinvertebrate drift and may reduce benthic
47
Potential impacts of salinity and turbidity in riverine ecosystems
densities as well as community structure. Experiments by Ryder (1989) showed a sudden
increase in the drift densities of stream insects when sediment was artificially introduced into
streams. Species that cannot travel long distances and are not able to avoid increased turbidity
and sedimentation may be most susceptible to these effects.
2.5.4 Food web interactions
Particulates settling on periphyton can result in the loss of food for invertebrates that in turn
has follow-on effects for the normal functioning of food webs. A study by Broekhuizen et al
(2001) employed a novel technique of contaminating periphyton with radio-labelled sediment
(containing C14) prior to its use in grazing tests. Groups of snails and mayflies grazed upon
different proportions of the radio labelled sediment and periphyton ratios. Growth rates were
found to be significantly lower and mortality higher at sediment ratios above 50 parts sediment
to 1 part periphyton. Intermediate levels of sedimentation were found to provide ideal growth
conditions for the snail and mayfly, as very low proportions of sediment were found to inhibit
growth rates, possibly due to reliance upon nutrients gained from sediment ingestion.
Other food web interactions from increased turbidity can include the disruption of normal
predator/prey interactions. Species interactions of predator/prey relationships are known
to play a major role in structuring aquatic communities (La Point et al, 1996). Suspended
sediments can inhibit the detection of food sources and predators making animals vulnerable
to predation and in some instances unable to locate food. Also in some cases, the ability of
predator and prey to detect each other is frequently impaired by turbidity (Abrahams and
Kattenfield 1997). Many fish species rely on sight to detect their prey (Abrahams and Kattenfield
1997). For these species prey must be within close range for successful feeding to occur. A study
by Granqvist and Mattila (2004) found that increased turbidity decreased the foraging success
of juvenile perch (Perca fluviatilis).
Likewise, predator avoidance can become ineffective at high concentrations for some animals
(Abrahams and Kattenfield 1997). The effectiveness of predator avoidance mechanisms
decreases the closer a predator is to its prey. Miner and Stein (1996) demonstrated that the
reactive distance of Bluegill Sunfish (Lepomis macrochirus) to their predator (Large Mouth
Bass, Micropterus salmoides) declined from more than 200 cm in clear water to 23 cm in turbid
water (10 NTU). In this case an increase in turbidity increased the risk of predation. At a local
scale the effect of increased turbidity on predator/prey interactions and its effect in structuring
communities is difficult to predict but is likely to be dependent on the morphological features
possessed by the species present to avoid predation and seek food.
48
Potential impacts of salinity and turbidity in riverine ecosystems
3.0
Determining acceptable concentrations
for Salinity and Turbidity
Elevated concentrations of salinity and turbidity are associated with a loss of biodiversity
and a decline in the health and integrity of aquatic ecosystems. Given a limited change
from background conditions their impacts may be subtle and difficult to determine without
undertaking an assessment of ecological condition. In circumstances whereby a change occurs
that is high magnitude when compared with background conditions, changes are likely to be
stark and easily distinguishable with the naked eye. However, once a change of high magnitude
has occurred it may be difficult to rehabilitate the system to its former condition. Therefore,
it is essential that the impacts from salinity and turbidity are detected at an early stage and
managed appropriately to prevent decline in the ecological health of waterways. The ability to
quantify subtle ecosystem changes and to determine acceptable concentrations will help to
provide relevant and biologically meaningful targets for salinity and turbidity.
3.1 Existing salinity and turbidity guidelines
The national water quality guidelines (ANZECC/ARMCANZ 2000) (i.e. ‘the guidelines’)
do not specify definitive values for salinity and turbidity. Rather, they provide general
concentrations for guidance at a regional scale. Regional default trigger values
recommended in the guidelines are classified into two zones relevant to Queensland.
These are southeast Australia and tropical Australia (refer to Table 7 and Table 8 in
Appendix 2). When considering these values it should be noted that, in circumstances
where these guidelines are exceeded it is recommended that this should not cause
maximum alert but rather it should trigger further investigations to establish if
impacts are occurring at the point of interest.
The regional reference ranges recommended in the national water quality guidelines are
not appropriate for target setting within the NAP context. The regional reference thresholds
provided in the guidelines have been derived using deviations of percentiles from broad
regional reference conditions. These have been grouped into like regions including upland and
lowland zones, lakes and reservoirs, and estuarine zones. This zonation reflects natural patterns
in salinity and turbidity within the same river or creek system. For example, in the upper part of
a catchment, salt concentrations are typically less than those found in the coastal plains. While
these ranges have broad-scale relevance they may not account for localised variability. In
fairness they are not meant to, rather they are meant to provide general broad‑scale guidelines
that should be refined for local relevance.
