Effects of Increasing Salinity on Freshwater Ecosystems in Australia

Abstract

Salt is a natural component of the Australian landscape to which a number of biota inhabiting rivers and wetlands are adapted. Under natural flow conditions periods of low flow have resulted in the concentration of salts in wetlands and riverine pools. The organisms of these systems survive these salinities by tolerance or avoidance. Freshwater ecosystems in Australia are now becoming increasingly threatened by salinity because of rising saline groundwater and modification of the water regime reducing the frequency of high-flow (flushing) events, resulting in an accumulation of salt. Available data suggest that aquatic biota will be adversely affected as salinity exceeds 1000 mg L–1 (1500 EC) but there is limited information on how increasing salinity will affect the various life stages of the biota. Salinisation can lead to changes in the physical environment that will affect ecosystem processes. However, we know little about how salinity interacts with the way nutrients and carbon are processed within an ecosystem. This paper updates the knowledge base on how salinity affects the physical and biotic components of aquatic ecosystems and explores the needs for information on how structure and function of aquatic ecosystems change with increasing salinity.

Full text

© CSIRO 2003 10.1071/BT02115 0067-1924/03/060655
www.publish.csiro.au/journals/ajb Australian Journal of Botany, 2003, 51, 655–665
CSIRO PUBLISHING
Effects of increasing salinity on freshwater ecosystems in Australia
D. L NielsenA,C,D, M. A. BrockB,C, G. N. ReesA,C and D. S. BaldwinA,C
AMurray-Darling Freshwater Research Centre, PO Box 921, Albury, NSW 2640, Australia.
BNSW Department of Infrastructure, Planning and Natural Resources, PO Box U245, Armidale,
NSW 2351, Australia.
CCo-operative Research Centre for Freshwater Ecology, PO Box 921, Albury, NSW 2640, Australia.
DCorresponding author; email: daryl.nielsen@csiro.au
Abstract. Salt is a natural component of the Australian landscape to which a number of biota inhabiting rivers and
wetlands are adapted. Under natural flow conditions periods of low flow have resulted in the concentration of salts
in wetlands and riverine pools. The organisms of these systems survive these salinities by tolerance or avoidance.
Freshwater ecosystems in Australia are now becoming increasingly threatened by salinity because of rising saline
groundwater and modification of the water regime reducing the frequency of high-flow (flushing) events, resulting
in an accumulation of salt. Available data suggest that aquatic biota will be adversely affected as salinity exceeds
1000 mg L–1 (1500 EC) but there is limited information on how increasing salinity will affect the various life stages
of the biota. Salinisation can lead to changes in the physical environment that will affect ecosystem processes.
However, we know little about how salinity interacts with the way nutrients and carbon are processed within an
ecosystem. This paper updates the knowledge base on how salinity affects the physical and biotic components of
aquatic ecosystems and explores the needs for information on how structure and function of aquatic ecosystems
change with increasing salinity.
BT02115
Effect of salini ty on freshwater ecosystems in AustraliaD. L. Nielsen
et al.
Introduction
Salt is a natural component of the Australia landscape and
has been deposited from a variety of sources over millions of
years. Salt enters aquatic systems from groundwater,
terrestrial material via the weathering of rocks or from the
atmosphere, transported by wind and rain (Baldwin 1996a;
Williams 1987). The relative contributions of these sources
depend on factors such as distance inland, climate and
geology (Williams 1987).
Under natural flow conditions in many wetlands and
rivers, periods of low flow resulted in the concentration of
salts in wetlands and riverine pools. Evaporation, combined
with intrusions of groundwater often caused natural salinity
levels to be high for periods of time (Close 1990; Williams
1999; Kay et al. 2001) (Fig. 1A). During these periods of low
flow/high salinity, biota that could not readily disperse
managed to survive either with little or no reproduction and
recruitment (Mills and Geddes 1980; Williams and Williams
1991) or as dormant propagules (Williams 1985; Brock et al.
2003). Biota that are unable to tolerate these periods either
perish or disperse to recolonise when more favourable
conditions occur (Williams 1985).
In many river system such as the River Murray, alteration
of flows, through modification of temporal and spatial
patterns, has reduced the periods of high flow/low salinity
and low flow/high salinity. The periods in which salt
concentrations exceed the critical thresholds of biota now
rarely, if ever, occur, but secondary salinisation, caused by
run-off from the terrestrial landscape, has increased the
amount of salt entering rivers. The reduction in the frequency
of high-flow (flushing) events is causing an accumulation of
salt in these river systems and a gradual increase in the mean
concentration over time (Close 1990; MDBC 1999; DLWC
2000). While the salinity threshold levels for mature biota
may no longer be exceeded, the mean salinity thresholds for
more sensitive life stages may eventually be surpassed
(Fig. 1B).
