Macroinvertebrate responses to
altered low-flow hydrology in
Queensland rivers
Yemaya Smythe-McGuiness, Jaye Lobegeiger, Jonathan
Marshall, Rajesh Prasad, Alisha Steward, Peter Negus, Glenn
McGregor and Satish Choy
Low flows report series, June 2012
NATIONAL WATER COMMISSION — Low Flows Report Series
ii
© Commonwealth of Australia 2012
This work is copyright.
Apart from any use as permitted under the Copyright Act 1968, no part may be reproduced by
any process without prior written permission.
Requests and enquiries concerning reproduction and rights should be addressed to the
Communications Director, National Water Commission, 95 Northbourne Avenue, Canberra
ACT 2600 or email bookshop@nwc.gov.au.
Online/print: ISBN: 978-1-921853-81-4
Published by the National Water Commission
95 Northbourne Avenue
Canberra ACT 2600
Tel: 02 6102 6000
Email: enquiries@nwc.gov.au
Date of publication: June 2012
An appropriate citation for this report is:
Smythe-McGuinness Y et al. 2012, Macroinvertebrate responses to low-flow hydrology in
Queensland rivers, National Water Commission, Canberra.
Disclaimer
This paper is presented by the National Water Commission for the purpose of informing
discussion and does not necessarily reflect the views or opinions of the Commission or the
Queensland State Government.
NATIONAL WATER COMMISSION — Low flows report series iii
Low flows report series
This paper is part of a series of works commissioned by the National Water Commission on
key water issues. This work has been undertaken by the Queensland Government on behalf
of the National Water Commission.
NATIONAL WATER COMMISSION — Low Flows Report Series
iv
Contents
Executive summary
Report context
1.
Introduction
2.
Background
2.1.
Queensland low-flow characteristics
3.
Pressure-Stressor-Response models
3.1.
Dams and weirs
3.2.
Water extraction
3.3.
Interbasin water transfer
3.4.
Disposal of coal seam gas industrial water into surface streams
3.5.
Climate change
4.
Case studies
4.1.
Walla Weir case study
4.2.
Wivenhoe Dam case study
5.
Current monitoring relevant to low flows: issues and recommendations
5.1.
Hydrology
5.2.
Ecology
6.
Discussion and conclusions
Shortened forms
References
viii
x
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Tables
Table 1: Pressure-Stressor-Response linkages of flow-related pressures
identified in Queensland. The expected macroinvertebrate responses to
each of the hydrological stressors are outlined, along with the case studies
included in this report to test and to demonstrate the expected responses. ........................7
Table 2: Flow metrics expected for two sites in the Dawson River under current
water resource development (ROP) and projected goal seam gas (CSG)
flow scenarios (from McGregor et al. 2011). ......................................................................16
Table 3: Site numbers and names of macroinvertebrate sampling sites within the
Burnett River catchment. ....................................................................................................21
Table 4: Time of sampling and habitats sampled within each site in the Burnett
River catchment. Sampling was conducted on five occasions: before, during
and after construction of the Walla Weir. Habitats sampled were E: edge,
and R: riffle. ........................................................................................................................21
Table 5: Gauging station numbers and site names of macroinvertebrate sampling
sites within the Brisbane River and Logan River catchments. ...........................................31
Table 6: Time of sampling and habitats sampled within each site in the Brisbane
River catchment. R: riffle/run; E: edge; P: pool; and M: macrophytes. Blank
spaces indicate no samples were taken (adapted from Choy et al. 2000). .......................33
Table 7: Results of pairwise comparisons between habitat types following twoway crossed ANOSIM (habitat x run, Global R = 0.191, p = 0.0043) using
abundance category data. This tests the null hypothesis that there is no
difference between habitats allowing for differences between runs. Shaded
cells indicate pairs of habitats that were found to be significantly different (p
< 0.05) ................................................................................................................................37
Table 8: Flow preferences and functional feeding group membership of
significant macroinvertebrate taxa identified within results, according to the
flow classification of Marshall and Marshall (in prep.). Flow preferences
correlate with the flow requirements, larvae and adult habitat of taxa as
outlined in Choy et al. (2000). ............................................................................................42
Table 9: The median Julian date of the one-day annual minimum flow for the
Brisbane and Logan rivers for the periods before (pre-1997) and after (post1985) construction of the Wivenhoe Dam. .........................................................................43
NATIONAL WATER COMMISSION — Low flows report series v
Table 10: Current DERM monitoring and research programs that may relate to
low flows and ecology. Relevant programs are listed, as well as the location
in which they are undertaken, the ecological indicators used, the frequency
of activities, and how these activities apply to our knowledge of low-flow
ecology. ..............................................................................................................................51
Figures
Figure S1: Context of reports produced for the Low Flow Ecological Response
and Recovery Project. Each circle represents the location of individual case
studies and the size of each circle represents the spatial extent of each case
study. .................................................................................................................................... x
Figure 1: Flow regime is of central importance in sustaining the ecological
integrity of aquatic ecosystems. The five components of the flow regime:
magnitude, frequency, duration, timing and rate of change influence
ecological integrity both directly and indirectly, through their effects on other
primary ecological regulators. Modification of flow thus has cascading
effects on the ecological integrity of rivers [source: Poff et al. 1997]. ..................................5
Figure 2: Hourly height and discharge data from GS113015A Tully River at Tully
Gorge over a 10-day period, demonstrating fluctuations relating to
hydroelectric water releases. ................................................................................................9
Figure 3: Surface water development status based on water diversion levels (as
a proportion of the long-term mean of annual pre-development discharge)
within National Land and Water Resources Audit (2002) catchment reporting
units. ...................................................................................................................................11
Figure 4: Conceptual model of the ecological impacts of surface water extraction
in conjunction with the impacts of dams and weirs in the case where they
reduce downstream flows [source: DERM 2008]. ..............................................................13
Figure 5: Conceptual model illustrating alterations to hydraulic conditions and
water quality resulting from coal seam gas wastewater disposal into surface
waters [source: Takahashi et al. 2011]...............................................................................17
Figure 6: Locations of sampling sites, numbered according to gauging station, in
the Burnett River catchment. The Walla Weir is situated between sites
GS1360069 and GS136023A. ............................................................................................20
Figure 7: Flow-velocity-preference group structure for edge habitat samples at
sites in the Burnett River in relation to the Walla Weir. Sample runs 1 and 2
refer to the pre-construction phase samples, 3 to the construction phase
samples, and 4 and 5 to the operation phase samples. .....................................................23
Figure 8: Flow-velocity-preference group structure for riffle habitat samples at
sites in the Burnett River in relation to the Walla Weir. Sample runs 1 and 2
refer to the pre-construction phase samples, 3 to the construction phase
samples, and 4 and 5 to the operation phase samples. .....................................................24
Figure 9: Dendrograms illustrating the separation of macroinvertebrate
communities based on site position from samples taken during the second
operation phase of the Walla Weir within a) edge and b) riffle habitats.
Samples were grouped for sites where samples were taken from each
habitat type. * indicates the impounded area at Ben Anderson Barrage
whereas other impounded areas were impounded by the Walla Weir. ..............................25
Figure 10: Map indicating the location of macroinvertebrate sampling sites within
the Brisbane River and Logan River catchments. Site numbers are gauging
station numbers. The Wivenhoe Dam is located at the base of Lake
Wivenhoe. Red triangles in the Brisbane River catchment indicate regulated
sites downstream of the Wivenhoe Dam, and blue triangles indicate
unregulated sites. Green triangles in the Logan River catchment indicate
upstream sites and yellow triangles indicate downstream sites. ........................................30
Figure 11: Ordinations showing the relationships between sampling periods
(runs) within riffle/run habitat, based on a) macroinvertebrate abundance
data allowing for differences between habitat types (stress=0.0000), and b)
NATIONAL WATER COMMISSION — Low flows report series vi
mean maximum flow velocity (stress = 0.0002) (adapted from Choy et al.
2000)...................................................................................................................................38
Figure 12: a) Ordination plot of average Faunal Abundance Category data with
arrows showing direction and magnitude of differences between regulated
(downstream) and unregulated (upstream) samples from riffle/run and edge
habitats in the Brisbane and Logan rivers (stress = 0.1790). B – Brisbane
River, L – Logan River, R – riffle/run, E – edge, r – regulated (downstream),
u – unregulated (upstream). b) Selected significant faunal correlation vectors
contributing to variation in macroinvertebrate community samples within the
Brisbane River (HPSY = Hydropsychidae, UACA = Acarina, COEN =
Coenagrionidae, VELI = Veliidae, ATYI = Atyidae, UCLA = Cladocera, CORI
= Corixidae, COPE = Copepoda). c) Selected significant environmental
correlation vectors explaining variance in macroinvertebrate communities
within the Brisbane River (MeanSed = mean sediment sizes, logVEMA = log
maximum velocity, logTOTN = log total nitrogen, logTURB = log water
turbidity) (from Choy et al. 2000). .......................................................................................40
Figure 13: Schematic representations of the Logan and Brisbane rivers showing
the proportions of macroinvertebrate functional feeding groups in samples
from the wet season and dry season at upstream and downstream sites in
each river (from Choy et al. 2000). .....................................................................................41
Figure 14: Comparison of median monthly flow volumes (ML day-1) in the
Brisbane and Logan Rivers in the periods before construction of the
Wivenhoe Dam on the Brisbane River (pre-1977) and after construction
(post-1985). ........................................................................................................................43
Figure 15: Change in median monthly flow volumes (ML day-1) for the Brisbane
and Logan rivers in the period following construction of the Wivenhoe Dam
(post-1985) compared with the preceding period (pre-1977). * indicates a
significant difference (P=0.05). ...........................................................................................44
Figure 16: Change in median low- and high-flow volume metrics for the Brisbane
and Logan rivers in the period following construction of the Wivenhoe Dam
(post-1985) compared with the preceding period (pre-1977). * indicates a
significant difference (P=0.05). ...........................................................................................44
Figure 17: Ecological responses occur in reaction to the physical and chemical
conditions experienced by biota, not directly to altered flow regimes. Altered
flows influence ecology by interacting with various other properties of the
setting to modify the conditions the biota experience. To understand
ecological responses to flow modification one must first understand the
relationships between physical/chemical conditions and the ecological
response, and secondly understand the relationship between flow and the
provision of these conditions. Stressors other than altered flow may also
influence the response, so the effectiveness of flow regime restoration or
protection cannot be judged directly by the occurrence or intensity of
response. Rather, it must be judged by evaluating the provision of the flowrelated conditions necessary for the response to occur (Cockayne et al.
2010)...................................................................................................................................49
NATIONAL WATER COMMISSION — Low flows report series vii
Executive summary
This report’s purpose is to provide information on the role of low flows in shaping aquatic
ecosystems, and how alteration of the naturally occurring low-flow regime can affect those
ecosystems, specific to Queensland waterways. Low flows are an important influence in
shaping aquatic ecosystems. As a result, aquatic biota are often adapted to the natural lowflow regime in the areas in which they occur. In many regions within Queensland low-flow or
no-flow conditions are common and biota have developed strategies to enable them to
survive annual dry seasons. This is in contrast with the impacts of low flows within more
perennial systems where biota are not as well adapted to low flows. Uncharacteristic low-flow
events within perennial systems can therefore act as a disturbance to biota and ecological
processes. Changes away from natural low-flow conditions can negatively impact freshwater
ecosystems – whether as an increase in the incidence of low flows within perennial systems
not adapted to these conditions, or as a reduction in the frequency or duration of low flows
within more ephemeral systems that are adapted to such conditions (especially when those
changes exceed thresholds of biotic resistance and resilience). Development and use of
water resources for urban, agricultural and industrial expansion have imposed a number of
changes on natural flow regimes, including the low-flow regime itself, and these changes are
often far beyond the range of variability experienced naturally. The main anthropogenic
pressures within Queensland that alter the low-flow regime were identified as:

barriers – dams and weirs, culverts

water extraction

disposal of excess industrial water

interbasin transfers

climate change.
Using a Pressure-Stressor-Response framework approach, a conceptual understanding of
the types of hydrological changes imposed by the above pressures and the likely responses
from aquatic communities were outlined. Two case studies using existing departmental
macroinvertebrate data in conjunction with new hydrological analyses were conducted to test
conceptual understandings of macroinvertebrate responses to dams and weirs.
The first case study investigated impacts of the Walla Weir on macroinvertebrate communities
through flow alteration. Reduced flow velocity both within the weir impoundment where riffles
were drowned out, and downstream where baseflows were reduced, led to a reduction in
high-flow-preference macroinvertebrate taxa. Similarly, the second case study investigated
the impacts of the Wivenhoe Dam on flow characteristics and resulting impacts on
macroinvertebrate communities. The dam altered the hydrology of downstream reaches by
decreasing flow variability and the duration and frequency of low-flow events and, in contrast
with the Walla Weir, increased flows through downstream aquatic habitats that naturally
experienced low-flow velocities. Shifts were observed in macroinvertebrate community
composition, with a higher proportion of high-flow-preference taxa downstream of the dam,
and altered functional feeding group composition.
While these case studies gave some indication of the response of macroinvertebrate
communities to alterations of the low-flow regime within Queensland aquatic ecosystems,
difficulties in identifying general low-flow-ecology relationships were encountered because
biota do not experience or respond directly to hydrology. Rather, flow interacts with other
environmental features to produce physical, chemical and biological conditions which are
perceivable to biota and which elicit or suppress particular ecological responses. Conceptual
models helped show the stressors likely to be intermediating the impact of low-flow alterations
NATIONAL WATER COMMISSION — Low flows report series viii
on aquatic biota. However, a pervasive lack of suitable data pertaining to low lows prevented
these conceptual understandings being tested further than was possible within the two case
studies.
A lack of data relating to biota, hydrology and the impacts of altering flow conditions therefore
inhibited a thorough assessment of low-flow-ecology relationships within Queensland. An
assessment of current management practices found that ecological problems relating to low
flows may be detected and dealt with through existing monitoring and research in instances
where they are locally relevant to anthropogenic pressures and ecological values. There are,
however, several difficulties associated with assessing the effect of low flows on ecology
based on existing hydrological and ecological data. To better inform flow management in
future and assess the impacts of altering the low-flow regime on aquatic biota, we advocate
developing conceptual understandings of the flow-related conditions required for biological
responses to be identified. Ecological risk assessment frameworks, where the risk to specific
values is linked to ecological responses with critical requirements for flow conditions, can be
used to account for both the potential influence of confounding stressors and the complex
relationships between flow, habitat and ecology.
NATIONAL WATER COMMISSION — Low flows report series ix
Report context
This report is part of a larger series of reports produced for the National Water Commission’s
Low Flow Ecological Response and Recovery Project (Figure S1). This report presents one of
11 hydro-ecological case studies. The purpose of the case studies is to test hypotheses that
relate ecological process and function and biological traits to key hydrological measures that
are affected by low flows. A summary of the findings in this report and the other case studies
are contained in Synthesis of case studies quantifying ecological responses to low flows
(Marsh et al. 2012).
Guidance on ecological response and hydrological modelling for low-flow
water planning
Low-flow hydrological classification of Australia
Review of literature quantifying ecological responses to low flows
Early warning, compliance and diagnostic monitoring of ecological
responses to low flows
Synthesis of case studies quantifying ecological responses to low flows
Figure S1: Context of reports produced for the Low Flow Ecological Response and Recovery
Project. Each circle represents the location of individual case studies and the size of each
circle represents the spatial extent of each case study.
NATIONAL WATER COMMISSION — Low flows report series x
1. Introduction
This report was prepared by the Queensland Department of Environment and Resource
Management (DERM) to provide state-specific information relating to the ecological
significance of low flows, as well as the potential ecological consequences of flow regulation
strategies, to inform improved water planning and management. Its approach is based on the
Pressure-Stressor-Response (PSR) conceptual model framework as explained in Marshall et
al. (2006) and is similar to the Driving forces-Pressure-State-Impacts-Response framework
used internationally (e.g. US Environmental Protection Agency and EU Water Directive) and
nationally (e.g. State of the Environment reporting). The PSR framework has been used in
several other ecological studies conducted by DERM to identify and explain how
Queensland’s aquatic ecosystems respond to particular human activities (i.e. pressures) and
the resulting physico-chemical changes (i.e. stressors) and subsequent biological changes
(i.e. responses) to the aquatic environment. For the purposes of this report, only the
pressures that relate to hydrology and particularly low flows have been considered. A broad
definition for ‘low flows’ has been applied here, encompassing a range of conditions including
baseflow, no-flow and completely dry. It is the change away from the natural flow regime that
is considered important, so the pre-existing conditions in a region will determine what is ‘low
flow’. Further, it is identified that dry spells are not necessarily an impact – many systems are
adapted for precisely these regimes – so in some cases additional flow, and thus the loss of
‘low-flow conditions’, is investigated.
Key flow regulation actions (pressures) within Queensland have been identified.
Environmental alterations (stressors) caused by the pressures within freshwater ecosystems
are outlined, along with the resulting ecological impacts (responses) within these systems.
Where possible, case studies based on existing biotic data collected by DERM and new
hydrological data are used to provide empirical illustration of these linkages and, where
empirical data is lacking, future monitoring is recommended. The results of case studies are
used to illustrate the ecological significance of low flows, along with the potential ecological
consequences of flow regulation strategies within Queensland – in accordance with the aims
of the Low Flow Ecological Response and Recovery Project. Current DERM water planning
projects operating within Queensland were assessed to consider how they include low-flow
requirements, and what future practices may be needed to better acquire data and
incorporate understandings of low flows within current water planning and management.
Findings can then be used to investigate the ecological significance of low flows within a
broader context, and inform guidelines regarding water developments.
NATIONAL WATER COMMISSION — Low flows report series 1
2. Background
Water resource managers face the ever-increasing challenge of balancing human demand for
water with environmental requirements (Naiman et al. 2002; Poff et al. 2003; Khan 2008).
Human interference has altered the natural conditions of rivers and streams worldwide, and
often in irreversible ways (Malmqvist & Rundle 2002; Naiman et al. 2002; Nilsson et al. 2005;
Morton et al. 2009; Walters & Post 2011). Freshwater ecosystems provide goods and
services of critical importance to human societies, yet they are among the most heavily
altered environments (Geist 2011). Alteration of natural flow regimes can have profound
effects because the natural flow regime, in synergy with water quality, are both essential
attributes for healthy aquatic ecosystems (Bunn & Arthington 2002; Olden & Naiman 2010;
Geist 2011).
The ecological significance of flow characteristics to aquatic biota are well documented (Poff
et al. 1997; Marshall et al. 2001; Bunn & Arthington 2002; Lloyd et al. 2004; Dewson et al.
2007; Leigh & Sheldon 2008; Monk et al. 2008; Naiman et al. 2008). Flow characteristics
directly control in-stream water quality and habitat conditions, and mediate biological
processes (Figure 1), such as spawning and the dispersal of water-dependent biota (Poff et
al. 1991; Dewson et al. 2007; Naiman et al. 2008). Low flows are an important aspect of the
overall flow regime, but are less studied compared with other aspects of the flow regime
(Rolls et al. 2010). Biologically significant aspects of the low-flow regime are a subset of the
characteristics of the wider flow regime, which include the magnitude, frequency, duration and
timing of flow events; the rate of change during a flow event (Figure 1); and the degree to
which these characteristics are temporally and spatially variable (Poff et al. 1997; Monk et al.
2008; Richter et al. 1998; Kennard et al. 2010; Rolls et al. 2010).
The water resource planning process within Queensland is designed to balance human and
environmental water requirements. The process is prescribed by the Water Act 2000, which is
aligned with the requirements of the National Water Initiative and relevant state and regional
strategies (Whittington 2000; DERM 2011a). Water resource plans have been undertaken for
all but a few catchments within Queensland, and are required to include water requirements
for future human consumptive and non-consumptive needs, as well as environmental needs.
The water allocation process for each catchment involves socio-economic and environmental
assessments, as well as community consultation and scientific review panels, to establish the
water requirements of all sectors involved (Whittington 2000). A water resource plan is then
developed which becomes the legal basis for the establishment of local water operations
within the specified water planning area (DERM 2011a).
Knowledge of natural flow conditions and the corresponding ecology within an area is needed
to understand the consequences of altering the flow regime and inform effective management
(Kennard et al. 2010; Rolls et al. 2010). Scientific assessments inform water resource plans
by determining the flow requirements of ecological assets such as in-stream and riparian
biota. This study aims to contribute to the understanding of low-flow ecology so that
consideration of low flows may be better incorporated into water resource planning and
management. This study also seeks to assess the degree to which current water resource
management within Queensland includes consideration of low flows, and determine how this
process can be improved.
Alterations to low flows can involve changes to low-flow characteristics over space and time.
Alterations to any of these flow aspects can cause changes to in-stream water quality and
habitat conditions, and affect in-stream biota (Figure 1) (Poff et al. 1997; Dewson et al. 2007;
Leigh & Sheldon 2008). Changes to the natural low-flow regime can be measured by
comparing hydrological data before and after flow modification. Where such data is not
available, a reference approach can be used whereby a nearby site with a similar flow regime,
NATIONAL WATER COMMISSION — Low flows report series 2
which is in a relatively natural state, can be used to infer antecedent conditions (Choy et al.
2000; Whittington 2000; Marshall et al. 2001).
Low-flow indices used to measure hydrological changes in the low-flow regime within this
study are based on the flow classification indices used by Kennard et al. (2010). A preliminary
low-flow classification performed by Mackay et al. (2012) built on the work of Kennard et al.
(2010) using only those flow indices relevant to low flows. The authors identified six of these
low-flow indices as the most useful in defining low-flow classes. Similar low-flow indices have
been used within this study to measure changes in the low-flow regime, and include
measures of the magnitude, timing and variability of low-flow discharge over a number of
temporal scales. Details of these indices are outlined within the case studies in this report.
Aquatic macroinvertebrates were used in this report as biological indicators to investigate the
ecological impacts of altering the low-flow regime. Aquatic macroinvertebrates are often used
to assess waterway health and gauge anthropogenic impacts on waterways (Chessman
1995; Growns et al. 1995; Metzeling et al. 2003; James & Suren 2009; Chessman et al.
2010). Aquatic macroinvertebrates are suitable biological indicators of waterway health for a
number of reasons: 1) they are sensitive to changes in flow, water quality and habitat
condition, and may therefore act as indicators of overall ecological health; 2) individuals tend
to have low mobility and so local impacts can be more accurately detected; 3) they are
important components of aquatic foodwebs, and are therefore indicative of wider aquatic
ecological processes, and 4) they are ubiquitous, relatively easy to sample, and are well
documented (Choy & Marshall 2000). Aquatic macroinvertebrate community composition has
been demonstrated to be responsive to varying flow conditions within Queensland and is a
common aquatic biological indicator used by DERM (e.g. Choy & Marshall 1999a; Marshall et
al. 2000; Marshall et al. 2001).
2.1. Queensland low-flow characteristics
The total land area in the Queensland jurisdiction is just under 1.74 million km 2 (ABS 2010).
Queensland is under the influence of a variety of climatic conditions consisting of low rainfall
and hot summers in inland western areas, a monsoon season in the north, warm temperate
conditions along the eastern coastal area and low minimum temperatures inland and around
the southern ranges (BOM 2011). As a result, the rainfall patterns, runoff and hydrology differ
greatly between different parts of the state.
Flow conditions in Queensland range between two extremes: the northern region is tropical
wet while the central and central west region is arid dry. A large number of streams are
characterised by variable and intermittent flow (Kennard et al. 2010; sensu Larned et al. 2010;
Mackay et al.2012). Streams in Queensland have been classified according to their low-flow
characteristics as predominantly ephemeral (Mackay et al. 2012). This is comparable with the
flow classification by Kennard et al. (2010) which was based on a wider range of flow
characteristics.
Kennard et al. (2010) classified the overall natural flow regime of rivers and streams in the
northern and north-western parts of Queensland as either ‘stable summer baseflow’ or
‘predictable summer highly intermittent’. Rivers and streams from the central-east to the
south-east were classified as either ‘unpredictable intermittent’ or ‘unpredictable summer
highly intermittent’. In the south-west, west and central west, rivers and streams were
classified as ‘predictable summer highly intermittent’, ‘unpredictable summer highly
intermittent’ or ‘variable summer extremely intermittent’.
According to Mackay et al.'s (2012) 35 metric classification, a large number of Queensland's
streams have highly ephemeral (class 6) or moderately ephemeral (class 5) low-flow regimes
NATIONAL WATER COMMISSION — Low flows report series 3
- many of these streams are located in the state’s inland areas. Other low-flow classes
represented include marginally ephemeral (class 3) and weakly perennial streams (class 2)
with a predictable low-flow period, which tend to be located near or in coastal areas,
particularly in the state’s north-east.
Low-flow-related stressors are expected to impact areas in different ways, depending on the
existing natural flow regime (Kennard et al. 2010; Rolls et al. 2010). Such large variability in
flow conditions across Queensland presents a wide range of scenarios and implications when
low-flow patterns are altered. For example, the same level of water extraction will affect
aquatic ecosystems in a perennial wet tropics stream in northern Queensland differently to an
ephemeral stream in the state’s west. This is due to underlying differences in hydrology,
habitat conditions and ecological characteristics, which are determinants of the aquatic
ecosystems’ ability to tolerate and recover from disturbances to the low-flow regime (Kennard
et al. 2010; Rolls et al. 2010).
Ephemeral channels in central and western Queensland often experience extended no-flow
periods followed by episodic high-magnitude flows and flooding. Systems fluctuate between
highly connected floodplains and highly disconnected waterholes (Mackay et al.2012).
Aquatic biota within ephemeral systems are adapted to these conditions, and have developed
flexible physiological and lifecycle characteristics that enable them to survive such large
fluctuations in flow conditions and habitat availability (Poff et al. 1997; Bunn & Arthington
2002; Dewson et al. 2007; Takahashi et al. 2011). Many aquatic species occurring within
ephemeral streams in Queensland depend on flow-related cues for critical life-history
activities, many of which are associated with low-flow events. Many species use drought
refugia, including waterholes and hyporheic zones; or have life-history stages that can resist
desiccation in dry river beds (Marshall et al. 2005; McGregor et al. 2011; Steward et al. 2011).
These characteristics, which allow biota to survive unpredictable and often harsh conditions
within ephemeral streams, also mean they are sensitive to flow modifications that alter these
conditions (Bunn & Arthington 2002). If flow conditions were altered so that the natural
patterns of variability were lost, flow-related cues for life-history activities may not occur. For
example, species with desiccation-resistant life-history stages may depend on periods of low
or no flow for a part of their lifecycle, and if flows were altered to disrupt the timing and
frequency of low-flow events, these species could be disadvantaged.
In contrast, aquatic biota within perennial streams in the north and coastal regions of
Queensland are adapted to constant and less variable flow conditions (Kennard et al. 2010;
Rolls et al. 2010). If flow conditions were altered such that flow variability increased – with
more low-flow events – biota within perennial streams would be less well adapted to cope
with such events when compared with biota within naturally ephemeral streams.
Water resource developments that alter the natural flow regime can therefore affect important
hydrological cues and habitat conditions on which biota depend. This has the potential to
significantly disadvantage native aquatic biota and impact the structure and function of
aquatic ecosystems within Queensland (Thoms & Cullen 1998; Bunn & Arthington 2002;
Richardson & Humphries 2010; McGregor et al. 2011; Takahashi et al. 2011). Disturbed
conditions may also promote invasion by exotic species (Koehn 2004; Takahashi et al. 2011).
The direction and magnitude of impact caused by water resource developments will vary
depending on the nature of the pressure, the existing natural flow regime, and corresponding
ecological characteristics of native biota.
This report investigates the nature of water resource developments within Queensland and
the direction of impacts expected as a result of flow alteration from these developments, with
emphasis on alterations to low-flow characteristics. The case studies in this report investigate
NATIONAL WATER COMMISSION — Low flows report series 4
the direction and magnitude of impact to aquatic biota caused by flow alteration from dams
and weirs in the context of two streams within south-east Queensland.
Figure 1: Flow regime is of central importance in sustaining the ecological integrity of aquatic
ecosystems. The five components of the flow regime: magnitude, frequency, duration, timing
and rate of change influence ecological integrity both directly and indirectly, through their
effects on other primary ecological regulators. Modification of flow thus has cascading effects
on the ecological integrity of rivers [source: Poff et al. 1997].
NATIONAL WATER COMMISSION — Low flows report series 5
3. Pressure-Stressor-Response
models
Water resource developments in Queensland have increased considerably during the past 50
years (DSEWPaC 2009). Medium to high levels of development are expected to significantly
affect aquatic ecosystems through flow alteration and resulting habitat modification (Bunn &
Arthington 2002); however, even relatively low levels of development can lead to local
impacts depending on the location and nature of the alteration. The water resource planning
process within Queensland has an important role in minimising the environmental impacts of
water resource developments, now and in the future.
Five main pressures relating to flow alteration have been identified as ecologically significant
within Queensland, based on water resource planning studies conducted by DERM and the
literature (e.g. Choy & Marshall 1999b; Choy & Marshall 2000; Choy et al. 2000; DSEWPaC
2009; Takahashi et al. 2011). Main flow-related pressures within Queensland include: 1)
dams and weirs, 2) water extraction, 3) disposal of excess industrial wastewater into surface
streams, 4) interbasin water transfer, and 5) climate change (Table 1).
The PSR models for each of the five pressures are summarised in Table 1 and discussed
below. The impacts of dams and weirs on in-stream biota as a result of flow alterations have
been documented in a number of studies, including reports produced by DERM (e.g. Choy &
Marshall 2000; Choy et al. 2000) and therefore can be demonstrated by in-depth case
studies. Two case studies are presented on the impacts of Walla Weir and Wivenhoe Dam.
Case studies were not conducted for the other identified pressures due to a lack of data, but
the possible implications of these pressures based on the findings of other studies are
discussed within PSR models. Coal seam gas water-disposal is a new emerging pressure
and, as such, data relating to its effects in the environment is not yet available. Modelling and
experimental approaches are being used to determine the magnitude and direction of likely
impacts (see McGregor et al. 2011; Takahashi et al. 2011) and initial findings from these
studies are briefly discussed in the PSR model. Similarly, with the pressure of climate change,
empirical flow-ecological data was not available because the most significant impacts are
likely to occur in the future. While climate change may have already had some effect on
aquatic ecosystems, the relatively gradual nature of the change means very long-term
datasets are required to detect trends and separate confounding effects of climatic variability.
As such, the current and future effects of climate change and coal seam gas water-disposal
on aquatic ecosystems require further research before PSR models can be tested.
Water extraction pressure occurs frequently in Queensland at a range of scales, and is
considered to be a particular threat to aquatic environments such as refugial waterholes and
riffle habitats during baseflow conditions. The ecological impacts of water extraction have
been studied in some areas of Australia (e.g. McKay & King 2006; Boulton et al. 2003; Nebel
et al. 2008) but few studies quantify the associated flow alterations. Thus little information is
available to describe the location, timing and magnitude of water extraction occurring in the
state, and for this reason a case study could not be conducted. With regard to interbasin
water transfer: while it is likely to cause altered flow conditions in both the source and
recipient systems, the more critical – and better studied and reported – threats of interbasin
transfer relate to the spread of exotic species and implications of translocation on the genetic
viability of native aquatic biota populations (e.g. Davies et al. 1992; Marshall et al. 2000; Page
et al. 2010). As such, case studies demonstrating the ecological impacts of water extraction
and interbasin water transfer in Queensland are not provided due to a lack of data specific to
the flow-related pressures and stressors caused by these pressures, however their possible
implications (based on the findings of other studies) are discussed within the PSR models.
NATIONAL WATER COMMISSION — Low flows report series 6
Table 1: Pressure-Stressor-Response linkages of flow-related pressures identified in Queensland. The expected macroinvertebrate responses to each of the
hydrological stressors are outlined, along with the case studies included in this report to test and to demonstrate the expected responses.
PRESSURE
Expected macroinvertebrate RESPONSE
Case study
Inundation within
impoundment, and
low/no-flow below wall.
Local loss of sensitive and riffle taxa, lower taxa
richness, reduced opportunities for recolonisation
(longer distances between habitats).
Walla Weir,
before & after
construction
Low flow/ baseflow
Increased frequency and
duration of medium/high
flow.
Loss of riffle taxa, change in flow-preference
groups in pools, change in functional feeding
group composition.
Wivenhoe
Dam operation
Variable/seasonal flows
replaced with stable or
unseasonal ones
Seasonality and
wider flow
variability
Change from variable to
stable flow.
Increase in taxa richness and sensitive taxa,
change in flow-preference groups.
Wivenhoe
Dam operation
Hydroelectric releases
interrupt baseflows with
frequent, brief, high-flow
pulses
Baseflow