3.2 Regional target setting
In developing targets the following points should be considered to ensure compliance
with regional Natural Resource Management (NRM) target setting guidelines (Queensland
Government 2004). Targets should be accurate and locally relevant using indicators that are
sensitive to changes in resource condition. They should have the capacity to provide early
49
Potential impacts of salinity and turbidity in riverine ecosystems
warning mechanisms and indicate where potential degradation is likely to occur. The indicators
used need to be clear, scientifically defensible, take a common sense approach and must be
capable of being monitored and suited to the resources available for such a task. Ideally they
would complement existing information and monitoring programs, and should be able to be
undertaken in a cost-effective way to monitor resource condition and trend whilst requiring
minimal technical difficulty to interpret.
3.2.1 Applying a risk assessment approach
A risk-based methodology of determining safe concentrations that link causes
with effects is supported by the national water quality guidelines (ANZECC/
ARMCANZ 2000). In principle this approach involves a determination of the
relative risk from salinity and turbidity concentrations by comparing existing
sensitivity data (consequence data) with the presence of salinity and turbidity
(likelihood data) (refer to Figure 17). In such a model it is anticipated that species
sensitivity distributions would be derived from toxicological studies using
selected individual species deemed to be representative of an ecosystem. These
results when combined would be used to predict ecological effects at various
exposures.
This risk-based approach is sound and is the approach used for most conservative
environmental contaminants. However, in the case of salinity and turbidity determining ‘safe
concentrations’ is made difficult by their highly variable nature that can vary considerably
spatially and temporally, and by the range of different effects observable at the individual and
the ecological level. Hence, there is currently no single dose-response model available that can
accurately depict all the effects likely to be observed in all ecosystems found in Queensland.
The use of a generic dose-response model that could be adjusted to suit all ecosystems and
biota within their local conditions would be ideal. In order to develop a generic dose-response
model, further investigation is required that quantifies the effect of all the factors that act to
compound or ameliorate the effects of salinity and turbidity at a local scale.
In addition to the use of individual tolerance ranges it is important to include effects that
may occur at an ecological level in an overall risk assessment model. As measures of EC and
NTU are integrative measures of salinity (total ions) and water clarity respectively and the
properties of a solution contributing to these measures can vary in their composition, some
correction or adjustment may be required to suit localised conditions. There may also be a
requirement for consideration of previous condition and the natural status of the ecosystem
as for example there are many riverine systems in Queensland having naturally high turbidity.
Exposure dynamics driven by flow rates can affect the duration and magnitude of exposure to
salinity and turbidity and hence should also be considered in such a model. Information that
characterises exposure dynamics could include hydrological and geochemical models of the
system of interest. Any other factors that characterise the risk of salinity and turbidity to aquatic
ecosystems should also be quantified where possible to be considered in the risk assessment
model. In this way, trigger values may be recalculated or adjusted at the catchment scale using
50
Potential impacts of salinity and turbidity in riverine ecosystems
locally relevant biota according to localised conditions. The risk approach should ideally be
able to be monitored and reviewed using localised biological studies that may include Direct
Toxicity Assessment (DTA) and/or other field based biological studies. An iterative risk based
approach such as recommended in the national water quality guidelines applies a logical and
defensible framework for setting targets for salinity and turbidity.
In the interim and prior to this information having been collected, a base model could be
constructed using the available information to which new information may be added as
it becomes available, improving the accuracy of the model over time. Also the regional
NRM target setting guidelines provide some guidance for the setting of targets for salinity
(Queensland Government 2004). The report shows a practical example of how targets may be
set for salinity based on the application of available condition and trend information to identify
desired targets. The report also provides guidance on appropriate approaches to be taken in
the absence of suitable condition and trend information.
