volume 2 - Research@JCU

Transcription

VOLUME 2
A Typological Basis for the
Assessment and Management of Wetland Water Quality
in the Dry and Wet-Dry Regions of Tropical Queensland
ACTFR Report No. 04/04
Appendices
May 2004
Prepared by Barry Butler
Australian Centre for Tropical Freshwater Research
James Cook University, Qld. 4811
Phone: (07) 47814262
Fax: (07) 47815589
email: [email protected]
APPENDIX A
1
A1
CASE STUDY: WATER QUALITY VARIATIONS THROUGHOUT THE
WATERWAYS OF THE HERBERT FLOODPLAIN
A1.1
Introduction
A1.2
Methods
Sampling entailed the collection of a discrete water sample from a depth of 30cm below the water surface
and at a lateral position that approximated mid-stream. Most sites were sampled on two or three occasions,
representative of different seasonal conditions, provided that water was present during field trips. Samples
were analysed for a wide range of parameters including nutrients, suspended solids (suspended particulate
matter), chlorophyll a, major ions, organic carbon, BOD and colour (see Appendix tables for further details).
Temperature, pH, electrical conductivity and dissolved oxygen were determined in the field using Hydrolab
multi-parameter probes and, in many cases, cool dark-equilibrated samples were returned to the laboratory
for further analysis of pH and conductivity. During the early stages of the project, field measurements were
confined to spot readings taken at the times when samples were collected. However, it became evident that
metabolism in the water body was too high and variable for spot measurement of this kind to be interpreted.
Hence, monitoring protocols were altered to ensure that measurements were conducted at least twice during
each sampling event – once during the early to mid-afternoon and once during the early morning
(temperature, dissolved oxygen and pH values reach daily maxima and minima at these times of day).
Detailed descriptions of prevailing and antecedent conditions of flow, weather, shade, riparian vegetation,
instream biomass, depth, etc. were recorded whenever samples or field readings were taken. The qualitative
or semi-qualitative observations for each biophysical variable were classified into nominal or ordinal
categories (e.g. high, medium and low). These category variables were then used to examine relationships
between biophysical conditions and water quality, and to generate key data summary groupings.
A1.3
Results and Summary Plots
A substantial database of biophysical records and associated category variables has been compiled; these
have not been tabulated here but are referred to where relevant in the text. The results for each monitoring
parameter are summarised and compared with data obtained from other components of the water quality
monitoring program in the box plots shown at the end of this appendix. These figures are clustered box
plots, each of which employs the same category and cluster variables.
Each plot displays the data obtained for one particular water quality parameter and values can be read off the
horizontal axis. Data analysis has shown that hydrographic (flow) state is the main factor influencing water
quality so, for summary purposes, data have been divided into separate groups along the vertical axis in
accordance with their allocated flow category. These nominal flow categories have been selected to fit the
conditions encountered during sampling exercises, and do not precisely match established hydrological
terminology. Categories are defined as follows:
1) Event Peak – applies to any situation where water levels were changing by more than 5% per day
relative to peak discharge levels. It incorporates both the rising and early falling stages of the
hydrograph.
2) Falling Hydrograph (falling limb) – includes cases where water levels were dropping by more than
5% per week but less than 5% per day. This stage is extremely brief or virtually absent from some
waterways in the study area, especially constructed drains. However, at study sites that support
significant base flows, this criterion represents the flows that are maintained very soon after surface
runoff is no longer observable (i.e., soon after rainfall stops).
3) Stable base flow – includes situations where water levels were falling by less than 5% per week, but
where there was still visible surface flow within, upstream or downstream of the monitoring site.
2
4) No flow – applies to cases where there was no visible flow greater than 0.01 m/sec in any part of the
water body that was protected from wind-induced currents. This category groups together data
collected at sites that were genuinely stagnant with those that were subject to substantial throughflow of groundwater. These types of sites could be discriminated using various methods (such as
comparing the rates at which water levels change, with the rates predicted from evapotranspiration
data); however, there were insufficient data to justify further categorisation.
These categories would ideally have been based on discharge rate measurements rather than water levels, but
the required data could not be collected.
Each box displayed within a particular hydrographic category contains data collected from a different type of
sampling site, each type being distinguished by its colour and position relative to other boxes. (Within each
hydrographic category, boxes are always displayed in the same order and relative position, with a gap being
left in the appropriate position if there are no data available for any particular combination of categories).
The sample types employed for general survey data are as follows:
1) Minimal inputs from cane farms – data from samples collected at times and places where it was
obvious that very little of the water had originated from farmland. In most cases the water came
from mountainous rainforest catchments. Notionally this category contains reference data, but most
of these sites were located high in the lowland watershed so the data are only indicative of one
wetland type (upper catchment stream) that is poorly represented in the more heavily farmed areas
lower in the catchment. Accordingly, direct comparisons with other sample types are not strictly
valid. There is in fact only one instance (on Ripple Creek) where sites are sufficiently similar to
allow direct comparisons. These cases are not displayed separately on the box plots but are
discussed individually where relevant in the text.
2) Moderate inputs from cane farms – data obtained at times and places where it was clear that a
significant proportion of the water originated from farmland but where the amounts originating from
other areas were either unknown or known to be significant.
3) High inputs from cane farms – data obtained from samples collected from water bodies with
catchments that are dominated by cane farms.
The variability displayed by the boxes for these sample types can be attributed mainly to between-site
differences but includes some cases where water quality has changed over time at an individual site. The
plots also summarise data collected during other water quality monitoring activities conducted as part of this
project. This includes three types of samples collected at the Lagoon Creek study site and storm-event
samples collected by an autosampler located on a communal farm drain in the Macknade area. Samples and
readings taken from Lagoon Creek were allocated one of the following sample-type categories:
1) routine sampling station/autosampler - collected from close to mid-channel and 30cm below the
water surface, and so directly comparable to the regional survey samples;
2) transect samples taken less than one metre below the water surface - taken at various points on one
of five lateral transects located at fixed intervals along the length of the lagoon; and
3) transect samples taken from depths greater than 1 metre below the water surface along the same
sampling transects as (2).
The boxes shown for Lagoon Creek sample types 2 and 3 simultaneously display spatial and temporal
variations, while the boxes for type 1 samples display temporal variability only. In most cases, samples were
collected only once on each day (or sampling trip), unless hydrographic conditions changed, but meter
readings were taken at different times of the day during most sampling trips. Where necessary (i.e., in cases
where doing so has a significant effect on the summary statistics) morning and afternoon field measurements
are plotted as separate parameters. This means that the variability displayed in the plots can be attributed to
differences between days and/or sampling trips.
3
In most cases, data were collected manually and at a relatively low sampling frequency, but the Lagoon
Creek routine sampling station and Macknade Drain were sampled intensively during and immediately after
the one significant storm event that occurred after autosamplers were installed. The data displayed in the
Macknade Drain box plot were all obtained from this single storm event. The Lagoon Creek event peak and
falling hydrograph data include samples that were opportunistically collected whenever anomalous flow
conditions were encountered, but data are still dominated by samples collected during the single autosampler
event. Accordingly, the plotted values should only be used as a rough guide to the kinds of concentrations
that can occur during a storm event. The results of the intensively monitored event are discussed in detail
separately in appendix B.
A1.3.1 Electrical conductivity (salt concentrations): Figures 1 & 2
•
Conductivity values were generally low throughout the floodplain regardless of flow conditions. Values
higher than 500 µS/cm were observed only during periods when flow was absent and even then, only in
Cattle Creek, suggesting that this is probably a natural characteristic of that system. Values at all other
sites were significantly lower than 350 µS/cm on most occasions.
•
In our experience the distributions of local species of freshwater macrophytes, invertebrates and fish are
not noticeably affected by variations in conductivity if concentrations remain below 1,000 µS/cm
(roughly equivalent to total dissolved salt concentration of 500 mg/L). This suggests that salt
concentrations are not a significant ecological issue on the Herbert floodplain. (Note that in some parts
of this region it is not uncommon to find that invertebrate communities correlate with conductivity but
this almost certainly stems from a close correlation between conductivity, flow and season).
•
Whenever flows were present, conductivities were generally highest in cane-affected areas, suggesting
that farmed catchments contribute more salt to the system than other parts of the catchment. However, a
significant proportion (if not all) of this effect can be attributed to the natural hydro-geology of the
floodplain and the fact that the sites that are most affected by farm runoff are located low on the
floodplain where they naturally receive less dilution from rainforest systems and higher levels of
subsurface input from floodplain soils and ground waters.
•
The above effect is evidenced by the somewhat unusual (for this region) relationship between
conductivity and flow. Almost without exception, our monitoring of intermittent inland streams in this
region has shown that conductivity levels fall dramatically during storm events, and rise gradually
during periods of base flow, achieving maximum values when flows stop. In contrast (Cattle Creek
aside), the conductivity of streams on the Herbert floodplain increases during storms and falls to
minimum levels when flows are low. There is little doubt that this is due to inputs of water from
headwater streams in the forested ranges which continue to discharge low salinity water long after the
lowland catchment areas have ceased producing runoff. These inputs of water dilute and/or disperse salt
and other contaminants and therefore have an important bearing on the vulnerability of individual water
bodies to adverse impacts from farm runoff. In general, all other factors being equal, higher inputs from
rainforests would be expected to reduce both conductivity concentrations and vulnerability to impact.
•
Statistical analysis of both pooled and hydrographically grouped data have shown that conductivity is
very strongly correlated (r2>0.97) with concentrations of sodium, calcium, magnesium, hardness,
chloride, bicarbonate and alkalinity. Accordingly, the concentrations of these salts have not been
individually plotted. It is, however, worth noting that (Cattle Creek aside) alkalinity and hardness levels
generally range from extremely low to low, seldom exceeding 40 mg CaCO3/L. This indicates that most
waters on the floodplain are poorly buffered (i.e., they have limited capacity to neutralise acids) and are
very soft. Water hardness reduces the toxicity of many commonly occurring toxicants, especially trace
metals. Soft waters can therefore be particularly vulnerable to adverse effects from such contaminants.
A1.3.2 Morning and afternoon dissolved oxygen concentrations: Figures 3 and 4
•
Figure 3 includes only data collected early enough in the morning to be indicative of the daily minium
values that would have been achieved if there had been cycling due to photosynthesis in the water
4
column. Such cycling was evident to some extent at most sites, the main exceptions being cases where
oxygen values were chronically low, either during the peak of the hydrograph (at which time increased
turbidity can prevent photosynthesis) or during periods of low flow at sites that are infested with floating
and/or emergent weeds (also preventing photosynthesis in the water column). In both these cases the
morning reading is still usually indicative of the daily minimum but the daily maximum is often not
significantly higher.
On rare occasions during our research in this region, we have encountered light-limited heterotrophic
(microbe dominated) water bodies that heat up sufficiently during the day to cause an oxygen sag
resulting in the daily minimum being observed in the afternoon. (A 10oC increase in temperature can
increase microbial and plant respiration rates by a factor of 2 to 3 fold). A few of the macrophytedominated wetlands on the Herbert floodplain can exhibit a subtle version of this effect during cloudy
days or periods of high turbidity but there are no cases where this was a significant factor in Figure 3.
•
Figure 4 includes only data collected between midday and 4pm. These are roughly indicative of the
maximum oxygen concentration achieved on each day that the site was sampled. The actual time of the
maximum depends on the light exposure period and the flow rate. The peak occurs near the middle of
the day at sites that become shaded in the afternoon and/or have high flow rates, and later in the
afternoon at sites that are shaded in the morning and/or are stagnant.
•
All sites were hypoxic (under-saturated) in the morning. Even reference sites that receive flows of good
quality water from forested headwater streams did not achieve the 90-120 %Saturation levels
recommended in ANZECC (2000) Guidelines for tropical wetlands, or the 85-105 %Saturation levels
recommended in Draft Queensland (2001) Water Quality Guidelines for lower catchment streams in the
Wet Tropics. The reference sites were low-productivity streams so diel cycling was moderate, and
although oxygen concentrations were higher in the afternoon, they still did not reach guideline levels. It
is therefore clear that more realistic local dissolved oxygen guidelines are needed.
This problem is not unique to the Herbert floodplain (or to cane-growing areas) – overall our research in
the region indicates that perennial rainforest streams and a few well-aerated reaches of large rivers are the
only freshwater habitats in the wet-dry and dry tropics where dissolved oxygen levels remain close to
saturation for prolonged periods (and even at these sites levels fluctuate during storms and/or prolonged
dry spells).
•
The effects of cane farm runoff are most evident when flows (and therefore farm runoff) are present.
When flows are low or absent, in-stream processes dominate, individual sites exhibit much more
independent behaviour, and between-site variations increase substantially.
•
In order to put these results into an ecological context two additional reference lines have been added to
the plot. These have been derived from the experimental findings presented in subsequent appendices,
and represent the best available estimates of critical ecological thresholds that are currently available. It
should be noted these are not guideline values but rather a priori indications of the kinds of critical acute
exposure thresholds that will need to be taken into consideration when developing guidelines.
The upper reference line (50 %Saturation) is an estimate of the concentrations below which even brief
exposure can induce significant sublethal responses such as avoidance and behavioural modification, in
sensitive species and/or age classes. It also coincides roughly with the point at which gill ventilation
rates in fish begin to increase non-linearly indicating that proportionately larger amounts of energy are
being consumed in order to breathe. Overseas studies have shown that the resulting energy deficit can
cause serious chronic problems such as failure to reach reproductive maturity as well as reduced growth
and loss of condition. Another consequence of the rapid increase in ventilation when oxygen
concentrations are decreased below 50 %Saturation is that animals pass much larger quantities of water
across their gills. This can substantially increase contaminant uptake, thus increasing their susceptibility
to toxins.
The second reference line (at 25 %Saturation) is an a priori estimate of the concentration below which
the acute lethal limits of sensitive species and/or age classes of fish are likely to occur. Further
5
reductions in oxygen concentrations below this level would be expected to severely stress or kill one
species and/or size class of fish after another, even if levels remain that low for only a short time.
•
It is evident from figure 3 that even at reference sites, levels sometimes fall below the 50 %Saturation
threshold and in one instance levels dropped to just below the lower limit. This suggests that oxygen
availability is one of many natural habitat factors that determine the composition of biological
communities in these ecosystems. There is little doubt that prior to European settlement there would
have been many floodplain wetlands that were chronically hypoxic and which would have provided
habitat for those species that are specially adapted to such conditions. However, in this study we have
selected sites which should provide water quality conditions capable of supporting all but the most
sensitive of local aquatic species. Accordingly, the low saturation levels reported for farm-affected sites
are cause for concern and deserve much closer investigation. It is particularly noticeable, for example,
that when stormwater was present, all of the sites categorised as having high levels of input from cane
farms reported minimum daily concentrations well below the lethal threshold for sensitive species and
within the range that our tests have shown to be acutely lethal to key species such as barramundi.
•
These findings leave no doubt that dissolved oxygen availability is the main driver of ecosystem
condition in freshwater wetlands in this region.
A1.3.3 pH: Figures 5, 6 and 7
•
pH is a measure of hydrogen ion concentration and indicates whether a solution is acidic, alkaline or
neutral (a pH value of 7.0 is neutral; lower than this is acidic; higher is alkaline). The pH of most
natural waters is primarily governed by free concentrations of carbon dioxide, carbonic acid, carbonate
and bicarbonate, each of which reacts with the others to establish an equilibrium pH value. As a rule
of thumb, increasing carbon dioxide concentrations cause pH values to fall and decreasing carbon
dioxide concentrations cause pH to rise. All aerobic organisms release carbon dioxide when they
respire (breathe), hence aquatic plant and animal respiration tends to reduce pH levels in the
surrounding water column. In contrast, plants consume carbon dioxide in the presence of sunlight to
manufacture organic substances such as sugars. Consequently, when submerged plants are exposed to
adequate light levels (i.e., in the day time) they consume large quantities of carbon dioxide, and pH
levels in the surrounding water rise. This results in cyclical daily pH fluctuations similar to those
observed for oxygen.
Each water body has a different characteristic pH when biological processes are absent and carbon
dioxide concentrations are at equilibrium with the air. Nevertheless, the pH levels of many natural
wetlands are a reflection of the balance between carbon dioxide inputs (from the air and from
respiration) and carbon dioxide consumption (by plants during daylight hours). The main exception to
this general rule is situations where acids are present in significant quantities, as can happen in natural
wetlands that produce large quantities of humic acids or when acid sulphate runoff is present.
Due to the above complexities, three different pH estimates have been plotted in the data summaries.
Figures 5 and 6 display morning and afternoon field measurements that essentially parallel the oxygen
samples discussed previously, in that they provide estimates of daily minimum and maximum values.
However, the factors that determine whether the lowest levels of pH occur in the morning or afternoon
are much more complex. In this dataset it can be assumed that when flows were low, pH values were
at a minimum in the morning. When flows were high the relationship between time of day and pH
level varied between sites, although in each case there is still a high probability that the reported value
is either a daily maximum or a minimum.
The third pH plot (Figure 7) shows laboratory measurements obtained from samples that were
equilibrated with the air in the dark at 25oC. This process removes many (but not all) biogenic pH
effects and provides some indication of what the pH level would have been in the absence of
biological activity (and/or short term physical influences such as the supersaturation of carbon dioxide
that can occur in rainfall and some groundwater discharges).
6
•
Current draft guidelines for pH levels in Queensland streams (EPA, 2001) suggest that values should
fall between 6.0 and 8.0, as indicated by the reference lines in the pH plots. It is clear from the plots
that field pH values complied with the guideline when flow was low or absent. During the falling
limb and baseflow stages of the hydrograph there was no significant difference between reference sites
and sites that receive farm runoff. The results suggest that a lower guideline limit of approximately 6.1
(median value) would be appropriate for those stages of the hydrograph. Field pH levels were
noticeably lower than the guideline value of 6.0 during events, median values being in the order of 5.8,
which is approximately equivalent to fresh rainwater. Reference sites were not sampled during events
in this project; however, it should be noted that the reported levels are very similar to the values that
we have previously obtained from rainforest streams located in the coastal ranges of the wet tropics,
and are higher than the levels that are often reported for Melaleuca wetlands. More data are required
but at this stage it appears likely that a value of 5.8 would be a realistic lower guideline level for
median field pH values during event peaks (i.e., at times when surface runoff is the main source of the
flow). Any values less than or equal to 5.5 could be indicative of problems and should be
investigated. All of the reported field measurements comply with this tentatively proposed guideline.
(It should be noted that in the monsoonal tropics runoff during storm events can be very dilute leading
to very low ionic strengths. The standard glass combination electrodes that are supplied with most pH
meters do not usually yield reliable measurements in low ionic strength waters. Most manufacturers
offer optional electrodes specifically designed for such situations and it is essential that these are used
when monitoring storm events in the wet tropics).
•
The lower pH values obtained during storm events are caused by super-saturation of carbon dioxide
(cold rain droplets take up large amounts of carbon dioxide which must be released when the rain
reaches the ground and heats up). Consequently, the equilibrated storm samples analysed in the
laboratory (which have had time to release excess carbon dioxide) yield pH levels that are not
significantly different from samples collected during the falling limb of the hydrograph. A minimum
median value of 6.5 is probably a realistic guideline for laboratory pH values, although based on our
experience with rainforest streams, a median level of 6.2, with an absolute minimum (for any individual
sample) of 6.0, can be expected in many relatively pristine streams.
•
The extremely low laboratory pH values reported for Macknade Drain occurred during the February
storm event. There is a high probability that these are indicative of farm-related water quality
problems. These results are discussed in detail in section 1.3.
•
The higher field pH levels observed during periods of stagnation are due to the combined effects of
increased alkalinity concentrations, decreased re-aeration and increased plant productivity. Both
morning and afternoon values generally complied with the water quality guidelines; however, afternoon
measurements taken within submergent macrophyte beds yielded values significantly higher than the
upper limit of 8. As would be expected, these higher levels are not evident in the laboratory analyses
(because samples have been kept in the dark and have had the opportunity to take up carbon dioxide
from the air).
•
It is tempting to conclude that an upper limit of 8 would be an acceptable guideline for both field and
laboratory pH levels in the Herbert floodplain. However, it is important to remember that the processes
that increase pH levels during the day are the same ones that produce oxygen. Most of the study sites
are prone to hypoxia and since they are poorly re-aerated when flow rates are low, production of
oxygen by plants appears to be one of the few natural means of ameliorating this undesirable condition.
In other words, higher levels of photosynthesis by submergent plant biomass may be desirable and this
would inevitably increase field pH values, probably to levels much higher than the guideline value of
8.0. This raises the question of whether such high levels of productivity would be sustainable or would
simply lead to other problems. Much more research will be necessary before this question can be
answered confidently, but at this stage it seems likely that outcomes will depend on which plant species
are involved. Notably, we have encountered water bodies in other parts of this region (e.g., the Ross
River Weirs) that seem to be capable of sustaining high levels of native submergent plant productivity
for prolonged periods without developing serious problems, other than some loss of recreational
amenity due to high plant biomass.
7
It currently seems likely, therefore, that amelioration of many of the existing hypoxia problems in local
wetlands will rely on rehabilitation measures aimed at reducing the competitive success of exotic water
weeds and enhancing the productivity of desirable native plant species. If this is accomplished,
afternoon pH levels are likely to rise well above current guideline values whenever flows are low. It is
likely that local aquatic animals are well adapted to cope with pH fluctuations of this kind, so unless
the levels of productivity causing these high pH values are attributed to problematic species, or plant
growth rates are shown to be unsustainable, these high values should not necessarily be considered
problematic in their own right. However, an additional issue that must be considered when evaluating
the pH requirements of wetlands is that high pH values greatly increase the toxicity of ammonia. Data
presented later demonstrate that potentially dangerous concentrations of ammonia can occur at certain
times and places in wetlands, especially those that receive runoff from cane farms or that are subject to
impacts from livestock. Any guideline for upper pH levels in such wetlands must take these situationspecific risks into account.
A1.3.4 Suspended particulate matter: Figures 8 and 9.
•
Suspended particulate matter (SPM) is also sometimes referred to as total suspended solids (TSS),
suspended sediment concentrations (SSC), or simply sediment concentrations.
•
It is clear from the plots that when flows are present, SPM concentrations increase in proportion to the
inputs from cane farms. This result confirms the findings of Bramley and Roth (2002) who reported
significant positive correlations between SPM concentrations and increased cane farming pressure in
the lower Herbert Catchments.
•
SPM levels during periods of stagnation are driven by site-specific factors and therefore varied quite
widely between sites.
•
As would be expected, SPM concentrations were highest during storm events. The event data shown in
these plots are discussed in detail.
•
SPM concentrations are influenced by a complex range of situation and site-specific factors.
Moreover the relative importance of different factors varies tremendously over the hydrograph. For
example, catchment characteristics are an important determinant of the concentrations that occur
during storm events, while in-stream biophysical factors such as hydrodynamics and biological
activity often play a leading role during periods of very low flow. Due to these complex natural
variations the SPM requirements and tolerances of wetlands vary enormously, making it impossible to
propose guidelines that are valid and meaningful for all local wetlands. This is problematic because,
while there are undoubtedly some wetlands that are remarkably tolerant of changes in SPM
concentrations, there are also many where anthropogenic changes in SPM concentrations can cause
very significant alterations to the ecosystem, and perhaps most importantly in this region, can modify
the metabolism of a water body leading to substantial changes in oxygen status. The experiments
provide some quantitative indications of the kinds of effects that can occur and suggest a methodology
for determining critical ambient levels. Unfortunately, each water body (indeed each part of each
water body) responds differently to SPM and has different water clarity requirements, so the methods
that have been developed must be applied independently to individual sites. Much more sophisticated
approaches (than have been developed to date) to inventory and classification of wetland systems and
identification of key habitats will need to be employed before it will be possible to apply these
methods at subcatchments scales.
A1.3.5 TOC, colour and BOD: Figures 10, 11 and 12.
•
TOC (Total Organic Carbon), colour and BOD (Biochemical Oxygen Demand) are related parameters.
TOC provides a measure of the total amount of organic matter contained in water samples. This
comprises fine particles of detritus, microbes, plankton, dissolved humic substances and other
8
dissolved substances such as sugars and alcohols. A variable proportion of this material (depending on
when and where samples are taken) is readily degradable and consumes significant amounts of oxygen.
Colour provides a measure of the intensity of yellow-brown colouration in the water. This is usually
caused by humic organic substances but can be affected by high concentrations of other substances,
especially iron complexes. In this dataset, colour is closely correlated to DOC (Dissolved Organic
Carbon).
BOD is determined by a standardised laboratory test designed to quantify the amounts of readily
biodegradable (i.e., oxygen demanding) organic matter present in a sample. The test was originally
designed for sewage testing and it is subject to many limitations and potential interferences when
employed in natural waters. Nevertheless, in the absence of any better alternative, it is widely
employed in this application.
•
Pristine waters generally have BOD levels less than 2 mg/L and all ten samples collected from
reference sites during this study reported these low values. This simply means that the concentrations
of rapidly degradable organic matter suspended or dissolved in the water column were low enough to
ensure that only small quantities of oxygen were consumed over the five-day duration of the standard
BOD test. It does not necessarily mean that the biological oxygen demand of the water body as a
whole was low (larger plants and animals, microbes and animals in the bottom sediments and benthic
algae can be the principal sources of oxygen consumption, especially in shallow waters), nor does it
mean that organic carbon concentrations were low (humic substances such as tannins are degraded very
slowly and therefore can be present in high concentrations without exerting much oxygen demand).
•
TOC concentrations were low at all reference sites, with values ranging from 0.4 to 3.2 mg/L, which is
typical of rainforest streams. In forest catchments, a high proportion of the organic carbon usually
comprises particulate organic detritus (e.g. very small particles of leaf material) so values do not
usually correlate with colour (which is indicative of dissolved organic substances). This absence of
correlation with colour was evident at reference sites. Paradoxically, colour concentrations were
somewhat higher than would normally be expected of other kinds of waters with similarly low TOC
concentrations. This is also typical of local rainforest streams and seems to indicate that the dissolved
organics from rainforest catchments have unusually strong colouration. The colour intensities obtained
at reference sites in this study are at the high end of the range that we have observed in pristine wet
tropics streams.
•
Concentrations of TOC, colour and BOD were significantly higher at sites that receive cane farm
runoff than they were at reference sites, and the differences were most consistent during events and the
falling stages of the hydrograph (i.e., at times when large amounts of farm runoff were present).
Natural wetland systems are known to be extremely variable with respect to both concentrations and
composition of TOC and colour. Natural Melaleuca wetlands, for example, can be expected to produce
tannin-stained waters with TOC and colour levels higher than those reported in farming areas, while
many submergent-plant-dominated wetlands consistently maintain low TOC and colour concentrations
similar to those observed at reference sites. It is likely that the large spatio-temporal variability of
colour (and therefore light attenuation) in natural wetland systems is an important mechanism for
maintaining their high biodiversity. Consequently, there is no simple numerical guideline that can be
applied when attempting to determine if colour and TOC concentrations are acceptable – requirements
vary immensely between water bodies and over time.
However, there is no doubt that the overall effect of wetland reclamations, clearing of land and riparian
vegetation, aquatic weed invasions and monoculture farming has been to remove much of the natural
diversity of carbon sources from these systems, creating a situation where both the composition and
concentrations of TOC in local waterways are now much more uniform than they once were. This
means that concentrations may be close to natural in some wetlands but completely deviant in others.
Bohl et al. (2001) have also shown that cane-farm runoff can at times (especially when runoff is
generated during the cane-cutting season) contain high concentrations of sugars and sugar degradation
products released from cane juice spilled during harvesting. These dissolved organic compounds,
9
which rarely occur in detectable quantities in natural systems, are highly bioavailable and exert much
higher oxygen demands than natural humic substances.
There were very limited opportunities for this study to sample runoff generated during the cane-cutting
season, so the results presented here are only indicative of the conditions that prevail at times when
there is a low potential for sugar to be released from farms. The presence of sugar would undoubtedly
increase BOD levels and decrease the correlation between TOC and colour (because sugar and its
degradation products are generally colourless).
•
In the absence of significant inputs of sugar and under the weather conditions encountered during this
study, BOD, TOC and colour reached maximum levels during the falling limb of the hydrograph (at
times when subsurface inputs would be expected to dominate). The concentrations maintained during
the event peak were still significant, and since discharge rates were much higher at that time, it is likely
that these surface-runoff-driven flows were responsible for the majority of carbon export from the
catchment. However, the comparatively smaller amounts introduced during the much more prolonged
periods when flows were driven by sub-surface inputs clearly have the greatest potential to affect the
receiving ecosystem. These inputs undoubtedly play a significant role in maintaining the hypoxic
conditions that have consistently been encountered at study sites.
•
Publications dealing with organic carbon in wetland systems often emphasise its direct role as a source
of energy (in food webs) and/or consumption of oxygen. However, in natural wetland systems the
creation of coloured water can be the principal means by which organic carbon influences the
productivity and diversity of the ecosystem. Colour absorbs significant amounts of photosynthetically
active light and therefore interacts with turbidity (caused by SPM) to determine when, where and how
different plant species can grow. This not only affects community structure but also impacts on
metabolism and therefore the oxygen status of the water body. In particular, it is important to
understand that even if coloured organic matter has negligible BOD, high levels can substantially
decrease the concentrations of DO in a water body simply by preventing plants from photosynthesising.
•
High BOD concentrations were reported from samples collected in the severely hypoxic bottom layer
of the water column at Lagoon Creek. This may be indicative of the generation of oxygen-demanding
by-products from anaerobic decomposition of organic detritus, but could also be attributable to inflow
of groundwater containing high concentrations of BOD. Based on consideration of the behaviour of
other contaminants, the former seems more likely than the latter, but more research is needed to reach a
definite conclusion.
A1.3.6 Chlorophyll a and phaeophytin: Figures 13 and 14
•
Chlorophyll a is used as an indicator of phytoplankton (planktonic algae) biomass. Phaeophytin is the
inactive form of chlorophyll a and is the first degradation product to appear when algae die. Some
algae store excess chlorophyll in the form of phaeophytin if it is not currently needed for some reason
(e.g. high light levels). Since algae have the capacity to transform one to another, chlorophyll and
phaeophytin concentrations must be interpreted very cautiously. But as a central rule, a predominance
of chlorophyll a is likely to be indicative of vibrant, growing algal populations, while a predominance
of phaeophytin suggests some form of stress.
•
Figure 13 shows that while chlorophyll levels are generally substantial most of the time, extreme
concentrations are comparatively rare, particularly in view of the fact that in many instances nutrient,
light, and flow conditions would have been conducive to very high levels of plant growth. Moreover,
most of the high concentrations shown on the plot are of little consequence: the few high values
reported for Macknade Drain (event peak) resulted when algae that had grown in the puddles of water
lying in the drain between events was flushed out, while the high concentration reported at Lagoon
Creek on the falling hydrograph were isolated to one small area at the upstream end of the lagoon and
never spread any further. The very high concentrations observed in the bottom waters of the Lagoon
Creek study site during periods of no flow were more significant as they persisted for long periods over
significant sections of the water body. These occurrences coincided with very high concentrations of
10
highly bioavailable ammonia and FRP. Notably, even though the phytoplankton biomass was
extremely high, oxygen evolution rates were not sufficient to increase dissolved oxygen concentrations,
suggesting that heterotrophic (microbial) productivity was very high at the time.