A similar pattern of salt accumulation occurs in wetlands.
Prior to the removal of the terrestrial vegetation, most of the
water carrying dissolved salts from the surrounding
catchments was trapped by vegetation and transpired or
evaporated. The salt that did wash into wetlands became
concentrated by evaporation, often exceeding tolerance
levels of sensitive biota. Once these wetlands dried the salt
accumulated in the sediments and was removed by flushing
during the next high-flow events. Removal of vegetation has
increased the amount of water entering the groundwater and
the amount of water and salt that enters wetlands. Many

656 Australian Journal of Botany D. L. N ielsen et al.
wetlands are no longer flushed, so the continual input of salt
increases concentration of salt in the sediments and this will
influence biota such as aquatic plants and benthic animals
(Bailey and James 2000). The increase in salt may also affect
the long-term viability of dormant eggs of micro-
invertebrates and seed of aquatic plants. When sediments
with raised salt concentration are wetted during subsequent
wetland refilling, salt concentrations in the water column
again can exceed the tolerance of wetland biota (Fig. 2).
A taskforce on salinity and biodiversity established by the
Australian and New Zealand Environment and Conservation
Council has predicted that by the year 2050 more than
40000 km of waterways and associated wetlands will have
significantly elevated salt concentrations (ANZECC 2001).
Management organisations such as the Department of
Infrastructure, Planning and Natural Resources (DIPNR) in
New South Wales (NSW) and the Murray–Darling Basin
Commission (MDBC), have set interim end-of-catchment
(or valley) targets for salinity on the basis of existing
knowledge. Available data suggest that aquatic biota will be
adversely affected as salinity exceeds 1000 mg L–1 (1500
EC) (Hart et al. 1991). For many plants and animals there is
information available on the threshold levels of salinity on
mature life stages. For most of these, there is very little
information on threshold levels for earlier developmental
stages (Fig. 3). The early stages of development for some
biota (i.e. fish) have been shown to be more sensitive to salt
than mature stages. For example, Macquarie perch
(Macquaria australasica) has been shown to have a salinity
tolerance of more than 30000 mg L–1 but if eggs are exposed
to salinity of only 4000 mg L–1 egg survivorship is reduced
by 100% (O’Brien and Ryan 1997). While some native
aquatic biota appear to be tolerant of increase in salinity
above 10000 mg L–1 (Williams and Williams 1991), early
life forms may be potentially most at risk from gradual
increases in salinity. There is even less information about
how salinity interacts with processes in aquatic ecosystems,
such as carbon and nutrient cycling.
The large spatial and temporal scales of salinity mean that
if our current best land-management practices were fully
implemented, salinisation would continue to increase in
aquatic ecosystems throughout Australia. Although the
A
Low
Salinity/Water level
Threshold
Water level
Salinity
Mean salinity
High
Low
Threshold
Mean salinity
Water level
Salinity
Salinity/Water level
B
High
Fig. 1. The relationship between flow and salinity in (A)
non-modified and (B) modified water regimes. Solid line indicates the
variation in salinity over time. Dotted line indicates the variation in
water level over time. Small-dashed line indicates biotic threshold level
above which loss of biota may occur or strategies to either avoid or
tolerate increased salinity available. Large-dashed line indicates mean
salinity over time.
Dissolved saltHigh
Low
Threshold
Salinity/Water level
Water level
Mean
Sediment salts
Fig. 2. The relationship between water regime and accumulation of
salts in wetlands. Solid line indicates the variation in salinity over
time. Dotted line indicates the variation in water level over time.
Small-dashed line indicates biotic threshold level above which loss of
taxa will occur. Large-dashed line indicates mean salinity over time.
Solid rectangle indicates the accumulation of salts in the sediment as
a consequence of reduced flushing.

Effect of salinity on freshwater ecosystems in Australia Australian Journal of Botany 657
effect of increasing salinity on aquatic biota has been
extensively reviewed, we do not understand the ecological
consequences of salinisation in Australian freshwaters (Hart
et al. 1991; Bailey and James 2000; Nielsen and Hillman
2000; Clunie et al. 2002). The aim of this paper is to review
and update the knowledge base on how salinity affects not
only aquatic biota but also the physical component of aquatic
ecosystems.
Physical and chemical environment
Salinisation of a freshwater body can potentially change both
the light climate and the mixing properties, which in turn
have an impact on the cycling of energy and nutrients.