Increased frequency and
magnitude of high-flow
events with steep rates of
change.
Fewer low-flow-preference taxa and a shift to
those that have adaptations for high flow (e.g.
hooks and suckers). Possible loss of shredder
functional feeding group if leaf litter is flushed.
Reduced flow leads to loss of
riffles and other flowing
habitats, reduced water
quality
Baseflow
Increased frequency and
duration of low and no
flow.
Loss of riffle taxa and sensitive taxa. Increase in
tolerant or opportunistic taxa. Change in
functional feeding groups, e.g. loss of filterer.
Loss of surface water through
extraction from pools
Presence of pool
refugia
Change from no flow to
dry.
Loss of aquatic taxa that do not have
desiccation-resistant stages. Some may disperse
to find refuges or suitable habitat. Colonisation of
the dry streambed by terrestrial taxa.
Interbasin
transfer
Loss of surface water in
source streams, increased
flows in recipient streams,
transfer of biota within water
Low flow/ baseflow
Reduced seasonality and
frequency of low flow in
recipient system, reduced
flows in source system.
Possible loss of riffle taxa and/or sensitive taxa,
loss of taxa dependent on low flows in recipient
streams, aquatic pests may be introduced to
recipient streams.
Disposal of
industrial
water (coal
seam gas)
Additional water into
temporary streams
No flow/ low flow
Reduced frequency and
duration of no and low
flows or change from
ephemeral to perennial.
Loss of invertebrates that require desiccation or
benefit from it. Possible increase in taxa richness
and sensitive taxa. Increase in taxa that prefer
high flow.
Climate
change
Overall reduction in flows
Baseflow
Increased frequency and
duration of low-flow
events.
Loss of high-flow-preference taxa, increase of
low-flow and no-flow-preference taxa.
Increased frequency of
extreme events, i.e. floods
and droughts
Baseflow
Increased frequency of
no-flow and/or high-flow
events.
Increase in tolerant or opportunistic taxa.
Dams and
weirs
Water
extraction
STRESSOR
Vulnerable facet of
Likely hydrological
hydrology
alteration
Barriers and impoundments
lead to loss of riffles through
drowning or drying
Baseflow
Dam releases replace low
flows with medium/high flows
NATIONAL WATER COMMISSION — Low flows report series 7
3.1. Dams and weirs
3.1.1. Pressure
Australia is largely dependent on water storage reservoirs given its highly variable precipitation
patterns and frequent long-term droughts (Ghassemi et al. 1995). Dams and weirs secure water
for irrigation, drinking water supply and hydroelectricity production (Ghassemi et al. 1995;
Arthington & Pusey 2003; Richter & Thomas 2007). As of the year 2000, approximately 446
large dams stored 88 000 GL of water in Australia (Kingsford 2000). At present 183 major water
storages are located in Queensland, with a total capacity of 13 389 GL, not including privately
owned off-stream storages (DSEWPaC 2009).
3.1.2. Stressors
A number of studies have illustrated the ecological impacts of dams and weirs (Brunke et al.
2001; Bunn & Arthington 2002; Nilsson et al. 2005; Almeida et al. 2009; Mueller et al. 2011).
Dams and weirs change the natural flow regime, modify the physical properties of upstream and
downstream habitats, and act as barriers to in-stream dispersal (Choy et al. 2000; Nilsson et al.
2005; Almeida et al. 2009). The impounded water upstream of dams and weirs forms large and
deep lentic systems, leading to a loss of riffles, runs, shallow pools, backwater pools, shallow
edges and other habitats. Impounded water also becomes stratified in terms of temperature and
dissolved oxygen (Olden & Naiman 2010). Downstream of dams and weirs, controlled release of
water changes the flow regime from its natural state. All aspects of the flow regime can be
altered, including the magnitude, frequency, duration and timing of flow events, the rate of
change during a flow event, as well as the degree to which these characteristics are temporally
and spatially variable (Brunke et al. 2001; Bunn & Arthington 2002).
Patterns of water release from dams and weirs vary. In some cases, water is released from
impoundments only periodically (Olden & Naiman 2010). As a result, baseflow is reduced and
high flows occur within a short time period, followed by long periods of low or no flows. Reducing
baseflow in this way can alter in-stream habitat availability by reducing the number of riffles and
runs, lowering water levels within pools, reducing the downstream movement of woody debris,
and increasing the distance between in-stream habitats in areas where flow ceases (Pusey &
Arthington 2003; Olden & Naiman 2010). The ecological impacts of dams and weirs in the case
where they decrease downstream baseflow are summarised in a conceptual model (
Figure 4), shown in conjunction with the impacts of water extraction in Section 3.2.
In other cases there is a continual release of water from dams and weirs, which increases
baseflow and removes flow variability. This acts to deepen existing pools and drown out riffles,
and increase the rate of nutrient movement downstream, thereby altering habitat availability. In
addition, the release of stratified water can alter downstream water quality, although this is
dependent on the method and timing of release (Olden & Naiman 2010).
In cases where flow is regulated for hydroelectricity generation, water is released in frequent
high-magnitude short-duration pulses. This significantly increases the frequency of high-flow
events, with a steep gradient of change between baseflow and high flows. For example, in the
Tully River in the Wet Tropics, daily releases lead to regular water level fluctuations downstream
(Figure 2) (DERM 2011b). This can act as a frequent disturbance to in-stream conditions by
scouring and causing siltation of the streambed, and can increase the rate of downstream
movement of organic material, therefore limiting in-stream habitat availability. Reduced organic
material, along with possible changes in fluvial geomorphology because of altered sediment
transport patterns, can limit habitat and nutrient availability within aquatic systems below dams
NATIONAL WATER COMMISSION — Low flows report series 8
and weirs (Gregory et al. 1991). Aquatic plants and animals may also have difficulty establishing
or maintaining their positions under such rapidly fluctuating conditions.
1.4
Height
4000
Discharge
3500
1.2
3000
1
Height (m)
2000
0.6
1500
Discharge (ML/day)
2500
0.8
0.4
1000
0.2
500
10/06/2010
9/06/2010
8/06/2010
7/06/2010
6/06/2010
5/06/2010
4/06/2010
3/06/2010
2/06/2010
0
1/06/2010
0
Figure 2: Hourly height and discharge data from GS113015A Tully River at Tully Gorge over a
10-day period, demonstrating fluctuations relating to hydroelectric water releases.
3.1.3. Responses
Modifications to flow characteristics and habitat conditions caused by dams and weirs can
negatively impact riparian and in-stream biota within impounded and downstream areas, causing
altered community composition and hence disrupting ecosystem function (Kingsford 2000;
Brunke et al. 2001; Bunn & Arthington 2002; Döll et al. 2009).
By acting as barriers to in-stream dispersal and increasing the distance between areas of
suitable habitat by drowning or drying out habitats, dams and weirs can significantly affect the
population viability of aquatic biota. Similarly, regular high-flow velocities from hydroelectricity
releases may also impede the upstream movement of biota who swim or crawl along the stream
bed, potentially leading to a loss of connectivity between populations.
Lowered habitat availability and loss of specific habitat types will impact biota dependent on
these habitats, and is likely to favour generalist species with wide habitat preferences. For
example, loss of riffle habitat can affect biota dependent on shallow high-flow-velocity habitat
such as filter-feeding aquatic macroinvertebrates that require high-flow velocity to collect food
from the moving water column. Similarly, increased flows through pools can impact biota
dependent on pool habitat such as aquatic macroinvertebrate ‘shredders’ that break down large
organic debris, some of which are dependent on still water conditions where leaf litter can
accumulate (Thoms & Cullen 1998; Bunn & Arthington 2002; Richardson & Humphries 2010).
The constant disturbance of the streambed caused by the increased frequency of high flows in
the case of hydroelectricity releases may cause a loss of taxa sensitive to physical disturbance,
and benthic biota dependent on stable streambed habitat.
NATIONAL WATER COMMISSION — Low flows report series 9
Reduced availability of coarse organic material through loss of downstream transport or
increased rate of transport through a reach can also impact aquatic community composition by
reducing the quantity of organic material settling onto the streambed, such as leaf litter and
woody debris. For example, a shift in the functional feeding group composition of
macroinvertebrate communities would be expected, with a loss of taxa belonging to the
‘shredder’ functional feeding group that consumes coarse organic material. An overall reduction
in nutrient availability would significantly impact trophic structure and ecosystem function.
Where baseflows are increased below dams and weirs, or where high flows from hydroelectricity
releases regularly interrupt baseflow conditions, macroinvertebrate taxa with a preference for
low flows may be disadvantaged. High-flow-preference taxa with physiological and/or
morphological adaptations to survive such conditions (e.g. hooks or suckers for attaching to
substrate) will be favoured. Similarly, in streams adapted to natural periods of low and/or no
flows, increased baseflows and reduced flow variability will disadvantage biota adapted to or
dependent on these conditions (e.g. species with desiccation-resistant lifecycle phases that rely
on dry benthic habitat). Reduced flow variability will also lead to the loss of flow-related lifecycle
cues related to dispersal and reproduction, hence disadvantaging a potentially large number of
native species. Loss of variability in the flow regime could lead to an overall loss of aquatic biota
with flexible life-history characteristics, with the consequence that ecosystems would lose
resilience. If flow regimes were to resume their natural patterns, biota may no longer be adapted
to survive natural seasonal extremes and unpredictable flow patterns (Choy et al. 2000).
As a result of water stratification, impounded or released water may differ in physical and/or
chemical properties to natural conditions. For example, it may be cooler or lower in dissolved
oxygen. The result of water impoundment and release could therefore be to disadvantage taxa
sensitive to water quality parameters, and disadvantage taxa tolerant of a wider range of water
physical and chemical conditions. The impact of dams and weirs may therefore be an overall
loss of sensitive taxa that depend on specific natural habitat and water quality conditions, and an
increase in generalist taxa with high disturbance tolerances (which often include exotic species).
The effect of changes to in-stream habitat and water quality conditions resulting from altered
flow conditions caused by dams and weirs can therefore reduce the abundance of taxa
dependent on specific in-stream habitats, flow conditions, or in-stream dispersal, which may lead
to the loss of affected taxa in reaches impounded by and below dams and weirs. Overall aquatic
community composition and diversity can therefore be affected within these systems, leading to
altered competitive interactions and foodweb dynamics, and ultimately disrupting ecosystem
function (Gregory et al. 1991; Bunn & Arthington 2002).
3.2. Water extraction
3.2.1. Pressure
Water is extracted from surface and groundwater sources within Australia for consumption and
to irrigate crops and water stock. It was estimated for the year 1996–97 that 2970 GL of water
was diverted from surface water sources in Queensland and 1623 GL extracted from
groundwater sources, which is likely to have significantly increased since then (DSEWPaC
2009). In dryland areas, considerable water is also harvested from floodplains during floods
(ANRA 2000). Additional water extraction occurs on privately owned land for stock and domestic
use, however these levels of extraction are not well documented.
Surface water extraction in Queensland is highest within catchments in the state’s south-east
(Figure 3). There is also considerable surface water use in the Condamine and Border Rivers
basins, which flow into the Murray-Darling Drainage Division. There are eight major irrigation
areas throughout the state. Areas of high water extraction in Queensland do not correspond with
NATIONAL WATER COMMISSION — Low flows report series 10
high rainfall areas. For example, there is high mean annual flow into the Gulf of Carpentaria but
little water extraction in this region at present (DSEWPaC 2009). This pattern of mismatched
supply and demand puts considerable strain on water resources within high-use regions.
Figure 3: Surface water development status based on water diversion levels (as a proportion of
the long-term mean of annual pre-development discharge) within National Land and Water
Resources Audit (2002) catchment reporting units.
The sustainable annual yield of groundwater for Queensland is estimated to be 2784 GL
(DSEWPaC 2009). Current statewide levels of groundwater use are below this sustainable level,
but levels of groundwater extraction within each discrete management unit are not sustainable in
all cases. Thirty-three per cent of management units experience extraction levels that are above
sustainable. The greatest yields were available in the Great Artesian groundwater management
unit, with 1017 GL allocated per year. Levels of groundwater development in this unit were
classified as low (less than 30 per cent of sustainable yield) to moderate (40 to 70 per cent of
sustainable yield). Allocations in the Tasman groundwater management unit, which spans much
of the state’s east coast, were estimated at 435 GL per year, with levels of development from
low to moderate. Levels of groundwater development within south-east Queensland were
classified as moderate. A total of 78 GL of groundwater was extracted per year within the
Clarence-Morton groundwater management unit located in the south-east (DSEWPaC 2009).
Growing industries, such as mining or manufacturing, can cause short- or long-term demands for
water extraction. Past experiences indicate that demands for water from urban, industrial and
agricultural users will increase. Unmanaged growth in demands will place considerable strain on
available water resources (ANRA 2000).
3.2.2. Stressors
Surface water and groundwater extraction can significantly alter flow characteristics within
affected streams (McKay & King 2006; Döll et al. 2009). Surface water extraction can reduce
baseflows and increase the incidence and magnitude of low flows and alter their timing,
NATIONAL WATER COMMISSION — Low flows report series 11
producing ‘artificial drought’ conditions within surface streams and connected hyporheic zones
(Richter et al. 1998; Boulton 2003; Deitch et al. 2009; Finn et al. 2009). Groundwater extraction
affects natural groundwater levels, which means that greater quantities of water are required for
aquifer recharge. There are many linkages between groundwater, hyporheic and surface water
zones, and extraction from either surface or groundwater sources will consequently impact all
three zones (Boulton et al. 2003; Hancock & Boulton 2005). Groundwater extraction therefore
has the potential to increase the duration and frequency of low- and no-flow periods in surface
streams. Water extraction from temporary streams during the dry season can cause increased
incidence and duration of no-flow periods. Extreme water extraction within perennial streams
can generate no-flow periods in systems not adapted to them (Mackay et al. 2012; Marsh et al.
2012).
Water extraction from surface streams reduces total water volume and water depth, and this
may cause narrowing of the stream channel (Richter et al. 1998; Finn et al. 2009). Reduced
water volume represents a loss of total aquatic habitat, and can also alter the types of habitats
present. Narrowing of the stream channel reduces the wetted area, which lowers hyporheic
habitat availability (Hancock & Boulton 2005; Stubbington et al. 2009). Loss of riffle habitats can
result from decreases in water discharge, and pools are likely to become the predominant
habitat type in such circumstances. This in turn may alter water quality, as stagnant pools are
subject to stratification and changing temperature, nutrient and ionic conditions. Still water also
promotes algal growth, which is associated with increased diurnal variability in dissolved oxygen
levels (Finn et al. 2009). Extended no-flow periods affect the persistence of refugial waterholes
(DERM 2010b), increasing the distance between available aquatic habitats (Pusey & Arthington
2003). Water evaporation during no-flow periods can concentrate in-stream nutrient and salinity
levels. Benthic sediment build-up may also occur due to a lack of flushing, which can further
decrease water quality by increasing in-stream turbidity and nutrient levels, and reduce the
availability of benthic habitats (Hancock & Boulton 2005).
Reduced area and volume of surface water reduces recharge and exchange with the hyporheic
zone and groundwater storages (Boulton et al. 2003; Hancock & Boulton 2005). Weakened
exchange between the stream and hyporheic zone can starve the interstitial environment of
oxygen and promote reducing conditions, altering the nutrient dynamics of surface stream
ecosystems (Hancock & Boulton 2005). High levels of water extraction from ephemeral streams
can reduce lateral connectivity between river channels and associated floodplains, and decrease
the frequency of longitudinal connectivity between refugial waterholes. This can severely impact
nutrient availability within affected systems and limit dispersal opportunities for aquatic biota
(Gregory et al. 1991; Thoms & Cullen 1998; Allibone 2000; Marsh et al. 2012).
3.2.3. Responses
The increased incidence, magnitude and duration of low- and no-flow periods resulting from
water extraction can severely impact aquatic ecosystem function, and a decline in aquatic
biodiversity is expected (Naiman & Latterell 2005; Rolls et al. 2010). Within perennial
ecosystems not well adapted to low and no flows, many aquatic species may lack the
physiological and life-history characteristics needed to survive such conditions. Within temporary
systems, biota are likely to be more resilient to the impacts of water extraction, and impacts are
dependent on their levels of tolerance, which will vary between taxa (Poff et al. 1997; Bunn &
Arthington 2002; Dewson et al. 2007). Similarly, alterations to water quality may cause in-stream
conditions to deteriorate beyond the tolerance levels of some species, particularly the sensitive
taxa such as Plecoptera (stonefly), Ephemeroptera (mayfly) and Trichoptera (caddisfly) (PET)
(Gregory et al. 1991). In general, water extraction is likely to more heavily impact biota
dependent on high-flow velocities, such as filtering macroinvertebrate species (Thoms & Cullen
1998; Bunn & Arthington 2002; Richardson & Humphries 2010).
NATIONAL WATER COMMISSION — Low flows report series 12
Loss of habitat availability and connectivity resulting from water extraction also has the potential
to significantly impact aquatic ecosystems. The overall reduction in habitat availability resulting
from water extraction concentrates individuals within smaller areas, heightening biotic
interactions. High mortality can result within contracted aquatic ecosystems, as the biotic
pressures of competition, predation and disease are increased (Arthington et al. 2010). The loss
of longitudinal connectivity inhibits the dispersal of aquatic biota, therefore reducing gene flow
within and between affected populations (Allibone 2000; Mackay et al. 2012). Any of the above
impacts can lead to a loss of taxa within affected streams, and hence cause lowered aquatic
biodiversity, disrupted foodweb dynamics, and lowered ecosystem productivity and function
(Gregory et al. 1991; Poff et al. 1997; Bunn & Arthington 2002; Arthington et al. 2010). Altered
habitat conditions may also favour exotic species (Koehn 2004). Lowered lateral connectivity
within temporary systems as a result of reduced high-flow and increased low-flow events could
also impact nutrient availability within streams, interfering with primary production pathways, and
hence disrupting foodweb dynamics and related properties such as system energetics and
carrying capacity (Gregory et al. 1991; Thoms & Cullen 1998).
The ecological impacts of surface water extraction, in conjunction with the impacts of dams and
weirs in the case where they reduce downstream flows, are summarised in the conceptual
model below (
Figure 4). At present no general conceptual models have been developed for the ecological
impacts of groundwater extraction.
Figure 4: Conceptual model of the ecological impacts of surface water extraction in conjunction
with the impacts of dams and weirs in the case where they reduce downstream flows [source:
DERM 2008].
NATIONAL WATER COMMISSION — Low flows report series 13
3.3. Interbasin water transfer
3.3.1. Pressure
The Queensland Government has developed several large-scale initiatives for the state’s southeast collectively known as the Water Grid. The Water Grid connects water supplies, storages
and treatment plants across south-east Queensland: from Noosa in the north to Coolangatta in
the south, and to Toowoomba in the west. This includes the transfer of water between different
basins in south-east Queensland. Water in the Wivenhoe Dam, in the Brisbane River catchment,
is now supplied to Toowoomba via the Wivenhoe-Toowoomba Pipeline. Similarly, the Southern
Regional Water Pipeline connects the Brisbane River catchment with the desalination plant at
Tugun on the Gold Coast; and the Northern Pipeline Inter-connector links water between the
Sunshine Coast and Brisbane (Queensland Water Commission 2010). Interbasin transfer
schemes also exist in other parts of the state, including the Pioneer irrigation area near Mackay
(Marshall et al. 2000) and the Mareeba-Dimbulah irrigation area near Cairns (Ryan et al. 2002).
These ‘interbasin water transfers’ can cause changes to the flow regime, and pose a risk to
aquatic ecosystems (Snaddon et al. 1998; Page et al. 2010).
3.3.2. Stressors
Interbasin water transfers affect source and recipient rivers differently (Rolls et al. 2010). Flows
in the recipient basin are supplemented, potentially reducing the duration of no- and low-flow
periods and increasing the duration of medium to high flows. In extreme cases, interbasin water
transfers could transform recipient rivers from ephemeral into perennial systems (Negus 2007).
In contrast, flows in the source basin are depleted and high and medium flows may be reduced
(Davies et al. 1992), with an increased frequency of low flows. The flow-related stressors and
responses in source rivers are expected to be similar to those experienced as a result of water
extraction (see Section 3.2). The flow-related stressors and responses in recipient rivers are
expected to be similar to those experienced by constant water releases downstream of dams
and weirs (see Section 3.1).
3.3.3. Responses
In addition to the responses shown in the water extraction and dams and weirs PSR models,
interbasin water transfer is also expected to affect source rivers by translocating biota within
transferred water. This may act to create pest species problems within recipient rivers. Exotic
pest species or diseases, which occur within source rivers but were not previously found within
recipient rivers, may be transferred. Native aquatic species found within source rivers may also
become pests within recipient rivers if they were not previously found there, and even co-existing
species can be affected by introducing individuals from differing genetic source populations
(Davies et al. 1992; Gibbins et al. 2000; Marshall et al. 2000; Page et al. 2010).
3.4. Disposal of coal seam gas industrial water
into surface streams
3.4.1. Pressure
Coal Seam Gas (CSG) is a natural gas adsorbed onto coal deposits and kept in place by
pressure from surrounding groundwater. Extraction of CSG is achieved by dewatering coal
seams to reduce the constraining pressure and release gas from the coal. This produces large
amounts of wastewater which then requires disposal. CSG extraction is a relatively new industry
in Queensland, and poses a new type of water resource issue. Whereas in the past water
NATIONAL WATER COMMISSION — Low flows report series 14
resource management has typically been concerned with the allocation and storage of naturally
occurring surface and groundwater supplies, the problem of excess water is relatively new. The
CSG industry is rapidly expanding over large areas of Queensland, and this has prompted the
need to develop appropriate guidelines for the disposal of wastewater. CSG production activities
are expected to steadily increase until about the year 2018, at which time the industry is
expected to produce up to 130 GL of wastewater a year (McGregor et al. 2011; Takahashi et al.
2011).
DERM released its CSG Water Management Policy in June 2010, in which the preferred option
for managing CSG wastewater was treatment followed by either injection into aquifers or
beneficial use. It is possible, however, that instances may occur, especially in the short-term,
where there is excess wastewater unable to be disposed of via the preferred methods. It has
been proposed that in such cases, excess water could instead be released into surface waters.
In such cases, strategies would be required to ensure water quality objectives are met and
ecological values of recipient streams are protected (McGregor et al. 2011; Takahashi et al.
2011).
Due to the novel nature of the CSG industry in Queensland, the impacts of wastewater release