•
In the majority of cases phaeophytin concentrations were higher than chlorophyll, suggesting that
phytoplankton were not experiencing optimal conditions.
•
Chlorophyll samples were not collected during all field trips, but visible phytoplankton blooms were
rare and were confined mainly to small isolated pools that formed as ephemeral wetlands began to dry
up.
•
The combined evidence of field observations, metabolism (oxygen and pH cycling) measurements,
chlorophyll sampling and laboratory experiments allow us to summarise the status of phytoplankton in
the Herbert wetlands over the course of this study as follows. The biomass of phytoplankton in
wetlands that receive cane farm runoff is generally higher than it is at reference sites but the differences
are more subtle and less consistent than they are for most other productivity indicators. Ambient
concentrations consistently remain substantial and in many other, less productive, types of water body
would rate as being high. However, in these wetland environments the productivity of other
components of the ecosystem (especially large plants and microbes) is so high that phytoplankton are
rarely a major contributor to overall productivity. Blooms occur occasionally but are usually brief and
isolated, and in the longer term phytoplankton are rarely able to out-compete other organisms. This
failure to achieve dominance means that algal blooms appear to be less of an issue than they might be
in other types of aquatic ecosystems. Nevertheless, significant levels of biomass are sustained and these
undoubtedly play some role in oxygen dynamics and support certain food webs. Hence, phytoplankton
cannot be ignored even when evaluating habitats that are strongly dominated by other producers.
11
Figure 1
REGIONAL SURVEYS
LAGOON CREEK STUDY SITE
MACKNADE CANE DRAIN
Minimal inputs from cane farms
Routine sampling station / autosampler
Samples collected by autosampler
Moderate inputs from cane farms
Transect samples taken <1m below the surface
High inputs from cane farms
Transect samples taken >1m below the surface
Minimum value that
Is not an outlier
Inter-quartile Range
25th Freshwater
Percentile
Australian Centre for Tropical
Research
Median
Maximum value that
Is not an outlier
75th Percentile
Outlier
12
Figure 2
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
13
Figure 3
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
14
Figure 4
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
15
Figure 5
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
16
Figure 6
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
17
Figure 7
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
18
Figure 8
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
19
Figure 9
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
20
Figure 10
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
21
Figure 11
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
22
Figure 12
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
23
AWQG 2000
Draft QWQG 2001
Figure 13
REGIONAL SURVEYS
MACKNADE CANE DRAIN
*Excludes an outlier (224 µg/L)
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
24
Figure 14
REGIONAL SURVEYS
MACKNADE CANE DRAIN
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
25
APPENDIX B
CASE STUDY: INVESTIGATION OF THE EFFECTS OF FLOATING WEED
MATS ON THE HYDRAULICS AND OXYGEN STATUS OF A BURDEKIN
FLODPLAIN LAGOON.
26
B1.
CASE STUDY: INVESTIGATION OF THE EFFECTS OF FLOATING WEED
MATS ON THE HYDRAULICS AND OXYGEN STATUS OF A BURDEKIN
FLODPLAIN LAGOON.
B1.1 Aim
To determine the effect of removing a dense assemblage of floating macrophytes from a permanent lagoon
on the Burdekin floodplain in terms of oxygen regime and hydrodynamic flow patterns.
B1.2
Site Description/background
These field experiments were conducted at Gorizia’s Lagoon, a permanent lagoon located on Sheepstation
Creek, approximately 5 kilometres north-west of the township of Brandon. The creek system is a floodplain
distributary channel that hosts a chain of naturally permanent lagoons. Prior to the commencement of surface
water irrigation on this section of the floodplain the creek flowed only intermittently and during the dry
season water in the lagoon was a surface expression of the local watertable. However, the system is now
used to distribute water for irrigation purposes and this has resulted in the establishment of an artificially
perennial flow regime. The altered flow regime has resulted in extensive invasions of aquatic weeds into this
system (initially water hyacinth followed by para grass and other emergent species).
These infestations have progressed to the point where most of the lagoons in the system are now completely
covered by a dense mat of floating vegetation, resulting in substantial loss of aquatic habitat value. Local
stakeholders are attempting to rehabilitate the lagoons by removing the weeds using mechanical harvesting
techniques, that provided a unique opportunity for us to examine the effects of weed removal on the ecology,
water quality and hydrodynamics of the lagoons. The study detailed here deals with the immediate effects on
dissolved oxygen and hydrodynamics. Studies of the longer-term effects on water quality, habitat values and
fish populations have been reported by Perna.
B1.3
Method
This field experiment was completed in two stages. The first stage involved conducting an initial assessment
of hydrodynamics and water quality in the lagoon (two months) prior to the commencement of weed
harvesting. The second stage involved conducting the same analyses after weed harvesting was completed to
assess the net gains for the system in terms of water quality and hydrodynamic efficiency.
The hydrodynamics in the lagoon were assessed using a fluorescent dye tracer (Rhodamine WTS). This
involved injecting the dye at the upstream extent of the lagoon and collecting regularly timed water samples
for dye analysis downstream at a transect to determine the detention time and general flow patterns through
the lagoon. Samples were also collected at the downstream extent of the lagoon to obtain a total detention
time for the lagoon under the prevailing conditions. Rhodamine WTS concentrations were analysed using a
Turner Model TD700 Flourometer.
Data collection for hydrodynamic assessment entailed:
The collection of water samples (for dye analysis) at regular time intervals from different points across a
transect located about one quarter of the way through the lagoon. Water samples were collected at different
points in the water column using a system of pipes connected to a peristaltic pump located on the adjacent
bank.
The collection of water samples at the downstream extent of the lagoon (using an ISCO 3700 autosampler)
for dye analysis to determine a total lagoon detention time.
Estimating inflow rates using the velocity-area method and a Swoffer model 3000 velocity meter, for the
duration of the tracer study.
Water quality assessments comprised:
Logging of physico-chemical parameters at the upstream and downstream extents of the lagoon using
Hydrolab multi-parameter probes throughout the term of the experiments and for a period of time afterwards.
Collection of physico-chemical data from the dye sampling transect points using a Hydrolab multi-parameter
probe and flow-through cell connected sequentially to the pipe-work via the peristaltic pump.
27
The collection of diel cycling data at predetermined locations on the lagoon at regular depths through the
profile post-harvest when access via a boat became possible.
Considerable effort was expended to remove the dense macrophyte assemblages. Initial trials with the weed
harvester proved difficult and it proved necessary to use an excavator to remove the bulk of the plant
material while the water harvester cut off pieces of the plant mat and maneuvering them over to the
excavator. A vast majority of the plant material was removed prior to the post-harvest assessments, with only
about 10 percent remaining at the upstream end of the lagoon. This could not be removed due to access
difficulties.
B1.4
Results
Figure 1 and 2 show images of the lagoon pre and post-harvest, respectively. It can be seen that the weed mat
consisted of a complex assemblage of several aquatic and semi-aquatic plant species including para grass
(Brachiaria mutica), bulrush (Typha domingensis), swamp rice grass (Leersia hexandra) and spiny mudgrass
(Pseudoraphis spinecens) while Figures 3 and 4 show the extent of the floating weed mat before and after
harvest. Also indicated on these figures are the locations of upstream and downstream monitoring stations,
the dye sampling transect and the location of the downstream dye sampling station.
Hydrodynamic dye analysis
The location for the dye sampling transect (See Figures 3 and 4) was chosen so that the flow of water
entering the main body of the lagoon could be traced through the water column. The three sampling points
were located at 1, 2 and 3 metre depths at approximately the centre of the lagoon (40 metres from the bank)
with a total depth of the lagoon at this point of 3.5 metres.
The dye was injected at the same time on respective days (1000 hours) for both the pre and post-harvest
analyses at the upstream culvert that represents the upstream extent of the lagoon (See Figures 3 and 4). The
results of the dye study pre-harvest (Figure 5) and post-harvest (Figure 6) show that a severe channelling of
the flow occurred midway through the water column when the water surface was totally covered with a layer
of floating plants (See Fig. 6). This figure shows that a large majority of the water that entered the lagoon
short-circuited through the centre of the water column with minimal surface mixing and exited the lagoon
relatively quickly. Figure 6 shows that once the plants were removed surface flows predominated which
resulted in greatly enhanced mixing and re-aeration of the water body.
The study also included the collection of dye samples at the downstream extent of the lagoon but this was
only successful for the post-harvest experiment. This was probably due an insufficient amount of dye used in
the pre-harvest experiment, which could not be measured once it had diluted through the whole lagoon. The
post-harvest dye trace experiment was, however, successful because a significantly higher concentration of
dye was used (3 times more than for pre-harvest) and we were able to produce a time-of-travel plot for the
entire lagoon (see Figure 7). It can be seen from this plot that the first significant dye concentrations were
measured about 54 hours (2.25 days) after injection with the average detention time being about 72 hours (3
days). Although we were not able to measure the pre-harvest detention time there is little doubt that it would
have been significantly less than 3 days (at existing inflow rates) due to the piped flow effects evidenced by
the results obtained further upstream (Figures 5 and 6). A feature of Figure 7 is the significant deviation from
plug flow, as evidenced by the lack of a significant single peak in the dye concentration curve. This is typical
of a system with complex mixing patterns within which incoming waters flow through a large majority of the
lagoon volume before exiting the system. This is a much-improved scenario for enhanced re-aeration of the
water body compared to the pre-harvest case.
Water quality analysis
Hydrolab multi-parameter probes installed at the extreme upstream and downstream ends of the lagoon
(Figures 3 and 4) collected physico-chemical data during and immediately after the dye tracer study. The
parameters measured included dissolved oxygen, pH, temperature and specific conductivity. In this study we
were primarily concerned with the dissolved oxygen levels as it the primary factor that dictates fish habitat
quality in these systems. The dissolved oxygen levels at the upstream and downstream locations (logged over
28
4 days) and the respective estimates of flow rates are shown in Figures 8 and 9 (pre and post-harvest
respectively). Remarkable improvements in dissolved oxygen levels due to weed removal are clearly evident.
Figure 8 shows that water that entering the lagoon pre-harvest was poorly oxygenated (with daily dissolved
oxygen cycling from 0 to a maximum level of 10% saturation during the study) the water became almost
totally be oxygenated (anoxic) once it reached the downstream end of the lagoon. After the weeds were
removed (Figure 9) the situation improved dramatically with minimum daily dissolved oxygen levels of 10%
saturation improving to 50% saturation once the water reached to downstream end of the lagoon. It is also
important to note that the observed improvement in the upstream dissolved oxygen levels (pre to postharvest) was primarily due to the rehabilitation of a major upstream lagoon in the interim.
Figures 10 and 11 show the dissolved oxygen levels that were measured at three different depths midway
across the lagoon at the dye sampling transect. These figures demonstrate that severely hypoxic/anoxic
conditions persist under dense macrophyte assemblages of this type but relatively healthy conditions can be
restored once the plants are removed. They also highlight how rapid recovery in dissolved oxygen levels can
be achieved when it is considered that the point at which these measurements were taken is only about 30
meters from the upstream water hyacinth clogged section of the lagoon (see Figure 4) under which the
dissolved oxygen levels would have been severely depressed.
Additional diel cycling and vertical profile data were also collected at 5 locations along the thread length of
the lagoon. Figures 12, 14 and 16 show the morning readings of dissolved oxygen, temperature and pH
respectively, measured at three different depths starting at the origin (representing the location of the dye
sampling transect) and extending almost 1 kilometre to the downstream end of the lagoon. Figures 13, 15 and
17 show the equivalent late-afternoon readings at these same locations. A general improvement in dissolved
oxygen levels can be seen going from the upstream to downstream end of the lagoon. The slight peak in the
late-afternoon readings (Figure 13) can be explained by the shallow depth (< 2 metres) and the dense
assemblages of submergent plants (consisting primarily of waternymph or Najas tenuifolia) found from
about 50 to 300 metres thread length during the time of the post-harvest experiments.
B1.5
Discussion
The results of the dye analysis show that the floating mats create an upper boundary on the waterbody which acts as a
confining surface that causes the flowing water to preferentially flow through the middle of the water column (see
Figure 5) instead of across the water surface, which was the case once the plants were removed (see Figure 6). This
piped flow caused a situation whereby the replenishment of the dissolved oxygen by atmospheric re-aeration was
severely inhibited. Combined with the total shading effect of the floating plants that resulted in minimal plant
photosynthesis within the water column, constant conditions of severe hypoxia/anoxia occurred throughout the entire
water body (see Figure 10).
Once the free water surface was restored to the lagoon the majority of the water flowed through the near
surface layers (see Figure 6). This greatly enhanced re-aeration and mixing within the lagoon and when
combined with the effect of prevailing winds on the water surface, which further enhances deep mixing, the
expectation that a greatly enhanced dissolved oxygen regime would result proved to be correct. The water
quality measurements taken during the experiment confirm this view. Figure 8 shows that a net deterioration
in the quality of water flowing through the lagoon was found to occur before the plants were harvested.
While Figure 9 shows a dramatic net improvement in water quality from the upstream to downstream extent
of the lagoon (note that the input water was also of an improved quality compared to the unharvested case
due to the rehabilitation of upstream lagoons during the course of the study). The results in Figures 12 and 13
provide clear evidence that there was rapid establishment of submergent plant growth within the shallower
(<2 m) section of the lagoon. This results in a healthy level of diel oxygen cycling which enhanced daytime
oxygen concentrations without inducing adverse levels of hypoxia overnight.
Long-term monitoring will be needed to determine: 1) if reinfestation of floating weeds can be prevented;
and 2) if improvements can be sustained in the long term. Experience in the area to date indicates that the
submergent macrophytes that re-establish after removal of weed mats are generally native species but the
potential for invasions by problematic weed species such as Hymenachne and Cabomba cannot be
disregarded. There is also potential for submergent plant biomass to increase to the point where respiratory
demand can exceed re-aeration capacity especially during periods of low flow supplementation and/or
29
cloudy weather (other studies (Perna, 2003) have reported significant oxygen sags in cleared lagoons under
such circumstances). The risks of adverse outcomes of this kind are exacerbated by the current water
management practice of allowing flow supplementation to fall to minimum levels when irrigation demand
drops during cloudy periods (farmers generally stop irrigating if rain appears imminent). There is theoretical
potential to provide an environmental flow allocation to boost re-aeration rates at such times but the
feasibility of doing this is yet to be investigated. The absence of riparian shade trees increases the risk of
such problems. Riparian shade could potentially create patches of edge habitat with sparse aquatic vegetation
assemblages that would act as refugia for aquatic organisms during brief oxygen sags.
30
Figure 1: Gorizia Lagoon before weed removal.
Figure 2: Gorizia Lagoon after weed removal.
31
Figure 3: Aerial view of Gorizia Lagoon before weed harvest indicating extent of weed infestation (green
hashed) consisting of a complex assemblage of aquatic plants and open water (blue solid).
Downstream dye
sampling point
Dye sampling transect
Downstream
logging
Hydrolab
location
Upstream
logging
Hydrolab
location
Dye
injection
point
Figure 4: Aerial view of Gorizia Lagoon after weed harvest indicating extent of remaining weed infestation
(green hashed) consisting primarily of water hyacinth and open water (blue solid).
Downstream dye
sampling point
Dye sampling transect
Downstream
logging
Hydrolab
location
Upstream
logging
Hydrolab
location
Dye
injection
point
32
Figure 5: Dye time-of-travel curve before weed harvest.
Dye Concentration, ug/L
15
12
9
6
3
0
0
5
10
15
20
25
30
35
40
45
50
45
50
Time, hours
1 metre deep
2 metres deep
3 metres deep
Figure 6: Dye time-of-travel curve after weed harvest.
15
12
9
6
3
0
0
5
10
15
20
25
30
35
40
Time, hours
1 metre deep
2 metres deep
3 metres deep
33
Figure 7: Dye concentrations recorded at the outlet end of the Lagoon after weed harvest.
2
1.5
1
0.5
0
0
20
40
60
80
100
120
Time, hours
Figure 8: Dissolved oxygen data logged before weed harvest with corresponding estimates of the upstream
flow rates.
3
90
2.5
80
70
2
3
60
Flow Rate, m /s
Dissolved Oxygen Concentration,% Saturation
100
50
1.5
40
1
30
20
0.5
10
0
12/03/02
00:00
0
12/03/02
12:00
13/03/02
00:00
13/03/02
12:00
14/03/02
00:00
14/03/02
12:00
15/03/02
00:00
15/03/02
12:00
16/03/02
00:00
Time
Upstream DO
Downstream DO
Upstream Flow Rate
34
Figure 9: Dissolved oxygen data logged after weed harvest with corresponding estimates of the upstream
flow rates.
3
90
2.5
80
70
3
2
Flow Rate, m /s
Dissolved Oxygen Concentration,% Saturation
100
60
1.5
50
40
1
30
20
0.5
10
0
04/02/03
00:00
0
04/02/03
12:00
05/02/03
00:00
05/02/03
12:00
06/02/03
00:00
06/02/03
12:00
07/02/03
00:00
07/02/03
12:00
08/02/03
00:00
Time
Upstream DO
Downstream DO
Upstream Flow Rate
Figure 10: Dissolved oxygen levels measured at the dye sampling transect before harvest.
Dissolved Oxygen Concentration, % Saturation
10
8
6
4
2
0
12/03/02
00:00
12/03/02
06:00
12/03/02
12:00
12/03/02
18:00
13/03/02
00:00
13/03/02
06:00
13/03/02
12:00
13/03/02
18:00
14/03/02
00:00
14/03/02
06:00
14/03/02
12:00
14/03/02
18:00
15/03/02
00:00
Time
1 metre depth
2 metre depth
3 metre depth
35
Figure 11: Dissolved oxygen levels measured at the dye sampling transect after harvest.
60
50
40
30
20
10
0
04/02/03
00:00
04/02/03
06:00
04/02/03
12:00
04/02/03
18:00
05/02/03
00:00
05/02/03
06:00
05/02/03
12:00
05/02/03
18:00
06/02/03
00:00
06/02/03
06:00
06/02/03
12:00
06/02/03
18:00
07/02/03
00:00
Time
1 metre depth
2 metres depth
3 metres depth
Figure 12: Early morning dissolved oxygen distribution along the Lagoon thread length after harvest.
100
90
80
70
60
50
40
30
20
10
0
0
100
200
300
400
500
600
700
800
900
1000
Thread Length, m
0.3 metres depth
0.5 metres depth
1 metre depth
36
Figure 13: Late afternoon dissolved oxygen distribution along the Lagoon thread length after harvest.
100
90
80
70
60
50
40
30
20
10
0
0
100
200
300
400
500
600
700
800
900
1000
Thread Length, m
0.3 metres depth
0.5 metres depth
1 metre depth
Figure 14: Early morning temperature distribution along the Lagoon thread length after harvest.
34
Temperature, Deg.C
32
30
28
26
0
100
200
300
400
500
600
700
800
900
1000
Thread Length, m
0.3 m
0.5 m
1m
37
Figure 15: Late afternoon temperature distribution along the Lagoon thread length after harvest.
34
Temperature, Deg.C
32
30
28
26
0
100
200
300
400
500
600
700
800
900
1000
Thread Length, m
0.3 m
0.5 m
1m
Figure 16: Early morning pH distribution along the Lagoon thread length after harvest.
8
pH
7.5
7
6.5
6
0
100
200
300
400
500
600
700
800
900
1000
Thread Length, m
0.3 m
0.5 m
1m
38
Figure 17: Late afternoon pH distribution along the Lagoon thread length after harvest.
8
pH
7.5
7
6.5
6
0
100
200
300
400
500
600
700
800
900
1000
Thread Length, m
0.3 m
0.5 m
1m
39
APPENDIX C
SPATIO-TEMPORAL WATER QUALITY VARIABILITY WITHIN
INDIVIDUAL WATERBODIES
40
C1
LAGOON CREEK AMBIENT MONITORING
Detailed monitoring of physico-chemical parameters was conducted at Lagoon Creek on the Herbert River
floodplain from November 1999 to March 2002 to gain an understanding of spatial and temporal variation of
these parameters under a variety of hydrological and seasonal conditions. In November 1999, this
monitoring coincided with a fish kill.
C1.1
Methods
Readings were taken along a series of five transects (set perpendicular to the water flow) which were
established over a 0.5-1 km section of the creek. The section of creek surveyed on any given trip depended
on accessibility; occasionally, dense mats of water hyacinth prevented access to some areas. At each
transect, morning and afternoon readings of dissolved oxygen, pH, temperature and conductivity were
recorded throughout the water column (from just below the surface to the bottom) at each of three locations
(left-hand-side edge, middle, right-hand-side edge).
C1.2
•
Results and Summary
The data collected during this study comprise more than 1600 measurements of dissolved oxygen, pH,
temperature and conductivity, which were used to evaluate the limnology and metabolism of the lagoon
in order to determine the key drivers of the oxygen status of the lagoon. The dissolved oxygen data are
summarised in two box plots:
1) Figure C1-1, which shows the effect of hydrographic conditions on vertical distribution and diel
cycling of oxygen concentrations;
2) Figure C1-2, which shows how the distribution of morning and afternoon oxygen concentrations
(indicative of daily minima and maxima) in the oxygenated near-surface layer of the water column
varied between sampling trips over the course of the study.
•
Each box in Figure C1-1 displays variations over both time (between sampling trips) and space
(different locations within the lagoon). The boxes in Figure C1-2 on the other hand display spatial
variations over the length and breadth of the surface layer of the lagoon.
•
It is immediately obvious from these plots that the precise timing (within minutes to hours) and
location (within ten to twenty centimetres) of a DO reading has a very large bearing on the results that
are obtained. Hence, the probability of any individual random spot measurement yielding a
representative result is very low. However, data of the kind presented in these figures can be analysed
in conjunction with biophysical data in order to determine the most likely times and locations of daily
maximum and minimum values, and this provides a basis for the collection of spot measurements that
can be meaningfully interpreted.
•
Some diel cycling was evident in the near-surface layer of the water column during storm events.
This is largely due to photosynthetic oxygen production by phytoplankton during daylight hours.
Oxygen concentration at depths greater than 1m below the surface were consistently low and generally
correlated with flow rates (i.e., concentrations fell as flow rates fell). This can be attributed to reduced
re-aeration (i.e., reduced rates of introduction of oxygen into the water from the air).
Even though there was measurable production of oxygen in the near-surface layer during events,
concentrations were consistently low, daily maximums rarely exceeding 20 %Saturation. Even in the
most highly oxygenated parts of the lagoon, daily minimum values were below 10 %Saturation most of
the time during events. Concentrations were frequently below the asphyxiation thresholds of oxygendependent fish species suggesting that these species, if present, would have needed to employ
specialised tactics such as aquatic surface respiration (ASR) in order to survive. Such responses to low
DO were observed on several occasions during this study.
41
•
The trend of decreasing dissolved oxygen concentration with decreasing flow continued during the
falling limb of the hydrograph. There was still some oxygen production in the near-surface layer but
afternoon dissolved oxygen concentrations were generally lower than they were during the event peak.
This appears to be due to increased colour (leading to light attenuation and reduced photosynthesis).
•
Early mornings during the falling stages of the hydrograph are by far the most stressful times for
oxygen-dependent aquatic organisms and this is when stressed or dying animals are most commonly
observed rising to the water surface. Our experiments indicate that in fresh waters, dead fish initially
sink and may take more than 12 to 24 hours to float to the surface, so the time of appearance of dead
fish is a poor indicator of when they died. Note that 80% of the 140 morning readings taken on the
falling limb of the hydrograph were below 5 %Saturation.
•
The lagoon was generally thermally stratified during daylight hours (i.e., the surface water layer was
warmer and more buoyant than the water beneath it, so the two layers did not mix). However, on some
occasions, especially during winter, the surface layer cooled sufficiently overnight for some mixing to
occur in the mornings. Mixing can introduce oxygenated surface water into the bottom layers but at the
same time carries poorly oxygenated bottom waters to the surface. This effect is evident in the falling
hydrograph data shown in Figure C1-1, with bottom oxygen levels tending to be higher in the mornings
than they were in the afternoon and near-surface morning values being lower than would otherwise be
expected.
•
During periods of stable baseflow, morning oxygen values remained very low but daytime oxygen
production was significantly higher, with dissolved oxygen concentrations greater than 10 %Saturation
being achieved 75% of the time during the afternoon. This could have provided some relief to
organisms capable of utilising the near-surface layer of the water column. It was evident in our
experiments (see appendix D) that juvenile barramundi and rainbow fish sleep during the night and
appear to require less oxygen at such times. It is likely that many species exhibit this characteristic.
Moreover, water temperatures fall during the night and this significantly reduces the metabolic oxygen
requirements of all cold-blooded animals. Low oxygen concentrations are therefore most stressful
when they occur during hot daylight hours).
•
During periods of stagnation, oxygen concentrations were far more variable but on at least 75% of
occasions were significantly higher than at other stages of the hydrograph. The higher concentrations
can be attributed mainly to photosynthetic production by submergent plants (in this case mainly
Ceratophyllum and its associated bio-film). Vertical stratification was present most of the time (i.e.,
there was a distinct oxygenated surface layer) even during winter, although the somewhat elevated
morning concentrations in the bottom layer indicate that there was some overnight mixing.
•
The high variability in dissolved oxygen concentrations in the surface layer was due to a variety of both
spatial and temporal factors. On individual trips, concentrations were highest within dense submergent
macrophyte beds and lowest in sections of the lagoon that were becoming overgrown with water
hyacinth. Between-trip variations were due to fluctuations in weather (levels were lower on cloudy
days than sunny) and water quality (high concentrations of BOD were occasionally present).
42
C1.3
Responses of Fish to Changes in Water Quality at Lagoon Creek
A fish kill was observed at Lagoon Creek during ambient monitoring in November 1999, six days after the
first significant rainfall event of the 1999-2000 wet season. Approximately 200 bony bream (Nemalatosa
erebi; 30-40 cm total length TL), six barramundi (Lates calcarifer; 25-90 cm TL), one eel-tailed catfish
(species unknown; 50 cm TL) and several thousand shrimp (Macrobrachium idae, M. latidactylus, M.
tolmerum; 2-8 cm TL) were observed floating dead on the water surface. In addition, many barramundi (3040 cm TL) and an eel-tailed catfish (species unknown; 25 cm TL) were seen performing aquatic surface
respiration (ASR), while several hundred shrimp were either “jumping” at the water surface or crawling out
onto emergent plants. Near-surface DO levels during this event were extremely low (Figure C1-2). Based
on known thresholds (see experimental results in Appendix d) it is probable that DO at this time would have
been sufficiently low to have caused the kill, regardless of any other water quality factor(s). High ammonia
concentrations (up to 1166 µgN/L) were also recorded during the fish kill. However, the low pH values at
this time (6.14 - 6.74) would have prevented the majority of this ammonia from occurring in the most toxic
(un-ionised) form.
Netting surveys conducted by the Hinchinbrook Landcare Group indicated that hypoxia-intolerant species
were present in the lagoon during the 2000 dry season. For example, both barramundi and bony bream were
caught during September 2000. Our monitoring at this time indicated that DO levels had improved
significantly since the previous wet season (Figure C1-2). No behavioural responses such as ASR were
observed during this time.
Within 24 hours of the first rainfall event of the 2000-01 wet season (November 2000), barramundi (30-45
cm total length) and bony bream (20-30 cm total length) were observed performing ASR but no dead fish
were seen. At the time, near-surface DO levels were very low during both the morning and afternoon,
although not as low as levels recorded during the 1999 fish kill (Figure C1-2). Several barramundi
performing ASR were collected and placed in stream-side tanks filled with Lagoon Creek water that was
constantly aerated (DO levels reaching as high as 62 %Saturation). All fish ceased performing ASR when
placed in the aerated Lagoon Creek water and showed no apparent distress behaviour. In one tank, the
aerator failed and DO levels fell to 7 %Saturation. The barramundi in this tank were observed performing
ASR, but ceased this behaviour once DO levels were raised to 9 %Saturation and higher. These observations
provide strong circumstantial evidence that DO was the main water quality stressor causing ASR behaviour
during the November 2002 event.
Although DO levels were clearly stressful during the 2000 event, they appear to have been just above lethal
threshold values. It is also possible that fish did die at our monitoring site during the 2000 event, but were
not observed because they sank and were washed downstream. However, the existence of a tidal barrier
immediately downstream of our site that impedes water flow, and the fact that not a single dead fish was
seen during intensive monitoring in the two-week period following the storm event, suggests this is unlikely.
It should also be noted that fish performing ASR during the 1999 and 2000 events were easily caught by
hand net. This would not be possible under less stressful water quality conditions, and highlights the
vulnerability of fish to predators (and other factors such as sunburn) while performing ASR.
Our experiments indicate a sharp transition from lethal to sublethal effects with respect to DO. From these
findings, we would predict that differences in concentration of only a few %Saturation would make the
difference between life and death for sensitive species. The observations at Lagoon Creek strongly support
this contention, with DO concentrations during the 1999 fish kill being only a few %Saturation lower than
the levels that the same fish species (and size classes) were able to survive in 2000. This illustrates that
management need only achieve relatively minor improvements in oxygen status to gain substantial potential
environmental benefits.
Very low DO levels were also recorded during and after the February 2002 storm event (Figure C1-2).
Based on observations in 1999 and 2000, responses such as ASR and mortality would have been expected
from hypoxia-intolerant species. However, no fish were observed performing ASR, and daily searches along
a 1 km stretch of the creek during the two-week period following the event failed to find any dead fish. Fish
surveys conducted using a boat-mounted electro-fisher indicated that the only fish present were tarpon and
empire gudgeons, both of which can tolerate very low DO levels. Thus, it appears that no hypoxia-intolerant
43
species were present in the creek at the time of the event. This highlights the fact that the absence of
behavioural stress response or mortality following runoff events (or at any other time) does not necessarily
indicate that water quality at that time is adequate for the survival of sensitive species – such species may be
absent from the water body. The absence of hypoxia-intolerant species in February 2002 is surprising given
that DO levels recorded during the preceding dry season should have been sufficient to ensure their survival
(Figure C1-2). It is possible that a mass mortality event occurred during this time but was not detected.
Unfortunately, heavy commitments to our experimental program precluded more intensive monitoring at
Lagoon Creek during this period.
In summary, the main water quality parameter associated with fish stress or fish kills at Lagoon Creek
throughout this study was low DO.
44
Figure C1 – 1:
Variation in dissolved oxygen concentrations with depth and time of day
DEPTH <1M BELOW SURFACE
DEPTH >1M<1.5M BELOW SURFACE
DEPTH >1M BELOW SURFACE
Afternoon
Afternoon
Afternoon
Morning
Morning
Morning
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
45
Figure C1 – 2: Temporal variations in the concentrations of dissolved oxygen in the near-surface
layer of the water column at the Lagoon Creek study site
TIME OF DAY
Afternoon
(E) – Storm Event
(F) – Falling Hydrograph
(B) – Baseflow
(N) – No Visible Surface Flow
Morning
Minimum value that
Is not an outlier
25th Percentile
Median
Maximum value that
Is not an outlier
75th Percentile
Extreme Value
Outlier
46
C2
LAGOON CREEK STORMEVENT
C2.1
Equipment and Sampling Protocols
ISCO-3700 automatic samplers equipped with conditioned polyethylene sample hoses and coarse solid
strainers were deployed at the lagoon creek study each site. These were fitted with level actuators
(constructed by the Electronics Section of the Advanced Analytical Centre at James Cook University) which
automatically initiated a pre-programmed sampling sequence whenever water levels in the stream exceeded a
pre-determined height.