Salt-induced aggregation and flocculation of suspended
matter is recognised as a major factor in the removal of
particles from the water column, resulting in an increase in
light penetration, and may increase photosynthesis. The rate
of clarification is enhanced by the presence of divalent
cations (particularly Ca2+ and Mg2+) common in saline
ground water (Grace et al. 1997). Increased water clarity, as
a consequence of saline groundwater intrusions has been
implicated in the formation of significant blooms of
cyanobacteria (Geddes 1988; Donnelly et al. 1997).
Alternatively, flocculation of colloids may also remove trace
elements and nutrients from the water column making them
less readily available to pelagic organisms (Donnelly et al.
1997).
Salinisation can alter the relative proportions of cations
and anions in water that can change chemical equilibria and
solubility of some minerals. The major cationic (Na+, K+,
Mg2+, and Ca2+) and anionic species (Cl–, SO42–,
HCO3–/CO3–) vary between locations in both abundance and
concentration. Freshwater biota are influenced as much by
the ionic composition and pH of water as by the total
concentration of dissolved substances (Frey 1993). The
relative proportions of the main cations and anions modify
the way biota respond to high salinities (Bayly 1969; Bailey
and James 2000; Radke et al. 2002). Bayly (1969) suggested
that the ratio of (Na+ + K+)/(Mg2+ + Ca2+) is important in
determining toxicity and suggested that the monovalent ions
are more toxic than divalent ions. This offers an explanation
as to why many species of copepods have been found across
a broad range of salinities in Australia (Hammer 1986). For
example, Boekella triarticulata is a freshwater species that
has been shown to survive in highly saline waters (Bayly
1969). Bayly (1969) hypothesised that the upper limit of
salinity tolerance of freshwater animals is determined by the
chloride content in the blood and that the suppression of this
by a regime of bicarbonate might permit survival in higher
than usual salinities.
Salt-dependent stratification can occur in freshwater
systems following groundwater incursions. Establishment of
a salt gradient can reduce mixing and solute transport within
aquatic ecosystems. The halocline is a barrier for transport of
materials between the surface and bottom strata and has
important implications for nutrient and carbon cycling. In
particular, it may become a barrier for the movement of
oxygen from the surface water to the bottom, causing the rate
of oxygen consumption in the bottom waters to exceed the
rate of replenishment from the surface, which ultimately
leads to anoxia and the death of benthic organisms (Legovic
et al. 1991). Anoxia can also alter the microbially mediated
cycling of nutrients. Anoxia of bottom waters has been
reported in rivers where intrusions of saline water occur
(Anderson and Morison 1989; McGuckin 1990; Donnelly
et al. 1997; Ryan et al. 1999). Salinity of the groundwater
intrusion does not need to be substantially higher than the
salinity of the surface water to induce stratification and
anoxia. Stratification in the Wimmera River has been
observed at a salinity gradient between 300 and 700 mg L–1
(Anderson and Morison 1989). Gribben et al. (2003)
reported the formation of a seasonal salinity gradient in a
High
Low
Salinity
Thresholds
Time
Eggs
Hatching
Juvenile
Salinit
y
Life span
A
High
Low
Salinity
Time
Eggs
Hatching
Life span
Juvenile
B
Fig. 3. Changes in life-history traits as a consequence of modifying
the delivery of salt. (A) Natural delivery. (B) Increased delivery. Solid
line indicates increasing salinity over time. Dotted lines indicate
tolerance levels for each life history phase. The time available for
completion of each stage is decreased as the rate of delivery of salt is
increased.

658 Australian Journal of Botany D. L. N ielsen et al.
shallow freshwater wetland of only about 70 mg L–1, which
they attribute to a ground-water intrusion during the drier
summer months. This gradient, coupled with a
corresponding thermal gradient, is sufficient to prevent
mixing between the surface and bottom waters, with
resultant anoxia in the bottom waters.
Saline ground waters can also lead to elevated levels of
sulfate, dissolved iron and nitrate (Nines et al. 1992). Sulfate
has been implicated in the cycling of phosphorus (Carraco
et al. 1989). Sulfate-reducing bacteria use the sulfate ion for
anaerobic respiration. The respiratory end product from
sulfate reduction is hydrogen sulfide, which is a reducing
agent that can facilitate the dissolving of iron minerals with
a release of phosphorus (Boström et al. 1988). It has also
been suggested that sulfide can displace P from insoluble
Fe2+ phases (Roden and Edmonds 1997). On the other hand,
if a saline groundwater intrusion has a high level of dissolved
iron, oxidation and subsequent precipitation of the iron can
lead to the removal of phosphorus from solution (Baldwin
1996b). Similarly, increases in the concentration of calcium
can also lead to the loss of phosphorus from solution through
precipitation (House 1999).