into surface waters are not yet completely understood, and empirical data is lacking. The
Australian Government has allocated research funding to ensure that possible environmental
impacts resulting from the CSG industry are understood and are kept below acceptable
thresholds of environmental change. DERM has and continues to contribute to this research
(McGregor et al. 2011; Takahashi et al. 2011) including a study to inform biological monitoring
(Takahashi et al. 2011), and development of guidelines for managing flow regimes within CSGaffected systems (McGregor et al. 2011). These two components were used to inform this PSR
model relating the likely impacts of CSG to flow alterations.
The state’s current CSG resources are concentrated in and around the Queensland portion of
the Murray-Darling Basin (QMDB) in the south-west. Freshwater systems in this area include the
Border Rivers, Moonie, Balonne-Condamine, Warrego, Paroo and Bulloo drainage basins. Most
of Queensland’s CSG resources underlie the Condamine-Balonne drainage basin in the QMDB
and the Fitzroy drainage basin located to the QMDB’s north (McGregor et al. 2011; Takahashi et
al. 2011). Natural flow regimes within the QMDB were described as being unpredictable and/or
intermittent according to Kennard et al. (2010). The predominant flow class within the QMDB
was ‘unpredictable summer highly intermittent’ (flow class 11). Only one naturally perennial
stream was identified in the QMDB (Kennard et al. 2010; McGregor et al. 2011).
3.4.2. Stressors
Takahashi et al. (2011) used a PSR approach to investigate the likely biological impacts of CSG
wastewater introduction into surface streams, and to inform appropriate monitoring guidelines.
They determined that changes can be expected to all of the important facets of the natural flow
regime, including flow duration, timing, variability, predictability, magnitude, and rate of rise and
fall. Two main flow-related stressors were identified: 1) alteration to hydrology leading to a
decrease in dry spells, and 2) alteration to hydrology leading to constant flow and decreased
seasonality. The introduction of CSG wastewater into surface streams therefore has the
potential to alter all aspects of the low-flow regime, particularly by decreasing the magnitude and
frequency of low flows, and altering the timing and variability of low flows. Within the QMDB,
increased baseflow through CSG wastewater introduction has the potential to change ephemeral
streams to perennial ones (McGregor et al. 2011). Flow changes will alter in-stream habitat
availability, with modifications in riffle-pool sequences and loss of dry river beds (Bunn &
Arthington 2002; Cockayne et al. 2010; McGregor et al. 2011). McGregor et al. (2011) modelled
flow data based on water resource availability under current resource operations planning (ROP)
and compared this to flow data from projected CSG water releases for the Dawson River,
NATIONAL WATER COMMISSION — Low flows report series 15
Queensland. Results indicate key changes in habitat and flow characteristics likely under current
CSG projections (Table 2).
Table 2: Flow metrics expected for two sites in the Dawson River under current water resource
development (ROP) and projected goal seam gas (CSG) flow scenarios (from McGregor et al.
2011).
Location
Flow metric
ROP scenario
CSG scenario
Baroondah
% time as a riffle
71
94
% time dry
24
0
% time drowned
5
6
Average duration (days) riffle-forming flows
19
85
Max. duration (days) of riffle-forming flows
390
873
Average dry spell (days)
8
0
Max. dry spell (days)
148
0
% time as a riffle
27
86
% time dry
60
0
% time drowned
13
14
Average duration (days) riffle-forming flows
8
53
Max. duration (days) of riffle-forming flows
87
426
Average dry spell (days)
45
0
Max. dry spell (days)
415
0
Taroom
3.4.3. Responses
Aquatic biota are adapted to the temporary flow regime of streams in the QMDB and are
therefore likely to be affected by the additional flow generated by CSG water releases. Increased
streamflow and the reduced incidence of no-flow and low-flow spells are expected to negatively
impact species reliant on dry or low-flow conditions for parts of their lifecycle, as well as increase
the potential for invasive species to colonise affected systems (Koehn 2004; Takahashi et al.
2011). Dry river beds are unique habitats colonised by a range of terrestrial biota, and these
biota will be displaced if this habitat is artificially inundated by CSG water (Steward et al. 2011).
Altered flow conditions resulting in more permanent availability of aquatic habitat and altered
light availability are also expected to increase macrophyte growth in these areas (Takahashi et
al. 2011), which will alter production pathways within affected systems (Gregory et al. 1991).
Loss of flow seasonality is expected to impact aquatic species dependent on flow-related cues
for reproduction and dispersal. Loss of native species may occur, resulting in altered species
richness, community composition, foodweb dynamics and overall ecosystem function (Gawne &
Scholz 2006; Takahashi et al. 2011). Impacts resulting from flow alteration will be further
compounded by stressors resulting from altered water quality as a result of CGS wastewater
(Takahashi et al. 2011).
Likely alterations to hydraulic conditions and water quality were summarised by Takahashi et al.
(2011) within a conceptual model ( Figure 5). The magnitude and level of risk posed by each of
these stressors and responses needs to be measured in areas where this pressure occurs to
gain an accurate insight into the significance of this pressure (Takahashi et al. 2011).
NATIONAL WATER COMMISSION — Low flows report series 16
Figure 5: Conceptual model illustrating alterations to hydraulic conditions and water
quality resulting from coal seam gas wastewater disposal into surface waters [source:
Takahashi et al. 2011].
NATIONAL WATER COMMISSION — Low flows report series 17
3.5. Climate change
3.5.1. Pressure
Australian water resource management needs to incorporate the risk of climate change, which
threatens to significantly alter future water availability (Khan 2008). Climate change in Queensland is
predicted to cause annual mean temperatures to increase, annual precipitation to decrease, and
annual runoff to decrease (Chiew & McMahon 2002; CSIRO & BOM 2007; DERM 2010a). Climate
change predictions also indicate an increase in extreme events, such as floods and droughts
(Hennessy et al. 2007; CSIRO & BoM 2007; DERM 2010a).
3.5.2. Stressors
The increased rates of evaporation due to higher temperatures, lower runoff and the increased
occurrence of droughts may result in a reduction in the magnitude, frequency and duration of
streamflows. Reductions in river flow may be represented by a decrease in the duration of natural
high to medium flows, and an increase in the duration of low-flow and no-flow events (Chiew &
McMahon 2002). This will lead to a loss of aquatic habitats, particularly high and medium flow habitats
such as riffles, and reduced persistence of aquatic habitat such as refugial waterholes during the dry
season (Bunn & Arthington 2002). Many of the stressors and responses will therefore be similar to the
impacts of water extraction (see the water extraction PSR model in Section 3.2). There may also be
an increase in the frequency of extreme high-flow events, which could significantly disturb stream
morphology and sedimentation rates, and alter in-stream habitat stability and availability.
3.5.3. Responses
Changes in flow and habitat conditions are likely to cause a shift in the community composition of
aquatic biota (Poff et al. 1997; Marshall et al. 2001). The availability of suitable habitats for taxa with
preferences for high or medium flows may be reduced, whereas habitat for taxa with no flow
preference, or a preference for low flow, may be increased. Reduced availability of refuges during
droughts would affect aquatic biota with limited dispersal capabilities or low resistance to desiccation
(Bunn & Arthington 2002). Extreme flood events could further disrupt community composition and
ecosystem function by disrupting habitat availability, particularly taxa dependent on stable benthic
habitat, similar to the impacts of dams and weirs in the case of hydroelectricity releases (see dams
and weirs PSR model in Section 3.1).
NATIONAL WATER COMMISSION — Low flows report series 18
4. Case studies
4.1. Walla Weir case study
4.1.1. Introduction
This case study analyses biological responses resulting from construction of the Walla Weir, located
on the Burnett River in south-east Queensland. Results suggest low-flow ecological linkages, which
are expressed via changes to habitat characteristics, are important within this region’s riverine
ecosystems.
The Burnett River catchment includes the Bundaberg Irrigation Area, and is classified as highly
developed, with 70 to 100 per cent of mean annual discharge diverted for use. Aquatic ecosystems
within the river are therefore expected to have undergone a high level of modification as a result of
these developments (Stazner & Higler 1986; Bunn & Arthington 2002). The Walla Weir is an
impoundment on the Burnett River, built for the purpose of supplying irrigation water to the South
Burnett region. Its construction was completed in 1998. Impounded flows and controlled releases are
expected to have altered habitat conditions and impacted aquatic biota downstream and immediately
upstream of the Walla Weir (Rolls et al. 2010; Bunn & Arthington 2002).
The ecological impacts of the Walla Weir’s construction and operation were assessed when
construction was proposed. This included an analysis of the impacts on aquatic macroinvertebrate
communities upstream and downstream of the weir. The study comprised several reports that
assessed the effects on macroinvertebrate communities before and during construction, as well as
during operation (see Choy 1997a, b; Choy & Marshall 1999a, b, 2000). These studies investigated
the implications of flow alteration on macroinvertebrate communities within the Burnett River.
We intended, for the purposes of this case study, to couple the previous analysis of macroinvertebrate
responses to the Walla Weir from the time of its construction with a new hydrological analysis of the
nature of flow alterations caused by the weir since its construction – with a focus on low flows – as
was conducted in the Wivenhoe Dam case study. Unfortunately, the gauging station located below
the weir (GS 136008A) ceased operation during January 2000, and no other active gauging stations
are present within the Burnett River downstream of the weir. This means that adequate data were not
available to assess flow characteristics downstream of the weir in the period after its construction.
Flow-ecological relationships were therefore only able to be inferred based on the responses of
macroinvertebrates documented in the aforementioned studies.
Based on the general PSR conceptual model for dams and weirs, a number of expectations about the
impacts of the Walla Weir on macroinvertebrate communities were formed. As a result of the
impoundment of water and release patterns designed to meet irrigation demand, it was expected that
riffle habitats would be drowned out above the weir and dried out below the weir at times and for
durations that differ from the pre-development case. Because the availability of riffle habitat is thus
likely to have been highly modified, riffle taxa were expected to be highly impacted, with reduced
abundance of riffle-dependent taxa, and lowered diversity within remaining riffle habitats. With respect
to hydrology, it was expected that flow downstream of the weir would change from natural conditions,
with a reduction in the frequency and duration of all flows, including low flows. It was expected that
the frequency and duration of cease-to-flow periods would increase and the timing of events would be
altered to match irrigation demand periods rather than rainfall patterns.
4.1.2. Methods
The methods used in the macroinvertebrate flow requirements analysis conducted by Choy and
Marshall (2000) are summarised below, as relevant to this report. A full account of the methodology is
given in the relevant reports (Choy 1997a, b; Choy & Marshall 1999a, b, 2000).
NATIONAL WATER COMMISSION — Low flows report series 19
Study sites
Study sites were chosen within the Burnett River catchment to represent sites that would be situated
upstream, downstream and impounded by the Walla Weir after its construction. A total of nine study
sites were sampled in the Burnett River catchment: three downstream, two impounded, and four
upstream of the Walla Weir (
Figure 6; Table 3). All sites were located on the Burnett River except the site Currajong Creek at
Wallaville (GS1360076), which drains into the river. One of the downstream sites, Burnett River at
Cedars Crossing (GS1360077), while on the Burnett River, was located in another impoundment: the
Ben Anderson Barrage.
Figure 6: Locations of sampling sites, numbered according to gauging station, in the Burnett River
catchment. The Walla Weir is situated between sites GS1360069 and GS136023A.
NATIONAL WATER COMMISSION — Low flows report series 20
Table 3: Site numbers and names of macroinvertebrate sampling sites within the Burnett River
catchment.
Position relative to the Walla Weir
Gauging station number
Site name
Downstream
1360077
Burnett River at Cedars Crossing
Downstream
1360078
Burnett River at Drinans Crossing
Downstream
1360069
Burnett River at Walla Weir tailwater
Impounded
1360076
Currajong Creek at Wallaville
Impounded
136001B
Burnett River at Walla
Upstream
1360071
Burnett River at Booyal Crossing
Upstream
136007A
Burnett River at Figtree Creek
Upstream
136012A
Burnett River at Mingo Crossing
Upstream
136002D
Burnett River at Mt Lawless
Sampling
Sampling was conducted on five occasions: twice before construction, once during construction, and
twice during operation of the Walla Weir. Each site was sampled between three and five times (Table
4). Aquatic macroinvertebrates were sampled from two habitat types – riffles and edges – where
present at each site. Wherever both edge and riffle habitats were present, macroinvertebrates were
sampled from both habitats. Riffle habitats were not sampled in impounded sites during the operation
of the weir because they were flooded and therefore no longer present (Table 4). Similarly, only one
downstream site (GS1360069) had riffles present during operation of the Walla Weir. Samples were
live-picked on-site and macroinvertebrates stored in ethanol. Macroinvertebrates were enumerated
and identified in the laboratory, mostly to family level.
Table 4: Time of sampling and habitats sampled within each site in the Burnett River catchment.
Sampling was conducted on five occasions: before, during and after construction of the Walla Weir.
Habitats sampled were E: edge, and R: riffle.
Sampling period
Site
Pre-
Pre-
construction
construction
Construction
Operation
Operation
number
Site position
(Oct ‘96)
(May ‘97)
(June ‘98)
(May ‘99)
(Nov ‘99)
1360077
Downstream
E
E
E
E
E
1360078
Downstream
E, R
E, R
E
E
E
1360069
Downstream
R
E, R
E, R
E, R
1360076
Impounded
R
E, R
E, R
E
E
136001B
Impounded
R
E, R
E, R
E
E
1360071
Upstream
E, R
E, R
E, R
136007A
Upstream
R
E, R
E, R
E, R
E, R
136012A
Upstream
E, R
E, R
E, R
E, R
E, R
136002D
Upstream
E, R
E, R
E, R
E, R
Data analysis
Analyses were used to assess the response of macroinvertebrate communities to assumed changes
in habitat and flow conditions resulting from the construction and operation of the Walla Weir. Analysis
of the flow-velocity-preference group structure of macroinvertebrate communities at each site was
performed, based on the flow-preference allocation of Marshall and Marshall (in prep.). Multivariate
NATIONAL WATER COMMISSION — Low flows report series 21
analyses were based on Bray-Curtis similarity matrices of the abundance of all taxa present. Cluster
analyses were conducted separately for edge and riffle habitats and used to create dendrograms – to
see if samples collected during the last sampling period (post-construction) separated according to
site position with respect to the weir.
4.1.3. Results
Flow-velocity-preference group structure
In the impounded sites, only the edge habitats could be sampled during operation of the Walla Weir
because riffle habitats were no longer present after it was built. No high-flow-preference taxa were
found in the impounded sites during operation of the weir. These sites had high-flow-preference taxa
present in both edge and riffle habitats before the weir was constructed (Figure 7, Figure 8).
In edge habitats upstream of the Walla Weir, taxa from all categories of flow preference (i.e. high-,
low-, and no-flow-velocity preference) were present. In contrast, sites downstream of the weir had
fewer high-flow-preference taxa, with either small proportions of, or no high-flow-preference taxa
present (Figure 7). This was, however, the case both before and after construction of the weir. There
was little change in the proportions of flow-preference-taxa present in edge habitats before and after
construction of the weir.
In the riffle sites upstream of the Walla Weir, high-flow-preference taxa were present on all sampling
occasions, but there were only small proportions of, or no low-flow-preference taxa present (Figure 8).
In contrast, there were higher proportions of low-flow-preference taxa present at the site (GS1360069)
immediately below the weir on all sampling occasions. This difference between upstream and
downstream sites was especially pronounced when the weir was in operation (sampling runs 4–5),
where there were no low-flow-preference taxa present in upstream sites. There was, however, little
change in flow-velocity-preference group structure in riffle habitats at downstream sites before and
after construction of the weir.
NATIONAL WATER COMMISSION — Low flows report series 22
Figure 7: Flow-velocity-preference group structure for edge habitat samples at sites in the Burnett
River in relation to the Walla Weir. Sample runs 1 and 2 refer to the pre-construction phase samples,
3 to the construction phase samples, and 4 and 5 to the operation phase samples.
NATIONAL WATER COMMISSION — Low flows report series 23
Figure 8: Flow-velocity-preference group structure for riffle habitat samples at sites in the Burnett
River in relation to the Walla Weir. Sample runs 1 and 2 refer to the pre-construction phase samples,
3 to the construction phase samples, and 4 and 5 to the operation phase samples.
Cluster analysis
Bray-Curtis measures of similarity in edge samples from the second post-construction period grouped
sites into distinct groups of impounded and non-impounded sites (Figure 9). The group of impounded
sites was more than 50 per cent different from other sites. This indicates that impoundment of habitats
by the Walla Weir significantly altered macroinvertebrate edge assemblages. Within the impounded
group, samples collected from sites above the weir were more similar to each other than the sample
from the site impounded by the Ben Anderson Barrage (GS1360077). The site impounded by the
barrage was approximately 36 per cent different from other sites, indicating that impoundment by the
weir had a larger impact on macroinvertebrate assemblages than impoundment by the Ben Anderson
Barrage.
NATIONAL WATER COMMISSION — Low flows report series 24
Bray-Curtis measures of similarity in riffle samples indicated that macroinvertebrates collected from
the one downstream site in which riffles were present after weir construction were significantly
different from upstream sites. The difference between the downstream and upstream sites was more
than 50 per cent (Figure 9). No other riffle samples were present in downstream or impounded sites.
This indicates that flow alteration by the weir not only altered the distribution of riffle habitats in
affected sites, but also significantly modified macroinvertebrate riffle assemblages within the
remaining riffles.
a)
b)
Figure 9: Dendrograms illustrating the separation of macroinvertebrate communities based on site
position from samples taken during the second operation phase of the Walla Weir within a) edge and
b) riffle habitats. Samples were grouped for sites where samples were taken from each habitat type. *
indicates the impounded area at Ben Anderson Barrage whereas other impounded areas were
impounded by the Walla Weir.
4.1.4. Discussion
Bray-Curtis measures of similarity in macroinvertebrate community composition from edge habitats
grouped the samples into two distinct groups: an impounded group and an upstream/downstream
group. When this result is compared with an earlier analysis by Choy (1997b) on macroinvertebrates
sampled in the same sites before the Walla Weir was constructed, there is an indication of change.
The pre-construction results (Choy 1997b) showed that all of the sites (except one) in the vicinity of
the proposed weir had similar taxa present; whereas post-construction results (Figure 9) showed that
macroinvertebrate composition at impounded sites had become different from non-impounded sites. It
is important to note, however, that the earlier analysis by Choy (1997b) included both edge and riffle
samples in Bray-Curtis measures of similarity, whereas the latter analysis presented in this study used
only edge samples – because after the weir was constructed there were no riffle habitats present
within impounded sites. Some of the difference between the two sampling periods may therefore be
attributable to this difference in data inclusion.
In the pre-construction results the site that differed from the other sites near the proposed weir was
the site impounded by the Ben Anderson Barrage (identified as Bingera Weir in Choy 1997b) –
indicating this impoundment had altered community composition at this site before the weir was built.
After construction of the weir, macroinvertebrate communities at this site were more similar to other
impounded sites than non-impounded sites, despite being located downstream of the weir. This
indicates that impoundment altered macroinvertebrate community composition in a consistent way.
NATIONAL WATER COMMISSION — Low flows report series 25
The composition of macroinvertebrate fauna based on Bray-Curtis measures of similarity from edge
habitats in sites impounded by the Walla Weir were different from those at non-impounded sites.
Choy and Marshall (2000) attributed these differences to a change in reach characteristics in the
impounded area from riverine to lacustrine. This is likely to reflect a change from flowing water to still
water, with macroinvertebrates responding to a loss of flow velocity in impounded habitat. This is
supported by analysis of the composition of macroinvertebrate fauna based on flow-velocity
preferences from edge habitats, which also indicated that sites impounded by the Walla Weir became
different from upstream sites after the weir was built, with a loss of high-flow-velocity-preference taxa.
Analysis of riffle habitat taxa post-construction showed that the site immediately downstream of the
weir (GS1360069: Walla Weir tailwater) was distinct from other sites upstream, in terms of
macroinvertebrate community composition. A 50 per cent difference in taxa between this site and
other riffle sites indicated the riffle at the tailwater site had been impacted by the weir (Choy &
Marshall 2000). This change may relate to either of two stressors. The isolation of this riffle from all
others by impoundment may have deprived it of macroinvertebrate colonists, particularly those that
use stream drift as a means of dispersal, leading to a different resident fauna. The likely alteration to
the frequency, timing and duration of low- and no-flow spells affecting this riffle may also have led to
the loss of riffle species sensitive to this stressor. However, there is no direct evidence to support
either mechanism, particularly in the absence of post-construction hydrology data.
The presence of high-flow-preference taxa in the edge habitats at sites upstream of the impounded
area of Walla Weir, and within impounded sites before construction of the weir, indicated that highflow-preference taxa were present in the parts of the river where flow was natural. This is because
natural flow conditions provide the flow velocity required by these species within their edge habitat.
The lack of high-flow-preference taxa within impounded sites after the weir was built could therefore
be directly attributed to a loss of flow velocity as a consequence of the altered flow regime imposed by
the weir. Season can be ruled out as a contributing factor, as high-flow-preference taxa were present
at upstream sites on all occasions.
The flow-preference structure of macroinvertebrate taxa in the riffle at the Walla Weir tailwater site,
just downstream of the weir wall, was expected to be similar to that of riffles upstream of the
impounded area due to similar habitat conditions. However, the proportion of low-flow-preference taxa
was higher than proportions observed at upstream sites on all sampling occasions. This indicates that
some level of difference between downstream and upstream sites exists naturally. However, the
difference in macroinvertebrate flow-preference structure between upstream and downstream sites
became more pronounced after construction of the weir. This was further evidence that the weir had
altered the flow-velocity-preference structure of macroinvertebrate communities. The likely stressors
that caused such a response could be reduced flow velocity and deteriorated quality of the water
released from the weir (possibly hypolimnetic water and/or water with low dissolved oxygen and/or
water being low in temperature). Once again, however, there is no direct evidence to support either
mechanism.
In a PSR context, the Walla Weir is the pressure in this instance. There are several hydrological
stressors emanating from the pressure, and they could be acting either individually or in combination
at various times. The stressors are:

The conversion of a riverine system into a lacustrine waterbody, with no flow velocity in the weir
pool.

Alteration of the natural low-flow regime downstream of the weir. This is likely to include alteration
of the timing of riffle-forming flows, and increased frequency and duration of periods when flows
are too low to form riffles that are functional as macroinvertebrate habitat.

Possible isolation by barrier of downstream riffles from macroinvertebrate colonisation via natural
dispersal pathways.
NATIONAL WATER COMMISSION — Low flows report series 26

Possible poor water quality downstream of the weir.

Low flows also often create another stressor in the form of sedimentation or siltation (McKay &
King 2006). Lack of flow means that sediments do not get flushed (Storey et al. 1991) and heavy
sedimentation on river beds leads to anoxic decomposition of organic matter and deterioration of
water quality.
The biotic responses to the above hydrological stressors that have been deduced in this study are a
high level of alteration in macroinvertebrate community composition within impounded sites, including
the total loss of high-flow-preference macroinvertebrate taxa in the impounded zone where flow
velocity is likely to be near zero, and the loss of riffle taxa in the impounded zone (as riffles no longer
exist). Downstream of the weir, the biotic response was a shift toward a higher proportion of low-flowpreference taxa. The reduced number of riffles throughout the Burnett River catchment due to
construction of the Walla Weir means a reduction in riffle biota in the catchment, and a reduction in
connectivity between the remaining riffle populations.
4.2. Wivenhoe Dam case study
4.2.1. Introduction
This case study analyses changes in flow characteristics resulting from operation of the Wivenhoe
Dam, and subsequent biological responses from aquatic biota within the Brisbane River. A
hydrological analysis was performed to identify the nature of low-flow alterations that have resulted
from dam operation. Biological responses to these changes in flow were investigated using
macroinvertebrates as biological indicators. Results indicate the nature of flow-ecological linkages
present within south-east Queensland riverine ecosystems.
The Brisbane River catchment is classified as highly developed, with between 70 to 100 per cent of
the catchment’s mean annual discharge diverted for use. A number of significant water resource
developments control water flow within the catchment, namely the Wivenhoe Dam, Somerset Dam,
and Mt Crosby Weir system (DSEWPaC 2009). Aquatic ecosystems within the Brisbane River
catchment are therefore expected to have undergone a high level of modification as a result of these
developments (Statzner & Higler 1986; Bunn & Arthington 2002).
The Wivenhoe Dam is currently the largest dam in south-east Queensland. The dam was designed by
the Water Resources Commission and completed in 1985. It provides most of Brisbane’s drinking
water and performs an important function in flood mitigation (Douglas et al. 2007; Seqwater 2009).
The dam has significantly altered the Brisbane River’s habitat and flow characteristics upstream and
downstream of the dam wall. The establishment of Lake Wivenhoe as a result of dam construction
has significantly altered habitat conditions upstream of the dam, inundating an area of approximately
33.75 km2 at full storage capacity (Seqwater 2009). The dam wall is a significant barrier to the
movement of biota through the system. For example, there has been no evidence of recruitment of
eels or mullet upstream of the wall since construction of the dam, as both are diadromous species
that require connection between the river and the sea (Arthington et al. 2000). Downstream of the
dam, controlled release of water has significantly altered the natural flow regime with increased and
more consistent flows (Choy et al. 2000; Seqwater 2009). All aspects of the natural flow regime are
expected to have been altered as a result of water releases, including the magnitude, timing,
frequency, seasonality and variability of flow, and similar low-flow characteristics (Lloyd et al. 2004;
Bunn & Arthington 2002; Rolls et al. 2010). The Wivenhoe Dam is therefore expected to have
impacted aquatic ecosystems within the Brisbane River downstream of the dam as a result of altered
flow conditions, and corresponding alterations to in-stream habitat conditions (Statzner & Higler 1986;
Choy et al. 2000; Bunn & Arthington 2002).
During development of the water allocation management plan for the Brisbane River catchment, an
analysis of flow requirements for aquatic macroinvertebrates was undertaken to investigate their
responses to flow alterations caused by the Wivenhoe Dam (Choy et al. 2000). Although the study
NATIONAL WATER COMMISSION — Low flows report series 27
was designed for the purpose of informing environmental flow requirements, the ecological
consequences of altering the natural flow regime that were investigated provide information about the
relationships between aquatic biota and low flows, and the ecological impacts associated with altering
the low-flow regime. This case study used results from the macroinvertebrate environmental flow
requirements study in conjunction with a new hydrological analysis at corresponding study sites to
investigate low-flow ecological linkages.
A reference site approach was used to determine the natural low-flow regime and ecological state of
the Brisbane River using low-flow indices similar to those outlined by the NWC (2011) and using
macroinvertebrates as biological indicators. In the hydrological analysis, the nature of flow alterations
specific to low flows caused by operation of the Wivenhoe Dam was investigated. The results of the
environmental flow requirements study were then used to analyse macroinvertebrate responses to
these alterations. Any differences between the actual and presumed natural ecological state of the
Brisbane River were analysed to determine their likely causes. Where the drivers of altered
macroinvertebrate community structure appeared to be in relation to changes in the natural flow
regime, it was reasonably assumed that these changes were related to the impacts of the Wivenhoe
Dam.
Based on the general PSR conceptual model for dams and weirs, a number of expectations about the
impacts of the Wivenhoe Dam on low-flow characteristics and macroinvertebrate community
composition were formed prior to analysis. As a result of consistent water releases downstream of the
dam, it was expected that the duration and frequency of low-flow events would be significantly
reduced compared with the natural flow regime. This was expected to affect in-stream habitat
availability by drowning out riffle habitats and deepening pools. Macroinvertebrate taxa dependent on
riffle habitat were expected to be the most highly impacted due to reduced habitat availability, as well
as lowered dispersal opportunities resulting from increased distances between areas of suitable
habitat. Lower abundance and diversity of riffle-dependent macroinvertebrate taxa were therefore
expected when compared with the reference condition.
4.2.2. Methods
Changes in macroinvertebrate community composition
A summary of the methods conducted in the study of macroinvertebrate flow requirements within the
Brisbane River by Choy et al. (2000) are presented below, as relevant to this report. A full account of
the methodology is presented in Choy et al. (2000).
Study sites
Study sites were chosen within the Brisbane River catchment to be representative of sites that were
both regulated and unregulated by the Wivenhoe Dam. A degree of water regulation was present in
two of the ‘unregulated’ sites, but to a much smaller degree than at sites influenced by the Wivenhoe
Dam, and usually with an opposite effect on flow – whereby water was extracted from the stream
rather than supplemented, as was the case downstream of the Wivenhoe Dam. Subsequent impacts
of water regulation within unregulated sites on macroinvertebrate communities were therefore
expected to be minimal, and if present, in the opposite direction of impacts to sites downstream of the
Wivenhoe Dam. These sites also grouped with unregulated sites during multivariate analysis, and are
therefore considered to be unregulated sites for the purpose of this study.
Sites unregulated by the Wivenhoe Dam are upstream of the dam, and regulated sites are
downstream. This introduces the possibility that any differences between regulated and unregulated
sites may be due to natural differences in relation to river position, in accordance with the river
continuum concept (Vannote et al. 1980). To aid in the separation of differences between sites and
identify which differences are natural and which can be attributed to the impacts of the dam, a
reference approach was used. The Logan River was chosen to represent the reference condition, as
it is a nearby river that drains into Moreton Bay as does the Brisbane River, and it is relatively
NATIONAL WATER COMMISSION — Low flows report series 28
unregulated compared with the Brisbane. At the time of the study, the Logan River had one dam and
two weirs, but these had altered in-stream hydrology by only a small amount, with flows estimated to
have been reduced by 5 to 10 per cent below the pre-development state (Ruffini & Pandeya 1996).
In comparison with the lower Brisbane River, flow regulation within the Logan River is of a much
smaller magnitude and consists of water extraction rather than supplementation. Sampling within the
Logan River was conducted (also by DERM) in an earlier study of macroinvertebrate flow
requirements, as published in Choy and Marshall (1997). Sampling methodology was consistent
between the Brisbane and Logan river sites.
A total of eight study sites – four regulated and four unregulated – were sampled within the Brisbane
River catchment (Figure 10, Table 5). A total of 10 study sites were sampled within the Logan River
catchment – five upstream and five downstream – at positions approximating the position of study
sites in the Brisbane River catchment (Figure 10, Table 5) (from Choy & Marshall 1997).
NATIONAL WATER COMMISSION — Low flows report series 29
Figure 10: Map indicating the location of macroinvertebrate sampling sites within the Brisbane River
and Logan River catchments. Site numbers are gauging station numbers. The Wivenhoe Dam is
located at the base of Lake Wivenhoe. Red triangles in the Brisbane River catchment indicate
regulated sites downstream of the Wivenhoe Dam, and blue triangles indicate unregulated sites.
Green triangles in the Logan River catchment indicate upstream sites and yellow triangles indicate
downstream sites.
NATIONAL WATER COMMISSION — Low flows report series 30
Table 5: Gauging station numbers and site names of macroinvertebrate sampling sites within the Brisbane River and Logan River catchments.
Unregulated sites
Regulated sites
Catchment
Gauging station number
Site name
Gauging station number
Site name
Brisbane River
1430050
Brisbane River at Crossing 26
143001C
Brisbane River at Savages Crossing
143921A
Cressbrook Creek at Rosentretors Crossing
1430060
Brisbane River at Atkinsons Crossing
143009A
Brisbane River at Gregors Creek
1430061
Brisbane River at Burtons Bridge
1430063
Pryde Creek below Splityard Creek Dam
1430062
Brisbane River at North Kholo
Upstream sites
Logan River
Downstream sites
Gauging station number
Site name
Gauging station number
Site name
1450044
Logan River at Peters Place
1450050
Logan River at Paynes
145003B
Logan River at Forest Home
145014A
Logan River at Yarrahappini
1450048
Logan River at Wyaralong
1450049
Logan River at Bromelton
1450043
Running Creek at Drynans
145020A
Logan River at Rathdowney
1450046
Running Creek at Campsite
1450045
Logan River at Maroon Dam
NATIONAL WATER COMMISSION — Low flows report series 31
Sampling
Sites within the Brisbane River were sampled between one and five times each, between
October 1994 and May 1997 (Table 6). At each site physical and water quality data were
recorded and macroinvertebrates collected, all within a 100 m reach. Physical data recorded
included single readings of stream width, depth, velocity, substratum composition, riparian
cover, adjacent land use, and proportion of algal and macrophyte cover. Water quality
measurements included water temperature, conductivity, pH, dissolved oxygen and turbidity.
Aquatic macroinvertebrate samples were collected from four different habitats, but these were
not always present at all sites. Habitats included riffle/run, edge, pool and macrophyte beds.
Each habitat was defined as follows: riffles were relatively shallow (<30 cm deep) and fastflowing areas of broken water, usually over stony substrates. Runs were similar to riffles but
tended to be deeper and over varying substrates. Edges were defined as habitats along the
bank where there is little or no current – some terrestrial vegetation or tree roots may be
present along the edges, but no aquatic vegetation. Pools were zones of relatively deep and
either stationary or very slow-flowing water, over varying substrates. Macrophytes were
regions of dense aquatic vegetation, preferably a distance from the banks. Samples were livepicked on-site and macroinvertebrates stored in ethanol. Macroinvertebrates were
enumerated and identified in the laboratory, mostly to family level.
NATIONAL WATER COMMISSION — Low flows report series 32
Table 6: Time of sampling and habitats sampled within each site in the Brisbane River
catchment. R: riffle/run; E: edge; P: pool; and M: macrophytes. Blank spaces indicate no
samples were taken (adapted from Choy et al. 2000).
Sampling period
1
2
3
4
5
6
7
Oct
May
Sept
July
Nov
Feb
May
Site no.
Site name
'94
'95
'95
'96
'96
'97
'97
1430050
Brisbane River
at Crossing 26
R, E,
P, M
R, E,
P, M
P, M
R, E,
M
R, E,
P, M
R, E,
P, M
R, E,
P, M
R, E,
P, M
R, P,
M
143921A
Cressbrook
Creek at
Rosentretors
Crossing
R, E,
M
R, E,
P, M
R, E,
P
R, E,
P, M
R, E,
M
R, E
143001C
Brisbane River
at Savages
Crossing
R, E,
M
R, E
143009A
Brisbane River
at Gregors
Creek
1430060
Brisbane River
at Atkinsons
Crossing
R, E,
M
1430061
Brisbane River
at Burtons
Bridge
R, E,
M
1430062
Brisbane River
at North Kholo
R, P,
M
1430063
Pryde Creek
ds Splityard
Ck Dam
R, P,
M
R, E,
P, M
Data analysis: comparison of regulated and unregulated sites within the Brisbane
River catchment
Univariate statistics (t-tests) were used to test for differences in mean values between
regulated and unregulated sites within the Brisbane River for a number of community
measures. These included: total taxon richness, mean taxon richness, total abundance,
Shannon-Weiner diversity, abundance of uncommon taxa (calculated for each sample as the
sum of the abundance categories of uncommon taxa in the sample, where uncommon taxa
were defined as those which contributed less than 0.5 per cent of the total sum of abundance
categories for the entire database), number of unique taxa (unique to an individual sample),
and abundance of pest taxa (in this case larvae of biting Diptera taxa, including Culicidae,
Simuliidae, Ceratopogonidae, and Tabanidae). These measures were used to test for overall
differences between sites with values grouped for all habitats, as well as for differences within
individual habitats. The significance value of all statistical tests was set at 5 per cent (α =
0.05).
Multivariate statistics based on Bray-Curtis similarity measures were also used to test for
differences in macroinvertebrate community composition between sites. Community
composition was recorded for each sample whereby each taxon was allocated to an
abundance category of 0–4, defined as follows: 0 = absent (no individuals); 1 = rare (1–2
individuals); 2 = scarce (3–4 individuals); 3 = common (5–8 individuals); and 4 = abundant (9
NATIONAL WATER COMMISSION — Low flows report series 33
individuals or more). Three datasets were constructed using this data. The first dataset
consisted of abundance categories with uncommon taxa removed (those contributing less
than 0.01% of the overall sum of abundance categories and less than 5 per cent of the sum of
abundance categories for any one sample). The second dataset represented relative
abundance, and the third dataset represented presence/absence data of each taxon within
each sample. During analysis of Bray-Curtis similarities all three datasets gave similar results,
and therefore only the first dataset (which maintained the most information) was used for all
subsequent tests. Abundances of taxa within each sample were therefore used to represent
macroinvertebrate community composition. Physical and habitat variables were range
standardised (x-min/range) and matrices of Euclidean distance calculated between samples.
These difference matrices and the Bray-Curtis matrix generated from the faunal abundance
category data with uncommon taxa removed were then used to conduct subsequent
multivariate analyses.
Multivariate statistical analyses based on the macroinvertebrate abundance dataset were
used to identify sources of variation. This included tests for differences in macroinvertebrate
community composition between habitat types, and between sampling times, and between
regulated and unregulated sites. Comparing regulated with unregulated sites in this way in
addition to univariate tests was beneficial because multivariate analyses had the ability to test
for differences between regulated and unregulated sites while still considering the impacts of
environmental factors, and thus maintaining additional sources of variation which univariate
statistics would be unable to account for. The physical and water quality variables recorded
for each site were therefore included within these analyses.
Differences between habitats and sampling periods were tested by two-way analyses of
similarity (ANOSIMs) with habitat crossed with sampling period. These tested the hypotheses
that there were differences between habitats allowing for differences between sampling
periods and that there were differences between sampling periods allowing for differences
between habitats.
Based on the results of the multivariate analyses, which indicated that habitat type was
significantly affecting macroinvertebrate community composition, subsequent analyses had to
also consider habitat type as a factor. Subsequently, the analyses of difference between
regulated and unregulated sites were conducted using two-way ANOSIMs which allowed for
differences between habitat types, as well as separate one-way ANOSIMs for individual
habitat types. The analyses within individual habitat types were, however, dependent on there
being enough samples of each habitat type from both regulated and unregulated sites. Only
riffle/run, edge, and macrophyte habitats had enough samples to allow this, as pool habitats
did not. Edge samples were, however, indicative of pool habitat as they were only sampled
from slow-flowing areas and therefore all major habitat types were represented in analyses.
Macroinvertebrate community composition also appeared to vary according to sampling time.
Due to the irregular sampling of sites across sampling periods, some sites were sampled later
than others. By chance there were more sites that were only sampled in the last two sampling
periods situated at regulated sites downstream of the Wivenhoe Dam, than there were
situated at unregulated sites upstream of the Wivenhoe Dam (see Table 6). Therefore the
differences in macroinvertebrate composition between sampling periods could have been due
to river position or water regulation; as opposed to temporal influence per se. To separate
these impacts, separate one-way ANOSIMs were conducted to test for differences between
sampling periods for regulated (downstream) sites and unregulated (upstream) sites.
Pairwise comparisons were made following globally significant ANOSIMs. If these differences
were significant, similarity percentages (SIMPER) were calculated from the faunal data, to
NATIONAL WATER COMMISSION — Low flows report series 34
indicate the taxa contributing the most to the differences. The matrices were ordinated in
three dimensions using semi-strong hybrid multidimensional scaling (SSHMDS), rotated to
maximise the variation in the plane of two dimensions and plotted in these two planes.
Correlation vectors were calculated from the faunal and corresponding environmental
datasets. The significance of correlations was calculated from 100 Monte Carlo
randomisations. The correlation vectors of selected significant variables (p < 0.05) were
plotted on the ordinations.
Data analysis: comparison of patterns of difference between the Brisbane River and
the Logan River catchments
A macroinvertebrate abundance dataset was compiled for samples collected within the Logan
River, calculated in the same manner as the first dataset used for the Brisbane River. Once
again, physical and habitat variables were range standardised (x-min/range) and matrices of
Euclidean distance calculated between samples. These difference matrices and the BrayCurtis matrix generated from the faunal abundance category data with uncommon taxa
removed were then used to conduct subsequent multivariate analyses. Multivariate analyses
were then used to test for differences in macroinvertebrate community composition between
upstream and downstream sites within the Logan River. For comparison with results from
similar tests conducted on samples from the Brisbane River, one-way ANOSIMs were also
conducted for individual habitat types within the Logan River. Only edge and riffle/run habitat
types had enough samples from both upstream and downstream sites within the Logan River
for analysis. Patterns of difference between upstream and downstream edge and riffle/run
habitats were compared between the two rivers. Any differences between the two rivers in
terms of macroinvertebrate composition and environmental (physical and water quality)
variables were then tested for significance, and correlation vectors were plotted using the
same methodology as was conducted for the Brisbane River.
Differences in the composition of macroinvertebrate communities were also analysed
according to functional composition. Taxa were assigned to trophic categories (functional
feeding groups) by reference to literature (Cummins 1973; Merritt & Cummins 1978;
Chessman 1986; Hawking & Smith 1997) (see Table 8 for allocation of functional feeding
groups to relevant taxa). Five functional feeding groups were used: collectors, filterers,
grazers, shredders and predators. They were defined as follows:

collectors feed on small particulate matter which they gather from the substrate or other
surfaces

filterers filter suspended particulate matter from the water

grazers scrape periphyton (attached algae, bacteria, fungi etc.) off the substrate

shredders chew large particles of plant matter such as leaves and twigs

predators feed on other animals.
The compositions of functional feeding groups within combined samples were compared
between upstream and downstream sites in both rivers. The composition of functional feeding
groups is expected to naturally fluctuate between seasons and along the river continuum in
response to changing longitudinal environmental conditions (Marshall et al. 2001; Vannote et
al. 1980). Comparisons were made between the two rivers, looking for differences in the
patterns of change in functional feeding groups between upstream and downstream sites, and
between wet and dry seasons.
NATIONAL WATER COMMISSION — Low flows report series 35
If the magnitude and direction of correlation vectors between upstream and downstream sites
was the same in both rivers, this would indicate that the differences were natural. If the two
rivers responded differently, however, the difference in the Brisbane River may be attributable
to the effects of water resource development.
Changes to hydrological conditions
To determine the nature of the stressor imposed by the Wivenhoe Dam, an analysis of the
change in flow regime downstream before and after dam construction was conducted, with
particular reference to low flows and variability. To separate the changes resulting from flow
management from those resulting from natural variation in flow over the study period,
analyses were also performed for the corresponding years at a reference gauge site on the
Logan River with no dam present.
Time series of flow (ML day-1) from gauging stations 143001C Brisbane River at Savages
Crossing and 145014A Logan River at Yarrahappini (Figure 10, Table 5) were used for the
analysis. Construction of the Wivenhoe Dam began in 1977 and was completed in mid-1985,
so data collected during this time was removed and the time series from both sites split into
pre- and post-dam periods. Part years at the start and end of datasets were removed so that
the analysis was conducted on full years only and data gaps were in-filled using simple linear
interpolation. The result was 19 years of pre- and 25 years of post-impact data for the
Brisbane River (1959–77, 1986–2010) and eight years of equivalent pre- and 25 years of
post-data for the Logan (1970–77, 1986–2010).
Hydrologic metrics were calculated, compared for before- and after-impact differences, and
tested for significance using the scorecard function of Indicators of Hydrologic Alteration (IHA)
Software V7.1 (Richter et al. 1996). Median measures were used, rather than means,
because the gauge data were non-normally distributed. Flow duration curves were produced
using the Time Series Analysis module of the River Analysis Package (RAP) V 3.0.4 (Marsh
et al. 2003).
4.2.3. Results
Changes in macroinvertebrate community composition
Differences between regulated and unregulated sites within the Brisbane River
catchment
When univariate measures were tested, there were no significant differences in total taxon
richness between regulated and unregulated sites (p<0.05) for either grouped or individual
habitats. There were, however, differences in the mean taxon richness of samples between
regulated and unregulated sites. Samples from unregulated sites had a higher overall mean
number of taxa when habitat types were grouped (37.1 v. 31.8 taxa), as well as a higher
mean number of taxa within macrophyte habitats (24.8 v. 18.8 taxa). The number of unique
taxa was also higher within macrophyte samples from unregulated sites compared with
macrophyte samples from regulated sites (2.2 v. 5.5 unique taxa). There was a similar trend
in grouped habitat samples and all other habitat samples except pools. The number of unique
taxa in pool samples was, however, assessed from only one sample each. Therefore, higher
sample numbers may have indicated more significant results. There were no other significant
differences in any of the other univariate community measures tested.
NATIONAL WATER COMMISSION — Low flows report series 36
When multivariate datasets were analysed, as previously mentioned, all three datasets gave
similar test results. Therefore the first dataset, comprising taxa abundance (minus rare taxa),
was used for all subsequent tests, and only results from this dataset are presented.
There were significant differences in macroinvertebrate composition between habitat types,
allowing for differences between sampling periods (Table 7). Macroinvertebrate communities
within riffle/runs were significantly different from all other habitats except pools, possibly due
to the low number of pool samples. Analyses of taxa contributing to the differences between
habitats consistently identified Simuliidae and Hydropsychidae as contributing to the
differences between riffle/runs and other habitats. Both taxa are associated with high-flow
velocities according to the flow classification of Marshall and Marshall (in prep.) (Table 8).
Similarly, ANOSIMs within habitats identified riffle/runs as associated with high abundances of
Hydropsychidae, and low abundances of Copepoda and Coenagrionidae, the latter of which
are taxa associated with low flow velocities (Table 8).
Table 7: Results of pairwise comparisons between habitat types following two-way crossed
ANOSIM (habitat x run, Global R = 0.191, p = 0.0043) using abundance category data. This
tests the null hypothesis that there is no difference between habitats allowing for differences
between runs. Shaded cells indicate pairs of habitats that were found to be significantly
different (p < 0.05)
Rocky pool
Macrophytes
Riffle
Sandy pool
Edge
R = 0.057
p = 0.308
R = -0.138
p = 0.932
R = 0.470
p < 0.0001
R = 0.116
p = 0.229
Rocky pool
-
R = -0.075
p = 0.608
R = 0.666
p = 0.005
R = 0.155
p = 0.281
Macrophytes
-
-
R = 0.194
p = 0.02
R = 0.228
p = 0.154
Riffle
-
-
-
R = 0.289
p = 0.118
There were also significant differences between sampling periods, allowing for differences
between habitat types. Sampling period two was significantly different from sampling period
four (R = 0.328, p = 0.029), and sampling period six was significantly different from sampling
periods two and five (R = 0.432, p = 0.007 and R = 0.421, p = 0.006). This may be due to
temporal effects, or river position as previously discussed. Further tests were conducted to
determine the likely cause. When regulated and unregulated sites were tested separately,
there was not a significant difference between sampling periods allowing for habitat
differences within regulated sites. In contrast, there was a significant difference between
sampling periods allowing for habitat differences within unregulated sites. Based on
macroinvertebrate abundances within unregulated sites, sampling periods two, four and five
grouped together in ordination space, while sampling periods one and three separated from
the group (Figure 11). Analyses of associated environmental variables were conducted to
determine what the likely underlying cause was for this separation. A high correlation between
sampling period and flow velocity was found. When ordination was based on maximum flow
velocity, sampling periods grouped in the same manner, with sampling periods one and three
separate from the group (Figure 11). This indicates that the variation in faunal composition
between sampling periods may be attributable to temporal differences in flow velocity. The
lack of temporal differences in macroinvertebrate composition within regulated sites indicates
NATIONAL WATER COMMISSION — Low flows report series 37
that much of the biologically significant variation in flow velocity may have been lost within
regulated sites.
Results of two-way ANOSIMs indicated significant differences in community composition
between regulated and unregulated sites, allowing for differences between habitat types.
Correlation vectors indicated which taxa and environmental variables were associated with
the differences between regulated and unregulated sites. With respect to unregulated sites,
regulated sites tended to have higher abundances of Hydropsychidae and Simuliidae (highflow-preference taxa), and lower abundances of Atyidae, Planorbidae and Copepoda (lowflow-preference taxa). Regulated sites tended to have higher values for flow velocity and
width, and lower values for conductivity, alkalinity and the proportion of the substrate
composed of boulders. Regulated sites also tended to have more macrophyte cover, less
detritus cover, and different adjacent land use compared with unregulated sites.
a)
b)
Figure 11: Ordinations showing the relationships between sampling periods (runs) within
riffle/run habitat, based on a) macroinvertebrate abundance data allowing for differences
between habitat types (stress=0.0000), and b) mean maximum flow velocity (stress = 0.0002)
(adapted from Choy et al. 2000).
Results of one-way ANOSIMs indicated there were also significant differences in community
composition between regulated and unregulated sites within all three of the habitat types
tested (riffle/run, edge and macrophyte). SIMPER identified the taxa contributing the most to
the differences between regulated and unregulated samples within each habitat type. Within
riffle/runs, regulated sites had higher abundances of Helicopsychidae, Pyralidae, Elmidae,
Simuliidae and Hydropsychidae; and lower abundances of copepods. Regulated edges were
characterised by higher abundances of Sphaeromatidae, Hydrobiidae and Hydropsychidae;
and lower abundances of Acarina and Leptophlebiidae. Regulated macrophytes were
characterised by higher abundances of Sphaeromatidae and Hydropsychidae; and lower
abundances of Coenagrionidae, Leptophlebiidae, copepods and Atyidae. This conforms with
the pattern of there being more high-flow-preference taxa in regulated sites compared with
unregulated sites (see Table 8). The only exceptions are Hydrobiidae and Leptophlebiidae
which are classified as having no flow preference, and Sphaeromatidae for which flow
preference is not defined according to Choy et al. (2000) and Marshall and Marshall (in prep.).
Patterns of difference within the Brisbane River catchment compared with the Logan
River catchment
One-way analyses of similarity indicated there were significant differences in
macroinvertebrate community composition between upstream and downstream sites within
the Logan River for both of the habitat types tested (riffle/run and edge). This was also the
NATIONAL WATER COMMISSION — Low flows report series 38
case in the Brisbane River, as outlined above. This indicates that some differentiation in
macroinvertebrate community composition between upstream and downstream sites is
natural. Some of the difference between regulated and unregulated sites within the Brisbane
River may therefore also be natural. Ordination plots were used to analyse the magnitude and
direction of the differences between upstream and downstream sites in both rivers to
determine how much of the differences observed in the Brisbane River were in accordance
with the differences in the Logan River.
Results indicated that differences in faunal composition between regulated and unregulated
sites within riffle/run habitats in the Brisbane River were in accordance with the Logan River,
but this was not the case within edge habitats. The magnitude and direction of difference in
riffle/run macroinvertebrate composition between upstream and downstream sites in the
Logan River was similar to the difference between regulated and unregulated sites within the
Brisbane River, with a difference in the angle of orientation between the two rivers of 10°
(refer to the similar size and orientation of vectors LR v. BR in Figure 12). In contrast, the
magnitude of difference in edge macroinvertebrate composition between developed and
undeveloped sites within the Brisbane River was much larger than the difference in the Logan
River (refer to the size of LE v. BE vectors in Figure 12). Similarly, the orientation of the
differences in edge macroinvertebrate composition between regulated and unregulated sites
within the Brisbane River was visibly different in ordination space to that of the Logan River,
with a difference in the angle of orientation between the two rivers of 45° (Figure 12). Also
note the similar direction of the Brisbane River edge vector to both of the riffle/run vectors in
Figure 12.
Correlation vectors of corresponding macroinvertebrate abundance and habitat variable data
gave some indication of the factors contributing to the differences between upstream and
downstream sites in both rivers. The average differences between upstream and downstream
riffle/run samples in both rivers and between upstream and downstream (regulated v.
unregulated) edge samples in the Brisbane River were associated with upstream samples
having higher average abundances of Hydropsychidae; lower average abundances of
Acarina, Veliidae, Corixidae, Atyidae, Copepoda and Cladocera; and higher average values
for flow velocity (Figure 12). The average differences between upstream and downstream
edge samples in the Logan River were not associated with the abundances of any particular
taxa but were characterised by samples from downstream sites having higher average values
for total nitrogen and water turbidity, and lower average mean sediment sizes (Figure 12).
These results indicate that edge habitats in the Brisbane River do not have the same
longitudinal variation in environmental variables that are seen in the Logan River. The
longitudinal differences in environmental variables and corresponding differences in faunal
composition within the Brisbane River therefore differ from the reference condition – which
indicates that they are not natural. This provides evidence that at least some of the
differences between upstream and downstream (regulated v. unregulated) sites within the
Brisbane River may therefore be attributable to hydrological disturbance, in this case the
impacts of the Wivenhoe Dam. This depends on the assumption that the Logan River is in a
comparatively natural state and that no other factors such as land use are responsible for the
differences between the two rivers.
NATIONAL WATER COMMISSION — Low flows report series 39
LRr
HPSY
BRr
LRu
MeanSed
logVEMA
BRu
BEr
LEu
a)
BEu
LEr
UACA
logTOTN
logTURB
COEN
VELI
ATYI UCLA
b) CORI
COPE HAEN
c)
Figure 12: a) Ordination plot of average Faunal Abundance Category data with arrows
showing direction and magnitude of differences between regulated (downstream) and
unregulated (upstream) samples from riffle/run and edge habitats in the Brisbane and Logan
rivers (stress = 0.1790). B – Brisbane River, L – Logan River, R – riffle/run, E – edge, r –
regulated (downstream), u – unregulated (upstream). b) Selected significant faunal correlation
vectors contributing to variation in macroinvertebrate community samples within the Brisbane
River (HPSY = Hydropsychidae, UACA = Acarina, COEN = Coenagrionidae, VELI = Veliidae,
ATYI = Atyidae, UCLA = Cladocera, CORI = Corixidae, COPE = Copepoda). c) Selected
significant environmental correlation vectors explaining variance in macroinvertebrate
communities within the Brisbane River (MeanSed = mean sediment sizes, logVEMA = log
maximum velocity, logTOTN = log total nitrogen, logTURB = log water turbidity) (from Choy et
al. 2000).
Differences in longitudinal variation between the Brisbane and the Logan rivers were also
demonstrated by results from the comparison of functional feeding compositions (Figure 13).
When comparing upstream sites, the Brisbane and Logan rivers had similar patterns of
change in functional feeding composition between wet and dry seasons. In both rivers, there
was a decrease in the proportion of filter feeders in the dry season compared with the wet
season, and an increase in predators. But when comparing wet and dry seasons at
downstream sites, the two rivers showed differing patterns of change. Functional feeding
composition at downstream sites within the Logan River changed considerably from wet to
dry seasons, with a decrease in the proportion of filterers and increases in collectors, grazers
and predators. In contrast, there was little difference in functional feeding composition
between wet and dry seasons at downstream sites within the Brisbane River. Similarly, when
comparing functional feeding composition between upstream and downstream sites, there
was a noticeable difference in the patterns of change between the two rivers. Within the
Logan River, there were considerable differences between upstream and downstream sites in
both seasons. There was also a noticeable difference in functional feeding composition
between upstream and downstream sites within the Brisbane River during the dry season but
little difference during the wet season. Within the Brisbane River, the functional feeding
composition of downstream sites in both seasons was similar to the composition of upstream
sites during the wet season. These results indicate that downstream sites within the Brisbane
River have been altered compared with the reference state represented by the Logan River.
There has been a change in the functional feeding composition of downstream sites with
macroinvertebrate communities consistently resembling those of upstream sites during the
wet season. There has also been a loss of variability in functional feeding composition
between seasons, and across the longitudinal gradient.
NATIONAL WATER COMMISSION — Low flows report series 40
wet season
Logan River
Brisbane River
Headwaters
Headwaters
dry season
wet season
dry season
Collectors
Filterers
Dam
Grazers
Shredders
Predators
wet season
dry season
wet season
dry season
Figure 13: Schematic representations of the Logan and Brisbane rivers showing the
proportions of macroinvertebrate functional feeding groups in samples from the wet season
and dry season at upstream and downstream sites in each river (from Choy et al. 2000).
NATIONAL WATER COMMISSION — Low flows report series 41
Table 8: Flow preferences and functional feeding group membership of significant
macroinvertebrate taxa identified within results, according to the flow classification of Marshall
and Marshall (in prep.). Flow preferences correlate with the flow requirements, larvae and
adult habitat of taxa as outlined in Choy et al. (2000).
Taxon
Flow-preference group
Functional feeding group
Elmidae
High flow
Collector
Helicopsychidae
High flow
Grazer
Hydropsychidae
High flow
Filterer
Planorbidae
High flow
Grazer
Pyralidae
High flow
Grazer
Simuliidae
High flow
Filterer
Atyidae
Low/no flow
Shredder
Cladocera
Low/no flow
Collector
Coenagrionidae
Low/no flow
Predator
Copepoda
Low/no flow
Collector
Corixidae
Low/no flow
Collector
Veliidae
Low/no flow
Predator
Acarina
No preference
Predator
Hydrobiidae
No preference
Grazer
Leptophlebiidae
No preference
Collector/grazer/shredder
Sphaeromatidae
Not determined
Shredder
Changes to hydrological conditions
At both the test and reference gauge sites, monthly flows were more seasonally variable in
the pre-1977 period, compared with post-1985 (Figure 14). In the earlier period, high flows
occurred in the first few months of the year, followed by a trough during winter and spring
before a slight rise at the end of the year. In the later period, flows were more stable
throughout the year at both locations. The dates of the minimum and maximum flow days
have not changed significantly (Table 9) so seasonality has not been altered; flow has simply
become less variable. This is likely to be largely due to climatic variation: the early period
experienced regular summer flooding, particularly in the early 1970s, and the later period was
dominated by drought for much of the 1990s and 2000s.
However, while flows became less seasonally variable at both locations, there appeared to be
differences in flow level: the Logan River experienced low-flow conditions post-1985, while
higher flow volumes were maintained in the Brisbane River. Median monthly flows have
decreased by more than 50 per cent in all months in the Logan River, while in the Brisbane
River there has been a decrease in flows in summer and autumn and an increase in late
winter and spring (Figure 15). Similar patterns are present in metrics measuring the median
minimum and maximum flow magnitude. In the Logan River, there has been a decrease in
both low-flow and high-flow magnitudes in the post-1985 data. In the Brisbane River, low-flow
magnitudes have increased since construction of the dam, while high-flow magnitudes have
decreased (Figure 16). This suggests that release strategies for the dam have led to an
‘averaging out’ of flow conditions at medium levels.
NATIONAL WATER COMMISSION — Low flows report series 42
Figure 14: Comparison of median monthly flow volumes (ML day-1) in the Brisbane and Logan
Rivers in the periods before construction of the Wivenhoe Dam on the Brisbane River (pre1977) and after construction (post-1985).
In terms of low flows, construction of the Wivenhoe Dam has decreased the occurrence of
natural low-flow events. Low-flow discharge (the 90th percentile of the flow duration curve)
increased at the Brisbane River gauge from 289 ML day-1 before dam construction, to 340 ML
day-1 after construction. For the same period in the Logan River, the metric decreased from
47 day-1 to 9 day-1.
Table 9: The median Julian date of the one-day annual minimum flow for the Brisbane and
Logan rivers for the periods before (pre-1997) and after (post-1985) construction of the
Wivenhoe Dam.
Brisbane River
Logan River
Pre-1977
Post-1985
Pre-1977
Post-1985
Julian date minimum
293
323
342
332
Julian date maximum
45
48
31
39
NATIONAL WATER COMMISSION — Low flows report series 43
Figure 15: Change in median monthly flow volumes (ML day-1) for the Brisbane and Logan
rivers in the period following construction of the Wivenhoe Dam (post-1985) compared with
the preceding period (pre-1977). * indicates a significant difference (P=0.05).
Figure 16: Change in median low- and high-flow volume metrics for the Brisbane and Logan
rivers in the period following construction of the Wivenhoe Dam (post-1985) compared with
the preceding period (pre-1977). * indicates a significant difference (P=0.05).
NATIONAL WATER COMMISSION — Low flows report series 44
4.2.4. Discussion
Differences in macroinvertebrate communities both within the Brisbane River catchment and
between it and the Logan River reference catchment indicated that in-stream biota responded
to altered flow conditions. The flow characteristics to which macroinvertebrate communities
responded were flow velocity and variability, and frequency and severity of low-flow events.
Sources of variation between macroinvertebrate communities within the Brisbane River
catchment were related to habitat type, river position and sampling time. Flow conditions
appeared to be an important factor contributing to the differences between habitat types,
upstream and downstream sites, and sampling times. Macroinvertebrate communities within
the Brisbane River catchment are clearly responding to flow conditions, with shifts in
macroinvertebrate communities observed in correspondence with shifts in flow velocity.
Differences between macroinvertebrate communities within the Brisbane River catchment
compared with the reference condition in the Logan River catchment also indicated that fauna
were responding to changes in flow conditions. The changes in flow conditions at downstream
sites caused by the Wivenhoe Dam were an ‘averaging out’ of flows with stable moderate flow
levels replacing annual fluctuations between high and low flows. The macroinvertebrate
response to these changes was a shift in macroinvertebrate community composition, without
a loss in diversity, overall abundance or uncommon taxa. There were two main shifts in
macroinvertebrate community composition associated with flow alteration caused by the dam:
1) A shift in macroinvertebrate community composition towards a higher abundance of highflow-preference taxa and a lower abundance of low-flow-preference taxa within pool
habitats (represented by pool-edge habitats). This suggests one of the impacts of the
Wivenhoe Dam has been a decrease in the frequency of low-flow velocities through pools
and edges. Taxa with a strong preference for low flows have been reduced in abundance,
and some particularly sensitive taxa may have been lost. Other studies have similarly
reported that increased frequencies of high flows downstream of impoundments have
adversely affected velocity sensitive taxa (e.g. Richter et al. 1997).
2) A shift in the trophic structure of macroinvertebrate communities, as indicated by functional
feeding group composition. Seasonal and longitudinal patterns of change in trophic
structure were altered. Stanford and Ward (1984) reported that sites downstream of dams
had functional compositions which resembled headwater sites rather than sites with similar
positions in unregulated rivers. This pattern was evident in the Brisbane River. Similarly,
Choy and Marshall (1999a) found that water development within the Burnett catchment,
Queensland, altered the physical and biological system from a temporally dynamic river
with seasonal patterns to a system with a generally fixed pattern.
The loss of seasonal variability in macroinvertebrate functional composition below the
Wivenhoe Dam is likely to have adverse consequences for the long-term maintenance of
invertebrate communities in the Brisbane River. Downstream assemblages may have lost
some of their ability to adjust to flow fluctuations and other environmental changes. As a
result of the more stable flow conditions, taxa downstream of the dam may be more sensitive
to environmental stresses than those at unregulated sites. Communities below the dam may
therefore be less able to adjust to any variations in flow and other environmental variables
should they occur in future. This effect is likely to become more pronounced the longer the
populations of macroinvertebrates are exposed to flows with reduced variability (Choy et al.
2000). Furthermore, loss of the natural hydro-ecological continuum, as seen by altered
longitudinal patterns of community composition, is likely to disrupt ecosystem function and the
movement of energy in a large section of the Brisbane River (Vannote et al. 1980).
NATIONAL WATER COMMISSION — Low flows report series 45
The lack of change in riffle macroinvertebrate communities within the Brisbane River
catchment compared with the reference condition is in contrast to the expectation that riffle
macroinvertebrates would be the most highly impacted habitat type as a result of being
drowned out by increased flows. It could be that riffle communities are more resilient to
increases in flow velocity than pool communities (Reice 1991), or it could be that the sampling
design was not able to capture changes in riffle communities (Marshall et al. 2001). The
sampling design was intended to detect changes within individual habitats, but not to detect
changes in the distribution of habitats. Changes in the flow regime may alter the distribution of
riffles rather than the conditions within the remaining riffles. Impacts on riffle communities may
therefore exist without detection (Marshall et al. 2001). Any riffles that were drowned out by
water release from the Wivenhoe Dam would not have been sampled as riffles. Pools, in
contrast, would have remained within the same habitat category for sampling despite potential
deepening and increased flow velocity within individual pools. The lack of response from riffle
macroinvertebrate communities therefore does not mean that riffle-dependent taxa were not
affected by the Wivenhoe Dam. Instead, the sampling techniques used within the study may
not have been adequate for identifying impacts on riffle taxa. To address this problem,
Marshall et al. (2001) suggested that during sampling sites be surveyed for the distribution of
biologically relevant habitat types and this information mapped. Changes in habitat availability
could then be compared between sampling times.
The response of macroinvertebrate communities to changes in low-flow conditions in the
Brisbane River catchment indicates that macroinvertebrates can be used as indicators of
alteration to the low-flow regime in this region. However, detected changes may or may not
have broader importance in terms of the ecological values of the region (see Section 5.2). To
measure macroinvertebrate community responses to low-flow conditions, a referential
approach using indices with conceptual links to low-flow hydrology, such as functional feeding
group composition and the proportion of flow-preference taxa, is most appropriate. This is
supported by a number of other studies by DERM (Marshall et al. 2001; Choy & Marshall
2000; Marshall et al. 2000; Choy & Marshall 1999a; Marshall & Marshall in prep.) in which
ecological-flow responses were demonstrated by changes in aquatic macroinvertebrate
functional feeding composition and flow-preference-taxa composition or abundance. Marshall
et al. (2000) used % high-flow-preference taxa, % low-flow-preference taxa, and % filterers as
indices to differentiate flow-related impacts on macroinvertebrate community composition
from other impacts. Marshall and Marshall (in prep.) provide a comprehensive list of the flow
preferences of common Queensland taxa. In addition to changes in functional feeding group
composition and the proportion of flow-preference taxa, taxa particularly sensitive to flow
conditions can be identified as indicators of low-flow response. Choy and Marshall (1999b)
stated that Hemiptera and Coleoptera are generally sensitive to increased flow conditions, as
well as pool-edge taxa in general, as they tend to be less tolerant of high flows than pool-bed
taxa. This study indicates that within south-east Queensland, the high-flow-preference taxa
Elmidae, Hydropsychidae, and Simuliidae seem to be particularly sensitive to flow changes;
as do the low-flow-preference taxa Acarina, Atyidae, Copepoda and Coenagrionidae.
The effects of the Wivenhoe Dam on macroinvertebrate communities indicate the types of
changes that occur when the low-flow regime is altered. Reduced occurrence of low flows,
probably through their effect on flow velocity and depth, resulted in a loss of low-flowpreference taxa and a shift toward a higher proportion of high-flow-preference taxa. Altered
trophic structure and loss of seasonality and natural hydro-ecology patterns of continuum
were also observed within the Brisbane River catchment in response to a reduced occurrence
of low-flow conditions. These changes are suggestive of wider changes in ecosystem function
caused by reduced low flows.
NATIONAL WATER COMMISSION — Low flows report series 46
5. Current monitoring relevant to low
flows: issues and recommendations
DERM has several freshwater ecology monitoring and research programs designed to
support policy making and water management (Table 10). These programs deal with a range
of management issues and questions, and thus use various methods and operate in different
locations.
At present DERM does not have a monitoring program explicitly for low-flow ecology. Lowflow issues may be detected and dealt with through existing monitoring and research where
they are locally relevant to anthropogenic pressures and ecological values (Table 10). There
are, however, a number of difficulties associated with assessing the effect of low flows on
ecology based on existing hydrological and ecological data. These difficulties, and the ways in
which some of them may be overcome in the context of Queensland water management, are
outlined in sections 5.1 to 5.2 below.
5.1. Hydrology