Figure 17 Risk assessment model
51
Potential impacts of salinity and turbidity in riverine ecosystems
Discussion
Although salinity and turbidity are natural components of aquatic ecosystems,
it is clear from the literature that excessive concentrations above background
levels may result in measurable ecological consequences in freshwater
ecosystems. Therefore they may be regarded as contaminants in freshwater
environments. They are known to have direct toxic effects and to have indirect
ecological effects on freshwater biota above certain thresholds. Direct impacts
are those that act at the organism level to reduce an individual’s chances of
survival and or its viability and can include mortality, reduced physiological
function, avoidance, increased susceptibility to disease or predation, depressed
growth rates, and reproduction and recruitment. Indirect effects are those
that can act at the ecosystem level to modify habitat and alter biotic interactions between
and within trophic levels, in turn reducing the chances of survival for individuals through the
interruption of food webs or the loss of habitat. These alterations can translate into changes
in ecosystem functioning, food web interactions and impacts on in-stream biodiversity. The
dose and duration of exposure to these contaminants are important factors that determine
their ecological scale and individual scale impacts. Salinity and turbidity impacts may also
be confounded by changes in physicochemical characteristics of streams that can co-occur
with increasing salinity and turbidity. They may also be confounded by the presence of
contaminants such as nutrients, herbicides and pesticides that can either ameliorate or
accentuate their impact.
As salinity and turbidity are naturally occurring ubiquitous components of aquatic ecosystems,
some aquatic organisms have developed behavioural and physiological mechanisms to tolerate
elevated concentrations. One of the difficulties in determining safe concentrations is separating
the effect of elevated concentrations from those of natural or background concentrations. We
have shown that for waterways having naturally very low turbidity or salinity, a slight increase
is likely to have a far more pronounced impact in that system than the same increase would in
a stream having a naturally high turbidity or salinity. As salinity and turbidity increase above
background concentrations, it would be expected that sensitive taxa would be replaced by
tolerant taxa, thereby altering community structure in affected streams. The Salinity Index
(SI) for macroinvertebrates in Queensland streams presented here confirms that a change in
community structure does occur with increased salinity. The SI also indicates that changes in
macroinvertebrate community structure may occur at salinity concentrations much lower (800
to 1000 µS cm-1) than previously suggested in the scientific literature (1500 µS cm-1). Broad
patterns in salinity concentrations in Queensland were found to align closely with regional
catchment boundaries. The salinity zones identified here may be used to identify sites or subcatchments where salinity is unusually high or low compared with the rest of the region.
It is also evident that as salinity is a representative measure of anions and cations in solution,
and that the composition of ions has an appreciable effect on biological impacts, measures of
salinity alone may not be adequate for the interpretation of biological impacts. Also, from the
52
Potential impacts of salinity and turbidity in riverine ecosystems
investigation into the ionic composition of surface waters in Queensland, it was observed that
most patterns were found to be consistent with geology and climate. The available literature
suggests that sodium (Na+) and potassium (K+) ions are more toxic than divalent calcium (Ca2+)
and magnesium (Mg2+) ions. Therefore, higher proportions of sensitive taxa may be found in
calcium bicarbonate dominated water rather than in sodium chloride dominated water under
equal conductivities, placing these ecosystems at a greater risk if conductivities increase.
As there are many different ways in which salinity and turbidity can impact on aquatic
ecosystems and given that these can occur at multiple scales, it is a challenging task to
determine ‘safe’ or ‘acceptable’ concentrations and to set appropriate targets for them. Given
this uncertainty, an ecological risk assessment model is an effective technique for assessing
the relative risk of salinity and turbidity in freshwater aquatic ecosystems. The development
of a risk assessment model has a high information demand and may require for example, a
knowledge of the aquatic biota that inhabit a given ecosystem, their maximum tolerance
ranges, characterisations of the composition of salinity and/or turbidity, characterisations of the
exposure concentrations and dynamics, and conceptualised models of ecosystem response.
The development of knowledge to meet these information requirements will be essential to the
successful implementation of such a model and the management of salinity and turbidity.