The increase in ionic strength as a consequence of
salinisation can also disrupt chemical equilibria between
dissolved and particulate phases, either through changes to
ion activity co-efficients or through salt ions blocking
mineral surface adsorption sites (Chang 1977; Stumm and
Morgan 1996). The activity co-efficient of phosphate
decreases with increasing salinity, suggesting that phosphate
should be more soluble in saline systems than in freshwater
systems. Surface chemistry may also be disrupted with
increasing salinity as cations present in salt compete with
other ions for adsorption sites on particle surfaces. For
example, Seitzinger et al. (1991) have shown that the
concentration of exchangeable ammonium in freshwater
sediments is significantly greater than in marine sediments.
They attributed this difference to cations out-competing
ammonium ions for adsorption sites on the sediment.
Biological communities
Some organisms are adapted for living in freshwater, others
for living in salt water. In general, freshwater biota do not
extend into saline or slightly saline water. Consequently, as
salinity increases, the species richness and growth of
freshwater biota is reduced (Hart et al. 1991). Freshwater is
generally defined as water in which salinity is less than 3000
mg L–1 and sea water as 35000 mg L–1. These are the world
average values for those systems (Boulton and Brock 1999);
3000 mg L–1 is often considered the lower limit for saline
waters (Hart et al. 1991). Water between 3000 mg L–1 and
10000 mg L–1 can be defined as saline as biotic effects are
well known within this range. Animals are divided on the
basis of their ability to regulate their internal osmotic
concentrations against the external environment: those that
regulate internal salt concentrations well can adapt to a wide
range of salinities (euryhaline regulators), whereas those that
are poor regulators cannot and are restricted only to a narrow
range of salinities (stenohaline regulators). Salt-tolerant
plants (halophytes) tend to prefer brackish or saline
conditions rather than freshwater, whereas most freshwater
plants (non-halophytes) do not tolerate increasing salt
concentration.
Changes in salinity can affect biota in freshwater directly
or indirectly. Toxic effects as a consequence of increasing
salinity cause physiological changes, resulting in a loss (or
gain) of species. Indirect changes can occur where increasing
salinity modifies community structure and function by
removing (or adding) taxa that provide refuge, food or
modify predation pressure. Other factors such as
water-logging or loss of habitat may interact with salinity or
have a more immediate impact on species richness (Savage
1979; Froend et al. 1987; Bailey and James 2000; Clunie
et al. 2002).
Over the past 12 years several reviews on the effects of
salinity in freshwater ecosystems have highlighted the
paucity of suitable information for making informed
predictions on what future aquatic communities will look
like as salinity increases (Hart et al. 1991; Metzeling et al.
1995; Gutteridge, Haskins and Davey Pty Ltd 1999; Bailey
and James 2000; Nielsen and Hillman 2000; Clunie et al.
2002). In this review, increases in salinity from less that 500
mg L–1 up to above 10000 mg L–1 are considered. This is the
most likely range of salinities that Australian freshwater
rivers and wetlands may experience in the next 50 years.
Pulses of higher salinity are also likely to be encountered in
some rivers and for some wetlands higher levels of salinity
may be experienced as they evaporate and dry out.
Ecological effects of salinity are likely to be observed within
these ranges (Hart et al. 1991).
Microbial function and community structure
Bacteria have a major role in carbon and nutrient cycling.
Our understanding of how microbially mediated processes
change with changing salinity has come from
cross-ecosystem comparisons, in which rates of various
processes have been measured in freshwater, estuarine,
marine and hypersaline environments. A less common
approach has been to examine bacterial populations along a
salinity gradient within rivers as they undergo transition
from freshwater to brackish at estuaries. Understanding of
the function, structure and diversity of microbial community
has recently been advanced with the availability of molecular
DNA methods to identify the presence and diversity of
microbes, and techniques to estimate in situ bacterial
production (growth) or the metabolic capacity of microbes.
In general, aerobic bacterial heterotrophic production in
different aquatic ecosystems has been found to be broadly
predictable, with no consistent differences existing between

Effect of salinity on freshwater ecosystems in Australia Australian Journal of Botany 659
marine and freshwater systems (Cole et al. 1988). Where
differences occur, factors such as carbon and nutrient input
and temperature are more important in regulating production
than salinity (Findlay et al. 1991). Similarly, Hobbie (1988)
concluded that although marine and freshwater microbes
have different physiological methods for tolerating high salt
concentrations, the ecology of marine and freshwater
microbes is virtually identical. As such, it has been assumed
that a process of species replacement will occur in salinised
freshwater systems, that is, increased salinisation of
freshwater ecosystems will simply select for new
physiological types that are able to tolerate given salt levels,
but possessing the same metabolic capabilities (Hart et al.
1991).