Availability of gauge data.
–
Gauges have generally been placed to collect data relating to water resource
management not ecology, therefore the location of gauges may not match biological
sampling sites or the location of impacts.
–
Sparse distribution of gauges in some catchments, particularly in western
Queensland, and in low-order streams in most catchments.
–
Improvements could be made by including ecological considerations into future
gauge-station network planning.
–
‘New enabling technologies’ such as networks of dataloggers (e.g. depth, water
quality) linked to central recording stations could be deployed in areas where gaugestation distribution is limited, to collect time series data for ecological assessments.
The technical limitations of gauging, estimating and/or modelling low flows make it difficult
to accurately measure low-flow conditions and detect the critical change from low flow to
no flow. When flow stops, waterbodies providing habitat to aquatic biota often remain,
which over time may contract into a series of separate waterholes before drying out.
Some waterbodies are more persistent than others, and each phase of drying has
implications for in-stream ecology. Gauging relies on rating curves that relate depth to
discharge for a defined cross-section and this information is seldom available once flow
stops. Currently depth and temperature loggers are placed in some waterholes to assess
no-flow conditions.
–
Gauging stations, mode of gauging (monitoring) and modelling of low-flow hydrology
may be issues covered in the hydrology component of this Low Flow Ecological
Response and Recovery Project.

There is a lack of data relating to the incidence of some low-flow pressures. For example,
localised water extraction, especially during dry times, is a major threat to aquatic
communities in some areas (e.g. small streams and waterholes), but information on
where and when extraction occurs and the volume of water removed is generally
unavailable.

Understanding groundwater contribution to stream baseflow and the supplementation of
refugial waterholes is essential to fully appreciate the low-flow environment.
NATIONAL WATER COMMISSION — Low flows report series 47
–
Hydrological modelling must incorporate groundwater contributions in a spatially
explicit way so that ecological responses can be modelled accordingly. This needs to
include the effects of groundwater extraction on surface water flow and aquatic
habitat persistence.
5.2. Ecology


Low-flow and no-flow conditions are natural occurrences in most Queensland streams
and the ecology is largely evolved to cope with these events. Changes to the ecology
exist where human pressures on water resources change the natural wetting and drying
of streams, including the reduction of the occurrence of low and no flows through
increases in flow (e.g. constant releases from dams, interbasin transfers).
–
It should be identified that low flows in many systems are not a stressor, but a natural
and necessary part of the flow regime. The alteration of low-flow regimes by
additional wetting is as equally detrimental as drying events in wetter regions. It is
important not to give the impression that dry is always bad and wet is always good in
terms of aquatic ecology.
–
Extremes and thresholds – while the ecology of Queensland streams is typically
adapted to survive periods with low or no flows, increases – particularly in the
duration of these spells, but also in their frequency and changes to their timing – may
none-the-less represent catastrophic events leading to population failure. Our
understanding of thresholds beyond which catastrophic changes occur is
rudimentary, and such instances are intrinsically difficult to predict. Modelling
provides a means to simulate these events, but needs to be based on realistic
approximations of system parameters.
Availability of ecological data.
–
Despite having a large dataset of existing AusRivAS-type macroinvertebrate
monitoring data, when we came to assess the ecological response to the pressures
identified in the low flows PSR framework, little data was available in suitable
locations and time periods. This is even truer of other biotic indicators (e.g. fish,
riparian vegetation) where samples often cover small areas and short time periods
and sampling methods differ between projects.
–
Existing data was collected for a specific purpose so it may not be appropriate for
answering questions for which it was not designed. For example, riffle habitat
macroinvertebrates should exhibit a response to changes in low-flow conditions, but
because many past monitoring projects aimed to sample a set number of riffle
habitats, rather than sample in the same place regardless of the nature of the habitat
at the time, trends may be difficult to detect (see Section 4.2.4). During a dry spell,
sampled macroinvertebrate communities in the remaining riffles are likely to be
similar, despite the fact that there may be only half the usual number of riffles in the
catchment, masking the impact. As such, in some cases, it is the spatial distribution
of habitats, not the communities within them, which indicate stresses.
–
Where low flows are identified as a potential threat to ecological communities,
specific sampling, monitoring or modelling designed for this purpose should be
carried out. This would include developing and testing a conceptual system
understanding; selecting measures of pressure, stressor and response; and
appropriate spatial and temporal sampling distribution. While this is often a high-cost
undertaking, it is the best way to ensure confidence in the assessment.
NATIONAL WATER COMMISSION — Low flows report series 48

The complex relationship between gauged flow and biological response – these often act
at different spatial and temporal scales and are mediated by other factors and impacts –
makes it difficult to draw direct conceptual links between the two (Figure 17). For
example, invertebrates living in a riffle experience the local hydraulics, substrate, water
quality and non-flow impacts (e.g. presence of invasive species, pollutants) and this is
what determines their community structure through time. These are indirectly related to
gross flow metrics but are likely to vary greatly over time and space.
–
Improved conceptual understanding of the flow requirements of vulnerable aquatic
ecosystem components and the links between anthropogenic pressures and
ecological responses is being developed within DERM through the Environmental
Flows Assessment Program (EFAP) and Steam and Estuary Assessment Program
(SEAP).
Figure 17: Ecological responses occur in reaction to the physical and chemical conditions
experienced by biota, not directly to altered flow regimes. Altered flows influence ecology by
interacting with various other properties of the setting to modify the conditions the biota
experience. To understand ecological responses to flow modification one must first
understand the relationships between physical/chemical conditions and the ecological
response, and secondly understand the relationship between flow and the provision of these
conditions. Stressors other than altered flow may also influence the response, so the
effectiveness of flow regime restoration or protection cannot be judged directly by the
occurrence or intensity of response. Rather, it must be judged by evaluating the provision of
the flow-related conditions necessary for the response to occur (Cockayne et al. 2010).