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Appendix 1 Glossary of Terms
62
additive effect
an additive effect is the overall consequence which is the result of two chemicals acting together and
which is the simple sum of the effects of the chemicals acting independently
adsorption
the adhesion of the molecules of gases, dissolved substances, or liquids in more or less concentrated
form, to the surface of solids or liquids with which they are in contact
alluvial
describes unconsolidated material such as sand, gravel, and silt which has been deposited by flowing
water
anion
a positively charged particle
antagonistic effect
opposing or neutralising or mitigating an effect by contrary action
aquatic ecosystem
any watery environment from small to large, from pond to ocean, in which plants and animals interact
with the chemical and physical features of the environment
aqueous
pertaining to, similar to, containing or dissolved in water
autotrophs
a group of organisms capable of obtaining carbon for synthesis from inorganic carbon sources such as
carbon dioxide and its dissolved species (the carbonates). This group includes plants and algae
bed load
sediment which moves along and is in contact with stream or river bottom
benthic
referring to organisms living in the sediments of aquatic habitat
bioaccumulation
accumulation of a substance in a living organism as a result of its intake in the food and also from the
environment
bioavailability
the degree to which an agent, such as a drug or nutrient, becomes available at the site of activity in
the body
biota
the sum total organisms of any designated area
bioturbation
the rearrangement of sediments by organisms that burrow through them and ingest them
cation
a negatively charged particle
colloid
a substance that remains suspended in a solution or fails to settle out of solution
colluvial
a loose deposit of rock debris accumulated through the action of gravity at the base of a cliff or slope
community
an assemblage of organisms characterised by a distinctive combination of species occupying a
common environment and interacting with each other
concentration
the quantifiable amount of a chemical, in for example, water food or sediment
contaminant
any physical, chemical, biological, or radiological substance or matter that has an adverse effect on
air, water, or soil
desorption
the opposite process of adsorption, the removal of the excess concentration of the adsorbate from the
vicinity of the solid surface
diadromous
describes or refers to fish that migrate between fresh and salt waters
diatoms
diatoms are a large and diverse group of single-celled algae
direct toxicity assessment
the use of toxicity assessment to determine the acute and/or chronic toxicity of the total or whole of
a substance
dose
the quantifiable amount of a material introduced into an animal
early-life-stage test
observable effects of exposure to a contaminant of the early life stage of a species, for example,
shortly after fertilisation, through embryonic development, larval and early juvenile stages of
development
effective concentration
the concentration of material in water that is estimated to be effective in producing some lethal
response
electrical conductivity
a measure of the ability of water to conduct electrical current
endocrine system
a system of ductless glands that regulates bodily functions via hormones secreted into the
bloodstream. The endocrine system includes the hypothalamus, pituitary gland, thyroid, adrenal
glands and gonads (ovaries and testes)
epithelial membrane
cell membrane covering internal organs
exposure
the amount of physical or chemical agent that reaches a target or receptor
euryhaline
tolerant of a wide range of salinity
fluvial
deposits of parent material laid down by rivers and streams
guideline trigger values
these are the concentrations (or loads) of the key performance indicators measured for the
ecosystem, below which there exists a low risk that adverse biological (ecological) effects will occur.
They indicate a risk of impact if exceeded and should ‘trigger’ some action, either further ecosystem
specific investigations or implementation of management or remedial actions
Potential impacts of salinity and turbidity in riverine ecosystems
guideline (water quality)
numerical concentration limit or narrative statement recommended to support and maintain a
designated water use
habitat
the place where a population (e.g. human, animal, plant, microorganism) lives and its surroundings,
both living and non living
heterotrophs
organisms that break down and use organic matter
hydrophobic
lacking affinity for water, or failing to adsorb or absorb water
hydrophylic
having an affinity for water, readily absorbs water
hyper-osmotic
animals are said to be hyper-osmotic when they are capable of surviving in ionic concentrations
greater than their internal concentration
hypersaline
salinities in excess of that commonly found in oceanic sea water, or greater than 35 parts per
thousand
hypo-osmotic
animals are said to be hypo-osmotic when they are capable of surviving in ionic concentrations lower
than their internal concentration
hyporheic zone
the hyporheic zone is the area under or beside a stream channel or floodplain that contributes water
to the stream
invertebrate
an animal without a backbone or spinal column
lentic
refers to standing or still waters such as lakes
LC50
The concentration of material in water that is estimated to be lethal to 50% of the test organisms.
The LC50 is usually expressed as a time dependent value
lotic
refers to flowing waters such as rivers
macrophyte
an aquatic plant that can be seen without the aid of a microscope
microbes
the term ‘microbes’ is a general term that is used to encompass viruses, bacteria, protozoa, rotifers,
fungi, slime, moulds, lichen and algae
olfactory system
the olfactory system is the sensory system used for olfaction. Olfaction, the sense of smell, is the
detection of chemicals dissolved in air (or, by animals that breathe water, in water)
osmoregulation
any mechanism in animals regulating a concentration of solutes within its cells or body fluids, or b)
total volume of solutes within its body
osmosis
diffusion of a solvent through a semi-permeable membrane into a more concentrated solution,
tending to equalise the concentration on both sides of the membrane
partitioning
splitting of target object into smaller units
pesticide
a substance or mixture of substances used to kill unwanted species of plants or animals
pH
value that represents the acidity or alkalinity of a solution, defined as the negative logarithm of the
hydrogen ion concentration of the solution
primary production
the production of organic matter from inorganic materials
secondary salinity
human induced salinity due to shallow groundwater from irrigation or other inputs
species
a group of organisms that resemble each other to a greater degree than do members of other groups
and that form a reproductively isolated group that will not produce viable offspring if bred with
members of another group
suspension
a system in which very small particles (solid, semi-solid, or liquid) are more or less uniformly
dispersed in a liquid or gaseous medium
synergism
a phenomenon in which the toxicity of a mixture of chemicals is observed to be greater than the
individual or additive effects of its individual components
tolerance
the ability of an organism to withstand adverse or other environmental conditions for an indefinitely
long exposure without dying
watertable
the level of groundwater, the upper surface of the zone of saturation for underground water
63
Potential impacts of salinity and turbidity in riverine ecosystems
Appendix 2
Table 7 Ranges of default trigger values for conductivity (EC, salinity), turbidity and
suspended particulate matter of slightly disturbed ecosystems in southwest Australia.