Molecular DNA techniques have established that distinct
differences occur in the phylogenetic make up of microbial
populations in freshwater and marine ecosystems (Nold and
Zwart 1998; Crump et al. 1999). Recently it was also shown
that shifts in microbial composition occur along fresh to
brackish gradients in riverine/estuarine systems (Bouvier and
del Giorgio 2002). Metabolic activities of planktonic bacteria
also are known to vary in space and time along a riverine/
estuary gradient (Schultz and Ducklow 2000; del Giorgio and
Bouvier 2002). In the latter study, salinity was an important
determinant in separating bacterial communities.
The sparse information on the response of cyanobacteria
to salinity indicates that some members of this group occur
at salinities greater than that of seawater (>35000 mg L–1).
Freshwater cyanobacteria appear to be inhibited by
variations in salinity (Hart et al. 1991) but may adapt to
gradual increases. Species of Anabaena have been found to
acclimatise to salinities of 7000 mg L–1 after several days’
exposure (Hart et al. 1991; Winder and Cheng 1995).
The relationship between salinity and specific bacterial
processes has been examined, although not extensively.
Nitrogen fixation and nitrification are known to occur in
environments with widely differing salt levels. Nitrogen
fixation by planktonic organisms generally is greater in
freshwater than in marine systems; however, within given
ecosystems, the rate of nitrogen fixation generally is
regulated by nutrient status and not salinity (Howarth et al.
1988). There appears to be no difference in the rate of
nitrogen fixation by benthic communities with respect to
different salinities. However, nitrifying and nitrogen-fixing
communities are known to vary significantly across such
systems (Affourtit et al. 2001; de Bie et al. 2001). Specific
linkages between structure and function of nitrifying and
nitrogen-fixing organisms have not been made.
A major difference between freshwater and marine
systems are the processes in anaerobic degradation of
carbon. In marine and estuarine systems sulfate-reduction is
the major step, whereas methanogenesis dominates in fresh
systems. (Capone and Kiene 1988). This difference is driven
by the presence of sulfate ions in sea water stimulating
sulfate-reducing bacteria, which in turn are able to
out-compete methanogens for substrates (Widdel 1988).
Marine and freshwater species of sulfate-reducing bacteria
and methanogens are known to exist (Postgate 1984;
Oremland 1988).
Denitrification occurs in all aquatic ecosystems; however,
it has been suggested that in general terms the range of rates
of denitrification in marine systems is greater than in
freshwater systems (Seitzinger 1988). Denitrification rates
tend to be limited by nitrate concentration, and salinity by
itself may not be the underlying regulating factor in nitrate
reduction. Molecular studies have been carried out on
denitrifying bacteria from freshwater and marine ecosystems
(Braker et al. 1998; Bothe et al. 2000; Scala and Kerkhoff
2000). However, no extensive cross-system comparisons of
denitrifying populations have been made.
Studies on rivers and estuaries continue to provide useful
insights as to how bacterial populations change across salt
gradients. Whether such comparisons can readily be
transferred to freshwater ecosystems that undergo long-term
increases in salinity remains to be tested.
Algae
There is only sparse information on the sensitivity and
tolerance of freshwater algae; however, the majority of taxa
do not appear to be tolerant of increasing salinity (Hart et al.
1991; Bailey and James 2000; Nielsen and Hillman 2000;
Clunie et al. 2002).
The majority of algae do not appear to tolerate salinities
in excess of 10000 mg L–1 (Bailey and James 2000). Field
observations indicate that as salinity increases, diatoms
decrease in both abundance and richness (Blinn 1993; Blinn
and Bailey 2001). Experimental flooding of sediments has
suggested that some phytoplankton emerge in substantial
numbers when exposed to saline water but diversity is
reduced (Skinner et al. 2001; L. Bowling, unpubl. data).
Some unicellular algae such as Dunaliella salina produce
resting cysts that allow them to survive high salinities.
Species such as D. salina also undergo morphological and
physiological changes that allow them to survive across a
broad range of salinities (Borowitska 1981; Brock 1986).
Aquatic plants
In general, freshwater aquatic plants are not tolerant of
increasing salinity. The majority of data on the response of
aquatic plants to increasing salinity come from field
observations. The upper limit of salinity tolerated by most
freshwater aquatic plants appears to be 4000 mg L–1. Above
this, non-halophytes such as Myriophyllum are replaced by
more tolerant halophytic species such as Ruppia spp. and
Lepilaena spp. which have been recorded in salinities several
times that of seawater (Brock 1981, 1985, 1986).
At salinities above 1000 mg L–1, adverse effects on
aquatic plants appear, with reduced growth rates and reduced

660 Australian Journal of Botany D. L. N ielsen et al.
development of roots and leaves. Both sexual and asexual
reproduction become suppressed (James and Hart 1993;
Warwick and Bailey 1997, 1998). The development of
below-ground tubers, necessary for growth in the following
year, and the development of flowers are also prevented
(Warwick and Bailey 1996).