Depending on their requirements, different ecosystem components may respond
differently to the same flow conditions. Therefore, understanding/predicting changes in
macroinvertebrate communities doesn’t necessarily translate to other taxa or processes.
Flows that provide a benefit to one taxon or process may impact another, meaning
choices need to be made about which to manage for.
–
Develop a transparent process for prioritising ecological assets and selecting the
most appropriate compromise where conflicting outcomes exist. Methods are
currently being developed and implemented as part of the Queensland Water
Resource Plan (WRP) development and review process.
NATIONAL WATER COMMISSION — Low flows report series 49

There appears to be a significant reliance on structural indicators of the ecology and how
they respond to low-flow conditions. The quantification of critical ecological processes is
likely to give a more direct or holistic measure of ecosystem response, particularly instream productivity which relies on the low-flow end of the hydrograph where bed
mobilisation risk is low, and microbial biofilms dominate the benthos.
–

Low-flow-ecology research should extend to hyporheic fauna and their critical
dependence on this part of the hydrograph – an area which is currently underrepresented in the scientific literature.
–

Instead of relying on traditional biological indicators, the system conceptualisation,
along with an understanding of the nature of the threat, should be used to select the
most appropriate measures of ecosystem response to impacts; for example, genetic
measures of population viability.
Work has begun in a number of Queensland catchments to identify the distribution of
stygofauna and the effect of groundwater use. This could extend to include hyporheic
fauna in the future.
Changes in an indicator such as macroinvertebrate assemblage composition do not
necessarily constitute an unacceptable change as a result of a stressor. Acceptable
change should be implicitly linked to the ecological values society holds for the
ecosystem.
–
Several DERM assessment programs use a risk assessment approach to present the
expected impacts of environmental change on ecosystem responses (WRP
assessments, SEAP). When these responses are linked to values, the outputs
represent the risk of a loss of values. For example, in WRP ecological assessments,
ecological assets are selected as indicators of ecological values and assessments
are made of the risk posed to the values from alternative flow management options.
This approach avoids the ‘so what?’ response to measured or predicted ecological
change, as the answer is implicit in the metric.
NATIONAL WATER COMMISSION — Low flows report series 50
Table 10: Current DERM monitoring and research programs that may relate to low flows and ecology. Relevant programs are listed, as well as the location in
which they are undertaken, the ecological indicators used, the frequency of activities, and how these activities apply to our knowledge of low-flow ecology.
Program
Location
Indicators
Frequency
Application to low flows
Stream and
Estuarine
Assessment
Program
(SEAP)
Sampling by freshwater
biogeographic province
(a group of catchments
clustered based on the
similarity of fish and
macroinvertebrate
communities).
Reporting at both
province and catchment
scales.
Random-stratified site
selection plus some
fixed long-term
monitoring sites.
Vary by province – selected to
respond to the anthropogenic
threats of importance in each
region, e.g. indicators for the
Wet Tropics bioregion included
riparian vegetation, feral pig
impact, macroinvertebrates and
freshwater fish.
One province sampled
per year.
Sampling repeated in a
province approximately
every eight years.
Long return interval so not suitable to detect incremental
responses.
Uses a PSR framework to select appropriate indicators to measure
threats of importance. Doesn’t focus on low flows, though
prioritised threats may relate to low flows (e.g. dams).
Provides the context of water management, including low flows, in
relation to other threats within each biogeographic province.
e.g. Change in the occurrence and stability of low flows by water
infrastructure was identified as an important threat in the Wet
Tropics bioregion. Pressure = dams/weirs, Stressor = change in
hydrology (based on IQQM modelling), Response = proportion of
invertebrate flow-preference groups.
An outcome of SEAP in the Wet Tropics bioregion was targeted
research to investigate the effects of changes to low flows on
populations of low-flow asset taxa, e.g. Mogurnda sp. which were
identified in the EFAP process.
Ecosystem
Health
Monitoring
Program
(EHMP)
135 sites within SEQ
catchments.
All sites visited on all
sample runs.
Water quality (pH, conductivity,
temperature, DO, nutrient
levels).
Macroinvertebrates (number of
taxa, PET richness, SIGNAL
index scores).
Fish (proportion of natives,
proportion of aliens, O/E).
Sites sampled twice per
year in autumn and
spring (post-wet and
pre-wet seasons).
Long-term dataset includes drought and subsequent flooding –
could be used to identify trajectories of decline and recovery from
related impacts.
Pre- and post-wet annual sampling may distinguish effects of
regular seasonal dry spells.
Raw data likely to be more useful for low-flow assessments than
calculated indices.
Program not designed for assessing the impact of low flows, which
may affect confidence in results if used for that purpose.
Surface Water
Ambient
Network
Subset of DERM
gauging stations
throughout Queensland
Flow and water quality measures
(temperature, pH, EC, total
nitrogen, total phosphorus,
turbidity, and spot grab samples
Manual water quality
sampling generally
occurs four times a year
(more frequently in
Collects ambient water quality data at a range of flow levels. This
is currently used for condition and trend analyses, but could be
used to inform relationships between flows and water quality for
ecological assessments, particularly if used in conjunction with
NATIONAL WATER COMMISSION — Low flows report series 51
Program
Location
Indicators
Frequency
Application to low flows
(SWAN)
(approx. 400).
190 of these are
automated water quality
sampling gauges and
manual spot samples
are collected at others
for laboratory analysis.
for major cations and anions).
some locations) along
with additional sampling
to capture variability.
Continuous automated
sampling of some
parameters at some
gauges.
sediment and nutrient load (during high-flow events) sampling
carried out in SEQ and GBR catchments.
As this project collects data on the changes in water quality
parameters in disconnected waterholes over the progression of a
dry spell, this could inform assessments of low-flow impacts, but it
is difficult to link this information to gauged data because of the
issues with gauging low flows.
Environmental
Flows
Assessment
Program
(EFAP)
Statewide.
Specific data collected
in WRP areas
(catchments or groups
of catchments).
Flow-dependent ecological
assets are selected specifically
for WRP areas as indicators of
the regions’ ecological values.
Prioritisation of values is based
on their vulnerability to the type
of hydrological modification
represented by the WRP/ROP.
e.g. for the Fitzroy Basin WRP,
assets included banana prawns,
barramundi, low-flow spawning
fish, refugial waterholes and
riparian tree communities.
Risks from flow management
were calculated by comparing
the water requirements of assets
(expressed as facets of the flow
regime) with modelled water
management flow scenarios.
Sampling varies by
ecological asset.
No monitoring as such,
rather data collection
where necessary to
define/refine the asset’s
flow dependencies. This
information is then used
to develop ecological
response models which
evaluate flow
management scenarios
at five- or 10-year
intervals.
The suite of assets selected for a WRP area will have
requirements across a number of different naturally occurring flow
bands, including low flows.
Responses of assets to changes in flows is modelled rather than
measured. Because of the influence of a range of non-flow
impacts, it is impossible to attribute biotic responses directly to flow
management.
Local research is being undertaken within WRP areas to better
understand flow-ecology relationships and responses to flow
management, including low-flow conditions, e.g. effect of low flows
on Tandanus tandanus movement in the Pioneer catchment, and
the persistence and connectivity of refugial waterholes in dryland
catchments.
Gauged flow
data
Approximately 400
surface water gauges
and 150 continuous
groundwater monitoring
bores statewide.
Water height and discharge.
Continuous monitoring.
Additional groundwater
bores provide manual
spot measurements.
Length of record varies and distribution is patchy and can be
sparse in some catchments. Difficulties measuring low levels of
flow, so often not possible to distinguish between low and no flow.
Low-flow gauging (monitoring) and modelling issues will hopefully
be captured by the hydrology component of this Low-flow
Ecological Response and Recovery Project.
NATIONAL WATER COMMISSION — Low flows report series 52
6. Discussion and conclusions
Flow regimes play a key role in shaping the ecology of freshwater systems. The low-flow
band of the hydrograph is particularly important in many catchments where it acts either as
the baseline condition to which ecosystems are adapted, or a disturbance imposed naturally
or anthropogenically (Bunn & Arthington 2002; Kennard et al. 2010; Rolls et al. 2010). As
such, the nature and effect of low-flow impacts on ecology need to be assessed by taking into
consideration the natural conditions in a catchment. While the reduction of flow volume in wet
catchments through drought or extraction may impact the specialised flow-dependent biota
that reside there, it is likely to have less of an effect in dryland catchments where the biota
have developed strategies to survive such conditions. On the other hand, dryland ecosystems
may be significantly altered by actions which reduce the number or duration of low- and noflow spells, such as water releases.
Aquatic biota have developed a range of strategies to survive in the environments they
inhabit. Many have life-history stages or processes linked to flow characteristics (e.g.
spawning and recruitment in Ambassis agassizii is linked to stable low flows – see Cockayne
et al. 2010). While for other biota, flow conditions contribute to their distribution and extent
(e.g. water level, velocity and substrate determine the extent of benthic algal mats; and flood
frequency and inundation area drive riparian forest communities) (Bunn & Arthington 2002;
Kennard et al. 2010; Rolls et al. 2010). Changes in flow conditions – from natural – will
therefore affect different organisms in different ways, and to varying extents (Arthington &
Pusey 2003). The best indicators of the effects of changes to the low-flow regime will vary,
depending on the region of interest and the nature of the flow change: which will determine
how the changes affect habitat in the given setting (Rolls et al. 2010).
In Queensland, the pressures which relate to the flow regime, and particularly to low flows,
are the construction and release strategies of dams and weirs, water extraction, disposal of
excess industrial water, interbasin transfers and climate change (Table 1). Within Queensland
it is predicted that climate change will result in reduced surface runoff due to higher
temperatures and evaporation rates, and decreases in rainfall (Chiew & McMahon 2002;
CSIRO & BOM 2007; DERM 2010a). This is likely to lead to increases in the number and
duration of low- and no-flow spells. Extreme droughts and storm events are also predicted to
become more common (CSIRO & BOM 2007; Hennessy 2007; DERM 2010a).
Interbasin transfers have different effects on the flow regime in source and recipient systems,
depending on the operation of the water network. In general, recipient streams will experience
increased flows, with reduced frequency of low- or no-flow periods, and lower flow variability.
Source streams may experience reduced flow and water levels, though in many cases source
streams are already impounded, and therefore the additional impact of reduced flows on local
hydrology and biota is likely to be minimal.
Disposal of excess industrial water already occurs relatively frequently in lowland and
estuarine river reaches along the Queensland coast (e.g. water used to cool machinery or
clean mining plants and equipment) and tends to be more of an issue of water quality than
quantity. However, with the emergence of the CSG industry in dryer regions such as the
Queensland Murray-Darling and Fitzroy basins, there is the potential for water disposal to
significantly alter flow regimes. Modelling of potential disposal scenarios conducted to date
points to a reduction in the number of zero-flow days and increased duration of flow events
(McGregor et al. 2011).
NATIONAL WATER COMMISSION — Low flows report series 53
Depending on their operation, dams and weirs can have a range of effects on low flows. In
the impounded area, velocity is reduced and stream depth and width are vastly increased.
Downstream, flow can be artificially increased by water releases, changing the variability of
flow conditions and drowning out some habitats that are created by baseflow conditions; or
conversely, flow can be stopped by the barrier, increasing the occurrence of low- and no-flow
conditions.
The case studies conducted for the Wivenhoe Dam and Walla Weir shed some light on the
response of biota to the changes in flow regime imposed by these structures. After the Walla
Weir was built, there was a reduction in the proportion of high-flow-preference
macroinvertebrate taxa both within the impounded area and below the weir wall due to
reduced flow velocities. As a result, fewer source populations of riffle taxa remain in the
catchment, with increased distances between populations, which may impact their resilience
at a broader scale.
Operation of the Wivenhoe Dam has significantly altered the hydrology of downstream
reaches by decreasing flow variability and the duration and frequency of low-flow events. This
increases the velocity of flow through habitats in the river that naturally had low velocities and,
as a result, the dam affected downstream aquatic macroinvertebrate communities. The
macroinvertebrate community composition within pool habitats shifted towards a higher
proportion of high-flow-preference taxa, in addition to changes in the functional feeding
composition, with a reduction in taxa that preferred low flows. These changes give meaningful
indications of the types of biological responses that occur in response to reduced low flows,
as mediated by their effects on the hydraulic conditions in the river.
While the Walla Weir and Wivenhoe Dam case studies go some way to explaining the
response of macroinvertebrate communities to the types of modification to low flows imposed
by dams and weirs, they also highlight many of the difficulties with undertaking such
assessments. At the Walla Weir, for example, there are no open gauges downstream of the
weir, so hydrology data that may help to confirm the nature of the stressor is unavailable.
There are a number of issues, both logistical and conceptual, in attempting to determine the
effect of low-flow hydrology on ecology (see Section 6). As in the Walla Weir example,
existing data about flow, ecological indicators and stressors is often unavailable for the time
and place of relevance; or is not suitable for the analysis because of the sampling methods or
design originally employed. There are also difficulties in understanding and interpreting
ecosystem responses to changes in low flow because of issues of scale, complex
interactions, confounding impacts, and differing responses from ecosystem components to
the same conditions. These issues, along with information gaps about the specific flow
requirements of biota and processes, make selection of the most appropriate ecological
indicators for a region problematic but also of critical importance. Lastly, the hydrology and
ecology of ground and subsurface waters, which are intimately linked to low-flow conditions,
must be included in a holistic assessment but are largely unknown.
The options for managing low flows, especially naturally occurring seasonal and supraseasonal dry spells, may also be limited. In regions where water resource development is
necessary to fulfill economic and social goals, a compromise must be reached between
human and environmental values. In addition, there are often limitations on the capacity for
management of flows, either because the necessary infrastructure doesn’t exist, or because
there simply isn't enough water available during dry times to meet demands. Ideally,
management of low-flow impacts will take into account the natural hydrologic conditions, the
NATIONAL WATER COMMISSION — Low flows report series 54
ecological values (and the most appropriate indicators/indices of these), the water use
requirements, and the intervention options within a region.
We suggest that in most cases it is not appropriate to seek direct relationships between
measures of flow regime change and ecological condition. Carrying out univariate or
multivariate statistical analysis of biological and environmental data do not provide strong
correlations, because of a combination of poor conceptual framework, inappropriate data,
confounding factors and possible lag effects. This recommendation is supported by the
outcomes of two large-scale studies conducted by DERM in the Condamine-Balonne and
Fitzroy River catchments (Negus et al. 2004). In these studies the responses of multiple
ecological condition indicators to gradients of flow alteration were investigated. Both studies
demonstrated that community and process-based ecological condition indicators did not
respond to flow change in a predictable way, and thus direct approaches cannot effectively
inform flow management decisions. This result is corroborated by a global literature review
(Poff & Zimmerman 2007) which concluded that general, quantitative patterns between flow
alteration and ecological responses are not strongly evident.
The absence of predictable ecological responses to flow modification is, in part, because of
the influences of confounding stressors (e.g. land use gradients often correlate with gradients
of flow alteration), but fundamentally it is because biota do not experience or respond directly
to hydrology. Rather, flow interacts with other features to produce physical, chemical and
biological conditions which are perceivable to biota and which elicit particular ecological
responses (Figure 17). Fish, for example, do not perceive mean annual discharge, but they
do perceive stream depth, velocity and water temperature, and react to these in predictable
ways.
To inform the management of flow regimes to achieve ecological outcomes, we must focus
our attention on understanding what conditions trigger ecological responses, and how flow
interacts with other aspects of the local setting to provide these conditions. The interactions
between particular responses and the conditions that trigger them should be general and
transferable. In contrast, provision of the necessary conditions by the flow regime interacting
with other influences is a function of setting and thus is not transferable.
Because in-stream ecological responses are influenced by stressors other than flow
alteration, it is inappropriate to assume that flow management will elicit the expected
responses. It follows that it is also inappropriate to evaluate the effectiveness of flow
management by measuring the occurrence or the intensity of ecological response. Even if
flow management is ideal, the influence of other stressors may prevent a response from
occurring. For this reason we advocate evaluating the provision of the flow-related conditions
required for the response as the means by which flow management should be determined
and evaluated. Through the EFAP program, DERM applies an ecological risk assessment
framework where the risk to specific ecological values is linked to ecological responses with
critical requirements for flow-related conditions. The provision of these critical requirements is
used as the measure of management success. This approach accounts for both the potential
influence of confounding stressors, and the complex relationships between flow, setting and
ecology.
NATIONAL WATER COMMISSION — Low flows report series 55
Shortened forms
ANRA
Australian Natural Resources Atlas
AUSRIVAS
Australian River Assessment System
BOM
Bureau of Meteorology
CSG
coal seam gas
CSIRO
Commonwealth Scientific and Industrial Research Organisation
DERM
Department of Environment and Resource Management, Queensland
DPSIR
Driving forces, Pressures, States, Impacts, Responses
DSEWPaC
Department of Sustainability, Environment, Water, Population and
Communities
EHMP
Ecosystem Health Monitoring Program
EFAP
Environmental Flows Assessment Program
EHMP
Ecosystem Health Monitoring Program
FFG
Functional Feeding Group
GS
Gauging Station
GBR
Great Barrier Reef
IHA
Indicators of Hydrologic Alteration
IPCC
Intergovernmental Panel on Climate Change
IQQM
Integrated Quantity and Quality Model
ML day-1
Megalitres per day
NWC
National Water Commission
PET richness
Plecoptera (stonefly), Ephemeroptera (mayfly) and Trichoptera (caddisfly)
species richness
PSR
Pressure-Stressor-Response
QMDB
Queensland Murray-Darling Basin
RAP
River Analysis Package
ROP
Resource Operation Plan
SEAP
Steam and Estuary Assessment Program
SEQ
South East Queensland
NATIONAL WATER COMMISSION — Low flows report series 56
SIGNAL
Stream Invertebrate Grade Number – Average Level
SIMPER
Similarity Percentages
SWAN
Surface Water Ambient Network
WRP
Water Resource Plan
NATIONAL WATER COMMISSION — Low flows report series 57
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NATIONAL WATER COMMISSION — Low flows report series 65