Ecosystem
type
Salinity
(µS cm-1)
Explanatory notes
Upland and
lowland rivers
120–300
Conductivity in upland streams will vary depending upon catchment geology. Values at the lower end
of the range are typically found in upland rivers, with higher values found in lowland rivers. Lower
conductivity values are often observed following seasonal rainfall.
300–1500
Values at the lower end of the range are observed during seasonal rainfall events. Values even
higher than 1500 µS cm-1 are often found in saltwater lakes and marshes. Wetlands typically have
conductivity values in the range
500–1500 µS cm-1 over winter. Higher values (3000 µS cm-1) are often measured in wetlands in
summer due to evaporative water loss.
Lakes,
reservoirs and
wetlands
Turbidity
(NTU)
Upland and
lowland rivers
Lakes,
reservoirs and
wetlands
Estuarine and
marine
10–20
Turbidity and SPM are highly variable and dependent on seasonal rainfall runoff. These values are
representative of base river flow in lowland rivers.
10–100
Most deep lakes and reservoirs have low turbidity. However, shallow lakes and reservoirs may have
a higher turbidity naturally due to wind-induced resuspension of sediments. Lakes and reservoirs in
catchments with highly dispersible soils will have high turbidity. Wetlands vary greatly in turbidity
depending upon the general condition of the catchment or river system draining into the wetland
and to the water level in the wetland.
1–2
Turbidity is not a very useful indicator in estuarine and marine waters. A more appropriate measure
for WA coastal waters is light attenuation coefficient. Light attenuation coefficients (log10) for
unmodified estuaries typically range 0.3–1.0 m–1, although more elevated values can be associated
with increased particulate loading or humic rich waters following seasonal rainfall events.
Table 8 Ranges of default trigger values for conductivity (EC, salinity), turbidity and
suspended particulate matter of slightly disturbed ecosystems in tropical Australia.
Ecosystem
type
Upland and
lowland rivers
Lakes,
reservoirs and
wetlands
Salinity
(µS cm-1)
Explanatory notes
20–250
Conductivity in upland streams will vary depending upon catchment geology. Values at the lower end
of the range are typical of ephemeral flowing NT rivers. Catchment type may influence values for Qld
lowland rivers (e.g. 150 µS cm-1 for rivers draining rainforest catchments, 250 µS cm-1 for savannah
catchments). The first flush of water following early seasonal rains may result in temporarily high
values.
90–900
Values at the lower end of the range are found in permanent billabongs in the NT. Higher
conductivity values will occur during summer when water levels are reduced due to evaporation.
WA wetlands can have values higher than 900 µS cm-1. Turbid freshwater lakes in Qld have reported
conductivities of approx. 170 µS cm-1.
Turbidity
(NTU)
Upland and
lowland rivers
Lakes,
reservoirs and
wetlands
Estuarine and
marine
64
2–15
Low values for base flow conditions in NT rivers. Qld turbidity and SPM values highly variable and
dependent on degree of catchment modification and seasonal rainfall runoff.
2–200
Most deep lakes and reservoirs have low turbidity. However, shallow lakes and reservoirs may have
a higher turbidity naturally due to wind-induced resuspension of sediments. Lakes and reservoirs in
catchments with highly dispersible soils will have high turbidity. Wetlands vary greatly in turbidity
depending upon the general condition of the catchment or river system draining into the wetland,
recent flow events and the water level in the wetland.
1–20
Low values are indicative of offshore coral dominated waters. Higher values are representative of
estuarine waters. Turbidity is not a very useful indicator in estuarine and marine waters. A move
towards the measurement of light attenuation in preference to turbidity is recommended. Typical
light attenuation coefficients (log10) in waters off north-west WA range from 0.17 for inshore waters
to 0.07 for offshore waters.