Information on sublethal effects of increasing salinity on
germination, growth or development of aquatic plants is
limited. Salt sensitivity may differ among various life stages
of a species, which may reflect exposure to different
environmental conditions (Bailey and James 2000). High
salinity is usually inhibitory or toxic to seed germination of
most freshwater plants (Ungar 1962; Williams and Ungar
1972; Baskin and Baskin 1998). For example, germination of
seeds from both Sagittaria latifolia and Ruppia megacarpa
decreases as salinity increases (Brock 1982; Delesalle and
Blum 1994). However, there are isolated cases in which
germination of a halophytic species has increased under
higher salinities (e.g. Ruppia tuberosa) (Brock 1982).
Results from the experimental inundation of sediments
from seven wetlands across inland New South Wales under
five salinities (300, 1000, 2000, 3000 and 5000 mg L–1)
indicated that salinity has a significant impact on the
germination of seeds of aquatic plants when it exceeds
1000 mg L–1. The greatest impact was on species richness
and abundance in communities developing from sediment
subjected to shallow flooding. Communities developing
from sediment subjected to deeper flooding showed a lesser
effect of salinity. This suggests that submerged aquatic plant
communities may be buffered from elevated concentrations
of salt, whereas those plants that live in the margins of
wetlands may be more susceptible to increases in salinity
(D. L. Nielsen and M. A. Brock, unpubl. data).
Invertebrates
It has been predicted that salinity exceeding 1000 mg L–1
will have adverse affects on invertebrates (Hart et al. 1991).
Results from field studies examining salinity gradients in
rivers or across wetlands indicate that as salinity increases
there is a loss of diversity. Diversity decreases rapidly as
salinity increases up to10000 mg L–1, but less rapidly above
10000 mg L–1 (Williams et al. 1990).
Invertebrates can be divided into the following two
groups: (1) microinvertebrates, comprising protozoan,
rotifers and micro-crustaceans (particularly copepods,
cladocerans and ostracods) (Shiel 1990) and (2)
macroinvertebrates in which the major taxonomic groups are
insects, worms, snails and macro-crustaceans (shrimp,
yabbies) (Bennison and Suter 1990).
Microinvertebrates
Microinvertebrates are generally considered to be of
non-marine origin (De Deckker 1983; Hammer 1986) and as
a group they appear not to be tolerant of increasing salinity.
As salinity increases, there is a general decrease in
abundance and richness of rotifers and microcrustaceans
(Brock and Shiel 1983, Campbell 1994). There is little
information on salt tolerance in protozoa, although they have
been recorded from Lake Gregory, Western Australia, when
the lake contains freshwater but not when it is saline (Halse
et al. 1998).
Field studies have shown that there is a decrease in the
number of rotifer species occurring in lakes at salinities
above 2000 mg L–1 (Brock and Shiel 1983; Green and
Mengestou 1991). In freshwater wetlands in Australia, over
200 taxa have been recorded from individual sites (Boon
et al. 1990), but at high salinities, taxon richness is
substantially reduced, often to as little as one or two taxa
(Timms 1981; Brock and Shiel 1983; Halse et al. 1998). The
rotifers and Brachionus plicatillis, Hexarthra fennica and
Trichocerca spp. have been recorded in saline lakes (Timms
1981, 1987, 1998; Brock and Shiel 1983) and many
ostracods also appear to tolerate a broad range of salinities
(De Deckker 1983).
Few studies have examined the effect of increasing salinity
on the emergence of microfauna from dormant eggs. It has
been shown that increases in salinity may inhibit emergence
from resting eggs (Skinner et al. 2001). High salinity has
been linked to blocking hatching of the rotifer Brachionus
plicatilis (Pourriot and Snell 1983), and the microcrustacean
Daphniopsis pusilla (Geddes 1976). Ostracods have also
been noted as emerging only in saline lakes when salinities
are low (De Deckker 1983). In the experimental inundation
of sediments from seven wetlands across inland New South
Wales under five salinities (300, 1000, 2000, 3000 and
5000 mg L–1), the majority of microinvertebrate taxa had
significantly reduced emergence at salinities of 2000 mg L–1
and above. For some taxa there was a significant reduction in
emergence below 1000 mg L–1 (Nielsen et al. 2003;
D. L. Nielsen and M. A. Brock, unpubl. data). Increasing
salinity may be reducing the viability of the eggs or it may be
blocking the required cues to trigger emergence.