The autosamplers were not refrigerated but had a central well, which held ice to keep samples cool. The
cylindrical ice well was surrounded by 24 one-litre polyethylene sample bottles. Preliminary tests indicated
that in summer conditions, ice in the well did not last for 24 hours and did not cool samples to less than 5oC.
However, it was discovered that, provided that the autosampler was not exposed to direct sunlight, samples
could be maintained below 5oC for 24 hours by using every third sample bottle as a cold pack to supplement
the ice in the well. This approach was adopted for the duration of the project.
The autosamplers were installed on high banks in quite open locations. It was therefore necessary to
construct sturdy housings capable of withstanding cyclonic winds or partial immersion during floods,
providing adequate shade to ensure the instrument and samples did not overheat during summer heat waves,
and providing adequate protection of personnel and samples while the sampler was being serviced during
storms. The equipment was also vulnerable to human interference so steps were taken to ensure that the
installation was as secure as possible and that it was impossible to tamper with samples without leaving
evidence.
Since the autosamplers were not refrigerated it was imperative that samples were iced down and/or removed
and properly preserved as soon as possible after collection. To assist in this regard the Lagoon Creek
autosampler was fitted with a mobile phone which would call laboratory staff whenever sampling was
initiated. As a back-up, local landholders also volunteered to phone if it started raining. At least one
laboratory staff member was on call at all times to ensure that samplers were attended to as soon as possible
after the commencement of an event, and every 24 hours thereafter to remove samples and replace ice.
Lagoon Creek is often poorly mixed (especially during the falling limb of storm hydrographs) and there was
a high probability that a significant proportion of the flow would be confined to the near-surface layer of the
water column. It was therefore necessary to employ a more complex installation to ensure that samples were
drawn from 30cm below the surface, regardless of water depth. This was accomplished by mounting the
sample hose inside a flexible polyethylene conduit, the end of which was attached to floats.
The autosamplers were installed during the 2001 dry season (while water levels were low) in order to
monitor any storm events that occurred during the 2001/2002 wet season and 2002 dry season. This proved
to be one of the driest periods on record for the Herbert floodplain, and during the twelve month monitoring
program there was only one rain event that generated sufficient runoff to induce stormwater flows.
C2.2
Results
Rainfall associated with the monitored event commenced on February 15, 2002, and peaked the next day
with Ingham registering 231mm for the 24 hours up to 9:00 a.m. on February 16. The rain event lasted just
under four days with a total precipitation of 434mm being recorded at Ingham. By wet-dry tropics standards
this was a moderate storm event and very little flooding was evident in the area. Because the event occurred
during an atypically dry period (monthly rainfall totals over the preceding six months ranged from just 1 to
98mm), the reported monitoring results are unlikely to be truly indicative of normal stormwater quality.
The dry conditions maintained over the twelve months of our study clearly affected catchment runoff
coefficients. For example, 160mm of rainfall was registered during April but there was no detectable surface
runoff into Lagoon Creek. By contrast, an overnight rain event of only 40mm which occurred on March 5
while catchment soils were still wet from February rains generated sufficient runoff to increase stream flows
47
noticeably. (The sampling event triggered by the February rainfall was still in progress on March 5, so the
small spate that occurred at that time has been treated as part of the main flow event).
The autosampler functioned satisfactorily during the course of the event, yielding 16 samples per day at
regular pre-programmed time intervals over the peak of the hydrograph. Sampling intensities were reduced
when flow conditions stabilised but, nevertheless, the numbers of samples collected far exceeded the
budgeted allocation for analyses. Accordingly, only selected samples were subjected to comprehensive
analyses. Most samples were analysed for pH, conductivity, colour and turbidity, and samples were chosen
for further analysis if one or more of these parameters exhibited a significant change. The results of these
preliminary analyses were plotted against the hydrograph, and additional samples were selected for analysis
in cases where concentration changes did not match hydrographic variations and/or where the difference
between consecutive data points was large.
Flows at Lagoon Creek are subject to complex natural regulatory mechanisms and this necessitates the
development of empirical stage discharge relationships. These curves have not yet been developed because
there has been insufficient rainfall since sampling commenced, so it is not possible to estimate discharge
rates from relative water levels. However, if stage discharge curves are developed in the future it will be
possible to retrospectively calculate discharge rates for the February 2002 storm event. It is important to
recognise that the relationship between water level and discharge rate is not linear. Water levels in a lagoon
can rise significantly before there is any substantial flow, while further increases in water level can result in
exponential increases in discharge rate. Accordingly, the water levels plotted in Figures 1a to 1f provide
only a very coarse indication of the shape of the event hydrograph and should be interpreted cautiously. It is
particularly important to note that the plot almost certainly under-estimates the relative height of the flow
peak. In fact, it is likely that discharge rates during the three days of peak flow would have been in the order
of ten times higher than at other stages of the event.
Figure 1a and 1b
•
There was a striking rise in total nitrogen concentrations up to levels in excess of 6,000µg N/L during
the rising stages of the event. High levels were maintained for 4 or 5 days with concentrations
decreasing gradually, roughly in proportion to flow, before stabilising at above 1000µg N/L during the
transition to base flow.
•
Concentrations of nitrogen, SPM and turbidity increased rapidly during the rising stage of the
hydrograph and reached a maximum at the peak of the hydrograph. SPM and turbidity concentration
fell rapidly within a day or so (soon after rainfall intensity decreased). In contrast, nitrogen
concentrations remained elevated for two more days before significant declines occurred. The
sustained nitrogen levels were attributable to an increase in nitrate concentrations during the early
falling stages of the hydrograph. The timing of concentration peaks of SPM and nitrate relative to daily
rainfall is plotted in Figure 1.3-1f.
•
Concentrations of SPM were very moderate for a storm event and did not exceed 110mg/L even at the
peak of the hydrograph. The samples collected during this event were, in fact, visibly less turbid than
runoff observed (but not sampled) during early wet season rain events in the late 1990’s. This issue is
discussed later in the context of particulate phosphorus export.
•
The nitrogen exported during the rising/peak stages of the hydrograph comprised approximately equal
proportions of particulate nitrogen and nitrate. However, more than 80% of the nitrogen exported
during the initial falling stages of the event was in the form of nitrate.
•
Flows during the early stages of the event can be attributed mainly to surface runoff. There are few
significant off-stream storages of surface water in the Lagoon Creek catchment so once rainfall ceases,
flows drop off rapidly and subsurface inputs are responsible for an increasingly higher proportion of
the flow. Initially this runoff would be expected to originate mainly from shallow soil layers, with
contributions from deeper groundwater increasing gradually until it became the major source of the
elevated base flow remaining in the aftermath of the storm event. The appearance of very high
concentrations of nitrate only during the early falling stage implicates soil-water through-flow as a
48
major source of nitrate export. Moreover, sudden declines in concentrations during the later stages of
the event suggest that deeper groundwater was not a significant source of nitrate on this occasion.
•
The concentrations of nitrogen, and particularly nitrate, reported during this event were extremely high
compared to the concentrations normally observed in regional waters during periods of base flow. In
the case of nitrate, they are one hundred times higher than the levels that would normally be expected
in storm water originating from rainforest catchments, and approximately ten times higher than the
concentrations that are normally reported during high flow events in inland rivers with catchments
dominated by grazing.
•
The Great Barrier Reef Marine Park Authority have identified riverine nitrate export as one of the
major existing catchment-based threats to the integrity of the Great Barrier Reef. From this
perspective, the high nitrate concentrations observed in this event are clearly cause for concern.
•
However, the significance of nitrate in freshwater ecosystems is much more difficult to judge. Nitrate
is not particularly toxic in freshwater; high concentrations adversely affect receiving ecosystems
mainly by stimulating excessive growth of plants and microbes. The principal cause for concern is
usually excessive growth of phytoplankton leading, for example, to algal blooms. However, this
happens only when nitrogen is the main factor limiting plant growth. In this instance, it is very likely
that nitrate was available in such an over-abundance and for such a short time (during which cloudy
weather conditions, turbidity and discoloured water would almost certainly have ensured that plant
growth was limited by light availability) that such effects are unlikely to have occurred in the creek.
However, this does not mean that the ecosystem was unaffected. Pearson and Connolly (2000) have
demonstrated that high levels of nitrate stimulate microbial decomposition in local rainforest streams.
There is also broad international consensus amongst wetland scientists that microbial biomass increases
in direct proportion to inputs of nitrate, provided that adequate organic carbon is present. Moreover,
nitrate in the presence of low oxygen concentration (as observed during this event) would be expected
to stimulate the growth of denitrifying bacteria. These are facultative anaerobes (they breathe oxygen
when it is present but use nitrate as an alternative source for respiration if oxygen is not freely
available), so the development of significant populations would be expected to hinder the recovery of
oxygen levels in the aftermath of the storm event. Potential impacts of this kind have not yet been
studied in tropical waters, but until there is evidence to the contrary it would be prudent to assume that
such processes are potential contributors to the chronic hypoxia problems that are regularly observed in
the waters of cane-growing areas.
Figure 1c
•
During this event, ammonia constituted only a minor proportion of the total nitrogen export. It is likely
that this is typical of events which occur after prolonged periods of dry weather during which urea and
other ammonia-based fertilisers would be oxidised to nitrate in the soil. However, during wet periods,
soils can become waterlogged (particularly in low lying floodplain areas) and it is possible for soils to
become severely hypoxic. Under such conditions, the conversion of ammonia to nitrate (which
requires oxygen) should be inhibited, and it is likely that ammonia would constitute a much larger
proportion of the nitrogen carried away in runoff, as was the case in the storm-water samples from the
wetter period during the early months of this study (see section 1.2).
•
The initial sharp peak in ammonia concentration during the rising hydrograph can almost certainly be
attributed to catchment inputs; however, the broader peak which developed on the falling hydrograph is
most likely due to instream production through ammonification (microbial decomposition of organic
nitrogen to form ammonia). Oxygen levels in the creek were too low for this ammonia to be oxidised
to nitrate, hence it accumulated. A second ammonia peak occurred about a week later. This coincided
with the cessation of diel oxygen cycling and establishment of constantly low oxygen values, again
indicating the involvement of in-stream generation processes.
•
Unlike nitrate, high concentrations of ammonia are toxic to aquatic organisms. The concentrations of
total ammonia encountered during this event approached the toxicological thresholds reported for free
ammonia in section 5 of this report. However, the low pH of the water would have ensured that most
49
of this was in the (much less toxic) ionised form, hence there was a low risk of direct toxicological
harm from ammonia.
•
Total phosphorus concentrations during the peak of the event (less than 400µg/L) were moderate. As
was the case with nitrogen there was a pulse of particulate P during the rising stage, followed by an
increase in dissolved inorganic forms (FRP) on the falling limb of the hydrograph, with the
contribution from FRP rising from <20% to a maximum of 75% before falling off to only 10% at the
bottom of the hydrograph. Despite this variability in speciation, total phosphorous levels remained
remarkably constant for more than a month during and after the event peak. This can be attributed to a
gradual rise in organic phosphorus concentrations, which compensated for the decline in FRP levels.
These patterns appear to be reasonably representative of what can be expected of storm events that
occur in the mid to late wet season when most fields are planted and ratoon crops are at a reasonably
advanced stage of regrowth. At such times the runoff from fields is very obviously less turbid than it
can be early in the wet season when many fields are bare and/or under cultivation. During the very
early stages of this project, prior to the commencement of sampling, we observed several instances
where very turbid water was discharging from cane drains. No early wet season rain events occurred
after our event sampling program was instigated, so events of that kind have not yet been sampled.
However, based on our observations and given the well established correlations between turbidity,
SPM and particulate phosphorus concentrations, it is very likely that such sampling would demonstrate
that particulate phosphorus concentrations are significantly higher when events occur during, or soon
after, the harvest.
Figure 1d
•
Colour was almost certainly due to the presence of filterable organic matter (based on hue and the
absence of significant concentrations of inorganic colorants such as ferrous iron) but was more closely
correlated to phosphorus than it was to TOC or DOC. This suggests that concentrations of colourless
organic matter fluctuated somewhat during the course of the event, and that organic phosphorus was
associated mainly with coloured organic constituents.
Figure 1e
•
In cane-growing areas, BOD concentrations can be expected to be highest when storm events occur
during the harvesting season (when there is cane juice on the ground) and/or during prolonged wet
spells (when anaerobic decomposition of organic matter in saturated soils can generate oxygen
demanding by-products). The February 2002 event occurred months after harvesting and marked the
end of a prolonged summer dry spell. This probably represents a best case scenario for BOD export so
it was not surprising to find that no extreme concentrations occurred during the event, the maximum
value reported being 7mg/L. Nonetheless, concentrations were still ecologically significant and were
maintained over the entire course of the 40-day sampling event, the minimum concentration reported
being 2.2mg/L.
•
The continuous curve labelled “Bottom DO” is a plot of the dissolved oxygen data obtained from a
CSIRO Land and Water data logger (Hydrolab) that had already been installed at the site prior to the
commencement of this study. This instrument was deployed at a fixed depth in the water column and,
prior to the event, was located in the hypolimnion (a stable layer of bottom water which does not mix
with oxygenated surface water layer and therefore tends to become anoxic). Consequently, DO values
close to zero were being recorded at that time. The inflow of storm water caused some initial mixing of
the water column, bringing some oxygenated surface water down to the depth of the sensor. However,
during the rising stages of the event, oxygen losses greatly exceeded inputs, and oxygen fell to less than
10 %Saturation levels through the entire water column.
•
Oxygen losses result from the combined effects of BOD (microbial oxidation of organic matter
suspended or dissolved in the water column), benthic respiration (consumption of oxygen by microbes,
plants and animals living on or in the bottom sediments), and transportation downstream. The main
sources of oxygen are inflowing water, re-aeration (from the overlying air) and, during daylight hours
only, photosynthetic production of oxygen by submergent plants.
50
•
The turbulent flows generated by storm events significantly increase the rate at which oxygen from the
air is delivered into, and mixed through the water column. Moreover, the rate at which oxygen from
the air enters the water increases tremendously (logarithmically) when the DO concentrations at the
water surface decrease. Accordingly, re-aeration rates at the Lagoon Creek site must have been
relatively high during the rising/peak stages of the hydrograph. Since DO concentrations did not rise
above 10 %Saturation (and were considerably lower in most cases), oxygen consumption rates must
have been even higher. This appears to be the result of several interacting factors:
1) BOD concentrations were undoubtedly a contributing factor, the minimum DO levels coinciding
with the BOD peak of 7mg/L.
2) The reaches of the creek upstream of the study lagoon were heavily infested with water hyacinth
and the water under these weed mats is known to be severely hypoxic. Through flow of this water
would undoubtedly have placed demands on the system’s re-aeration capacity, especially during the
very early and late stages of the event.
3) The combined effects of cloudy weather, turbid water and a sudden increase in water depth would
have severely inhibited the capacity of submerged plants to produce oxygen via photosynthesis.
Such effects have been examined experimentally in this study and are discussed in subsequent
sections of this report. It is noteworthy that the maximum oxygen values observed during the event
coincided with the sudden fall in turbidity levels during the early falling stages of the event.
The processes discussed above occur over quite small scales of time and space and, in many cases,
make only subtle contributions to the overall picture. However, from the experimental work reported
elsewhere in this report, it is clear that aquatic organisms can asphyxiate within minutes if they are
exposed to DO concentrations below their survival threshold; and that asphyxiation thresholds for all of
the species tested to date are so sharp that survival probability can be reduced from 100% to 0% by a
DO concentration change of only 1-2 %Saturation. This means that a single, very brief oxygen sag can
have overwhelming effects on the ecosystem, sometimes causing such severe damage that it no longer
matters what happened in the longer-term. The DO concentrations encountered during and
immediately after this event were low enough to have caused a fish kill if DO-dependent fish species
had been present prior to the commencement of the event. In this case, however, the lagoon was in
such poor condition leading up to the event that no such species were present.
•
On the early falling limb of the hydrograph, turbidity and BOD concentrations dropped and DO
levels rose to their maximum value of 5 %Saturation. Concentrations in the bottom layer then fell in
direct proportion to flow (due to reduced re-aeration and mixing) until base flows were re-established.
By this time, a stable hypolimnion (bottom layer) had formed and DO levels remained close to zero.
•
Manual spot measurements taken 30cm below the water surface at the routine monitoring site
(“Surface DO”) and at the upstream end of the lagoon (“DO @ top of lagoon”) show that there was
significant diel cycling of DO throughout the surface layer of the water column during the falling limb
of the hydrograph. During this time, DO levels rose to as high as 25 %Saturation in the afternoons but
fell overnight. This is indicative of photosynthetic oxygen production by phytoplankton and
submerged plants. Cycling remained evident in various places in the lagoon for a few weeks until
water hyacinth gradually covered the water surface, shading out submerged plants (and preventing
them from photosynthesising). Regardless of these changes, minimum daily DO concentration in the
surface layer never rose above 5 %Saturation during the 30 days of base-flow monitoring.
•
Cycling of this kind in the surface water layer can be crucial to the survival of some organisms,
especially DO-dependent fish. Observations conducted during hypoxia tolerance testing (section 3)
indicate that fish which sleep during the night substantially reduce their oxygen requirements. These
fish appear to be pre-adapted to cycling oxygen levels, sleeping when DO concentrations are lowest
and becoming active during the day when they are highest. Our experimental observations suggest that
the cycling observed during the falling limb of the event under discussion would have been adequate to
ensure the survival of many oxygen-dependent fish (if they had been present).
51
•
The results leave no doubt that DO availability is by far the most important water-quality-related
factor governing the ecological condition of Lagoon Creek.
52
Figure 1a to 1f:
1a:
Lagoon Creek storm event, February 2002. Plots show temporal variations
in relative water depth and contaminant concentrations during and
immediately after the event
Total nitrogen, nitrate, suspended particular matter (x40) and turbidity (x40) over the full
sampling period (41 days)
7000
1800
6000
1200
4000
3000
600
2000
1000
0
13/02/2002
0:00
0
18/02/2002
0:00
Total N (µg/L)
23/02/2002
0:00
28/02/2002
0:00
Nitrate (µg N/L)
05/03/2002
0:00
10/03/2002
0:00
SPM x40 (mg/L)
15/03/2002
0:00
20/03/2002
0:00
Turbidity x40(NTU)
25/03/2002
0:00
Water Level
53
Relative Water Level (mm)
Concentration (see legend)
5000
1b:
Total nitrogen, nitrate, suspended particular matter (x40) and turbidity (x40) during the first
8 days of the flow event
7000
1800
1600
6000
1400
1200
4000
1000
800
3000
600
5000
2000
400
1000
200
0
0
15/02/2002 0:00 16/02/2002 0:00 17/02/2002 0:00 18/02/2002 0:00 19/02/2002 0:00 20/02/2002 0:00 21/02/2002 0:00 22/02/2002 0:00 23/02/2002 0:00
Total N (µg/L)
1c:
Nitrate (µg N/L)
SPM x40 (mg/L)
Turbidity x40(NTU)
Water Level
Total phosphorus, filterable reactive phosphorus (FRP), ammonia and colour over the full
sampling period
600
1800
400
1200
300
200
600
500
100
0
13/02/2002
0:00
0
18/02/2002
0:00
Total P (µg/L)
23/02/2002
0:00
28/02/2002
0:00
FRP (µg P/L)
05/03/2002
0:00
10/03/2002
0:00
Ammonia (µg N/L)
15/03/2002
0:00
20/03/2002
0:00
Colour (TCU)
25/03/2002
0:00
Water Level
54
1d:
Total organic carbon (TOC), dissolved organic carbon (DOC) and colour(÷6) over the full
1800
35
1600
30
25
1200
20
1000
800
15
600
1400
10
400
5
200
0
0
13/02/2002
0:00
18/02/2002
0:00
23/02/2002
0:00
TOC (mg/L)
28/02/2002
0:00
05/03/2002
0:00
10/03/2002
0:00
DOC (mg/L)
15/03/2002
0:00
20/03/2002
0:00
Colour (TCU/6)
25/03/2002
0:00
Water Level
sampling period
1e:
Biochemical oxygen demand (BOD) and dissolved oxygen (DO) over the full sampling period
DO values include measurements taken by a data logger deployed at a fixed depth in the bottom water
layer adjacent to the autosampler (Bottom DO), spot measurements taken in the near-surface water layer
adjacent to the autosampler (Surface DO) and spot measurements taken in the near-surface water layer at
the upstream end of the lagoon (surface DO at top of lagoon)
30
1800
20
1200
15
10
600
25
5
0
-5
13/02/2002
0:00
BOD
0
18/02/2002
0:00
23/02/2002
0:00
28/02/2002
0:00
Bottom DO (% Sat)
05/03/2002
0:00
Surface DO (%Sat)
10/03/2002
0:00
15/03/2002
0:00
20/03/2002
0:00
Surface DO @top of lagoon (%sat)
25/03/2002
0:00
Water Level
55
1f:
Concentrations of nitrate and SPM in relation to daily rainfall over the full sampling period.
1800
350
1600
300
1400
200
1000
800
150
600
Rainfall and Concentraion
250
1200
100
400
50
200
0
15/2/02 0:00
17/2/02 0:00
Water Level
19/2/02 0:00
Daily Rainfall @Ingham (mm)
21/2/02 0:00
Nitrate ÷20 (µg N/L)
0
23/2/02 0:00
SPM x3 (mg/L)
56
APPENDIX D
EXPERIMENTS
57
D1
D1.1
EFFECTS OF REDUCED CLARITY DUE TO SUSPENDED PARTICULATE
MATTER ON THE UNDERWATER LIGHT CLIMATE OF AQUATIC SYSTEMS
Aim
We aimed to determine empirical relationships between the concentration of suspended particulate matter
(SPM) and the amounts of light available to support the growth of aquatic plants at different depths in the
water column of freshwater wetlands.
D1.2
Methods
This experiment was conducted in two main stages. The first step was to conduct laboratory experiments in a
specially built light attenuation column to enable us to test a wide variety of different water clarities very
quickly. The second step was field calibration which entailed comparing in-situ field measurements of light
attenuation to measurements obtained in our laboratory light attenuation column, in order to derive validated
estimates of the numerical relationship between SPM concentrations and light attenuation in natural systems.
Sample testing
The experiment was conducted using a specially constructed light attenuation column which was designed to
measure the intensity of light after it has passed through the overlying column of water. This was used to
examine the effects of water depth on light attenuation by simply varying the quantity of water in the
column. It was designed to allow us to use natural sunlight (using a flat reflecting mirror to direct parallel
light beams down the column) as the incident light source and to simulate natural water column conditions. It
consisted of a 150 mm diameter by 1 metre long PVC pipe which had a circular glass sheet sealed into its
base (Figure D1–1). This column slotted into a base fitting that housed a LI-COR Quantum sensor (Model
No. LI-192SB) as shown in Figure D1–2.
To determine the light attenuation capacities of water with different SPM concentrations, specifically
prepared water samples were progressively placed into the column and a large flat mirror used to focus
parallel sunlight beams directly down the column and into the sample. The sample water would scatter the
incident light and a proportion would reach the light sensor located at the base of the column. The received
light intensity was recorded as Photosynthetically Active Radiation (PAR) in units of micro-Einsteins per
metre squared per second.
Sample preparation involved the collection of erodable soil material from drains in the Herbert River
floodplain area. Nine soil samples were collected, ranging from light clays to dark organic soils (see Table
D1-1). Each of these soils was analysed for moisture content, organic/inorganic composition and the less
than 63 micron size fractions. Preparation of the stock water suspensions included the following steps:
1) Approximately 500 g of each of the soil samples was soaked in R-O water overnight and
aggregations broken down by manual agitation.
2) The saturated samples were then transferred to a 15 L settling column filled with R-O water,
vigorously agitated and then allowed to settle for 24 hours.
3) The supernatant suspension was decanted and collected (precipitates were discarded).
4) This stock suspension was subjected to serial dilution in order to prepare a series of test suspensions
(total of 5 different SPM concentrations) with concentrations ranging from the maximum obtained in
the stock suspension (see Table D1-2) down to less than 10 mg/L.
Each of the test suspensions prepared (A-I) was then tested in the light attenuation column as follows:
1) The column was thoroughly cleaned.
2) The incident PAR light intensity was determined by directing sunlight down the empty column using
the flat reflective mirror.
3) Aliquots of each test suspension were incrementally added to the column in order to progressively
increase the depth of suspension in the column. After each addition, sunlight was directed down the
column and the PAR light intensity at the bottom of the column and the water column depth were
recorded. Whenever possible, additions continued until the light at the bottom was less than 1% of
the incident light intensity (i.e., until the euphotic depth was exceeded).
58
Accordingly for each soil sample, the experiment yielded data indicative of the changes in PAR light
intensity with water depth over a range of SPM concentrations. These data were used to calculate extinction
coefficients and corresponding euphotic depths.
Table D1-1 Properties of 9 soil samples collected for experimental light attenuation tests.
Sample Moisture
(%)
A
7.5
B
4.3
C
8.6
D
46.5
E
47.2
F
1.7
G
5.3
H
4.5
I
0.7
Table D1-2
Organic
Composition
(%)
3.3
6.1
6.5
11.3
10.4
3.8
4.7
5.8
2.0
Inorganic
Composition
(%)
96.7
93.9
96.5
88.7
89.6
96.2
95.3
94.2
98.0
Less than
63µm Fraction
(%)
25.3
66.9
94.2
93.4
82.9
56.9
48.2
98.5
19.3
Greater than
63µm Fraction
(%)
74.7
33.1
5.8
6.6
17.1
43.1
51.8
1.5
80.7
Properties of 9 stock water samples prepared for experimental light attenuation tests.
Sample Colour
TCU
A
10
B
20
C
10
D
2
E
20
F
15
G
10
H
600
I
25
Turbidity
NTU
100
924
922
295
354
514
1120
5445
337
Filterable
Turbidity
(GF/C)
NTU
21.3
58.7
19.1
12.0
25.9
15.4
31.6
1200
24.9
TSS
mg/L
72
544
434
204
267
277
629
2790
205
TDS
mg/L
51.1
27.7
26.3
47.1
32.7
28.5
23.3
107.3
48.5
Organic SPM
Composition
(%)
23
13.2
10.4
8.3
12.9
19.9
14.1
18.5
17.9
Inorganic
SPM
Composition
(%)
77
86.8
89.6
91.7
87.1
80.1
85.9
81.5
82.1
Calibration
The attenuation characteristics of the test column used in these experiments are proportional to but not the
same as those of natural waters. To correct the experimental results for natural conditions it was necessary to
measure the light attenuation characteristics of a number of different natural systems. Different types of SPM
occur in receiving waters at different times and places depending on a range of factors and these variations
have to be incorporated into the calibration process.
Measurements of SPM concentrations and turbidities (as well as secchi depths) were taken at several local
water bodies with different clarities, and the in situ light attenuation properties determined. Composite
samples of the water were also taken and tested in the laboratory column to determine the light attenuation
correction factor. It was then possible to correct the experimental values for the “real world’. Limitations of
this calibration are discussed in Section D1.5.
D1.3
Results
Figure D1-3 shows the soil samples used to produce the SPM stock solutions (top row) and the filter disks
from the TSS analysis (middle row). Even though these soils were quite different in composition and looked
very different, the fines which remained suspended in our test column were surprisingly similar in
59
appearance. When the organic fraction of the fines was removed from the filter paper using an oxidizing
furnace (set at 600 oC) the residue was even more similar in appearance (bottom row). This may be
attributable to the fact that all samples were collected from the Herbert River floodplain; that is, it is likely
that the fine fraction of all floodplain soils were historically derived from the same source – the Herbert
River. Most of the observed differences are due to variations in particle size distribution (a function of
localized hydro dynamic factors) and autochthonous organic matter.
Figure D1-4 shows an example of how one particular soil (Sample D) appeared before the preparation of the
SPM solutions, after emersion followed by 24 hours of settling and the corresponding texture of the
suspended material after filtration (filter disk). This figure displays the similarity between the colour of the
SPM in the settling column and the colour of the SPM contained on the filter disk (Figure D1-3) and the
stark difference between the colours of the soil sample (left) and the SPM (right).
The light attenuation curves for all nine soil samples were obtained (soil sample D is shown as an example in
Figure D1-5) and the experimental light extinction curves were derived using the average value obtained at
each successive depth for all nine soil samples (Figure D1-6). The experimental euphotic depth (depth at
which 1% of the incident light intensity is received) was also calculated and is shown in Figure D1-7.
So far only seven sites have been examined (Table D4-3), all located in the Burdekin catchment because
freshwater sites with open water surfaces (i.e., not weed infested) and sufficient inorganic turbidity and depth
to perform the tests are rare on the Herbert floodplain during the dry season. The water clarity associated
parameters analysed during the field testing phase and the actual values obtained are shown in Table D1-3.
The results of the in situ light attenuation study for the three Burdekin floodplain field sites (i.e., sites 1, 2 &
3) are shown in Figure D1-8. The laboratory column measurements obtained from water samples collected at
these field sites are shown in Figures D1-9.
From the results of the field and water sample tested light attenuations, we derived a light extinction
calibration curve for the sites tested (see Figure D1-10). Note that the light extinction values in the higher
end of the scale (> 30 m-1) are estimates based on the theory that at higher turbidities (< 0.15 m euphotic
depth), which is the diameter of the experimental column, the experimental light attenuation would perform
similarly to that of the field sites. The gap in the calibration curve (between field k= 1.71 m-1 and field k=30
m-1) will be completed once suitable field sites are found that have clarity values within this range.
Table D1-3
Parameters measured at each calibration test site.
Secchi Depth
Suspended Solids
Turbidity Concentration
(m)
(NTU)
Calibration Test Sites
Location
1. Clare Weir
Lower Burdekin River
0.65
18
7.0
2. Payards Lagoon
North Burdekin floodplain
0.90
13
8.0
3. Church's Lagoon
North Burdekin floodplain
7.00
5
4.2
4. Muckibulla Waterhole
Upper Burdekin
0.22
78
85
5. Gorman's Lagoon
Upper Burdekin
0.01
2570
2590
6. Lake Amelia
Upper Burdekin
0.09
364
197
7. Pajingo Waterhole
Upper Burdekin
0.23
73
86
(mg/L)
The curve of best fit for the light extinction coefficient calibration (Figure D1-10) is a power function with
the following equation (R2 = 0.972):
Kfield = 0.5957 Kcolumn1.0167
……....… (1)
Using this calibration equation we obtain the corrected light extinction coefficients and euphotic depths as
shown in Figures D1-11 and D1-12.
D1.4
Discussion
60
Due to the severely dry conditions encountered during our experiments it has proven impossible to field
calibrate our laboratory findings in the Herbert. The interim calibrations presented here are based on
Burdekin data and should be validated on the Herbert floodplain as soon as sufficient rainfall occurs.
Nonetheless the light attenuation column response should not be greatly affected by sediment composition,
so the interim calibration should yield close approximations of the true values.
Even though the nine soil samples displayed very different properties, in terms of their organic contents and
colour, they all fit the general trend shown by Figures D1-11 and D1-12. Using this relationship it has been
possible to derive a curve of best fit that can be used to predict the light attenuation coefficient, for any given
freshwater wetland within the study region (i.e., Herbert River floodplain)(see Figure D1-13). We have used
SPM concentrations to derive these relationships. It is theoretically feasible to derive similar relationships
using other clarity-related parameters such as turbidity, secchi depth, black disc visibility, apparent colour,
etc but these have not been used in this study because it is much more difficult to link such parameters
quantitatively to land management indicators such as soil erosion and sediment delivery rates.
The basis for this work is the exponential light attenuation equation:
IZ = IO e-kd
………………. (2)
Where: IZ = Light intensity at depth z, µEinsteins/m2/s;
IO = Incident light intensity, µEinsteins/m2/s;
k = Light extinction co-efficient, m-1; and
d = Water depth, m.
This equation indicates that light passing through a water column will be scattered and absorbed (attenuated)
at an exponential rate determined by the light extinction co-efficient, k. Each water body has a unique
attenuation characteristic that depends on many factors with just one being the concentration of suspended
solids in the water column. Other factors include colour, phytoplankton and colloidal concentration.
However, SPM concentration can play a dominant role in light attenuation, particularly during and soon after
catchment runoff events.
With accurate calibration within a study area, it is possible to estimate the light attenuating properties of a
water body by determining the SPM concentration and using a calibration equation (similar to Figure D1-13)
to estimate the extinction coefficient. This coefficient can then be used to calculate the euphotic depth for
that system using equation (2). Additionally, the light intensity can be calculated for any depth within that
water body for any given incident light intensity.
D1.5
Summary and Conclusion
The purpose of this experiment was to quantify the effects on light attenuation due to SPM discharged from
farm drains. Soil samples were collected from close proximity to drain invert points and are representative of
the material which would be expected to be discharged from the drains during the next storm event. The
gravel, sand and coarse silt portion of this material is too heavy to remain suspended in the water column
long enough to affect water clarity over biologically relevant time scales. Accordingly, samples were pretreated to collect only particles which (in the absence of flocculation) could remain suspended in a 1 metre
deep water column for more than 24 hours (i.e., medium to fine silt, clay and colloids). This is the portion of
the soil material introduced to the aquatic environment from catchment areas that is responsible for persistent
turbidity. In most water bodies, this material is either gradually flushed from the system by base flows or
slowly flocculates and/or settles during periods of prolonged dry weather. Hence, in many areas, including
the Herbert floodplain, there is seldom much of this material left in the water column by the end of a
prolonged dry spell (some wetlands in the Burdekin are exceptions to the rule because they contain colloids
which are resistant to flocculation and therefore remain turbid even during droughts).
The Herbert district soils tested in these experiments were quite different in colour and composition.
However, the suspended particulate material they liberated was much more similar in appearance and the
light attenuation characteristics of the suspensoids obtained from samples were very similar (see Figure D13). All samples were collected from the floodplain so many of these similarities could be attributable to the
61
fact that most of the fine inorganic soil material ultimately originated from Herbert river. Accordingly it
would be advisable to test soils and waters from other catchment areas before accepting the general
applicability of the relationships obtained in these experiments.
This experiment was conducted during a severe dry spell, so the turbidity in most local water bodies was
either at a minimum or could be attributed to in-stream production of SPM that is very different in
composition from that originating from cane farms. Accordingly it was not possible to properly field
calibrate the results of laboratory experiments. The distributary channels of the Burdekin floodplain, due to
the presence of chronically turbid Burdekin Dam water, are the only waterways in local farming areas that
have maintained high turbidity levels during this dry spell. Even these waters are approaching the clearest
they have been since the Burdekin Dam was commissioned well over a decade ago, so the full range of
turbidities required for field testing could not be found in this area. Moreover, the SPM composition of
Burdekin Dam water is unique so it is possible that field calibrations conducted in the Burdekin floodplain
would not be applicable to any other cane growing area. The only natural water bodies in the local region
that maintained high turbidity levels during the dry spell were inland lagoons which are quite remote from
the floodplain. As an interim measure and in the absence of any alternatives, field calibrations were
conducted in lagoons of this kind. Results provide first order indications of what can be expected in the
floodplain waterways but accurate quantitative measures will only be obtainable when it rains in the study
area.
62
Figure D1 – 1: Picture of the light
attenuation column.
Figure D1 – 2: Picture of the base
piece with light sensor
installed and connected
to the Li-Cor quantam
meter.
Figure D1 – 3: Soil samples and corresponding filter disks (before and after oxidation) of the 9 SPM
water samples.
Soil A
Soil B
Soil C
Soil D
Soil E
Soil F
Soil G
Soil H
Soil I
Soil sample
Suspended
solids
Residue after
oxidation
Figure D1-4: Soil sample D and the
corresponding SPM sample in
the settling column prior to
extraction.
63
Figure D1 – 5: Experimental column light attenuation profiles of SPM water sample D.
0
Water Depth, m
0.2
0.4
0.6
0.8
1
1.2
0
500
1000
1500
2000
Light Intensity, µE/m2/s
204 mg/L
102 mg/L
51 mg/L
25.5 mg/L
12.75 mg/L
Figure D1 – 6: Experimental (uncorrected) light extinction coefficients for the 9 soils and the 5
different suspended solids concentrations for each soil type.
Extinction Coefficient, m
-1
1000
100
10
1
1
10
100
1000
10000
Suspended Solids Concentration, mg/L
Soil A
Soil B
Soil C
Soil D
Soil E
Soil F
Soil G
Soil H
Soil I
64
Figure D1 – 7: Experimental (uncorrected) euphotic depth plots for the 9 soils and the 5 different
suspended solids concentrations for each soil type.
1
Euphotic Depth, m
0.9
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
1
10
100
1000
10000
Soil A
Soil B
Soil C
Soil D
Soil E
Soil F
Soil G
Soil H
Soil I
Figure D1 – 8: Field results obtained during the light attenuation calibration experiment (sites 1,2 &
3).
0
0.5
Depth, m
1
1.5
2
2.5
3
3.5
4
4.5
0
500
1000
1500
2000
2
Light Intensity, µE/m /s
Clare Weir
Payards Lagoon
Church's Lagoon
Log. (Clare Weir)
Log. (Payards Lagoon)
Log. (Church's Lagoon)
65
Figure D1 – 9: Laboratory column results obtained during the light attenuation calibration
experiment (sites 1,2 & 3).
0
0.2
Depth, m
0.4
0.6
0.8
1
1.2
1.4
0
500
1000
1500
2000
2
Clare Weir
Payards Lagoon
Church's Lagoon
Log. (Clare Weir)
Log. (Payards Lagoon)
Log. (Church's Lagoon)
Field Measured Light Extinction Co-eff., m-1
Figure D1 – 10: Interim calibration curve derived from field testing and experimental results.
1000
100
10
1
0
0
1
10
100
1000
-1
Column Tested Light Extinction Co-eff., m
66
Figure D1 – 11: Corrected light extinction plot for the 9 soil samples.
-1
1000
100
10
1
1
10
100
1000
10000
Soil A
Soil B
Soil C
Soil D
Soil E
Soil F
Soil G
Soil H
Soil I
Figure D1 – 12: Corrected euphotic depth plot for the 9 soil samples.
Euphotic Depth, m
3
2.5
2
1.5
1
0.5
0
1
10
100
1000
10000
Soil A
Soil B
Soil C
Soil D
Soil E
Soil F
Soil G
Soil H
Soil I
67
Figure D1 – 13: Interim prediction curve for the light extinction co-efficient.
-1
1000
y = 0.412x
0.6607
2
100
R = 0.9678
10
1
1
10
100
1000
10000
68
D2
EFFECTS OF LIGHT INTENSITIES ON THE METABOLISM OF SUBMERGENT
AQUATIC PLANTS
D2.1
Aim
We aimed to quantify the metabolic response of submergent aquatic plant species to changing light intensity
levels.
D2.2
Methods
Preliminary experiment
A preliminary experiment was undertaken to determine the approximate metabolic responses of various local
submergent plant species to different light intensity levels with a view to selecting two plant species for use
in the main replicated experiment. We also used this experiment to select appropriate light intensity levels
for use in the main experiment.
This preliminary experiment involved the collection, acclimation and preparation of four common species of
submergent plants from the Ross River weir system. The four most common species found in Ross River
were: Hydrilla verticillata, Ceratophyllum demersum, Cabomba caroliniana and Vallisneria gracilis. Note
that, in natural waterways, these plants usually have a biofilm of algae and/or microbes growing on their
surface. No attempt was made to remove this associated biota from the plant specimens used in the
experiments. Consequently, it is not possible to separate the relative effects of the experimental treatments
(in this case light intensity) on the target plant species from effects on the associated biofilm. Nonetheless,
the use of plants with associated biofilm in the experiments makes the results more realistic and ecologically
relevant than if the biofilm had been removed.
A light exposure chamber capable of maintaining the desired light and temperature conditions was
constructed for use in both the preliminary and main experiments (see Figure D2-1). This consisted of an 80
litre storage bin, which was painted black on the inside. An access sleeve was installed at one end to allow
insertion and retrieval of Winkler sample bottles without interfering with the light levels within the chamber
during the experiments. Layers of translucent plastic were then placed over the chamber to obtain the
required light intensities within the chamber. Light levels where measured using a LI-COR Quantum sensor
(Model No. LI-192SB) which was mounted vertically within the chamber, and positioned to ensure that
accurate readings of the incident light levels were obtained.
Samples were prepared the day before by inserting a pre-weighed specimen of each plant species into a
standard Winkler bottle containing filtered tap water and plant nutrient. Each bottle was sealed and placed in
a constant temperature chamber (at 28 o C), in the dark overnight.
The following day, when oxygen levels were low enough to begin the preliminary experiment, each sample
was exposed for a certain period (1-5 minutes) to a number of different light intensity levels starting from
almost complete darkness (1-3 µE/m2/s) to full sunlight intensity (1200-1500 µE/m2/s) while constantly
logging dissolved oxygen levels and temperature. Samples were subjected to constant stirring throughout the
experiments via a magnetic stirrer placed immediately below the metabolism chamber.
Main experiment
In this experiment we aimed to quantify the metabolic responses of submergent macrophytes to different
incident light intensities. To achieve this, it was necessary to determine respiration and photosynthesis
independently.
Respiration rates were determined in a specifically designed dark room which was equipped with:
•
•
constant temperature water bath with cover;
magnetic stirrer;
69
•
•
dissolved oxygen and pH measuring equipment (WTW CellOx 325 DO probe and a WTW SenTix
41 pH probe with associated WTW Universal Pocket Meter); and
a very low intensity light source.
The photosynthesis phase was conducted using the metabolism chamber positioned in an open area, fully
exposed to the sun.
The two species of macrophytes (Hydrilla verticillata and Ceratophyllum demersum) selected for the main
experiment are the most common and abundant submergent native plant species that occur in coastal lowland
wetlands in the Herbert to Burdekin region. Specimens were collected from the freshwater portions of Ross
River, in Townsville (Gleeson’s and Aplin’s weir sections) and placed in a holding tank at James Cook
University in an open area and supplied with sufficient nutrients for them to thrive. They were allowed to
acclimate for approximately one week before representative samples were taken for use in the experiments.
This procedure allowed the plants to become accustomed to the light conditions that they would be exposed
to in the photosynthesis experiment and also ensured a relatively healthy biomass of plants.
Samples of each of the two plant species were prepared the day before the commencement of metabolism
measurements. This involved weighing 10 standard samples of each species (approximately 4 g) and placing
each into a standard Winkler bottle containing filtered tap water with added plant nutrient. These sample
bottles were then sealed and placed in the dark at a constant temperature of 28 oC overnight to obtain
dissolved oxygen levels of between 20 and 40 %Saturation prior to the experiment commencing.
A total of ten sample bottles where used for each of the two plant species. These were divided into two sets
of five bottles with one bottle from each set used as the temperature reference. This enabled us to conduct
four replicates at each light intensity while only using each set of samples for two light exposure events. This
was necessary to ensure that dissolved oxygen within the samples did not rise to levels higher than the
desirable maximum of 90 %Saturation.
The experiments were carried out on two cloudless days in July 2002. This was to ensure that the metabolism
chamber received a relatively constant light intensity over time scales ranging from 10 to 30 minutes. The
bottles were then carefully sealed and placed in the dark chamber to respire for 20 minutes. Subsequently the
samples were unsealed, reread and placed directly into the light exposure chamber for the required exposure
time (see Table D2-1). After exposure they were again placed in the dark chamber for a further 20 minutes to
determine if any changes in the respiration rate had occurred due to the light exposure. The four replicate
samples were treated in this same way for each light intensity level. Initial pH, temperature and dissolved
oxygen levels were determined in each sample bottle immediately prior to the commencement of the
experiments.
Sample bottles were immersed in water to a depth of 5cm while in the light exposure chamber in order to
maintain a relatively constant temperature of between 27 and 29 oC. By constantly monitoring the reference
bottle temperature using a WTW CellOx 325 combined DO & temperature probe it was possible to maintain
the sample bottles within the acceptable temperature range for the duration of the tests. When temperatures
began to drift during the tests, the water in the chamber was exchanged as shown in Figure D2-1. During
dark respiration, the sample bottles were placed in a covered water bath (within the dark room) which
maintained a constant 28 oC temperature within the sample bottles.
At the conclusion of each set of experiments, readings of the final pH were taken, the sample bottles emptied
and the oven-dry weights of the sample plant material determined using standard laboratory procedures.
Table D2-1: Exposure Times at Different Light Intensities.
Hydrilla verticillata
Light Intensity,
2
µE/m /s
0
105.5
181
600
1450
Exposure Time,
minutes
20
30
15
10
10
Ceratophyllum demersum
Light Intensity,
Exposure Time,
minutes
µE/m2/s
0
20
100.5
30
272
15
760
15
1350
15
70
D2.3
Results
Preliminary experiment
The results of the preliminary metabolism experiment are shown in Figure D2-2. While Ceratophyllum and
Vallisneria displayed very similar photosynthetic responses under these constantly stirred conditions,
Hydrilla performed significantly better while Cabomba performed unpredictably and erratically, displaying a
very heightened photosynthetic response at certain times, but not responding to light exposure at all at other
times. We are yet to determine the reasons for this peculiar behaviour.
This experiment suggests that the three native species (Hydrilla verticillata, Ceratophyllum demersum, and
Vallisneria gracilis) each respond to changes in light intensity in a similar predictable manner while under
these specific experimental conditions Cabomba did not. Since Cabomba is a noxious weed and emerging
pest species in this region, this deserves closer investigation.
For these experiments, we focused on Ceratophyllum and Hydrilla, for the following primary reasons:
•
•
•
these two species are the most common native aquatic submergent macrophytes in this area;
both display predictable responses to a wide range of light climates; and
Vallisneria is not nearly as prolific in this area as the other species.
The other aim of the preliminary experiment was the selection of appropriate light intensity levels for the
main experiment. Four different light levels were sufficient to provide the critical points on the metabolism
curves; they were 100, 200, 600-700 and 1300-1500 µE/m2/s.
We found from our preliminary experiments that different plant species respire at different rates. At a
constant temperature of 28oC in a constantly stirred state, Hydrilla and Ceratophyllum respired at
approximately the same rate while Cabomba had a substantially higher rate (see Figure D2-3). These results
show that Cabomba has a higher metabolism than the native plant species and that the native species may be
at a competitive disadvantage in terms of growth rates, at least under the conditions used in these tests.
We also examined the effect of different water temperatures on the respiration rate of Ceratophyllum. Figure
D2-4 shows the results of this experiment where Ceratophyllum was subjected to different water
temperatures (i.e., 18, 28 and 38 oC) and the respiration rate measured under constantly stirred conditions.
This suggests that plant respiration rate increases at an exponential rate as the ambient water temperature
increases from 18 to 38oC. The Q10 (temperature coefficient) for respiration (i.e., ratio of increase due to a 10
o
C increase in ambient temperature) varied from 2.11 (18 to 28 oC) to 2.44 (28 to 38 oC) in this experiment.
These values are higher than the traditionally accepted Q10 value of 2 stated in many scientific publications.
Main experiment
The results of the main experiment for the two species tested (Ceratophyllum and Hydrilla) are displayed
graphically in Figures D2-5 and D2-6. These plots show how photosynthesis rate increases with light
intensity. It can be seen that a threshold point (that we have termed the Optimum Light Utilization Point or
OLUP) is reached whereby the rate of increase in photosynthesis with light intensity dramatically decreases
at a certain light level. For Hydrilla this threshold point is approximately 207 µE/m2/s and for Ceratophyllum
it is 265 µE/m2/s. The effects of high light intensities on Ceratophyllum are more complex than Hydrilla (see
Figure D2-6).
The shaded bars at the side of the plots in Figures D2-5 and D2-6 provide an indication of the types of
natural conditions that would typically produce the different light intensities shown. Of particular note is that
both plant species are capable of net oxygen production (and growth) under very low light conditions (heavy
shade) if they exist at or near the surface of the waterway. At 1 metre depth they are still capable of some
production even in waters with turbidities greater than 40 NTU in clear, sunny and unshaded conditions.
The compensation point (the point at which the plants exactly balance their rate of oxygen production and
consumption) is an important factor in determining the success of any given species. This depth is a
71
boundary within the water column that dictates the depth at which plants can actually grow and accumulate
biomass. The euphotic depth (i.e., depth at which the light intensity is 1% of the incident) is widely used as a
surrogate for the compensation depth because the actual compensation point varies widely between species
and even within the same species. It was determined during this experiment that the compensation point of
Hydrilla is approximately 11 µE/m2/s while for Ceratophyllum it is about 32 µE/m2/s for these particular sets
of conditions. This effectively means that Hydrilla is capable of existing in deeper and more turbid waters
than Ceratophyllum and is likely to be most competitive immediately after flushing events when water
clarity is typically at a seasonal minimum.
The calibration curves from experiment D1 were used to calculate the depths at which threshold light levels
for Hydrilla occurred under different combinations of ambient SPM concentration and incident light
intensity. The calculated values are plotted in Figure D2-7. This plot shows, for example, that in direct
sunlight the critical threshold depth varies from approximately 1.75 metres at 10 mg/L SPM concentrations
to over 10 meters at 1 mg/L. These depths decrease substantially when the incident light intensity falls due to
shading (either by cloud cover or riparian vegetation). Over this same SPM concentration range, the euphotic
depth varies from 2.4 metres to over 21 metres. As an example, if the SPM concentrations of a waterway
always remained greater than 10 mg/L then these aquatic plants would not be able to colonize depths greater
than 2.4 metres at any time. Parts of the water column than are deeper than the euphotic depth are known as
the aphotic zone. Here no net photosynthesis occurs and in most waterways this critical depth varies
(sometimes on very short time scales) depending upon the light attenuation characteristics of the water
column. This light attenuation can be caused by many factors (eg. SPM, plankton, dissolved colour, colloids,
etc.) but during major runoff events SPM from the catchment is usually the main contributor and was the
primary focus in this study.
Considerations of this kind can be used to classify different parts of water bodies based on variations in light
climate. This is illustrated by example in Figure D2-8 which shows that three different light zones can be
identified in the hypothetical situation portrayed in the figure. Zone A represents the upper water column
where excess light is generally available for use by submergent plants. Plants located in this zone are
potentially light inhibited but photosynthesise at high levels.
Zone B is bounded by the OLUP above and the euphotic depth below (the euphotic depth is used as an
estimate for the compensation point in this example). Plants in this zone are able to utilize light very
efficiently to produce biomass and oxygen (if not nutrient limited), but can be expected to respond very
strongly to subtle changes in light intensity (and/or clarity).
Zone C (aphotic zone) may or may not exist depending upon the clarity and depth of the water body. It
represents the area where plants are not able to grow (i.e., respiration is greater than photosynthesis).
D2.4
Discussion
The preliminary experiment showed that the photosynthetic performance of submergent plants differed
greatly between species. Under the test conditions Cabomba in particular proved to be extremely erratic in its
responses to incident light intensities and did not exhibit a consistent photosynthesis rate over the range of
light intensities used (see Figure D2-2). The other three species Hydrilla, Ceratophyllum and Vallisneria
performed more consistently. The former two species were used for the main experiment, primarily due to
their prevalence in the wetlands of the Herbert to Burdekin floodplain wetlands.
A primary objective of the main experiment was to identify critical light intensity levels that induce a change
in response by the plants. The experiment with Hydrilla was particularly successful in this regard with an
obvious threshold light level (OLUP) of 207 µE/m2/s obtained, that represented a point where
photosynthesitic response to increased light intensity was severely inhibited (see Figure D2-5).
Ceratophyllum was more complex, exhibiting a less pronounced decline in the rate of increase in
photosynthesis at 265 µE/m2/s and partial recovery in photosynthetic response at light levels above 760
µE/m2/s and up to 1360 µE/m2/s (see Figure D2-6). However, rates of increase were still significantly
retarded at light levels higher than the initial threshold of 265 µE/m2/s.
72
These experiments represent a limited range of conditions compared to the natural environment. Many other
variables affect plant responses to ambient conditions, such as season, growth phase, condition, acclimation,
acclimatisation, etc., so these results are only intended to be indicative of the performance of these plant
species in freshwater wetlands in this area. Much more extensive testing and replication would be needed to
quantify natural variability fully.
D2.5
Figure D2-8 illustrates some important concepts. For a given set of conditions (level and type of light
exposure and bathymetry) there are critical ranges of SPM concentration that will impose fundamental
restrictions on the spatial distribution of primary productivity. At concentrations that are above a critical
limit, photosynthesis will be severely retarded throughout the entire water column. Under these conditions
emergent/floating plants are the only autotrophic species that can grow and productivity within the water
column must be heterotrophic (i.e., based on consumption of organic carbon by microbes, zooplankton,
invertebrates, etc.). At intermediate SPM levels there is sufficient light penetration to allow photosynthesis to
occur in the upper water column and in benthic habitats in the shallow littoral zones. This creates situations
where the balance between heterotrophy and autotrophy varies spatially throughout the water body leading to
very heterogeneous metabolic characteristics. However, in most large lagoon systems that are not choked by
floating weeds, the majority of the water column becomes autotrophic with productivity being mainly
planktonic. Below a lower SPM threshold concentration, light penetration all the way to the bottom may be
possible enabling the establishment of benthic autotrophic communities (rooted macrophytes, benthic algae,
biofilms, etc.) Resulting in the activation of new benthic food webs.
Changes in SPM concentrations within the upper concentration range have little impact on autotrophs;
however, once levels fall below the upper threshold, further reductions can greatly affect primary production
rates in the water column (nutrient limitation is unlikely to be a factor in the wetland environments under
consideration). Similarly, below the lower threshold, variations in SPM can be expected to substantially
affect benthic productivity.
The process-level changes from heterotrophy to mixed productivity to autotrophy play a pivotal role in
determining the types of biological communities that develop in the water body. In other words, SPM
changes within the bounds of critical limits can be expected to affect the productivity of existing biological
communities while changes which cross threshold values are much more likely to bring about fundamental
changes in the types of biological communities that colonize the water body.
73
Figure D2 – 1: Light exposure chamber design.
Direction of
light quanta
l
Layers of
translucent
plastic
sheeting
Water level
(approx. 5
cm deep)
Access
Submersible
pump & water
discharge pipe
Temperature
probe
and
meter
Light sensor
& meter
Sample
bottles
Figure D2 – 2: Net oxygen evolution at different light intensities by four common species of
submergent aquatic plant.
1400
2
1600
1200
1000
800
600
400
200
0
0
5
10
15
20
25
30
O2 Evolution Rate, mg O2/min/kg of wet plant
Ceratophyllum
Cabomba
Hydrilla
Vallisneria
74
Figure D2 – 3: Aquatic plant respiration rates of three local species (Hydrilla, Ceratophyllum and
Cabomba) at 28oC under constantly stirred conditions.
0.025
Respiration Rate,
mg O2/min/g oven-dry plant
0.02
0.015
0.01
0.005
0
Hydrilla
Ceratophyllum
Cabomba
Figure D2 – 4: Aquatic plant respiration rates of Ceratophyllum at different ambient water
temperatures under constantly stirred conditions.
Dark Respiration Rate,
mg O2/min/g oven-dry plant
0.035
0.03
0.025
y = 0.0013e
0.0819x
2
R = 0.9973
0.02
0.015
0.01
0.005
0
16
18
20
22
24
26
28
30
32
34
36
38
40
Temperature, Deg.C
75
Figure D2 – 5: Photosynthetic response of Hydrilla to different light intensity levels.
In direct sunlight in summer
2000
Relative light
levels incident
on water
surface
Relative light levels at
different turbidities
and at a constant
depth of 1 m
Dec 22
Light Intensity
2
Incident Light Level, E/m /s
Annual light intensity range
on clear days
Direct
Sunlight
1500
June 22
different depths and
with a constant
turbidity of 20 NTU
y = 86375x - 16478
2500 µE/m2/s
1500 µE/m2/s
0 - 0.5m depth
0 - 10 NTU
Sunset
Sunrise
Time of Day
1000
Part
Shade
Threshold point
x = 0.192, y = 207
500
0.5 - 1m depth
10 - 20 NTU
Medium
Shade
Deep Shade
Heavy Shade
0
-0.05
1 - 1.5m depth
1.5 - 2m depth
20 - 40 NTU
y = 1075.3x
> 40 NTU
> 2m depth
0
0.05
0.1
0.15
0.2
0.25
0.3
0.35
0.4
Photosynthesis Rate, mg O2/min/g oven-dry plant
Figure D2 – 6: Photosynthetic response of Ceratophyllum to different light intensity levels.
In direct sunlight in summer
2000
Relative light
levels incident
on water
surface
different turbidities
and at a constant
depth of 1 m
different depths and
with a constant
turbidity of 20 NTU
Direct
Sunlight
Dec 22
Light Intensity
2
Incident Light Level, E/m /s
Annual light intensity range
on clear days
1500
June 22
2500 µE/m2/s
1500 µE/m2/s
0 - 10 NTU
y = 4867.8x - 180.17
Sunset
Sunrise
Time of Day
1000
Part
Shade
Threshold point
x = 0.152, y = 265
y = 12158x - 1588.3
500
Medium
Shade
10 - 20 NTU
y = 1737.7x
20 - 40 NTU
Deep Shade
Heavy Shade
> 40 NTU
0
-0.05
0 - 0.5m depth
0.5 - 1m depth
1 - 1.5m depth
1.5 - 2m depth
> 2m depth
0
0.05
0.1
0.15
0.2
0.25
0.3
0.35
0.4
Photosynthesis Rate, mg O2/min/g oven-dry plant
76
Figure D2 – 7: Plots of the depths at which the Optimum Light Utilisation Points (OLUP’s) occur for
different incident light levels and different suspended solids concentrations for
Hydrilla. Incident light intensities used were: 2000 µE/m2/s (direct sunlight), 800
µE/m2/s (part shade) and 400 µE/m2/s (medium shade).
10000
OLUP for Hydrilla = 207 µE/m2/s
(Ceratophyllum = 265 µE/m2/s)
1000
100
10
1
0
2
4
6
8
10
12
14
16
18
20
22
Depth, m
Direct sunlight
Part shade
Medium shade
Euphotic Depth
Figure D2 – 8: Plot of the curves for the OLUP (Hydrilla) and euphotic depth in direct sunlight (2000
µE/m2/s) with the photic zones identified for a wetland with a nominal suspended
solids concentration of 10 mg/L.
10000
Zones within a wetland with
SS Concentation = 10 mg/L
Light Inhibition Zone (A)
1000
Optimum Plant Efficiency Zone (B)
Aphotic Zone (C)
100
Estimate of Plant
Compensation
Point
C
10
Critical
Threshold
Curve
Depth of zone
A = 1.6m 1
Euphotic Depth Curve
OLUP Curve
0
2
A
4
B
6
C
8
10
12
14
16
18
20
22
Depth, m
77
D3
D3.1
THE EFFECTS OF SEDIMENTATION ON DISSOLVED OXYGEN, PH AND
PLANT GROWTH IN AN AQUATIC MICROCOSM DOMINATED BY THE
SUBMERGENT PLANT CERATOPHYLLUM DEMERSUM
Aim
An experiment was conducted to determine how moderate levels of sedimentation affect water quality and
the metabolism of Ceratophyllum demersum, a submergent native aquatic plant that commonly dominates
waterways in local cane-growing areas. The experiment focused on the effects of sedimentation (the settling
of silts and flocculated clays and colloids) and excluded the effects of increased turbidity from suspended
particles such as unflocculated clays and colloids and also of burial by coarse sediment. Water clarity effects
were dealt with as a separate issue in Experiments D1 and D2 while the effects of coarse material are
discussed
D3.2
Methods
The experiment was conducted in an outdoor compound at James Cook University, Townsville using 30-litre
glass tanks. The tanks were positioned such that they were exposed to direct sunlight each day from
approximately 0915 to 1630 hours. Each tank was stocked with 300 g (wet weight) of Ceratophyllum
demersum which had been collected from a local waterway (Ross River). This plant species commonly has a
biofilm of algae and/or microbes growing on its surface. This bio-film was not removed from the plant
specimens used in this experiment to ensure that the results are applicable to natural waterways. Plants were
submerged in each tank by entwining them through weighted plastic mesh. When submerged, the plants
extended through the mesh at the bottom of the tank and almost reached the water surface. The density of
plants used in the experiment was visually similar to that observed in moderate stands of C. demersum in
natural waterways.
Experimental Conditions
Plants were exposed to 1 of 3 sediment treatments: (1) plants buried under a light to moderate layer of
flocculated silt, (2) plants placed on top of a layer of silt, and (3) plants not exposed to silt. The experiment
was a randomised block design with each treatment being replicated 3 times. Treatments and plants were
allocated randomly to tanks. The sediment used in the experiment was collected from a local dam (Lavarack
Upper Dam) whose catchment consists of 100% natural vegetation (in order to minimise the risks of
introducing anthropogenic contaminants).
Preparation of Sediment Suspension
Coarse particles (gravel and sand) were removed by suspending the sediment and allowing it to settle for a
few minutes. At the end of the settling period, the supernatant was collected and the material that had fallen
out of suspension was discarded. The supernatant was then filtered through 1mm fibreglass mesh to remove
coarse particulate organic matter. Dispersed clay and colloid particles were then removed by suspending the
resultant sediment in 30 litres of filtered tap water. Ninety drops of flocculent (AquaclearTM) were added and
the sediment was allowed to settle. After 2 hours, the supernatant was decanted and discarded while material
that had fallen out of suspension was collected. This last process was repeated several times to ensure the
removal of clay particles from the sediment.
Experimental Treatments
Treatment 1 – Plants buried under sediment
Treatment 1 tanks were established by filling them with 26 litres of filtered tap water and adding plants. One
litre of concentrated sediment was then added to each tank. Care was taken to ensure that sediment was
added over the entire surface area of the tank. After 1 hour, 90 drops of flocculent were added to each tank
and the sediment was allowed to settle overnight. The following morning, it was observed that the majority
of sediment had settled out of solution and was covering the plants. The appearance of sediment covering the
plants was very similar to what we have observed in local waterways subject to moderate sediment loads.