Food availability may also influence the ability of animals
to tolerate increased salinity. The estuarine copepod
Sulcanus conflictus has been shown to have lower survival at
high salinities when the available food is of poor quality
(Rippingale and Hodgkin 1977). However, in the case of
rotifers, decreases in numbers have been linked more to
specific physiological tolerances rather than food availability
(Green and Mengestou 1991). The effects of salinity may
also be sex-dependent. Females of the copepods Boekella
hamata and Acartia tonsa are larger than males and exhibit
higher survival at increased salinities (Hart et al. 1991;
Cervetto et al. 1999; Hall and Burns 2001).
Macroinvertebrates
A large proportion of Australian macroinvertebrates has a
marine ancestry (Hart et al. 1991) and as a group they appear

Effect of salinity on freshwater ecosystems in Australia Australian Journal of Botany 661
to be more tolerant of increasing salinity than the
microinvertebrate group.
There is more information on salinity effects on
macroinvertebrates than other biotic groups, as they have
been widely used in the monitoring of the health of aquatic
system. Data are generally from field surveys comparing
taxon presence with conductivity (salinity) collected as an
environmental parameter. In some cases, there is limited
monitoring of community changes at a site over time so
ranges of tolerance for some taxa can be inferred from
within-site as well as between-site data. Much of these data
have been collated into a database (Boon et al. 2002).
In river ecosystems, macroinvertebrate diversity and
salinity are not closely correlated. While salinity may cause
the loss of some taxa and facilitate the intrusion of estuarine
taxa upstream, increasing salinity may not be a catastrophic
event. The macroinvertebrate fauna of rivers appear to be
tolerant and relatively resilient to increasing salinity
(Williams et al. 1991; Metzeling 1993; Metzeling et al.
1995). Data from wetlands confirm this view. Substantial
changes in diversity of wetland macroinvertebrates are not
likely to occur until salinities exceed 10000 mg L–1, after
which substantial loss of diversity and changes in
community composition may occur (Suter et al. 1993; Halse
et al. 2000). The groups most sensitive to increasing salt are
the structurally simple, often soft-bodied animals such as
hydra, insect larvae and molluscs (Hart et al. 1991). Data
from acute 72-h toxicity tests (LC50) of 59 macroinvertebrate
taxa indicate that the salinity tolerance ranged from 5000 up
to 50000 mg L–1, with baetid mayflies the least tolerant
(LC50 = 5500 mg L–1) and macrocrustaceans the most
tolerant (LC50 = 38000 mg L–1) (Kefford et al. 2003).
Although the adults and larvae of many
macroinvertebrates appear to be tolerant of elevated salinity,
there is little information on modifications to egg
development or early instar and juvenile development.
Fish
Most adult native and introduced fish are tolerant of
increasing salinity, but juveniles and eggs of some species
are susceptible (Clunie et al. 2002).
The majority of native Australian fish are derived from
relatively recent marine ancestors. Only the lung fish
(Neoceratodus forsteri), spotted barramundi (Scleropages
leichardti and S. jardini) and the Western Australian
salamanderfish (Lepidogalaxias) have long evolutionary
histories in freshwater (Merrick and Schmida 1984). Studies
have shown that the majority of native and introduced fish in
Australia appear to be tolerant of salinities exceeding 3000
mg L–1 (Chessman and Williams 1974; Hart et al. 1991;
Williams and Williams 1991; O’Brien and Ryan 1997;
Whiterod 2001).
There has been only limited examination of the affect of
salinity on juveniles and eggs, although evidence suggests
some are susceptible to increased salinity. Eggs of the native
Macquarie perch have only 50% survival when exposed to
3000 mg L–1 and juveniles that hatched in this salinity were
smaller than the controls. In a similar experiment, trout cod
egg survival was reduced by 50% at 4500 mg L–1 (O’Brien
1995; O’Brien and Ryan 1997). Eggs of silver perch are not
affected until salinity exceeds 9000 mg L–1; however,
juveniles hatched at 6000 mg L–1 had better survival than
those hatched in freshwater, possibly resulting from
decreased mortality as a consequence of salt-inhibiting
diseases that commonly affect larvae (Guo et al. 1993). The
Australian grayling, which is found in coastal rivers of
south-eastern Australia and spends part of its life cycle in
estuaries, produces eggs that are tolerant of salinities up to
5000 mg L–1 (Bacher and O’Brien 1989).
Discussion
There is a general acceptance that freshwater ecosystems
undergo little ecological stress when subjected to salinities
up to 1000 mg L–1. However, much of our understanding of
the effects of salinity on freshwater ecosystems comes from
lowland rivers where exposure to significant salt
concentrations already occurs; other systems may be more
sensitive. Hence, this view could lead to the
misinterpretation that freshwater ecosystems below 1000 mg
L–1 are ‘healthy’, and there will be no adverse effects on
biota or ecosystems. For many taxa, sublethal effects may not
be apparent at the community level for many generations.