78
Treatment 2 – Plants above layer of sediment
Treatment 2 tanks were established as per the treatment 1 tanks with the exception that plants were not added
until after the sediment had settled overnight. The result was that a layer of sediment covered the bottom of
the tanks but the plants in the water column were not covered with sediment. The purpose of this treatment
was to separate the effects of smothering (treatment 1) from the effects of sediment per se.
Treatment 3 – Control
Treatment 3 tanks were established by initially allocating a duplicate tank to each experimental tank. The
duplicate tanks were filled with 26 litres of filtered tap water at the same time as the treatment 1 and 2 tanks.
Sediment and flocculent were added to the duplicate tanks as per treatments 1 and 2. Following overnight
settling of the sediment, 24 litres of the supernatant was siphoned from each duplicate tank into the
corresponding experimental tank. Two litres of filtered tap water was added to each experimental tank to
give a final volume of 26 litres. Plants were then added to treatment 3 tanks. The procedure of filling
treatment 3 experimental tanks with supernatant from duplicate tanks was necessary to ensure that the
composition of the water was identical in all treatments.
Addition of Nutrients
Nutrients (N as NaNO3 and P as K2HPO4) and trace elements (Seachem™) were added to each tank on the
day following the addition of sediment. Initial concentrations of N and P were similar to those found in
unmodified waterways during periods of good water quality.
Adjustments to Experimental Conditions
Tanks were exposed to a range of light regimes and levels of re-aeration during the experiment. Light
regimes were created by covering the tanks with layers of translucent colourless or black plastic. Light
attenuation categories ranged from low (1) to high (5). Categories were: (1) 2 layers colourless plastic, (2) 4
layers colourless plastic, (3) 2 layers black plastic, (4) 3 layers black plastic, and (5) 4 layers black plastic
since natural sunlight was being used the timing of adjustments was dictated by weather conditions, and the
practical need to maintain dissolved oxygen levels within normal limits. (For example during consecutive
cloudy days oxygen levels gradually became hypoxic and it was necessary to admit more light in order to
prevent anoxia). Re-aeration was inhibited (to enable changes in oxygen concentration to be used as a
measure of photosynthesis rates) by placing a layer of clear plastic on the water surface of each tank.
Monitoring of Responses
The experiment ran for 22 days. During this time, dissolved oxygen, pH and temperature were measured at a
standard position and depth in each tank every morning (0800-0900 hrs) and afternoon (1530-1630 hrs)
using a WTW Multiline meter. Datalogging subsequent to the experiment verified that these were the times
of daily minima and maxima for these parameters. Observations on plant condition and weather were also
made daily.
D3.3
Results
Daily variations in dissolved oxygen concentrations and pH are plotted in Figures D3-1 to D3-6. Data are
either indexed as a proportion of mean control values or are presented as actual values. For simplicity of
presentation and interpretation, dissolved oxygen data are plotted as mean index values in Figures D3-1 and
D3-2. These data are discussed further below. Ranges of dissolved oxygen and pH recorded in each tank are
plotted as index values and actual values in Figures D3-3 to D3-6.
The addition of sediment reduced (i) dissolved oxygen concentrations (Figures D3-1 and D3-2), and (ii) pH
and the degree of pH cycling (Figures D3-5 and D3-6). Trends in dissolved oxygen concentrations in tanks
where sediment was added (either on top of, or below, plants) were similar during both mornings and
afternoons (Figures D3-1 and D3-2). On at least one occasion per day during days 2 to 16, dissolved oxygen
79
concentrations in tanks where plants were smothered with sediment were lower than in tanks where sediment
was present but plants were not smothered. Reductions in dissolved oxygen concentrations were particularly
evident during this period when plants were exposed to low light levels (light attenuation category 5; periods
of cloudy days).
On day 14 it was noted that plants that had been smothered with sediment had begun producing new
branches that were not covered with sediment. These plants, as well as plants in the other two treatments,
continued to grow during the remainder of the experiment. At this time, dissolved oxygen concentrations in
both sediment treatments increased and approached the levels recorded in control tanks. During the last few
days of the experiment, mean dissolved oxygen concentrations in tanks where plants were smothered
exceeded levels in tanks where sediment was present but plants were not smothered. Plant growth appeared
to be most prolific in the “smothered” treatment tanks at this time.
D3.4
Discussion
The addition of sediment to the experimental tanks brought about a substantial reduction in dissolved oxygen
and pH levels, and suppressed diel pH cycling. This suggests that the sediment contained significant
quantities of organic matter which was acting as a substrate for microbial respiration, leading to increased
consumption of oxygen and production of carbon dioxide. Examination of the results obtained at times when
light levels were not severely limiting (attenuation categories 1 to 4 and cloudless days) suggest that effects
due to the presence of sediment (regardless of whether plants were covered by it or not) diminished gradually
over time and would have been undetectable after 10 to 15 days if adequate light conditions had been
maintained. This is almost certainly indicative of constant reduction in the concentrations of organic matter
in the sediment and consequent reduction in microbial respiration rates. The results suggest that, by the end
of the experiment, virtually all of the sedimentary organic matter had been consumed and that microbial
respiration rates were the same as that in the control tanks. The effects of smothering plants with sediment
were far more subtle and less predictable. The main initial effect was an increase in the intensity of light
required for optimum photosynthesis and hence an increase in the compensation point (the light level at
which oxygen production is equal to oxygen consumption). Accordingly, when light levels were higher than
the optimum there was no readily discernable effect from smothering, but at lower light levels (light
attenuation category 5 and/or cloudy weather) reduction in oxygen production and pH cycling was evident.
This is a very important finding because in the real world sediment is normally introduced into water bodies
during rain events, so sedimentation rates are often highest at times when light levels are suppressed by
turbidity in the water, increased water column depth and cloudy weather.
Suppression effects were relatively short-lived and the smothered plants were able to grow through the
sediment layer within about 13 days. Once this had happened the smothered plants began to photosynthesise
vigorously and towards the end of the experiment appeared to be photosynthesising more efficiently than the
plants in other treatments. This probably indicates that the plants were utilising the nutrients present in the
sediment that they were in intimate contact with.
The effects observed in this experiment are governed by great variety of variables – sediment added, depth of
smothering, organic composition of sediment, incident light intensity, water clarity, water depth, re-aeration
and mixing rates, and temperature – which all play important roles in determining outcomes. The conditions
chosen for this experiment simulate just one of the many scenarios that can occur in natural waters, so
findings must be interpreted cautiously. Nevertheless, some important qualitative outcomes are evident:
•
The organic matter and microbes contained in sediment are the principal source of impact from settlable
sediments. The maximum impacts are likely to become evident soon after the introduction of sediment
(within hours to a few days depending on initial microbial abundances) and are likely to gradually
diminish over periods of several days to a few weeks if conditions do not become severely hypoxic
(allowing microbes to decompose organic matter to carbon dioxide). Recovery is likely to be
significantly delayed if oxygen levels remain low enough to force microbes to respire anaerobically
leading to the production of oxygen demanding by-products such as organic acids, alcohols, methane
and hydrogen sulphide.
80
•
Ceratophyllum is clearly well adapted to cope with the smothering effects of silts and flocculated
sediments and it would be reasonable to assume that other species with a demonstrated ability to
compete in environments that are periodically subject to pulse inputs of sediment would possess similar
adaptations. Nevertheless, their capacity to tolerate smothering is clearly finite and is heavily reliant on
the maintenance of adequate light and oxygen levels in the surrounding water column. Specifically it is
clear that smothering increases the optimum light utilisation point (OLUP smothered), so smothered
plants require higher light levels in order to photosynthesise efficiently, and are therefore much more
susceptible to adverse effects from decreased clarity and cloudy weather.
81
Figure D3 – 1: Mean treatment values for dissolved oxygen concentrations recorded each morning
during the experiment. Data are indexed as a proportion of mean control (Plant Only
treatment) values. Light attenuation categories range from 1 (highest light level) to 5
(lowest light level).
Light Attenuation Category
1 2 3
4
5
4
Re-aeration Inhibited
Re-aeration Not Inhibited
1.2
Plant Only (No Sediment)
Plant Buried Under Sediment
Plant on Top of Sediment
Morning Data
Mean index of dissolved oxygen (% saturation)
1.1
1.0
0.9
0.8
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0.0
0
1
2
3
4
5
6
7
8
9
10
11
12
cloudy
13
14
15
16
17
18
19
20
21
22
cloudy
Time (days)
82
Figure D3 – 2: Mean treatment values for dissolved oxygen concentrations recorded each afternoon
during the experiment. Data are indexed as a proportion of mean control (Plant Only
treatment) values. Light attenuation categories range from 1 (highest light level) to 5
(lowest light level).
1 2 3
4
5
4
1.2
Afternoon Data
Mean index of dissolved oxygen (% saturation)
1.1
1.0
0.9
0.8
0.7
0.6
0.5
0.4
0.3
0
1
2
3
4
5
6
7
8
9
10
11
12
cloudy
13
14
15
16
17
18
19
20
21
22
cloudy
Time (days)
83
Figure D3 – 3: Dissolved oxygen concentrations recorded each morning and afternoon during the
experiment. Data are indexed as a proportion of mean control (Plant Only treatment)
values. Light attenuation categories range from 1 (highest light level) to 5 (lowest light
level). Numbers are experimental tank numbers.
1 2 3
4
5
4
1.4
4
1.3
44
Index of dissolved oxygen (% saturation)
1.2
4
1.1
4
4
4 711
4
1.0 11
9
9
6
12
11
5
6
8
8
7 1210
10
5
7
5 12
9
10
0.8
8
6
0.9
7
11
9
5
10
12
6
8
0.7
4 44
119
11
9 911
8
10
6
6
10
10
78
7 8
5
12
5 612
5
127
0.6
0.5
0.4
0.3
0.2
4
4
4
44
119
11 119
911 1111
11
9
11
9
9 12
6
6
10
10
8
7
9
8
12
10
512
5
8
67
95
7
10
5
710
7
8
66
8
5 1010
7
8
6
5
1212
8
712 10
7
6
8
6
5
12
12
5
7
8
0.1
4 4
9
4
9 11
1111 11
910 9
6
6
7 10
8
5 10
7
1012 7 8
8
7
6
65
8
5
5
12
12
4
9
44 49
4
9
4
4
9
1111 11
911 11
9 11
9
9
6 11
6 11
7
10
10
6 107
77
8
10 7 5 10
712
5
7
105
610
10
8
12
8 12
7
6
6
8
6
86
5
8
8
5
5
8
5
5
12 1212
12
12
4
12
49
4
49
4
9 4
4
9
4
9 4
4
7
114
912
11 11
6
6
12
10
6
9
9
7
9
11
7 1111
11
11 10
911
5
5
8
11 10
7
10
7
12
6
5 108 12
5
8
77
10 7
8
710 7
5
6
10
10
12
6
10
6
8
5
5
612
8 12
5
68 5
8 12
5
12
6
8 12
8
49
712
4
11
7
6
910
11
5
8
10
12
5
8
6
15
20
4
12
9
4
76
910
11
7
11
5
128
10
5
8
7
4
9
12
11
10
8
5
6
6
0.0
0
1
2
3
4
5
6
9
10
11
12
cloudy
13
14
16
17
18
19
21
22
cloudy
Time (days)
84
Figure D3 – 4: Dissolved oxygen concentrations recorded each morning and afternoon during the
experiment. Data are actual values. Light attenuation categories range from 1 (highest
light level) to 5 (lowest light level). Numbers are experimental tank numbers.
1 2 3
4
5
4
200
4
9
11
190 12
6
8
180 10
7
170
5
11
4
9
6
10
8
5
7
12
160
Dissolved oxygen (% saturation)
150
140
130
120
110
100
2
3
90
70
60
4
50
7
11
5
12
9
10
8
6
40
4
8
7
5
12
11
9
4
6
8
10
12
2
5
3
1
7 27
5 3
12
1
4
11
9
7 10
7
6
5
8
8
12
10
5
6
12
11
9
6
8
10
2
1
33
2
1
80
6
10
2
1
3
4
9
5
11
7
10
12
6
8
4
9
11
2
3
1
210
6
1
3
5
8
12
47
11
9
7
8
12
5
10
6
30
20
10
4
2
4
2
1 2
9
3
211
1
3
11
1
9
1
3
3
4
4
1110
5
6
97
8
11
10
7
8
12 9 5
6
10
7
6
12
8
5
12 10
8
7
6
5
12
9
11
12
6
10
8
2
4
9
11
10
6
2
5
1
23
7
3
7 25
8
1
12
3
1
3
1
4
4
11
9
11
10
8
7
6
5
12
9
9
9
4
4
9
4
11
4
9
11
6
10
2
7
25
3
8
1
12
3
4
1
11
9
9
4
2
2
6
211
10
7
5
3
3 212
1
8
1
3
6
110
3
1
8 4
47
11 11
95 9
1012 10
7
7
6
6
8
8
5
5
12
12
10
7
10
12
6
8
5
12
4
9
6
11
10
7
11
6
7
10
8
2
5
12
3
2
1
3
4
1
11
9
7
10
6
8
5
12
9
4
12
6
10
7
11
5
8
8
2
3
3
2
9
12
4
6
7
11
10
5
8
47
1110
96
5
712
8
10
6
5
8
12
47
11
910
6
5
12
78
10
3
2
12
5
2
21 21 21
3
3
8
4
1
4
3
3
1
1
7
1
7
9 11
9
4 11
4 11
10
9 10 12
5
8
6
12
9
5
6
11
8
7
10
5
6
12
7
10 8
6
12
5
8
5
6
8
12
15
16
17
9
4
22
4
2 29
9
2
24
11
3
31
3 31
3
1
1
11
311 1
7
410 1
9
11
6
8
75
6
1012
8
5
12
4
6
12
11
7
10
5
10
7
8
6
12
5
9
12
4
6
10
7
11
5
8
3
2
1 3
2
1
7
4
9
7
12
9 11
11
10
12
8
5
5 10
6
8
6
2
3
1
4
0
0
1
2
3
4
5
6
7
8
9
11
cloudy
13
14
18
19
20
21
22
cloudy
Time (days)
85
Figure D3 – 5: pH values recorded each morning and afternoon during the experiment. Data are
indexed as a proportion of mean control (Plant Only treatment) values. Light
attenuation categories range from 1 (highest light level) to 5 (lowest light level).
Numbers are experimental tank numbers.
1 2 3
4
5
4
1.15
4
1.10
4
1.05
Index of pH
4
1.00 11
4
6
9
8
10
5
0.95
7
12
4
54
11
5
76
9 11
6
11
910
8
12
10
8 6
5 12
7
9
7
8
10
12
0.90
4
4
4
4
4
11
4
4
4
4
4
4
4
4
4
11
5
5 11
5
11
7
7
9 11
9
9
6
7 10
9 10
6
6
8 9 10
8
12
12
12
811
9 11
5
10
9
6
6
10
7 12
7
8
5
6
8
8
5 10
6
8
10 12
0.85
5
12
7
4
4
4
11 5 11
5
6
9 11
9 9 911
5
6
7 10
7 10
7
6
8
8
10
12
9 11 8
1211
129
5
10
7
6
8
12
10
6
5
7
6
8
12
5
10
8
12
7
12
11
5
6
99
711
10
8
12
5
6
10
7
8
12
9
4 4
44 4
9
9 11
9 1111 11
99
11
5
6
711
11
10
9
6
6
5
5
7 10
7 10
7
6
8 10
12
5
8
8
9 8
7
12
1210
5
6
11 12
8
12
6
10
12
5
7
8
7
4
4 4
9
11
9
511
7
6
10
8
6
7
5
1210
11
9 9 11
9
7
611 10
7
6
5
5
10
8
5 12
8
1210
7
6
8
8
12
12 9
10
6
5
7
8
12
6
10
7
5
6
10
5
7
2
3
4
5
6
7
8
9
10
11
4
9
11
9
6
5
7
10
8
12
11
12
cloudy
13
4
4
9
11
5
7
6
9
10
8
12
11
10
7
7
6
5
12
5
6
8
8
12
4 4
49
9
11
7
5
9
6
10
8
12
11
6
7
5
12
10
8
5
6
7
11
9
8
12
10
5
7
11
9
6
12
10
8
7
12
11
6
10
5
8
8
12
8
12
1
4
44
10
0.80
0
4
11
4
14
15
16
17
18
19
20
21
22
cloudy
Time (days)
86
Figure D3 – 6: pH values recorded each morning and afternoon during the experiment. Data are
actual values. Light attenuation categories range from 1 (highest light level) to 5
(lowest light level). Numbers are experimental tank numbers.
1 2 3
4
5
4
10.0
9.5
4
11
9
6
8
10
9.0 5
7
12
4
11
6
9
10
8
5
7
10
6
12
8
5
8.5
pH
4
11
7
12
4
8.0
7.5
4
5
7
6
11
12
9
8
10
7.0
11
5
6
7
12
9
10
8
4
4
11
9
4
9
4
4
9
4
9
11
5 4
12
7 12
7
5
5
9
7 11
7
6 10
6
12
12
8
8
10
11
11
4
11
9
4
11
4
11
9
11
9
4
9
7
55 4
11
6 11
9
10
6
7
12
5
10
812
9
8
6
11
5
10
9
6
8
7
712
10
8
12
6.5
11
9
11
5
5
910
6
10
7
76
8
8
1212
6
10
5
8
412
7
5
11
6
9
7
10
12
8
410
6
5
7
8
1112
9
5
7
6
10
8
12
7
8
9
4
4
9
9
9
9
6
8
5
10
4
9
4
10
6
8
4
4
4
11
4
10
6
45
7
9
411
9
411 410
6
10
7
5
5
6 1112
8
11
9
5
5
9
7 10
6
7
68
7
10
8
812 12
12
10
45
6
7
8
11
912
6
5
10
7
8
12
8
1112
9
6
10
7
5
8
12
10
7
6
11
95
5
68
7
10
12
8
12
10
12
13
14
4
4
9
411
5
6
10
7
98
11
5
7
612
10
8
12
9 10
6
5
7
411 4
12
8
5
11
7
910
9
6 11
10
5
8 5
7
6
10
7
612
8
8
12 12
15
16
11
10
11
7
12
7
11
10
6
6
7
5
10
12
5
5
6
8 4
12
48 48 4
10
6
7
5
412
8
11
9
5
6
10
7
12
8
5
11
9
7
6
10
8
12
11
5
9
7
6
10
8
12
18
19
20
5
11
9
7
6
12
10
8
5
7
9
11
6
12
10
8
21
22
6.0
0
1
2
3
4
5
6
11
cloudy
17
cloudy
Time (days)
87
D4
EFFECTS OF LIGHT DEPRIVATION AND HYPOXIA ON A SPECIES OF
SUBMERGENT AQUATIC PLANT (CERATOPHYLLUM DEMERSUM)
D4.1
Aim
To provide preliminary indications of the capacity for Ceratophyllum (a native submergent plant that
commonly dominates local floodplain wetlands with unobstructed water surfaces) to cope with light
deprivation and different levels of hypoxia, and to obtain initial estimates of critical exposure times for
survival under these conditions.
D4.2
Methods
In this experiment plant samples (and associated bio-film) were placed in small aquaria in a constant
temperature room for a period of several weeks under various fixed ambient light and oxygen conditions.
The metabolic performance of test specimens was monitored at regular intervals to determine how the health
of the plants was being affected. The experiment was conducted in an aquarium room that maintained an air
temperature of approximately 27 oC. It was important to replicate natural conditions as much as possible for
this experiment so that the plant material functioned in a natural manner for the duration of the tests. A
functioning ecosystem was established within each tank by inserting unpolluted sediment sourced from the
local area (strained to remove large debris) and aged aquarium water (to ensure the inclusion of healthy
aquatic microbes). A flocculent was added and the water allowed to settle for two days until a constant level
of high clarity was achieved. Each of four identical 30-litre aquaria were subjected to one of the following
experimental treatments:
1)
Plants exposed to a daily photoperiod of 7 hours creating diel oxygen cycling between 10 and 120
%Saturation. This mimics the conditions encountered in natural macrophyte assemblages located in
wetlands with significant background respiration. This is essentially a control to confirm that the plants
can thrive in the experimental tanks if exposed to favourable light and oxygen conditions.
2)
Plants kept in complete darkness at dissolved oxygen concentrations between 20 and 30
%Saturation. In many places these oxygen levels would be considered to constitute severely hypoxic
conditions; however, 20 to 30 %Saturation levels are typical of the concentrations encountered in the
mid-depths of local wetlands during periods of stagnation and throughout most of the water column in
the aftermath of flow events. Accordingly, by local standards, these conditions could be described as
being moderately hypoxic.
3)
A tank of water exposed to complete darkness and low oxygen concentrations. This is a control tank
used to determine background respiration levels due to organisms suspended in the water column.
4)
Plants kept in complete darkness under severely hypoxic conditions (dissolved oxygen < 5
%Saturation). This treatment simulates the conditions that frequently occur in the hypolimnion of most
wetlands deeper than 1 to 2 m (depending on size and location, etc.). Monitoring in Lagoon Creek
(Herbert floodplain) has demonstrated that a hypolimnion of this kind can form at a depth of less than
1.0 m even during the falling stages of a storm hydrograph (i.e., long before the lagoons stagnate).
Moreover, on occasions, anoxia can persist throughout most of the water column for periods of several
days to weeks. Research conducted in the Burdekin wetlands has demonstrated that Water Hyacinth
mats can also create dark severely hypoxic conditions similar to those employed in this treatment.
Plant material was obtained from the Aplin’s weir section of the Ross River in Townsville and immediately
placed in the tanks (1, 2 & 4) along with sufficient plant nutrient to sustain the plants for the duration of the
experiment. Approximately the same biomass of plant material was carefully placed in each tank (1, 2 & 4)
with minimal sediment resuspension and an initial metabolism test performed on all four tanks.
Metabolism tests involved placing samples of aquaria water and plant (when present) from each tank in clear
glass Winkler bottles and then determining oxygen evolution and consumption rates during alternative
periods of photosynthesis (20 minutes) and dark respiration (30 minutes). This was done by determining
oxygen concentrations in each bottle (using a WTW CellOx 325 combined DO and temperature probe)
88
before and after it was placed either within a photosynthesis chamber (equipped with two (2) Sun-Glo full
spectrum aquarium lamps) or in a dark respiration chamber. After the completion of each test the sample
water was returned to the respective tank and the oven-dry weight of the plant material determined. The tests
were performed in triplicate on samples collected from each tank every 1-2 days until practically all the
viable plant material had disappeared from tanks 2 and 4.
D4.3
Results
The results of the oxygen production/consumption tests are shown in Figures D4-1 and D4-2. These figures
show the mean and standard errors of the three replicate tests performed on each day. The plants contained in
tank 1 remained viable throughout the experiment and were actively producing new biomass from day one.
Some variation in the oxygen production and consumption rate plots can be seen (Figures D4-1 and D4-2
respectively) but averages of about 200 and 25 mg O2/minute/ kg of oven-dry plant material were obtained.
In tank 2, which was subjected to moderately hypoxic conditions, the plants survived for a period of 31 days
before they totally disintegrated. Prior to this the plants went through stages of yellowing (after about 10
days) and fragmentation (after about 20 days) indicating that they were under severe stress. The oxygen
production/consumption plots (Figures D4-1 & D4-2) show that the plant photosynthetic performance
decreased after about 10 days and gradually declined (with a slight recovery seen at 25 days) to about half
the peak after 31 days. Importantly it was found that the plant fragments (even after 31 days) were still
capable of significant photosynthesis and were able to recover if exposed to sufficient light. Respiration
stayed relatively constant during the initial stages of the experiment (up to 20 days) then gradually increased
in the final stages indicating that significant microbial decomposition was occurring.
In tank 4, which was subjected to severely hypoxic conditions, the plants survived for a period of 21 days
before they fragmented and then disappeared. The plants underwent the same visible stress reactions as the
plants in tank 2 but the timeframes were shorter. The yellowing started after only 7 days and fragmentation
occurred after about 13 days. A rapid fall in the photosynthetic response occurred after 13 days (see Figure
D4-1) with a net negative value recorded on the 21st day indicating that the respiration within the bottle
exceeded photosynthesis.
D4.4
From the results of this experiment we have been able to estimate in broad terms the length of time that
Ceratophyllum can remain viable when subjected to conditions of little or no light, and depressed oxygen
levels – conditions that occur regularly in freshwater floodplain habitats. This experiment was designed to
give us an understanding of the ability of this submerged plant species to survive such conditions. We
examined the photosynthetic and respiratory responses of plant samples, to brief periods of exposure to
bright light and total darkness, to determine the point at which plant viability was affected.
At 27oC in dark anoxic conditions this species of plant is able to survive without detectable loss of
physiological condition for about seven days. Signs of stress in the form of progressive yellowing,
fragmentation and reduced photosynthetic response developed from this time on. After 21 days practically all
plant material had either disappeared or decomposed to a severe degree and fragments were no longer
capable of recovery.
In dark and hypoxic conditions the plants maintained physiological condition for a period of 10 days, after
which time photosynthetic response dropped off and plants began to yellow. Plants began to fragment (after
20 days) and after 31 days little viable plant material remained. These conditions of low light and oxygen
availability are typical of those that occur in many local wetland systems particularly in deeper sections
and/or in the aftermath of runoff events. An important finding of this experiment was that Ceratophyllum
remains viable until almost all of its biomass has completely decomposed and is therefore capable of rapid
regeneration if sufficient light levels are restored in the interim.
89
Figure D4 – 1: Photosynthesis oxygen production of Ceratophyllum samples.
Oxygen Production, mg
O2/min/kg oven-dry plant
Tank 1: Normal Conditions
300
250
200
150
100
Good plant health
maintained throughout the
experiment
50
0
0
5
10
15
20
25
30
35
Day Number
Tank 2: Hypoxic Conditions
300
250
Plant expiry
after 31 days
200
150
100
50
0
0
5
10
15
20
25
30
35
30
35
Day Number
Tank 4: Anoxic Conditions
300
250
Plant expiry
after 21 days
200
150
100
50
0
-50 0
5
10
15
20
25
Day Number
90
Figure D4 – 2: Respiratory oxygen consumption of Ceratophyllum samples.
Oxygen Consumption, mg
Tank 1: Normal Conditions
60
50
40
30
20
10
0
0
5
10
15
20
25
30
35
25
30
35
25
30
35
Day Number
Tank 2: Hypoxic Conditions
60
50
40
30
20
10
0
0
5
10
15
20
Day Number
Tank 4: Anoxic Conditions
60
50
40
30
20
10
0
0
5
10
15
20
Day Number
91
APPENDIX E.
EXAMPLES OF CATCHMENT CLASSIFICATION OR TYPOLOGY
92
E1
DRAFT TYPOLOGY FOR RIVERINE WETLANDS IN THE BURDEKIN
CATCHMENT
Table E1.1 lists reaches that have been provisionally allocated to one of four major types.
E1.1
Type 1 Reaches
Most type 1 watercourses in the Burdekin generally rise in rainforested coastal ranges that generate large
quantities of exceptionally high quality runoff. Rainfall intensity, duration and frequency are far greater in these
upland subcatchments than in any other part of the Burdekin catchment. Precipitation is distinctly seasonal,
peaking during summer months, but even in the dry season there is usually enough rainfall to ensure that
headwater streams are virtually perennial and that flow spates are not rare. In low-order rainforest stream subtypes suspended particulate matter (SPM) concentrations streams can rise to 100-500mg/L very briefly during the
peak of a flood event but levels decline rapidly on the falling stage of the hydrograph, seldom exceeding 20mg/L
once rainfall ceases. SPM contains an unusually high proportion of organic (humic) material and appears to be
dominated by silt-sized particles which settle rapidly in calm waters. Soil throughflows contain exceptionally low
SPM concentrations (<1mg/L) and are low in nutrients (total nitrogen <150µg/L; total phosphorus <10µg/L).
Additionally, the productivity of receiving streams is light-limited (preventing phytoplankton growth) and bottom
substrata are very coarse-textured (i.e. contain few readily suspendable particles), consequently low flows are
often exceptionally clear, attaining underwater visibilities in excess of 20 metres.
93
Table E1.1: Preliminary classification of reaches based on hydrological regimes
Type
1
2
Characteristics
•
Baseflows from rainforest streams are
adequate to displace most lowland surface
runoff prior to stagnation
•
Numerous permanent waterholes (often
small)
•
Upper reaches virtually perennial
•
•
•
3
•
•
4
•
•
Ephermeral
Few permanent waterholes
Insufficient baseflow to displace surface
runoff from semi-permanent and ephemeral
waterholes
Streams contain large permanent pools
(instream waterholes and/or off-channel
lakes) that are maintained by ground water
Two distinct classes
a) Baseflow adequate to displace surface
runoff
b) Baseflow inadequate to displace surface
runoff
Artificially supplemented baseflows create
perennial flow conditions
Includes irrigation water distributaries and
irrigation tailwater drains.
Location/Reach
North-East
•
Upper Burdekin (above Valley of Lagoons)
•
Michael Ck
•
Running R.
•
Star R.
•
Keelbottom Ck
South-East
•
Broken R.
•
Urannah Ck.
•
Grant Ck.
•
Emu Ck.
North
•
Dry River
North-East
•
Douglas Ck
Western
•
Gray Ck
•
Clarke R.
•
Maryville Ck
•
Upper Lolworth Ck.
South-East
•
Broughton R.
•
Rollston R.
•
Campaspe R.
•
Upper Cape R.
•
Warrigal Ck
•
Amelia Ck.
North
•
Wyandotte Ck?
•
Reedy Brook Ck
•
Valley of Lagoons
•
Fanning R.
West
•
Upper Broken (Clarke) R.
•
Basalt R.
•
Clarke River (headwaters only)
•
Lower Lolworth Ck
•
Toomba Ck
•
Fletcher Ck
•
Lower Cape R.
South
•
Lower Belyando R.?
•
Lower Mistake Ck?
•
Lower Suttor R.?
•
Burdekin R. below BFD
Coastal Floodplain
•
Sheepstation Ck
•
Plantation Ck
•
Saltwater Ck (Inkerman)
•
Cassidy Ck
•
Barratta Ck.
94
The lowland reaches of the Type 1 tributary streams listed in Table E1.1 benefit tremendously from the
dilution, dispersion and flushing capacity provided by flows from the low-order rainforest streams in their
headwaters. Rains in the lowland catchment areas are brief and infrequent, therefore the poorer quality runoff
from these watersheds is almost inevitably displaced by the much more persistent flows of high quality water
from the mountains. Channel scouring and bank erosion in the lower reaches cause increases in turbidity and
nutrient concentrations but water quality is still far better than in most other inland streams. In the aftermath
of very large-scale rain events (eg the exceptionally wet months associated with Cyclone Joy) baseflows can
be sustained for a year or more, but in most years surface flows cease by the middle of the dry season.
In these more typical years lowland stream reaches fragment into a series of relatively small waterholes that
often contain relatively high quality water. Permanent water holes of this kind are the only reliable refugia for
aquatic life during prolonged dry spells and provide habitat types that are poorly represented in the main river,
hence they have high conservation significance.