Much of our knowledge of the impacts of salinity on aquatic
ecosystems comes from field sampling along a gradient of
salinity, from which it is difficult to attribute cause of
ecological change. Other underlying factors such as habitat
modification, loss of food resources or modification of
predation pressure may also be causing changes within these
systems (Blinn and Bailey 2001).
Some biotic groups are more tolerant of salinity than
others. Communities of adult fish and macroinvertebrates
appear to tolerate increasing salinity because they either
comprise salt-tolerant remnants left after salt-sensitive
species have been eliminated or reflect an evolution from
marine ancestors (Williams et al. 1991; Bunn and Davies
1990; Mitchell and Richards 1992; Metzeling 1993; Kay
et al. 2001). The freshwater algae, aquatic plants and
microinvertebrates, appear to be less tolerant of increased
salt. For these groups the general trend is to reduction of
species richness as salinity increases with either a loss (or
gain) in abundance. The freshwater taxa in these groups
appear restricted to below 3000 mg L–1, which may reflect a
non-marine recent ancestry.
Life cycles of aquatic organisms generally are controlled
by the presence of water, in association with other triggers
(e.g. temperature) that cue the onset of processes such as
germination of seeds, hatching of invertebrates from
diapausing eggs or spawning of fish. Although specific

662 Australian Journal of Botany D. L. N ielsen et al.
information on impacts of increasing salinity is limited, we
do know that life-history traits related to fitness, such as
survival, growth and reproduction, can be reduced by stress
(Hoffman and Parsons 1991). Hence, stresses such as
long-term exposure to salinity may lead to reduction in
reproduction, recruitment and ultimately depletion of biotic
reservoirs, reducing the sustainability of communities and
their ability to respond when a flush of freshwater occurs.
The current rate of change of salinity in freshwater
ecosystems may be much faster than freshwater biota can
evolve or adapt. Although lowland river biota may have
mechanisms that allow survival during periods of extreme
salt concentrations, upland rivers potentially have
experienced lower natural variation in salinity and therefore
biota in these systems may be less salt-tolerant. Induced
changes in salinity in upland systems may be too rapid for
taxa to adapt, suggesting that freshwater taxa may be lost and
communities will become dominated by salt-tolerant taxa.
Pulses of salt into freshwater ecosystems will influence
survival of a range of biota and although such increases in
salt may be rapid and short-lived, the consequences to the
freshwater biota are unknown.
We need to know how salinity changes ecosystem
functioning through alteration of biotic and abiotic processes: do
changes to ecological processes change community
composition? Managers need to know more about the
relationship between flow patterns, salt concentrations and
environmental damage to predict consequences of management
actions. How a combination of changes in flow and salt affect
river and wetland communities is also relevant to management
predictions. We have many systems that are naturally variable in
both salinity and hydrology, yet we do not know how increasing
salinity will affect the biota or ecosystem integrity. Linking
salinity levels directly to mortality or recruitment potential of
aquatic biota is not sufficient to predict the outcome of
increasing salinity on freshwater ecosystems. Second- and
third-order effects must also be taken into account in describing
the full effect of salinity on aquatic ecosystems. Of particular
interest are the effects of increasing salinity on primary and
secondary production, nutrient dynamics and food-web
structure. Once we understand these interactions, links and the
flow on consequences, managers and researchers will be in a
better position to predict the condition of aquatic ecosystems
under modified salinity and move towards focusing on effective
rehabilitation. For example, the use of environmental water
allocations (environmental flows) could be considered as a tool
in managing salinity in aquatic ecosystems, once the
relationships between hydrology, salinity and environmental
damage are further delineated. Use of this relationship could
enhance effective disposal of salt-contaminated water, with
minimal damage to the environment.
If ecosystem health and salt can be related, then tools such
as water allocations, river operation, engineering
intervention and catchment management programs can be
designed to manipulate salt loads to increase the health of
aquatic ecosystems. Innovative experimental science,
together with imaginative predictive management can work
together to underpin salinity management issues on both
broad and local scales.
Acknowledgments
The authors acknowledge the support of the CRC for
Freshwater Ecology project and the NSW Department of
Infrastructure, Planning and Natural Resources. We thank
Sebastien Lamontagne, Katharine Crosslé, Ken Harris and
Michael Healey for discussions on the effects of salinisation
on freshwater ecosystems. The authors also thank Ben
Gawne, Terry Hillman, Paul Humphries, Helen Gigney and
Brian Wilson for their constructive comments on this
manuscript.
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Manuscript received 13 December 2002, accepted 22 May 2003