These types of streams are very efficiently scoured by storm flows which leave behind clean bottom sediments
that are seldom finer than medium sand and which contain little bioactive organic matter of any kind – living
or dead. Hence events of this kind ostensibly reset the aquatic ecosystem and mark the beginning of a
recovery period that involves numerous successional changes in both the ecological community and water
quality. Recovery may take many months to several years depending on the severity of the flood event and the
length of the ensuing dry spell. The key features of these successional changes are summarised in the
following points:
•
•
•
•
•
•
•
•
•
•
•
Water quality is generally at its best when pools first become isolated.
Instream benthic biomass and productivity are minimal after a major storm flow.
Salinity levels rise over time but changes are independent of evaporation suggesting that there are still
significant flows of water through the basal sands. This indicates that the waterholes can have much
higher dilution and dispersion capacity than their size might otherwise suggest.
During prolonged periods of baseflow fine sediments (which form a thin layer over the coarse basal
sands) and coarse particulate organic matter (eg leaf litter) are deposited on the benthos, leading to the
gradual establishment of benthic heterotroph communities (microbes, macroinvertebrates, etc.).
Over time decreased water depth and increased clarity gradually enable benthic autotrophs (plants) to
become established.
The quality of water in isolated pools is governed by instream biophysical processes that are driven by
very localised, site-specific factors. Consequently, consecutive waterholes in the same stream reach can
develop strikingly different water quality characteristics if they are of different sub-types.
Water quality may be adversely affected by small, localised events and especially congregations of
livestock and other animals. The probability of such events occurring increases substantially with time
until flows are re-established.
Waterholes usually become isolated while the catchment is still relatively green and numerous ephemeral
drinking sources are still available to terrestrial animals, hence animals tend to be dispersed widely
During dry spells waterholes gradually lose dilution capacity and accumulate biomass. Re-aeration rates
fall to low levels, reaching a minimum in sub-types characterised by low surface area to volume ratios
and/or low wind exposure. Even in sub-types with high surface areas, windy conditions and significant
wind fetches are often required to achieve re-aeration rates high enough to compensate for respiratory
oxygen consumption. At the same time oxygen consumption rates approach a maximum, hence dissolved
oxygen concentrations are often low and are heavily reliant upon the production of oxygen by submerged
plants. This makes the waterholes naturally susceptible to acute impacts from instream disturbances.
(For example resuspension of bottom sediments by wading animals can rapidly increase turbidity to the
point where in-stream oxygen production ceases and aquatic animals die of asphyxiation).
The frequent prolonged dry spells that increase the vulnerability of aquatic habitats to localised
disturbances also affect the catchment. Pasture grasses dry out, (increasing the water requirements of
livestock), temporary drinking water supplies disappear and terrestrial animals are forced into riparian
areas in the vicinity of permanent watering points. This results in an exponential increase in livestock
pressure around permanent waterholes at the time of the year when the water bodies are least able to cope.
Waterholes are particularly vulnerable to the development of water quality problems during hot, dry
summer months (due to high productivity rates and thermal stratification of the water column). These
conditions also increase the water requirements of animals so there is an attendant increase in livestock
pressure.
95
•
Rain events that occur before baseflows have been re-established can induce major, often catastrophic,
water quality deterioration because waterholes retain very poorly diluted runoff from the local watershed.
This is a major source of between-site water quality variations in– the probability of such an event
occurring at particular sites must be taken into consideration when conducting between site comparisons.
Intrinsic differences in natural runoff quality (dependent on catchment topography and lithology etc) must
also be taken into account, but in practice some factors are more important than others. For instance the
proportion of dispersible clays and colloids in catchment soils can have a larger bearing on turbidity
outcomes than the total erosion potential (silts and coarser suspended sediments settle, so turbidity does
not persist if these are the dominant suspensoids contained in runoff). For example, available data
indicate that conventionally tilled farms are a major source of SPM losses (i.e. a major source of sediment
export), while trash blanketed cane-fields have much lower total sediment yields. However, trash
blanketed fields retard runoff and preferentially yield fine dispersible sediment. Consequently runoff
entering receiving waters from blanketed farms on the tail of the hydrograph can (perhaps ironically)
contain higher and more persistent turbidity levels than runoff from tilled farms. Similarly the potential
for serious oxygen sags to result from effects other than turbidity changes is dependent on the amounts of
bio-available organic matter in the catchment soils.
•
These problems can occur in pristine environments but are much more common and severe in grazing
areas especially when livestock and/or feral animals are allowed to congregate in the riparian zone, and in
irrigated farming areas where tailwaters are discharged into sluggish or stagnant streams.
The total quantities of sediment and nutrient retained within tributary streams and waterholes are
negligible compared to the quantities that pass through them during high flow periods, and the amounts of
contaminant required to induce major episodic or periodic water quality problems in vulnerable
waterbodies are even lower. This is particularly true of livestock excrement which exerts significant
oxygen demand and only needs to be present in very small amounts to exacerbate natural hypoxia
problems.
•
The maintenance of natural aquatic ecosystems in these streams is reliant on the provision of swift flows with
adequate power to reset the ecosystem (by flushing out biomass and fine sediments), and frequent, relatively
prolonged inputs of high quality baseflows from the rainforested headwaters. Regulation of the upstream
reaches of such systems clearly has the potential to adversely affect both of these attributes. The only existing
dams of this type in the Burdekin catchment (Paluma and Eungella) are relatively small and are located high
enough in the system to ensure that measurable impacts on natural stream hydrology are confined to a few
relatively small stream reaches. It would however, be extremely difficult to construct larger dams on higher
order streams in these systems without significantly impacting upon downstream reaches. Even if releases
from the dam were used to artificially supplement baseflow it is highly unlikely that the quality of the
impounded water would be high enough to serve as an adequate substitute for natural flows. The dam waters
would almost certainly develop thermal, oxic and chemical stratification during summer months. Surface
waters would be exposed to sunlight and would be significantly warmer than normal for shaded rainforest
streams. They would be much more biologically active and would contain elevated concentrations of
plankton, nutrients and potentially, cyano-bacteria. Deeper waters would almost certainly be anoxic and
contain even higher concentrations of bioavailable nutrients, organic carbon, dissolved metals and potentially,
viable phytoplankton. Such water would not offer the same dilution capacity as natural baseflows, and would
be likely to stimulate entirely unnatural ecological processes, and consequently, further water quality changes.
In Type 1 lowland streams, long term ambient water quality is largely dependent on the condition of the upper
catchment and local riparian environments; the condition of the greater catchment is much less relevant in this
regard. Stormwaters on the other hand, can originate from any watershed in the catchment so during storms
water quality depends more on overall catchment condition. The vast majority of the contaminants exported
to downstream environments are mobilised during brief but intense storm events. These flow through the
receiving stream so rapidly that they are seldom of consequence to the freshwater ecosystem, provided that
acutely harmful conditions do not develop. (Acute episodic water quality effects are too complex to
summarise here, but are discussed in other sections of this report). In fact it is common to find that stream
reaches that maintain good water quality most of the time, do so largely because the contaminants contained
in poor quality stormwaters are flushed away so efficiently. Hence many of the tributaries and reaches that
have been identified as high risk with respect to their capacity to transport fine sediments to the sea (for
example by CSIRO’s Sednet model) have a somewhat paradoxical tendency to maintain good water quality
most of the time (except briefly during storm events).
96
E1.2
Type 2 Reaches
Type 2 streams sustain negligible baseflows and consequently contain very few permanent or even semipermanent aquatic habitats. A few drought-resistant waterholes, which are presumably surface expressions of
locally developed ground water tables, persist in a few places (for example two locations on the Rollston
River and four in the lower Campaspe River). The importance of these as refugia for aquatic organisms is
unknown, however, they are rare aquatic habitats and arguably deserve protection on that basis alone. The
water quality of such habitats has not yet been studied in detail but it is currently assumed that these few small
reaches will behave similarly to Type 3b streams.
In many respects Type 2 streams function as biologically inert drains that intermittently discharge pulses of
quite concentrated stormwater from their respective subcatchments. The ways in which streams of this type
contribute to downstream water quality are dependent on the timing and intensity of rain events in their
respective catchments relative to others.
Most of the larger Type 2 streams drain relatively dry catchment areas and do not flow as frequently as Type
1 systems. Many drain watersheds with bare erodible soils that are subject to severe gully erosion, so the
quantities of sediment discharged when they do flow, can be substantial. On rare occasions, especially during
drought years, individual inland catchments of this kind may receive localised rainfall while the rest of the
catchment remains dry. In such cases discharges from type 2 streams enter the river while it is at low
baseflow and poor quality stormwaters may remain in the river and estuary long enough to affect ambient
water quality and ecosystems condition. However, Type 2 streams are much more likely to flow during wet
season months when other wetter catchment areas are already generating swift flows. Historically,
contaminants delivered to the river system at such times would have been flushed downstream to the sea quite
rapidly. There is little doubt that this still occurs in the Bowen/Broken catchment but the fate of the
stormwater contaminants that currently enter the river upstream of the BFD is less certain.
Much of the sediment and nutrient mobilised by small/localised storm events, especially those that occur
when the dam is below FSL, will remain in the freshwater system. These will contribute to the persistent
turbidity in the dam (which tends to act as a very large type 3b waterbody). However, our observations
indicate that the vast majority of the fine sediment discharged during large storm events pass through the dam
without being trapped. This contention is not currently supported by CSIRO’s Sednet model which predicts
that 90% of the fine sediment carried into the dam will be trapped there. (This model has not yet been
calibrated for the Burdekin and the authors note that the trapping capacity of the dam may have been over
estimated). It will not be possible to confidently identify the main sources of sediment and nutrient exports to
the marine environment until this discrepancy is resolved.
E1.3
Type 3 Reaches
Type 3 stream reaches have not been as well studied as Type 1 and are considerably more diverse.
Consequently, many sub-types will need to be identified. They include a few streams such as Toomba Creek
and the lower parts of Lolworth Creek that receive sufficient inputs of ground water (from the Chudleigh
Basalts) to flow perennially; however, most systems of this type eventually stop flowing and retract into a
series of permanent instream waterholes and/or off-channel lakes. The quality of water retained within such
systems depends on the capacity of baseflows to dilute and disperse surface runoff which, because it is
generated from (often disturbed) lowland catchments, is generally turbid and of relatively poor quality
compared to other potential water sources.
Type 3a systems receive sufficient baseflow to displace stormwater runoff before surface flows cease, hence
water quality directly reflects the quality of the aquifers from which baseflows originate. When the water
comes from basalt (or limestone, as is the case in the upper reaches of Broken River in the Clarke River
catchment) this water can be expected to be quite clear and moderately saline by local surface water standards
(conductivity 250 to 800 µs/cm). These waters have some capacity to slowly flocculate fine sediments and
this helps reduce the persistence of the turbidity that inevitably results if bottom sediments are disturbed. By
contrast, in cases where flows originate from lower salinity sources, (eg shallow alluvial aquifers) the water
starts out clear but can rapidly become turbid and remain that way for long periods if bottom sediments are
disturbed (for example, by wading livestock).
Low baseflow (Type 3b) systems such as upper Fanning River and the lower reaches of Cape River (with its
associated floodplain wetlands) are not capable of completely displacing surface runoff and hence seldom
97
achieve the very high water quality levels that are evident in many Type 1 systems at the beginning of the dry
season. Nevertheless, ground water throughput can be sufficient to gradually dilute and/or exchange the
water retained in permanent waterholes so that by the end of prolonged dry spells, they can potentially contain
better quality water than many Type 1 systems.
Waterholes in Type 3 systems experience post-flood successions that are similar to those discussed for Type 1
systems, but the following differences exist:
•
•
•
•
Type 3 waterholes generally have superior ground water throughput. This provides higher dilution
making them less vulnerable to adverse effects from localised contaminant inputs and/or instream
disturbances.
Channel scouring is not usually as efficient as it is in Type 1 and 2 systems so significant benthic
biomass and fine sediment may still be present after storms. Interactions between water quality and biota
can therefore be significant even during the falling limb of the storm hydrograph.
Background water quality (i.e. in the absence of instream and riparian influences) is subject to greater
spatial variability than equivalent Type 1 systems due to the natural diversity of ground water sources.
By contrast, long term temporal variations tend to be less pronounced because biophysical conditions do
not fluctuate as severely as they do in many Type 1 systems.
In situations where waterholes are sustained entirely by ground waters, it is not uncommon for salinity to
rise to levels as high as 1500 µS/cm. (In rare cases isolated waterholes in the vicinity of mineralised
ground water formations may develop conductivities in excess of 6,000 µs/cm. However, effects are
very localised and will not be discussed here.) Our field observations suggest that conductivity levels up
to 1500µs/cm do not greatly affect the aquatic fauna in this region. However, it does seem likely that
salinity changes of this kind would play a significant role in determining which aquatic plant species
become dominant. (Note that regional survey data often show strong correlations between conductivity
and macroinvertebrate community structure, but our research suggests that this is an artefact created by
the even stronger correlations between flow and conductivity). There is also usually a concomitant
increase in water hardness whenever salinities rise, and (because hardness promotes flocculation)
sediment settling rates often increase accordingly. The resulting improvement in water clarity stimulates
plant growth, further complicating the relationships between salinity and plant community dynamics.
Since they occur mainly in drier catchments, Type 3 systems can generally be expected to export stormwaters
to downstream environments less frequently and efficiently than Type 1 streams.
The Belyando/Suttor system is not sufficiently well understood to be able to confidently propose a water
quality model at this stage. It is understood that due to its low relief and the existence of extensive inland
floodplains, this system has a much less peaky hydrograph and significantly lower contaminant transport
capacity than the Burdekin River. The system is known to be chronically turbid but it is uncertain if this is a
natural condition. We suspect that it may be, because the combined effects of low sediment transport capacity,
low baseflow and low conductivity water (all natural features) would be expected to create conditions where
fine sediments are not flushed from the system and cannot settle (due to lack of flocculation). For similar
reasons, we speculate that the Belyando River may be much more vulnerable to certain anthropogenic impacts
than most rivers in the Burdekin catchments. (The chronic turbidity of the Belyando system has often been
attributed to the existence of dispersible clays and colloids in catchment soils but our research indicates that
the composition of suspended sediments entering the BFD from the Burdekin and Bellyando/Suttor arms of
the impoundment on the tail of storm hydrographs are remarkably similar, and that the total concentrations in
the Burdekin arm are sometimes higher than the Belyando arm – this evidence supports the contention that
detention of stormwater is a more important determinant of ambient water quality outcomes than stormwater
quality per se)
At this stage it is appears that most aquatic habitats in this catchment would be Type 3b systems that are
maintained by very low throughputs of low salinity ground water, but this assumption needs to be tested. The
lower reaches of the Belyando River, Mistake Creek and Suttor River support an extensive network of large
permanent waterholes and off-channel wetlands. These aquatic habitats in particular demand research
attention.
E1.4
Type 4 Reaches
Type 4 systems comprise waterways that were once mixtures of Types 1, 2 and/or 3, but which now maintain
artificially elevated baseflows due to constant inputs of irrigation water from the BFD. The Burdekin River
98
main channel and irrigation water distributaries on the floodplain, exhibit water quality characteristics that are
uniformly similar to the BFD. The water is reasonably good in quality except that it is usually undesirably
turbid, but the lack of spatial and temporal water quality variability has contributed to a loss of aquatic habitat
diversity. The ecological condition of the main channel has not been examined in great detail but current
indications are that the artificial water quality regime is capable of sustaining a functional, albeit significantly
altered, ecosystem.
This is not the case in floodplain distributaries such as Sheepstation and Plantation Creek. Here the aquatic
habitats created by the permanent presence of flowing water have been invaded by massive infestations of
floating and emergent aquatic weeds. These have severely de-oxygenated the water, effectively eliminating
many dissolved-oxygen-dependent animal species. It is likely that weed invasions of this kind would have
occurred regardless of the quality of the water being introduced into these systems (due to the constant
presence of water), but there is little doubt that the chronic turbidity of the BFD water has enhanced the
competitive success of the invading weeds (all of which are emergent species that can obtain light regardless
of water clarity). Moreover, even though the concentrations of nutrients in BFD water are quite moderate,
constant inputs of the water provide a virtually inexhaustible nutritional supply for aquatic weeds, making
them much more difficult to manage. Experimental projects have demonstrated that water quality suitable for
the maintenance of most local aquatic species can be restored if weeds are removed, but it is yet to be
determined if these improvements can be maintained in the long term.
Notably the hypoxic waters in the rehabilitated parts of these systems rely quite heavily on flow to provide reaeration, and stressful conditions can develop if flows stop. This is a significant management issue because
flows currently fall to a minimum when irrigation demand drops off during the onset of rainy weather. At
such times cloudy weather also reduces instream oxygen production by plants, so oxygen sags can occur.
These might be avoided or at least minimised, by providing an environmental flow allocation to maintain reaeration until adequate stormwater flows are achieved.
99
APPENDIX F
REVIEW OF PUBLICATIONS RELATING TO THE FACTORS THAT
CONTRIBUTE TO RE-AERATION AND MIXING IN WETLANDS AND
THE MODELING AND ESTIMATION OF OXYGEN FLUXES
100
F.
REVIEW OF PUBLICATIONS RELATING TO THE FACTORS THAT CONTRIBUTE
TO RE-AERATION AND MIXING IN WETLANDS AND THE MODELING AND
ESTIMATION OF OXYGEN FLUXES
F1.
INTRODUCTION
The world literature on the topic of re-aeration and mixing is basically divided into two categories that
deal with distinctly different waterbody types. On one hand are the lotic (or flowing) systems such as
rivers and streams that are very well mixed that do not stratify readily and are generally very efficient at
distributing surface oxygen through the water column. The re-aeration of these systems is primarily a
function of flow rate, depth and bed slope. The other category are the lentic (or non-flowing) systems
such as lakes, lagoons, reservoirs and palustrine wetlands that are not so efficient at oxygen distribution
and which have to rely primarily upon wind and wave action to distribute oxygen to deeper waters. They
stratify readily, particularly in areas with high solar irradiance levels and low wind strengths, creating
physical barriers (eg. thermoclines or haloclines) to the movement of dissolved oxygen through the water
column. In both cases the re-aeration rate (across the air/water interface) is governed by the same
mechanism and is basically a function of the difference between the dissolved oxygen concentrations in
the surface film and the saturation level. However, it is the transport of this dissolved oxygen away from
this surface film to deeper waters that determines the oxygen regime of the water body and the factors
that affect this transport (or mixing) is the main focus of this review.
F2.
Lotic Systems
Atkinson, J.F., Blair, S., Taylor, S. and Ghosh, U. (1995) Surface aeration. In Journal of
Environmental Engineering, 121, No. 1, pp. 113-118.
An analytical approach to calculating the surface oxygen flux is presented that enables the modeling of
DO in a density-stratified river. A two-dimensional (vertical plane) model is used that is based on a twofilm approach that describes the mass transfer process at the air-water interface. This model assumes the
presence of two adjacent laminar films in intimate contact (ie. air and water) with the gas transfer being
governed by diffusion within these films and relatively well-mixed turbulent zones on either side of these
interfaces (see figure below).
where: Kh = Henry’s constant;
δ = film thickness.
The method presented is based on the assumption that the mass flux of oxygen is proportional to the DO
deficit within the water body, ie.
J = KL(Cs – C)
where: J = mass flux;
C = DO concentration;
Cs = saturated DO concentration; and
KL = bulk mass-transfer coefficient.
For a mixed water body of depth H, the rate of change of DO due to atmospheric flux is,
101
dC
A
= KL
(C s − C )
dt
Vol
where: A = surface area;
Vol = volume = AH.
H = water body depth.
This method estimates a value for the liquid film thickness δ, because turbulence controls gas transport
and the point at which turbulence just starts to be important, compared with molecular diffusion
determines the critical film thickness. So the re-aeration coefficient is given by:
K2 =
Dm
K
= L
δH
H
where: Dm = molecular diffusivity for DO
and the film thickness, δ is estimated to be:
⎛ν 3 H ⎞
δ ≅ A2 ⎜⎜ 3 ⎟⎟
⎝ U ⎠
1
4
where: ν = kinematic viscosity;
U = mean velocity; and
A2 = a derived co-efficient.
Using these equations it is possible to estimate the re-aeration coefficient K2, and then calculate the bulk
mass-transfer coefficient KL, which can be used to model the mass flux of oxygen across the air-water
interface. So in principle the larger the water film thickness (which is basically a function of depth and
mean flow velocity) the lower the flux of oxygen either into or out of the water body.
Wilcock, R.J., Nagels, J.W., McBride, G.B.Collier, K.J, Wilson, B.T. and Huser, B.A. (1998)
Characterisation of lowland streams using a single-station diurnal curve analysis model with continuous
monitoring data for dissolved oxygen and temperature. N.Z. Journal of Marine and Freshwater Research,
32, pp 67-79.
A single-station diurnal curve model, DOFLO (Dissolved Oxygen at Low Flow) was tested on DO and
temperature data sets collected over several days from 23 lowland streams in New Zealand. The model
produced reach-average values for:
• Reaearation coefficient at 20oC, K2(20);
• Maximum daily rate of photosynthesis production of oxygen at 20oC, Pmax;
• Daily respiration rate at 20oC, R20; and
• Ratio of respiration rates10oC apart, Q10.
Additionally, 24 hour average estimates for the ratio P/R were also calculated for the sites.
The range of values calculated for these stream were:
• K2(20) of 0.05-40 d-1 and median of 6.0 d-1;
• Pmax of 1.75-86.5 g m-3 d-1 and gross primary production over a day of 0.5-29.2 g m-2 d-1
(calculated from Pmax);
• R20 of 3.5-55.0 g m-3 d-1; and
• P/R ratios of well below 1 representing heterotrophic systems with higher amounts of
respiration compared to photosynthesis.
Diurnal curve analysis using DOFLO yields parameter values that describe existing stream conditions at
the time of monitoring and also quantify the stream’s potential diurnal DO variation. The DOFLO model
assume the diurnal DO is affected by three fundamental processes: re-aeration, plant and bacterial
respiration and photosynthesis, ie.
102
dC
= K 2 (C s − C ) + Pi − Ri
dt
The model assumes homogeneous conditions extending upstream for a distance of at least:
3U
where:
K2
U = mean stream velocity (m s-1) in the reach;
The re-aeration coefficient K2 is calculated from equations that are a function of mean water depth (H)
and mean water velocity (U) ie. K 2 ( 20) = 5.61
U
H3
The authors suggested that classification of the streams using cluster analysis of the three DOFLO
calibration parameters (K2(20), Pmax, R20) utilizing experimental variables such as DOmin and Tmax (see
diagram below) gives a method of grouping sites according to similarity of processes affecting their DO.
The clusters identified and the features of these systems were:
Cluster 1: large diurnal DO ranges with very low re-aeration coefficients (average K2(20) values of 1.1 d1
);
Cluster 2: low re-aeration rates and low values of Pmax and R20;
Clusters 3 and 5: high respiration rates balanced by moderate to high re-aeration coefficients;
Cluster 4: high re-aeration coefficients and low productivity and respiration rates.
Charpra, S.C. and Di Toro, D.M. (1991) Delta method for estimating primary production, respiration
and re-aeration in streams. Journal of Environmental Engineering, 117, 5, pp. 640-655.
The delta method is a tool used to estimate stream re-aeration, primary productivity and respiration rates
on the basis of diurnal dissolved oxygen measurement. Features of the diurnal curve are used to
determine (by a graphical method, see figures below) the key DO related parameters detailed below:
• The time of minimum deficit (Dmin) (relative to solar noon) is used to estimate the re-aeration rate;
• The deficit range (∆) is used with the reaearation rate to predict photosynthetic production; and
•
_
The average deficit ( D ) can be used in conjunction with the re-aeration rate and production rates to
compute respiration rates.
103
Plant primary productivity and respiration
(mg L-1 d-1) versus time.
Dissolved oxygen deficit
(mg L-1) versus time.
Dissolved oxygen concentration (mg L-1)
versus time.
A series of graphs (detailed in the
reference) based on several Fourier series
solutions are then used to calculate the re-aeration and photosynthesis rates for the stream, while the
equation:
_
R= Pav +ka D ,
is then used to calculate the respiration rate.
A serious limitation of the method is that it is assumed that photosynthesis can be adequately represented
as a half-sinusoid. As such this method is not appropriate for streams that are shaded for any time during
the day due to overcast conditions and/or riparian/bank shading. Other assumptions are that stream
temperatures are constant and longitudinal oxygen gradients are negligible (ie. a uniform distribution of
plants upstream of the diurnal monitoring station). The assumption of constant temperature and therefore
respiration is entirely unrealistic in the dry and wet-dry tropics.
104
McBride, G. (2002) Calculating stream re-aeration coefficients from oxygen profiles. Journal of
Environmental Engineering, 128, 4, pp. 384-386.
This paper builds on the work of DiToro (1991) by presenting an alternative method of determining the
re-aeration coefficent ka(20). A logistic curve is fitted to the numerical solution developed by DiToro that
enables direct calculation of the stream re-aeration rate, ie:
k a ( 20)
⎡ 5.31η − φ ⎤
= 7.48ϕ ⎢
⎥
⎣ ηφ ⎦
0.85
withϕ = 1.0241
20 −T
⎡f ⎤
andη = ⎢ ⎥
⎣14 ⎦
0.75
where: ϕ = temperature correction factor;
η = photoperiod correction factor;
φ = time lag of minimum deficit relative to solar noon (hr);
f = photoperiod (hr); and
T = water temperature (oC).
This equation can be used to estimate the stream re-aeration coefficient given only the time lag,
photoperiod duration and daily average water temperature.
Wilcock, R.J. (1988) Study of river re-aeration at different flow rates. Journal of Environmental
Engineering, 114, 1, pp. 91-105.
A study was performed in a fast-flowing, turbulent river in New Zealand to assess the effect that varying
flow rates had on the measured re-aeration coefficient. A gas tracer method (methyl chloride with
rhodamine WT used to correct for dispersion and other dilution effects and to calculate time of travel)
was used to measure the re-aeration coefficient (K2) using a number of equations that are temperature
dependent. The figure shows how the re-aeration coefficient varies with temperature for this river. The
upper curve is fitted to measurements made at ambient temperature and the lower curve is fitted to data
corrected to 20oC.
The equation of this lower curve is:
K2(20) = 22.4Q-0.324
where Q is the river flow rate.
105
G3.
Lentic Systems
Imberger, J. (1987) Hydrodynamics of lakes, Australian Water and Wastewater Association, Technical
paper, pp. 401-423.
A brief review of the dynamic processes operating in a lake or reservoir is given that is focused on the
factors that influence their thermal structure. Australian lakes generally stratify in spring and over
summer since the potential energy loss induced by the heating at the surface far exceeds the mechanical
energy introduced by the wind. Seasonal stratification in lakes can be disturbed in several ways,
including:
• The diurnal mixed layer (surface layer) is influenced by the daily weather changes in temperature.
Vertical mixing is enhanced by the surface waters cooling during the night and plunging to a depth of
neutral buoyancy thus deepening the surface layer (the layer in which there is active mixing) until
solar heating re-commences early in the morning.
• Barotropic and baroclinic waves are produced when wind stress at the surface is transmitted through
the mixed layer and acts to set up a series of basin seiches or oscillations. On the commencement of a
wind event the water body is deformed and must suddenly resist a surface stress. This disturbance is
transmitted through the mixed layer and may cause a range of different waves (barotropic/baroclinic,
Kelvin, Poincare, shelf and surface waves). These waves set the fluid into motion so that, on average,
the stresses imposed by the wind are balanced by friction and advective inertia, ie. a steady
circulation is generated.
• For strong winds seiche oscillations may actually cause water from some depth to surface or upwell
(see figure below). Wind induced thermocline tilting occurs and as it proceeds the thermocline
spreads at the upwind end and compresses at the downwind end. At the upwelling end water from
some depth is exposed to the turbulence induced by the wind, which mixes it with the remaining
surface water.
106
•
•
•
The heat capture at the surface of a lake is often uneven causing differential heating and cooling to
occur. This can cause longitudinal temperature gradients to occur, particularly in shallow sidearms
and embayments were waters change temperature more rapidly due to their smaller depth and
volume. These gradients can cause sizable water exchanges to occur between these shallow zones
and the main body of the lake.
The differential adsorption of the waters caused by variations in clarity can cause patches of warmer
water surrounded by cooler water (see figure below). This causes mixing to occur whereby deeper
water is constantly circulated to the surface.
Other circulations include: differential deepening of the thermocline by inconsistent wind stresses,
internal waves that propogate randomly throughout the density stratified portions of the water
column, internal mixing in the hypolimnion, boundary mixing close to the extremities of the lake and
intrusion, inflow and outflow generated disturbances.
All these compounding factors interact to produce a very complex mixing regime that is impossible to
fully predict.
Florida Lakewatch (2001) A beginner’s guide to water management-lake morphometry. Department of
Fisheries and Aquatic Sciences, University of Florida, Part 4, pp. 20-28.
107
The movement of water (or mixing), caused by wind effects, is explored as well as the role of lake
morphometry. Wind travelling across a water surface creates underwater currents that tend to move water
particles horizontally through the water column. At the same time, water particles are being distributed in
an irregular swirling motion known as turbulence. When waves are created they cause circular movement
of water particles (see figure below).
The radius of the orbits gets smaller as they move downward and become negligible at a depth of about
one-half the wavelength (mixing depth) or Zmix=0.5L. The longer the fetch length the greater the
wavelengths and wave heights and hence greater the mixing depth. Similarly, the greater the wind speed
the greater the mixing depth. There is potential for bottom sediment resuspension if the mixing depth is
greater than the water depth. The table below was presented that can be used to determine the mixing
depth (ft) given a fetch length (ft) and wind velocity (miles/hr). This table was produced using the wave
theory above and standard engineering equations.
A quick method of estimating the impact of wave mixing on a whole lake was also presented which
involves the calculation of the dynamic ratio,ie,
Dynamic ratio (DR) = square root of the lake area (km2)/lake mean depth (m).
Lakes with DR >0.8 were subject to wave disturbance at all areas of the lakebed at least some of the time
while lakes with DR<0.8 showed linear decrease in areas disturbed at one time or another. For example if
108
the ratio were 0.4, only about 50% of the lakebed would be disturbed at one time or another. This
relationship was found to occur during a study in Florida lakes and it is not clear whether it would be
valid elsewhere and in smaller waterbodies. Also the calculation assumes that the lake does not contain
significant amounts of aquatic plants (which would reduce the degree of disturbance by wave action).
Robertson, D.M., Imberger, J. and Boland, K. (1990) Lake number, an indicator of reservoir mixing: a
water quality management tool. Water, pp. 29-33.
The lake number (LN) approach was developed to enable an easily measured parameter to be calculated
that reflects the dynamic physical behavior of density-stratified lakes and reservoirs and which can be
used to predict changes in specific water quality parameters. It is defined as the ratio of moments about
the center of volume of the water body of the stabilizing force of gravity to the destabilizing force
supplied by wind stress. Changes in the lake number correlate to changes in deep water dissolved oxygen,
iron and manganese concentrations and surface chlorophyll concentrations in two Australian (Western
Qld) and one USA lake. The equations used to calculate LN are as below.
⎛
Z ⎞
gS t ⎜⎜1 − t ⎟⎟
⎝ Zm ⎠
LN =
⎛ Zg
ρ o u* 2 Ao 3 / 2 ⎜⎜1 −
⎝ Zm
⎞
⎟⎟
⎠
where: g = gravity;
St = Schmidt stability;
Zt = thermocline height from bottom;
Zm = maximum depth of water body;
ρo = water density at surface;
u* = friction velocity;
Ao = surface area;
Zg = height to center of volume of reservoir;
St = ∫
Zm
0
(
(z − Z )* A( z ) * ρ ( z ) * dz
g
u* = 1.612 × 10 −6 * U w
)
where: z = depth from bottom;
A(z) = area of the lake at depth;
ρ(z) = density of the water at depth z.
2 1/ 2
where: Uw = wind velocity at 10m above the water surface.
A LN of 1 indicates that the wind is just sufficient to force the thermocline to be deflected to the surface at
the upwind end of the lake. For LN >> 1 the stratification will be strong and dominate the wind stress. For
LN <<1 the stratification will be weak and mixing through the water column will occur.
Hodges, B.R., Imberger, J., Laval, B. and Appt, J. (2000). Modeling the hydrodynamics of stratified
lakes. Hydroinformatics 2000 Conference, Iowa Institute of Hydraulic Research, 23-27 July 2000.
The wind, surface thermodynamics and inflow dynamics provide the energy sources for transport,
turbulent kinetic energy (TKE), internal waves and mixing. A conceptual type model is shown below that
encompasses these transport factors associated with wind energy.
The energy from the wind and thermodynamics directly influences the TKE and stratification in the upper
layer (epilimnion) while indirectly influencing the mixing across the thermocline (metalimnion) through
entrainment and generation of internal waves.
109
Stefan, H.G., Fang, X., Wright, D., Eaton, J.G., and McCormick, J.H. (1995). Simulation of
dissolved oxygen profiles in a transparent, dimictic lake. Limnology and Oceanography, 40(1), pp. 105118.
A study was performed on a small (6.6 ha), highly transparent and wind sheltered lake in northeastern
Minnesota, called Thrush Lake. A deterministic one-dimensional model was developed to simulate
vertical dissolved oxygen profiles on a daily basis. The model uses the standard re-aeration equation for
calculations of oxygen fluxes and uses the equation of Wanninkhof (1991) for calculations of the bulk
surface oxygen transfer velocity, ke, ie:
ke = 0.108U101.64(600/Sct)0.5
where: U10 = wind speed 10m above the lake;
Sct = Schmidt number of oxygen at the surface water
temperature = µ/ρ*Dv;
µ = water viscosity;
ρ = water density;
Dv = diffusivity.
Hence the rate of oxygen transfer is a function of these parameters as well lake morphometry, lake
trophic state and weather parameters (air temperature, dew point, solar radiation, windspeed and sunshine
percentage) and are inputs into the simulation model presented.
Ambrosetti, W. and Barbanti, L. (2002). Physical limnology of Italian lakes. Relationships between
morphometric parameters, stability and Birgean work. Journal of Limnology, 61(2), pp. 159-167.
The relationship between morphometric parameters with maximum stability and maximum Birgean work
(quantity of external energy required to produce a certain distribution of density) are analysed for 31
110
Italian lakes. It was shown that in deep lakes the depths are more correlated with the stability and fetch
length is more correlated with work. In shallow lakes the depths are correlated with both stability and
work. Parameters that were considered in the study were: volume, area, maximum depth, mean depth,
fetch area (Area1/2), effective fetch (maximum effective fetch + maximum effective width/2), depth of
epilimnion, depth of thermocline and other stability related terms (total, chemical, thermal).
111
APPENDIX G
IMPORTANT BACKGROUND INFORMATION ABOUT DISSOLVED
OXYGEN AND ITS BEHAVIOUR IN DRY AND WET-DRY TROPICAL
WETLANDS
112
G1.
IMPORTANT BACKGROUND INFORMATION ABOUT DISSOLVED
OXYGEN AND ITS BEHAVIOUR IN DRY AND WET-DRY TROPICAL
WETLANDS
Reduced dissolved oxygen (DO) availability is by far the most important and widespread water quality
issue in tropical wetlands, especially in dry and wet/dry catchments. These systems are naturally prone to
the development of hypoxia (oxygen deficiency). Even at relatively pristine reference sites,
concentrations of DO rarely exceed 80 %Saturation, and can at times fall to levels low enough to be
stressful to some aquatic organisms. This suggests that DO availability has always played some role in
determining which biological communities inhabit certain wetlands. Nevertheless, there are also
unequivocal indications that wetlands that receive anthropogenic contaminant inputs suffer from much
more severe, frequent and prolonged episodes of hypoxia than sites that receive runoff from less
disturbed catchments. (See for example the case studies shown in Appendix A and B)
Few of the freshwater wetlands that we have surveyed in disturbed coastal catchments to date are capable
of supporting the biological communities and ecological values which in our experience are desired by
the local community. There is often insufficient oxygen for highly valued fish species such as
barramundi and mangrove jack to survive and, even where they can survive, conditions may still be too
stressful for them to prosper. Many wetland systems that once provided prime habitat for such highly
regarded fishery species are now populated by less valued, hypoxia-tolerant species such as tarpon,
gudgeon and eels.
Presence of tolerant species can be misleading. During our investigations it has been common for land
holders to point to highly visible display of fish activity as evidence that the waterways are healthy.
However, tolerant species such as tarpon can survive in low oxygen conditions by gulping air. Tarpon do
this by darting rapidly to the surface, creating the visual impression that they are feeding; and it only
takes a small number of tarpon to create the impression of high fish abundance, so such displays can be
evident even after extensive fish kills. They really only provide evidence that the water is hypoxic, and
do not indicate presence of a healthy fish community.
Some wetlands can maintain adequate oxygen levels for most of the year and therefore support
reasonable populations of DO-dependent fish. If wetlands of this kind experience a catastrophic oxygen
sag (due, for example, to inputs of oxygen-demanding organic matter from the catchment) fish kills
commonly ensue. Acute mortality events of this kind are relatively rare but they are a highly visible
manifestation of existing problems and often attract much attention from the local community. Such
events are cause for concern but they also provide some grounds for optimism because they demonstrate
that the affected wetlands are capable of maintaining reasonable fish stocks most of the time. This means
that managers can achieve useful environmental outcomes by implementing remedial measures to reduce
the severity of oxygen sags during particular types of storm events. While this is by no means simple, it
is often achievable with existing and emerging technology.
Ironically, many of the wetlands that do not currently experience fish kills are of far greater concern and
present much more complex challenges to managers because they are simply incapable of sustaining high
enough oxygen levels for sensitive species to ever become established. It is particularly worrying that
more than half of the wetlands we have surveyed in the Herbert-Burdekin region over the past four years
fall into this category, suggesting that the major factor preventing fish kills from occurring is the
exclusion of all sensitive fish species (due to poor water quality and/or the existence of physical barriers
to recruitment).
Dissolved oxygen concentrations are influenced by a range of interdependent natural processes. Our
current research on wetland water quality is demonstrating that in slowly flowing waters there are very
few, if any, biological, chemical or physical processes that do not ultimately affect oxygen. This means
that while dissolved oxygen is a very effective indicator of overall condition, it can also be exceptionally
complex to interpret.
It is not possible to properly appreciate the ways in which human activities influence the oxygen status,
and ultimately the ecological condition of receiving waters, without first becoming familiar with the
113
major sources of natural DO variability and some relevant biophysical concepts and terms. These are
summarised in the following points.
•
Oxygen is only slightly soluble in water, and rarely reaches concentrations higher than 14 mg/L in
natural waters. The amounts which can be dissolved in the water increase when the temperature and
salinity of the water fall, or the barometric pressure in the overlying air rises. A water that contains
the maximum amount of oxygen that can be obtained from the overlying air under the prevailing
conditions of temperature, salinity and air pressure is said to be saturated and the dissolved oxygen
concentration is 100 %Saturation. It is important to understand that even at the saturation point
(when no more oxygen from the air can dissolve in the water) there is constant exchange of oxygen
between the water and air. However, oxygen moves much more slowly through water than it does
through air, unless the water is being vigorously mixed. Therefore, if one of the factors that
determines the saturation concentration (e.g., temperature) changes, dissolved oxygen concentration
can take a substantial time to adjust. During this adjustment period, concentrations can deviate
significantly from saturation. For example, if cool water (capable of holding large amounts of
oxygen) is suddenly warmed up, the solubility of oxygen decreases, the saturation concentration
falls, and some of the oxygen in the water must escape into the overlying air. However, it can take
hours for the oxygen in the bottom of a water body to reach the overlying air and, until this happens,
concentrations in the water will be higher than 100 %Saturation. This water is said to be supersaturated with oxygen. The reverse happens when temperatures fall: oxygen enters the water from
the air and concentrations remain below saturation until the oxygen is distributed through the entire
water volume. In this case the water is said to be under-saturated.
•
The above processes occur in water bodies that heat up during the day and cool down at night.
Hence cyclical daily fluctuations are evident in many natural water bodies. This is referred to as diel
cycling (i.e., over 24 hours; note that “diurnal” is sometimes incorrectly used in this context, but
“diurnal” refers to daylight hours only – cf. “nocturnal”).
•
Most field meters can read oxygen concentrations as both mg/L (sometimes referred to as parts per
million or PPM) and %Saturation. If the water temperature, salinity and barometric pressure are
known, either of these can be calculated from the other. There is confusion in the literature
regarding which is the better measure to employ in water quality investigations. In practice both
measures are needed to fully interpret data, but %Saturation is the most readily interpreted and
ecologically relevant of the two. It is the saturation level that directly indicates how much oxygen is
available for aquatic organisms to breathe, not the amount that is dissolved in the water. As a rule of
thumb, a fish in water that is 100% saturated with oxygen is able to gain access to an amount of
oxygen equivalent to that in the overlying air. If the concentration falls to 50 %Saturation then it can
only obtain half the amount of oxygen that is in present in the overlying air. The mg/L
concentrations of dissolved oxygen required to achieve these saturation levels vary enormously,
particularly with temperature, so results expressed in terms of mg/L are much more difficult to
interpret. (Technically minded readers should note that it is the partial pressure of oxygen in the
water that is of primary concern. At saturation the partial pressure in the water is equal to that in the
thin layer of moisture-saturated air at the surface boundary layer. Meters that read partial pressure
are available and are widely used in some scientific disciplines, but they are not generally used in
water quality investigations).
•
There is confusion in the water quality literature regarding the significance of the reduced solubility
of oxygen in warm waters, and it is not uncommon to encounter statements suggesting that this is the
reason oxygen levels are low in tropical waters. Such statements are misleading because, while
warm tropical waters do usually tend to be under-saturated in oxygen, this has little to do with the
solubility of oxygen. The lower solubility of oxygen in warm water simply means that less oxygen
needs to dissolve in the water in order to achieve any particular saturation level and it is the
saturation level that determines how much oxygen is available to organisms in the water. The fact
that more oxygen can dissolve in cold water is not an indication that there is more oxygen available
for animals to breathe, but rather that much more oxygen must dissolve in the water in order for
animals to be able to breathe. This is evidenced by the fact that asphyxiation (death due to lack of
114
•
oxygen) occurs at quite similar saturation levels in both cold and warm water fish even though the
mg/L concentrations at which this occurs are much higher for the cold water species.
All aerobic organisms consume oxygen when they breathe (a process called aerobic respiration).
Some fully aquatic organisms breathe air but most extract dissolved oxygen from the water they live
in. Consequently, natural bodies of water that contain many aquatic organisms constantly consume
oxygen. The amounts of oxygen consumed depend, in part, on the total mass of organisms (i.e.
biomass) respiring in the water. In many water bodies this biomass comprises mainly microscopic
organisms (microbes) that live either in the water column or in the bottom sediments. Submerged
aquatic plants (including planktonic algae) can also occur in very large quantities in some wetlands
and these also contribute to the total oxygen consumption of the system. (Emergent and floating
plants obtain oxygen mainly from the air and do not directly contribute to oxygen consumption in the
water. They do, however, provide habitat and sustenance for other organisms which do consume DO
from water).
•
The total amount of oxygen consumed by all the organisms in a water body is sometimes referred to
as its biological oxygen demand. This should not be confused with the water analysis term,
Biochemical Oxygen Demand (BOD), which refers to the amount of oxygen consumed by a sample
of water when it is subjected to a standardised laboratory incubation test. The BOD test measures
only one of many factors that contribute to overall biological oxygen demand.
•
Plants that are exposed to adequate sunlight take up carbon dioxide and transform it into sugars
during photosynthesis. This process produces oxygen as a by-product. Emergent and floating plants
generally release oxygen into the air but all submerged plants (and photosynthesising microbes)
release oxygen directly into the surrounding water. Since photosynthesis can only occur during
daylight hours this can result in enormous DO fluctuations, with plants producing much more
oxygen than they consume during the day, and consuming oxygen at night. The resulting diel
cycling of DO concentrations is usually much more pronounced than the cycling induced by
temperature variations, and in extreme cases levels can vary from zero to greater than 100
%Saturation over the course of a single day. Provided that plant growth is not limited by other
factors (such as light or nutrient availability), rates of photosynthetic oxygen production are far more
rapid than the plants’ respiratory consumption rates. Therefore, in plant-dominated systems, DO
levels seldom fall to zero overnight, although concentrations may become very low if conditions
suddenly change and reduce photosynthesis during the day (e.g., if it is very cloudy). However, only
a very brief period of photosynthesis is required (e.g., a break in the clouds) for higher
concentrations to be restored.
•
Regardless of the processes that cause DO concentrations to deviate from saturation, exchanges of
oxygen with the overlying air (re-aeration) continually act to return the system to saturation. The
rate at which oxygen enters (or leaves) the water increases exponentially as concentrations move
further away from saturation; that is, rates increase disproportionately when concentrations are
extremely low or extremely high. This mechanism is important as it protects water bodies against
the development of extreme concentrations.
•
Re-aeration rates also depend on the surface area of the water body and the amount and type of
mixing that is present. Typically, re-aeration rates are highest in shallow, turbulent flowing waters.
Wind is also an important facilitator of re-aeration, especially in stagnant wetlands. Re-aeration
processes do not necessarily increase oxygen concentrations – in situations where waters are supersaturated they reduce oxygen concentrations to saturation level.
•
The DO concentrations at a particular location and time are usually the sum of inputs from upstream,
oxygen consumption and production within the water body, and re-aeration. Since oxygen
consumption is usually highest at the bottom and exchanges with the air occur at the surface, there is
a natural tendency for DO concentrations to vary with depth. Moreover, plants (and the microbes
and animals attached to them) tend to favour particular parts of a wetland (for example, shallow
sections), introducing further spatial variability. The nature of these variations change over the
course of the day as light levels change. The situation is further complicated by the effects of
115
passing clouds and shade from trees, which modify the rates at which plants in different parts of a
wetland produce oxygen at different times.
•
If conditions are very turbulent re-aeration rates can become so high that they overwhelm all of these
other processes such that DO concentrations throughout the entire water column deviate little from
saturation. However, rainforest streams are the only freshwater environments on the Queensland
tropical floodplains where such conditions occur regularly. Some very open wetlands may approach
this condition during very windy weather but, in the majority of cases, biological processes have a
measurable effect on the distribution of DO concentrations throughout the water body. As a result, it
can be exceptionally difficult to determine and/or describe the oxygen concentrations in a water
body. Spot measurements (typically used in monitoring programs) are usually meaningless, and
even large numbers of measurements are difficult to interpret if they are collected from a single point
in the water column.
•
In most local wetlands, oxygen consumption can usually be attributed entirely to the respiration of
microbes, plants and other organisms, including fish. However, in some cases, chemical oxygen
consumption processes also play an important role. (This is sometimes referred to as chemical
oxygen demand and should not be confused with an analytical test, which goes by the same name.
This test, often abbreviated as COD, was originally developed as a more rapid substitute for the
standard 5-day BOD test. It utilises a chemical oxidation method to obtain a rapid estimate of the
total organic matter present in a water sample. Oxygen consumption due to inorganic reactions may
also contribute to the final result, however, most standard methods include steps to minimise such
effects. The result is reported in terms of the amounts of oxygen that would ultimately be consumed
in the process of decomposing organic matter in the sample even though some of it may not be
biodegradable). Ferrous iron, which is present in some ground waters, and is a major constituent of
acid sulphate runoff, is the most important inorganic source of oxygen demand in cane growing
areas. Ferrous iron reacts with oxygen to form insoluble ferric oxyhydroxides, which are visible as
orange-brown precipitates when present in high concentrations. This reaction not only consumes
oxygen but also produces hydrogen ions sometimes resulting in reduced pH levels.
•
Water that is devoid of oxygen is said to be anoxic, that which is oxygen deficient is hypoxic and
that which contains adequate oxygen levels to support most oxygen-requiring functions is said to be
oxic. “Hypoxia” and “hypoxic” are somewhat subjective terms that must be interpreted in the
context within which they are used. In this report hypoxic can be taken to imply that oxygen
concentrations are low enough to affect key ecological functions.
•
At temperatures above 4 ºC (ie all of the time in the tropics), warmer water is less dense than cool
water, so when the surface waters of a wetland are exposed to sunlight they heat up and tend to float
on top of the cooler waters underneath. This can lead to the formation of two separate layers of
water that do not mix with one another – a process called stratification. The bottom layer does not
have access to oxygen from the air but experiences high levels of respiration because it is in contact
with bottom sediments, so it can become severely hypoxic or anoxic. Tropical wetlands can be
stratified most of the time, especially if they are deep and/or sheltered from wind. During cooler
months diurnal stratification is quite common (i.e., separate layers form during the day but mixing
occurs during the night).
•
The tendency for DO concentrations to be higher near the surface of a water body is particularly
pronounced in the microscopically thin layer of water that is in direct contact with the air. Some
species of fish can exploit this layer very efficiently by using aquatic surface respiration (ASR) (not
to be confused with the gulping of air that is used by some other species). Most local species do not
perform ASR as efficiently as many of the exotic species that have been introduced into our
wetlands, but it is still an important short-term survival mechanism, especially for small individuals.
Even large adult barramundi, with mouths and gills far too large to enable proper ASR, can be
observed lying on their sides at the surface in an attempt to obtain a little extra oxygen when
conditions are very stressful. Most laboratory determinations of hypoxia tolerance use artificial lowoxygen atmospheres, so ASR is not an available option for test specimens. Accordingly, reported
116
asphyxiation thresholds are somewhat higher than might be expected in natural waters where ASR is
available. However, animals that only use ASR when they are forced to, often expose themselves to
increased risks of predation and sunburn, and cannot participate in normal behaviours such as
feeding and breeding. Accordingly, asphyxiation thresholds determines by laboratory tests provide
meaningful indications of the levels at which there is high risks of acute harm to animals, even
though actual asphyxiation will not necessarily occur in natural environments until levels become
significantly lower.
•
The amounts of oxygen produced and consumed by photosynthesis and respiration in the water
column depend not only on the amounts and types of biomass that are present, but also on the
activity levels of the individual organisms that make up that biomass. This is governed by several
factors, some of which are species-dependent, but the most important are light availability and
temperature. The relationships between light intensity and photosynthetic oxygen production in
some of our most common local submergent plant species have been examined quantitatively in this
project (see Appendix D). Our results show that local species behave in much the same way as
plants that grow in other parts of the world, in that each is adapted to perform optimally at different
light levels. This means that although it is always true that higher light levels (due, for example, to
reduced turbidity or colour in the water) will result in increased oxygen production, the effect is far
from linear, may have an upper threshold, and depends on which species are present. Consequently,
the effects of shade, turbidity and colour, all of which influence the amounts of light reaching aquatic
plants, are extremely complex and not easily predicted; in some cases changes in these factors may
significantly alter productivity levels while in others a change of similar magnitude may have no
detectable effect.
•
Appendix D shows how critical light threshold values can be determined for particular plant species
and how this information can be used in conjunction with depth profiles and data on incident light
intensity to predict the effects of light-attenuating parameters such as turbidity or colour on a
particular water body. This work clearly demonstrates that the capacity of water bodies to tolerate
changes in ambient concentrations of light-attenuating parameters (such as Suspended Particulate
Matter (SPM), turbidity and colour) varies enormously between wetlands, making it impossible to
propose any simple guideline values that could be meaningfully applied to all situations. Currently,
guidelines can only be developed for individual water bodies on a case-by-case basis. The BLC
classification proposed in this report addresses this problem by identifying classes of water bodies
with similar susceptibilities to changes in light attenuation as a basis for the development of typespecific rather than site-specific guidelines.
•
The effects of temperature on the rates of biological processes are commonly expressed in terms of a
standardised temperature coefficient known as the Q10 value. The Q10 is the factor by which a rate of
a biochemical process changes in response to a 10oC increase in temperature. Published data from
around the world indicate that Q10 values for respiration range from 2 to 4 for most types of aquatic
organisms, including microbes, plants, invertebrates and fish. Our experiments found that Q10 values
for local aquatic plant species were within the range of 2.4 to 3. That is, a temperature increase of
10oC will cause oxygen consumption rates to increase by as much as 3-fold. (Note that Q10 values
refer to temperature changes that occur over periods in the order of hours; processes such as
accumulation and acclimatisation allow organisms adjust their biochemical systems to accommodate
more gradual and/or long-term temporal changes, such as seasonal ones, so effects are not usually so
pronounced).
Many of the large, well-mixed water bodies that are typically the principal focus of water quality
investigations have relatively stable water temperatures on a day-to-day basis, even though seasonal
variation can be high. Consequently, short-term temperature variations are not of great concern in
understanding physical dynamics. The situation is very different in the small, shallow, poorly mixed
wetland systems of tropical Queensland. Temperatures in these water bodies commonly fluctuate by 8 to
12oC each day. Moreover, small-scale spatial variations can be just as large – a deep shaded backwater,
for example, can be strikingly cooler than a shallow section that is exposed to the sun. Variations of these
kinds have an enormous effect on the distribution of DO concentrations over time and space and must be
taken into account when modelling or interpreting biophysical processes.
117
APPENDIX H
TECHNICAL CONSTRAINTS TO THE SUCCESSFUL
IMPLEMENTATION OF EXISTING NATIONAL AND STATE WATER
QUALITY GUIDELINES IN THE DRY AND WET-DRY TROPICS
118
H1.
TECHNICAL CONSTRAINTS TO THE SUCCESSFUL IMPLEMENTATION OF
EXISTING NATIONAL AND STATE WATER QUALITY GUIDELINES IN THE
DRY AND WET-DRY TROPICS
Water quality evaluation requires knowledge of the water’s physical and chemical characteristics but this
in itself provides a very limited basis for assessing whether water quality is “acceptable”. In order to
reach such conclusions data must be evaluated against various benchmarks; the choice of benchmark
being dependent on the purpose of the assessment.
Prior to 1992 there were no authoritative national guidelines or standardised benchmarks for water
quality assessment in Australia. Various international and state guidelines were available but these
related almost exclusively to a few specific beneficial human water e\uses. The situation changed in
1992 with the signing of the COAG Agreement on Ecologically Sustainable Development, which
committed the states the tenets of the National Water Quality Management Strategy (NWQMS,
ANZECC, 1992). – a Commonwealth initiative that included the development of Australia’s first national
water quality guidelines (ANZECC, 1992). These guidelines, which have since been revised by
ANZECC and ARMCANZ (2000), called for managers and investigators to identify all of the protectable
environmental values (PEV’s) that should be attributed to particular bodies of water, and to then select
indicators and benchmarks appropriate to each PEV. Notably PEV include, not only human water uses
(such as drinking, irrigation and recreation) but also the integrity and health of natural aquatic
ecosystems.
The 1992 ANZECC Water Quality Guidelines (WQG) related only to the assessment of long term
ambient water quality in natural surface water bodies under low flow conditions. Assessments of the
quality of stormwaters, effluents and attenuation zones, and the issues of contaminant loading, export
rates and the potential for acute effects resulting from brief episodic water quality disturbances, were not
addressed. Similar constraints apply to the recently revised guidelines (ANZECC and ARMCANZ
2000), although for the first time they include some advice on how to develop situation-specific
benchmarks for contaminant export loads (which requires monitoring of stormwater flows) and detail
direct toxicity assessment methods which could be used to evaluate acute water quality effects in
particular situations. The guidelines cover a range of water quality parameters and indicators that is as
broad as possible within existing ecotoxicological data constraints. However, due to information
deficiencies, guideline values suitable for use as benchmarks in water quality assessments are not yet
available for many parameters and/or PEV’s. (For example guideline values for the protection of aquatic
ecosystems have not yet been developed for the most of the pesticides currently being used by farmers in
North Queensland).
The authors point out that it is not possible to develop guideline values that are applicable in all
situations, so the default values presented in the current guidelines only provide a starting point for risk
assessment – if these values are exceeded it will not necessarily mean that environmental values are being
harmed but rather that investigators should use to develop locally relevant water quality benchmarks, and
provide advice relating to the evaluation and management of water quality problems.
The water quality evaluation programs implemented by water resource managers fall into two distinct
categories:
1) Ambient assessments, aimed at determining if the quality of water contained within a particular
body of water is adequate to maintain the PEV’s identified for that water body; and
2) Assessments of downstream effects, which attempt to determine the extent to which fluvial
contaminant exports influence ambient conditions in downstream receiving waters.
The new national WQG contain substantial quantities of advice relating to ambient assessments but most
of it is generic in nature. The existence of situation-specific modifying factors is noted, but methods for
dealing with them have not been elucidated or are explained by way of case studies and examples that are
not applicable in the dry and wet-dry tropics. Given the sheer size of Australia and the diversity of its
waterways it would be unrealistic to expect national guidelines to contain more specific advice, and
particularly, information relating to the more sparsely populated and poorly studied parts of the continent.
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The WQG document comprise several hundred pages of text and substantial amounts of supporting data
and information, so it is already a substantial undertaking to read them, and there is little doubt that
incorporation of more specific advice would serve to confuse rather than inform. Accordingly the
authors’ insistence that local jurisdictions will need to develop their own guidelines is more than
justifiable.
It seems likely that the paucity of detailed guidance relating to the assessment of downstream effects will
be addressed in future revisions of the guidelines but in view of the relative remoteness of the dry and
wet-dry tropics, it would be unrealistic to expect specific advice relevant to this region to be incorporated
in the foreseeable future, if at all. Hence it is probable that managers will be heavily reliant on locally
developed guidelines.
Most of the advice contained in the national guidelines is based on experience gained from research in
large perennial water bodies located in temperate areas of the world, and there are sound grounds to
doubt that this is directly applicable to the intermittent fluvial systems of the dry and wet-dry tropics of
Queensland, and especially to the relatively small wetland systems that are typical of the region. This
chapter summarises the approaches advocated in the national guidelines and identifies some barriers to
successful implementation in the dry and wet-dry tropics. Most of the identified problems relate to
ecological water quality assessments (ie water quality monitoring programs aimed at determining if the
ecological function and diversity of the aquatic ecosystem is being adequately protected). However,
there are also some indications that the designs of monitoring programs primarily aimed at water resource
assessment may not adequately contend with the hydrological and biophysical conditions encountered in
this region.
H1.1
Resource Assessment
Resource assessments are more straightforward than ecological assessments, mainly because there are a
variety of authoritative water quality criteria available (eg. NHMRC 1996 Drinking Water Guidelines,
(ANZECC and ARMCANZ 2000) Australian and New Zealand Water Quality Guidelines) and it is
generally unambiguous which of these constitute appropriate benchmarks for evaluating various aspects
of a water’s resource value. The current ANZECC AND ARMCANZ (2000) Guidelines represent a
departure from historical convention in that they specify trigger values rather than simple non-exceedance
values. Compliance with a trigger value can be taken to indicate unqualified suitability for use while a
breach of the trigger value indicates that closer investigation would be required before a definite
conclusion can be reached. For example in the case of irrigation use, investigations might take into
consideration the types of crops to be irrigated and areal soil and drainage properties. The guidelines
therefore make allowance for the fact that a particular water may be unsuitable for irrigation in some
situations but well suited for use in others. Other triggered investigations simply require a closer analysis
of available water quality data. For example a water body with salinity levels above the trigger value
may still be quite acceptable for use if concentrations of troublesome ions such as sodium or sulphate
remain within acceptable limits.
Despite these minor complications, assessing the suitability for human use is still a conceptually
straightforward procedure and in most cases published criteria are just as applicable in dry tropical
catchments as they would be anywhere else. However, this does not mean that the unique characteristics
of each catchment does not need to be taken into consideration in resource evaluations.
It is particularly noteworthy that the vast majority of water quality data available for these systems
comprise salinity and major ionic concentration values obtained from samples collected mainly under
baseflow conditions. However, in the dry tropics there are few tributary streams or river reaches where
the volume and reliability of baseflows are inadequate to be of value for most human uses except
livestock watering (although there may be some limited exploitation by small-scale users in some cases)
and most water supplies are derived from spears installed in the bed sands or adjacent alluvial aquifers.
The baseflows in perennial reaches are more widely used but the volumes of water available for major
resource development are extremely limited compared to the quantities that are available from storm
flows. Moreover, baseflows are usually extracted either from near the bottom of deep river reaches and
waterholes, (or from spears), and since most waterbodies in the dry tropics are vertically stratified
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(discussed later), the quality of water at the surface (where most water samples are currently collected) is
a poor indicator of the quality of the water that is actually being used.
Even during quite prolonged dry spells, the water retained in large dams is usually dominated by surface
runoff that is carried into the reservoir during small-scale storm events or the falling stages of major flood
events; the quantities of baseflow entering large impoundments are very commonly inadequate to
significantly affect water quality. Smaller reservoirs (i.e. most weirs) are much more significantly
influenced by baseflows, but relationships between the quality of water entering the storage and the
quality of water being extracted and/or released from it, are exceptionally complex and not easily
predicted, even if inputs are closely monitored.
It is therefore concluded that data indicative of the quality of baseflows in unsupplemented river reaches
are an extremely dubious tool for resource assessment and development. Data indicative of the rising and
falling stages of the event hydrographs are arguably more relevant and can be used to assess the quality
of water that can be captured by various methods (eg water harvesting or reservoir construction) but these
would need to be interpreted in conjunction with data obtained from impoundments in order to examine
and predict storage effects.
Reports on water quality seldom point out that the ionic concentration data that dominate most existing
datasets in the tropics are useful for evaluating suitability for irrigation and livestock watering but are of
limited use for potability assessment. These parameters provide some insights into the water’s suitability
for human consumption but they are mainly of only aesthetic concern. For example, high levels of salts
and hardness may adversely affect the taste of water and may cause nuisance problems such as formation
of scales in plumbing fixtures and kitchen items, but they will not harm the consumer.
People in rural areas throughout the dry tropics routinely drink ground water with hardness and/or salinity
levels well in excess of water quality guidelines. (In many cases they do this even in situations where
there is soft, low-salinity surface water available nearby, because they feel that the ground water is less
likely to be microbiologically contaminated.) Biological contamination and physical and chemical
factors that affect the efficiency of disinfection processes are far more likely to be the principal
constraints on potability and water treatment for drinking. For example, high levels of faecal
contamination from livestock, which can carry human pathogens such as Cryptosporidium, could be a
major problem in some areas. Linked with high concentrations of organic matter, colour, suspended
solids, turbidity, and dissolved iron and manganese, all of which reduce the efficiency of standard
disinfection methods, water treatment costs could be increased to the point where the water is no longer
considered a desirable resource. Currently, there are virtually no data available for any of these
parameters , except for turbidity, in most dry and wet-dry tropical catchments.
H1.2
Ecological Assessment
H1.2.1 Available guidelines
Ecological water quality evaluations are far more complex than others because it is not always clear
precisely what benchmarks should be used. It is generally accepted under the National Water Quality
Management Strategy (1992) that the principles of ecologically sustainable development allow for a
certain amount of deterioration in both water quality and ecosystem condition (in some circumstances at
least) provided that a functional aquatic ecosystem is maintained. This poses a variety of very complex
philosophical questions. For example, how much change is acceptable and what constitutes a functional
ecosystem? And how do we incorporate systems that are subject to natural periodic, seasonal or episodic
water quality deterioration?
ANZECC and ARMCANZ (2000) Water Quality Guidelines (WQG’s) are currently the authoritative
source of guiding principles relating to these issues. Queensland WQG’s (currently in draft) are due for
release and these will contain more specific advice on how to apply the ANZECC principles in
Queensland catchments. The ANZECC and ARMCANZ (2000) guidelines ushered in a new (and
somewhat experimental) era for water quality assessment and management in Australia. Notably, for the
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first time the National guidelines are based on ecological risk assessment principles and promote the
concept of holistic issue-based water quality management.
They strongly advise that managers take account of all of the factors influencing the condition of an
ecosystem as a whole, in order to ensure that the ecological issues of highest priority are addressed and
that the most appropriate indicators are used. (For example resources should not be invested in
monitoring nutrient concentrations unless it has been established that (i) nutrient inputs are likely to be a
major factor influencing the condition of the ecosystem in question and (ii) that nutrient concentrations
are the most accurate indicator of effects). The capacity to arrive at such decisions is contingent on a
well-developed understanding of the aquatic system and a knowledge of how various factors interact to
determine ecological outcomes. The recommended means of achieving this is through the development
of holistic conceptual models.
The guidelines suggest that the concentration and bioavailability of contaminants in the water column can
often vary so much over time and space that water samples are not necessarily a reliable indicator of risk
potential. Accordingly, ANZECC advocates the use of a variety of alternative indicators that have never
previously been included in water quality guidelines. These include benthic sediment residues, biological
tissue residues, biological condition indicators (such as macroinvertebrate diversity and stream
metabolism), and direct toxicity measurements. The guidelines do not yet recommend specific
benchmarks for most of these indicators but they do include interim guideline values for sediment
quality. Additionally, it is expected that the Queensland Guidelines will include reference data for both
benthic sediments and biological tissues. However, in both cases the range of contaminants dealt with is
limited mainly to commonly occurring heavy metals and a few artificial organic toxicants, (most of
which are not widely used in the tropics) so in many cases individual investigators will still need to
obtain their own reference data.
Arguably the most significant innovation in the new guidelines is the abandonment of the simple
inflexible limit-based criteria that have traditionally been the cornerstone of water quality guidelines, in
favour of very flexible risk-based trigger values. The approaches used to derive trigger values vary
somewhat depending on whether the parameter being dealt with is a toxicant or a stressor (i.e. a nontoxic, naturally occurring chemical or physical characteristic that affects the health and survival of
aquatic organisms).
Toxicants
Trigger values for toxicants have been derived from statistical analysis of available ecotoxicological data
and are designed to stipulate the contaminant concentration at which a particular percentage of species
will be protected with a specified level of certainty (eg. 90% of species protected with 50% certainty).
The guidelines recommend certain default combinations of protection and certainty levels for use in
various types of water body, but the selection of these is ultimately left to the discretion of local
jurisdictions. To this end the guideline package includes a CD-ROM that contains raw toxicological data
and software which will allow individual users to calculate a situation-specific (eg. local or regional)
trigger value using any combination of protection and certainty level they deem to be appropriate.
Notably it is also feasible to set aside toxicity data for species that are not present in the local area and/or
add data for local species if it comes available. Compliance with the selected trigger value is taken to
indicate that the minimum level of protection desired by local water quality managers is being achieved.
Exceedances of the trigger value do not constitute a true compliance failure but rather indicate that more
detailed investigations will be needed to determine if the risk of harm is unacceptable.
The investigations that are triggered depend on the toxicant being dealt with and are clearly delineated in
the guidelines. They are required because the toxicological data from which the initial trigger values are
derived relate to the behaviour of bioavailable forms of the toxicant under standard laboratory conditions,
and cannot be validly applied to natural waters where a diversity of site-specific biophysical factors
interact to substantially modify bioavailability and therefore toxicity. Triggered investigations are to be
carried out in a step-wise fashion, and each step results in an adjustment to the trigger value to take
account of a particular toxicity-modifying factor. (Following this process to its logical conclusion would
ultimately result in the development of a site-specific water quality guideline value that accurately
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reflects actual ecological risks.) At the completion of each step, data are re-evaluated against the new
modified trigger value and further steps are required only if they are still in exceedance. At each step
local managers have the choice of accepting that there is a potential risk of environmental harm and
instituting remedial measures, or proceeding to the next step in order to evaluate risks more accurately.
Each subsequent step involves more complex and costly studies so the relative costs and benefits of
proceeding with further investigations will need to be weighed up carefully.
These approaches attempt to strike a balance between the scientific need to conduct very detailed sitespecific investigations to quantify toxicological risks accurately and the pragmatic consideration that such
undertakings are costly and should only be implemented in situations where risks are demonstrably high
enough to provide an imperative.
Stressors
Physical and chemical stressors (including parameters such as water clarity, sediment and nutrients,
which have been the principle focus of most recent research in regional aquatic systems) are not normally
considered to be toxic but can exert very strong influences on aquatic ecosystems by modifying
productivity, food webs, inter-species competition, etc. Natural temporal and spatial variations in stressor
levels play an important role in determining the diversity and function of natural ecosystems. (For
example some water bodies are characteristically deficient in nutrients and therefore support biological
communities and processes that are very different from naturally eutrophic water bodies). Different
stressor levels are required by different ecosystems so it is not feasible to set generic limits that would
apply in all situations. Accordingly ANZECC trigger values are to be derived from reference data
collected from comparable types of water bodies that are considered to be in an acceptable condition.
The guidelines provide generic advice regarding the selection of reference data sources and determination
of acceptable levels of departure from reference, and statistical decision rules for data evaluation are
documented. These procedures can also be adapted to other parameters such as biological and habitat
indicators. Default trigger values are suggested for some combinations of parameters and water body
type, but ultimately the choice of reference source, acceptable level of departure and therefore trigger
value is left to the discretion of local jurisdictions.
The recommended statistical decision rules all relate to the central tendency of the water quality data –
the median being preferred if the data are not normally distributed (as is usually the case) and the mean
being favoured if they are. (No attention has been given to the potential effects of extreme values,
changes in variance or other potentially relevant statistics). A guideline for nitrogen might, for example,
stipulate that the median concentration should not exceed the 80th percentile of the reference range. If the
80th percentile nitrogen concentration of the selected reference data is 800µg/L then the trigger value for
nitrogen is set as a median concentration of 800µg/L. Again exceedances of the trigger value indicate the
need for more detailed, usually site-specific, investigations to determine if the ecosystem is actually being
modified to an unacceptable degree.
Traditionally reference data are obtained by monitoring undisturbed sites. However, such sites are absent
for certain types of water bodies in many parts of Australia. Moreover, many waterways are now
permanently altered to the extent that they could no longer be expected to behave like a natural water
body. In such circumstances, the use of undisturbed reference sites could easily create unachievable
water quality expectations. Accordingly, ANZECC leave open the possibility of local jurisdictions
electing to use as a reference source, partially degraded sites that are in a condition that is considered to
be acceptable to the local community. This requires a great deal of subjectivity and has more to do with
socio-economics and politics than science.
Experience suggests that most of the default (generic) trigger values contained in the ANZECC and
ARMCANZ (2000) guidelines (for both toxicants and stressors) are very unlikely to be directly
applicable in the dry tropics. This means, that the prescribed trigger modification and development
processes will probably need to be implemented to obtain realistic local water quality criteria. In the case
of toxicants default trigger values are most likely to be overly conservative because factors such as
turbidity commonly reduce toxicity. Hence, compliance with the default values would be expected to
guarantee that the ecosystem is being protected. However, there may be important exceptions to this
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general rule and these are discussed later. In the case of stressors, default trigger values are derived from
statistics gathered from a narrow range of regional reference sites, and it is inconceivable that these could
function as a valid and meaningful reference for all types and situations
H1.2.2 Application of water quality guidelines in Queensland
Management of water quality in Queensland is regulated by the Environmental Protection Act (1994) and
the Queensland Environmental Protection (Water) Policy (1997) (EPP). The EPP Water states that water
quality management in Queensland shall be consistent with the management processes expressed in the
National Water Quality Management Strategy (which includes the ANZECC and ARMCANZ 2000
Australian and New Zealand Water Quality Guidelines). The EPP also states that locally derived
guidelines (sanctioned by the environmental authority) will take precedence over state guidelines, which
in turn take precedence over national guidelines. The same policy is supported in the National
Guidelines. No local guidelines have yet been developed for the Queensland dry tropics so the State
guidelines will take precedence (once they are released). The national guidelines for ecosystem
protection do not actually provide water quality criteria but rather a prescriptive framework that local
jurisdictions should use to develop their own criteria. The large number of choices left to the discretion
of local authorities means that the ANZECC guidelines cannot be fully implemented until the Queensland
Guidelines are finalised.
The National WQG’s provide default trigger values for some parameters (such as nutrients) in freshwater
streams, marine waters and parts of estuaries (but not wetlands). However, the freshwater defaults are
derived from reference data that are representative of only a few stream-types and in some cases, limited
sets of flow conditions. They are unlikely to provide ecologically meaningful benchmarks for most
regional aquatic systems.
The draft Queensland Guidelines (March 2001) confirm that the approaches recommended in the national
guidelines are to be adopted in Queensland. They accept ANZECC’s default trigger values and
acceptable risk limits for toxicants and the default statistical decision rules for physical and chemical
stressors.
These are as follows:
•
Pristine Systems
No detectable change to toxicant levels.
°
°
No departure from reference conditions.
ANZECC recommend similarly stringent measures for waters with high conservation values but the
status of World Heritage Areas, national parks and other designated conservation areas are not
specifically discussed in the available drafts of the Queensland Guidelines.
•
Slightly to Moderately Modified Systems (most waters in Queensland)
°
Toxicants in water: Toxicant concentration trigger values should be set to protect 95% of
species with 50% certainty.
°
Toxicants in sediments: Trigger values should protect 90% of species.
°
Physical-Chemical Stressors: The median value of the indicator should not move outside of the
20th to 80th percentile reference range.
°
Biological Indicators: To be determined at a regional or local level.
•
Highly Disturbed Systems
To be determined at a local level. Options include selecting disturbed reference sites with acceptable
ecological condition characteristics or relaxation of the allowable effect sizes for slightly modified
sites (eg. use of 10th and 90th percentile reference values instead of 20th and 80th percentile). Local
jurisdictions can amend any of these benchmarks if they have reasonable justification.
In the case of toxicants only two types of water are recognised:
1. Fresh
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2. Marine
There are no specific guidelines for other categories such as brackish, estuarine or ground waters.
In the case of other parameters the draft Queensland Guidelines consider six water types (ANZECC also
recognise lakes and wetlands as separate water types but these are not included in the current draft of
state guidelines):
• Upper catchment (upland stream)
• Lower catchment (lowland stream)
• Upper estuarine
• Mid-estuarine
• Enclosed coastal
• Marine
The Queensland Guidelines divide water types into the following bioregions:
•
•
•
East Coast:
Catchments between the NSW border and the wet tropics excluding inland areas of
large catchments such as the Fitzroy and Burdekin.
Wet Tropics: East Coast catchments between Ingham and Cape York
Inland Rivers: Intermittent slow flowing streams in lowland areas away from the coast, which are
characterised by high levels of turbidity.
Reference statistics and default trigger values for selected parameters have been collated for several of
the above combinations of water type and bioregion. The stated intention is to incorporate higher levels
of water type division into future versions of the guidelines. Notably it is intended that future reference
data sets will include sub-divisions representative of different conditions of flow and season, but this
requirement has not yet been addressed. The use of local reference data is strongly encouraged provided
that quality assurance is acceptable.
Lower catchment water classes are currently defined as streams with reduced riffles, larger pools and
significant areas of muddy substrata. The latter requirement essentially excludes many streams in the
lower catchment and floodplain of wet/dry and dry tropical rivers which, in their natural state, often
maintain coarse sandy substrata all the way to the estuary.
H1.2.3 Implementation in the dry and wet/dry tropics
Local interpretations and implementation of the new generation water quality guidelines are likely to
have a significant bearing on the types of monitoring and management strategies that are developed well
into the foreseeable future. Many of the principles and approaches recommended in the new generation
water quality guidelines are largely untested in the dry and wet/dry tropics and until that happens it will
not be possible to answer fundamental questions about the acceptability of water quality in local aquatic
systems. Moreover, many of the recommendations in the guidelines are based on assumptions that are
unlikely to be valid for most water bodies in dry and wet/dry tropics, so there are sound reasons to doubt
that any useful improvements could be gained by implementing the guidelines in their generic form
without adding extensive refinements to account for local conditions. The guidelines place a great deal of
onus on the local jurisdiction to carry out such refinements but the available advice regarding how this
should be done is currently inadequate to guarantee successful outcomes in local catchments.
Our current understanding of water quality processes in the tropics is insufficient to recommend the
precise refinements that are needed (that would be a substantial undertaking in its own right) but we do
know enough to be able to identify some potential pitfalls and recommend preferred lines of investigation
aimed at developing more suitable alternatives.
Toxicants
Toxicants are not a major catchment-wide issue in sparsely populated catchments where free range
grazing is the dominant land use but are potentially a sub-regional issue in areas that are subject to more
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intensive human development. Most of these developments have limited potential for widespread
impacts but high potential for adverse effects on initial receiving waters. It is important to recognise,
however, that much of the degradation that has occurred in the systems as a whole is probably the result
of cumulative effects from many localised impacts of this kind.
Notably potential toxicant sources such as urban centres, industrial processing plants, ports, sugarcane
farms, horticulture and aquaculture are generally concentrated in near-coastal subcatchments. This has
resulted in the loss of and/or irreversible alterations to, many of the unique freshwater and marine
wetlands that once formed an extensive mosaic of highly productive and exceptionally diverse aquatic
habitats along the coastal plains. In many cases the remnant habitats are not only the primary receiving
waters for anthropogenic toxicant inputs, but have also been altered to the extent that their capacity to
tolerate exposure to toxicants has been significantly modified.
Moreover, many toxicants are particle-reactive (i.e. under certain environmental conditions they either
adhere to the surface of sediment particles through processes such as sorption, or precipitate to form
insoluble particles, while under other conditions they can occur in free dissolved form) so their mobility
and fate in the environment is often governed by sediment transport processes. Other factors being equal,
sediment export efficiency decreases with increasing distance and time of travel (due to deposition within
the drainage system). Consequently the close proximity of near-coastal catchments to the sea greatly
increases the probability of toxicants being delivered to the marine environment. For example
reconnaissance scale modelling using the CSIRO SedNet model (Prosser et al 2001) indicates that
although sugarcane farms produce only 0.8% of the total catchment soil erosion they occupy
subcatchments that are responsible for 9% of the total quantities of fine sediment delivered to the coast.
Current water quality guidelines focus quite heavily on evaluating the concentration of toxicants in water.
However, due to the biophysical instability of most aquatic systems in this region the concentrations and
bioavailability of most toxicants can be expected to fluctuate enormously over time and space, making it
extremely difficult to obtain accurate or representative results. Moreover, even if quantification proves
feasible there are several constraints to accurate data interpretation:
• The toxicity data upon which trigger values are based include few local species and notably, lack
information on local keystone species (such as barramundi and sooty grunter, for example).
• The system of setting guidelines to protect a certain percentage of species is unlikely to be effective
unless it is proven that this is adequately protective of valued keystone species (eg. fishery species
and major components of food webs). It is doubtful if there is sufficient available information to
identify key species other than perhaps fish and their prey in the unusually (by national standards)
diverse aquatic systems of the tropics.
• The guidelines are based on the premise that toxicants are only bioavailable if they are freely
dissolved and in a form that can pass directly through biological membranes such as gills and cell
walls. It is generally assumed that toxicants attached to suspended sediment particles (and that is
where most toxicants are normally found) are not sufficiently bioavailable to be of concern.
However, there is evidence (acknowledged in the guidelines) that some sediment-feeding organisms
take up toxicants from ingested sediment particles. There is equally clear evidence that this is not
always the case, but in chronically turbid aquatic environments, sediment feeders (such as microcrustacea and other planktonic organisms) could play a vital role in natural food webs and, unless
proven otherwise, it would be prudent to check for such effects. Our current inability to predict the
importance of sediment feeders highlights the urgent need for process-level ecological research in the
tropics.
•
Many of the principles outlined in the guidelines are mainly applicable to large well-mixed water
bodies such as estuaries, lakes and large perennial rivers where biophysical conditions vary only
moderately over time and space. As discussed in elsewhere, most of the time rivers in the dry tropics
behave like small, poorly-mixed streams and drainage systems are fragmented into numerous
ostensibly separate small water bodies. Each water body functions independently to some extent and
biophysical conditions within them are commonly subject to enormous variation. For example, in
extreme cases, over the course of a normal day dissolved oxygen concentrations in a water hole can
fluctuate from zero to supersaturation, temperature can vary by more than 12oC, and pH can rise from
around 7 to values in excess of 10. Furthermore the range of values obtained by sampling at different
126
•
depths in the water column at one instant in time can be just as large. Many of the procedures
recommended in the guidelines for developing site-specific trigger values entail the use of chemical
equilibrium models and published toxicological data to predict the effects of parameters such as
dissolved oxygen, pH and temperature on the bioavailability and toxicity of contaminants. In the
waters of dry tropical river systems it would often be a substantial undertaking to simply quantify the
status of these modifying parameters, let alone predict their effect on the toxicity of other substances.
One of the initial steps recommended for the derivation of metal trigger values is to adjust for the
antagonistic (i.e. toxicity reducing) effects of hardness, and for some common metals, algorithms are
provided which allow a modified trigger value to be calculated. The algorithms are a function of
hardness, but are said to also account for the effects of pH and alkalinity. This is based on an
assumption, which is valid in many situations, that hardness can be attributed almost entirely to
calcium carbonates. It is important to recognise though that the geochemical diversity of natural
catchments is such that this assumption will not be true in all water bodies. More importantly, inputs
of metals from mines, industrial effluents, mineralised subcatchments, acid sulphate soils and some
farms are likely to be associated with acids and sulphates. Thus, problematic metal concentrations are
often most likely to occur in situations where the assumptions behind the hardness-modifying
algorithms are least likely to be valid. Particular caution is needed in near-coastal wetlands with an
elevation less than 5m AHD that have not been surveyed for acid sulphate soil and/or where such
soils are known to be present.
Accordingly, it is concluded that collection and meaningful interpretation of data relating to toxicants in
water is likely to be a substantial and challenging undertaking and is unlikely to be achieved by
employing the kinds of routine monitoring methods that are currently used throughout most of the dry
tropics. An option suggested in the national guidelines for routine monitoring of particle-active toxicants
is to examine residues in bottom sediments and/or the tissue of aquatic biota.
Physical-Chemical Stressors
ANZECC (and Draft Queensland) WQG include in this category most of the parameters that are currently
of primary concern in most rural and many urban catchments (nutrient, sediment, turbidity, oxygen, etc).
The basic premise behind the approaches recommended for stressors is that their effect on ecosystems
accrue gradually over long periods and hence that “average” levels are a useful indicator of the potential
for adverse effects. Ostensibly it is assumed that since these parameters are not toxic that occasional
extreme values have little practical significance and hence extremes receive little statistical weight in the
analyses used for data interpretation.
Recent research findings (summarised in other sections of this report) cast serious doubts on the validity
of this assumption in the dry and wet/dry tropics and especially in wetland environments. There are clear
indications that the condition of many wetland ecosystems in the region may be largely governed by the
frequency and severity of episodic water quality disturbances during which certain stressors either reach
acutely lethal levels (eg. dissolved oxygen, pH or temperature) or induce such rapid increases in
biological activity (e.g. nitrogen, phosphorus, turbidity, organic matter) that acute problems such as fish
kills occur. As discussed elsewhere such episodes occur naturally but it is apparent that in many cases
human water and land-use practices have substantially increased their frequency and severity. Most
episodes occur during dry times when there is minimum connectivity in the system so the effects of
localised extinction events can be quite prolonged and full recovery may not be possible until the next
flow event. For example, a single brief event that causes oxygen levels in a water hole to fall to zero for
just a few hours, may be the single factor preventing the aerobic organisms living in a water hole from
surviving a drought.
Available data also indicate that the variations in oxygen, temperature and pH levels that can occur over
the course of a single day can be almost as large as those that occur over an entire season, and in some
cases spatial variations within a water body at a particular instance in time can be just as significant. (See
figures 1 and 2 for example). Dissolved oxygen levels, in particular, can fall below lethal thresholds
briefly at night and/or at certain depths in the water column long before conditions deteriorate to the point
where daytime surface values indicate the potential for harm. For this reason we are convinced that
127
random daytime spot measurements of parameters such as oxygen, pH and temperature are at best
uninterpretable, and at worst can be completely misleading.
These considerations lead to the following conclusions:
• The collection of meaningful data for parameters such as dissolved oxygen (DO2), pH and
temperature, which can fluctuate rapidly in time and space, requires the adoption of specialised
monitoring techniques and modified indicator parameters should be considered. For example,
minimum daily dissolved oxygen or maximum daily temperature can be meaningful indicators, but
randomly determined estimates of dissolved oxygen and temporal are not.
• In cases where stressor levels can potentially cause acute mortality it is very doubtful that guidelines
based on mean or median levels would be capable of achieving the level of protection being sought.
Long-term medians may be a useful indicator of some effects and should be retained, but improved
consideration of extreme values and the attendant risks of acute effects need to be incorporated into
the data assessment process. There are many alternative approaches that could be used to achieve this
end and these need to be investigated. For example, threshold values could be derived from local
tolerance test data and then statistical analyses could focus on analysing the frequency of compliance
breaches (note that breaches can occur from time to time even at pristine reference sites).
• The fact that variations in stressor levels can be severe even over very small spatial scales (eg.
microhabitats surrounding individual plants) has important ramifications for biological monitoring, as
does the fact that many ecosystems are ultimately limited by oxygen availability. Benthic
invertebrates for example, often utilise quite different microhabitats from, say, small fish. Planktonic
species on the other hand are exposed to the physical and chemical conditions in the greater water
column and large fish have limited capacity to utilise small microhabitats (most large native fish
species appear to have limited aquatic surface respiration capabilities). For this (and other) reasons
one biological community (eg. macroinvertebrates) is not necessarily a sensitive indicator of the
status of other organisms. The national WQG provide some basic generic advice on the selection of
biological indicators but they do not consider differences in the susceptibility of different indicators
to different water pressures in different environmental settings. Basically there is a need to review
the selection of biological indicators to ensure that individual water body characteristics, dominant
water quality issues and habitat preferences are taken into consideration.
•
Many anthropogenic impacts (such as flow regulation, increased turbidity, introduction of organic
matter, loss of riparian vegetation) increase the tendency for oxic stratification to develop in water
bodies. In some cases (eg. weir pools) a substantial percentage of the benthic habitat in a water body
can become severely hypoxic for most of the year essentially eliminating all aerobic productivity in
those parts of the water body that are deeper than a certain depth (in some cases considerably less
than 1 metre). Most of the organisms which once inhabited these bottom waters are still present in
the waterbody, but they are now confined to shallow sections and this undoubtedly limits the
potential productivity of the community. In such circumstances, biological indicators of the kind
recommended in the WQG, which rely mostly on presence-absence data are unlikely to detect effects
even if the total productivity of the macroinvertebrate community has become severely limited.
Supplementary indicators such as pool habitat value assessment or biomass/productivity
measurements would be needed to detect such effects.
Regardless of the statistics used for data analysis, reference data are required in order to obtain
meaningful interpretations. Selection of appropriate reference sites (and reference conditions) is crucial
if accurate evaluations are to be accomplished – the use of an excessively degraded reference site will
result in water quality objectives being set too low to be adequately protective, while an unusually good
quality site may lead to unachievable objectives being set.
Based on the research findings discussed in and personal experience, it is clear that the available
reference data, including those contained in draft Queensland water quality guidelines, cannot validly be
used to derive meaningful trigger values for the majority of dry tropical wetlands including riverine ones.
The refinements proposed for future versions of the Queensland guidelines, which include the
incorporation of a larger number of water types and division of data into different flow regimes and
seasons, could improve this situation, but the methods used will need to be substantially more
sophisticated than those that have traditionally been employed in water quality data analyses.
128
At a minimum, flow divisions need to be capable of discriminating between the discrete stages of the
hydrograph identified in the conceptual models proposed in Section 2 of this report. Note that the
nutrient and sediment concentrations observed during these flow states indicate quite different things, so
there are valid grounds for questioning the logic of combining the data together (as is almost always
done). Note, for example, that under higher flow conditions low nutrient concentration would normally
be considered to be an indication of good quality; however, under lower flow conditions very low
nutrient concentrations may be obtained from heavily impacted eutrophic sites where productivity is
dominated by attached plants and algae (which assimilate nutrients from the water column very
efficiently).
Strikingly different states can exist even at this level of flow division. For example our research on the
Townsville Field Training Area (TFTA) has identified three sets of conditions that result in very
significant differences in water quality expectations in stagnant water holes (i.e. three subdivisions of low
flow conditions). This project also demonstrated that seasonal division of data was an ineffective water
quality discriminator in the Burdekin study area.
The above approaches provide a potential framework for meaningful data analysis and for assessing the
inherent vulnerability of sites to particular kinds of impacts under different sets of conditions. In our
experience one of the most important factors to take into consideration when selecting reference sites is
that the quality of water in a river reach or wetland that normally receives infrequent, sluggish flows of
runoff from a semi-arid lowland catchment could never be expected to be as good as the water in a stream
that is frequently flushed out by high flows generated from undisturbed coastal ranges. Significant
differences of this kind can occur over quite small spatial scales. For instance, Keelbottom Creek and
Fanning River, adjacent tributaries in the upper Burdekin catchment, exhibit such differences.
Clearly there is an urgent need to develop a more sophisticated typology for water bodies if the
approaches advocated in existing and emerging water quality guidelines are ever to become an effective
management tool. ANZECC and ARMCANZ Guidelines acknowledge this need and identify the
development of an improved typology as a priority for future research.. An improved system is essential
not only for reference data selection but also as an aid to holistic assessment of vulnerabilities, threats and
issues.
129
Figure 1
Physico-chemical data logged at 20 minute intervals at site S25 on Keelbottom
Creek from 12:00 on 06-10-99 to 19:00 on 07-10-99
8.0
32
110
7.9
31
100
7.8
30
7.7
29
1.6
1.4
90
80
1.2
70
pH
7.5o
7.4
Temperature ºC
60
7.3
27
50
26
40
25
1.0
0.8
Conductivity (mS/cm)
28
Dissolved Oxygen (%Sat)
7.6
30
7.2
24
20
7.1
23
10
7.0
22
0
12:00
16:00
20:00
00:00
04:00
08:00
12:00
16:00
0.6
0.4
20:00
Time
0
Depth, m
1
2
3
4
0
10
20
30
40
50
60
70
80
90
100
110
120
DO, % Sat.
6:58AM
4:34PM
Figure 2
8:23AM
5:24PM
10:00AM
6:17PM
12:50PM
2:14PM
3:33PM
Inkerman Lagoon daily cycling, Site 3 (downstream) on 28-03-02
130
APPENDIX I
OVERVIEW OF WATER QUALITY PROCESSES AND MANAGEMENT
STRATEGIES IN WATERWAYS OF THE BURDEKIN REGION
131
I1.
OVERVIEW OF WATER QUALITY PROCESSES AND MANAGEMENT
STRATEGIES IN WATERWAYS OF THE BURDEKIN REGION
I1.1
Contaminant Exports from the Catchment
The overwhelming majority of water quality-related research and monitoring in the dry tropics (and
indeed most Great Barrier Reef catchments) have focused on the issue of riverine contaminant inputs into
the Great Barrier Reef (GBR) lagoon. Efforts to date have concentrated mainly on monitoring and/or
modelling end-of-river sediment and nutrient export loads. Little attention has been devoted to
determining rates of export of other contaminants, managers electing instead to monitor residues in the
bottom sediments and biota or changes in ecological indicators in receiving waters. However, it is
generally accepted that the majority of toxicants released into the GBR lagoon (with the major exception
of dissolved species of nitrogen) will be attached to very fine sediment particles when they first enter the
marine environment. This significantly increases the potential importance of sediment export
management.
The following general conclusions can be drawn from the available research literature:
• Riverine runoff is the largest external source of nitrogen and phosphorus to the GBR lagoon and is
the main source of sediment and nutrients that is directly influenced by human land use and water
management practices.
• Over the long term the majority of water, nutrient and sediment discharged into the GBR lagoon
originates from the dry catchments of the Burdekin and Fitzroy rivers.
• Even in the wet tropics, the majority of export occurs during storm events, usually associated with
cyclones and tropical lows. In the Burdekin, Fitzroy and other dry catchments, the quantities
discharged at any other time are negligible. Accordingly ambient water quality data (which are
currently gathered almost exclusively during low flow periods) do not provide useful information
about end-of-river export rates.
• Flood pulses generated by runoff from the wetter northern (Burdekin) and south-eastern
(Bowen/Broken) subcatchments of the Burdekin River pass through the river system with remarkable
speed and as a consequence the majority of contaminant export typically occurs over the course of a
few days to a week. The flood hydrographs generated by runoff from the drier, lower gradient
Belyando/Suttor subcatchment rise and fall more gradually, but most sediment discharge still
generally occurs over no more than a month or so. On rare occasions prolonged rainfall or
consecutive storm events can extend storm hydrographs out to one or two months but even in these
cases the majority of sediment export is generally associated with a few brief flow peaks of no more
than a few days’ duration. Many of the shorter rivers and creeks that rise in the coastal ranges have
been flashier hydrographs and can often be expected to maintain high discharge levels for periods in
the order of hours rather than days once rainfall intensity in the catchment declines. Even low
gradient systems that rise on the coastal plains 89e.g. Lagoon Creek) exhibit rapid hydrographic
responses to intense rain events, discharge rates falling by one or two orders of magnitude within a
few days of the cessation of rainfall it is impossible to obtain meaningful contaminant export data
from such sites without implementing automated sampling programs.
•
Nitrogen export rates are much less well understood as there are very limited monitoring data
available for large scale flow events in the dry tropics. Available data indicate that the correlations
between particulate and dissolved forms of nitrogen are significantly different from those observed in
Wet Tropics rivers. Accordingly it does not appear to be valid to extrapolate findings from more
intensively studied rivers such as the Johnstone.
Despite the erratic episodic nature of discharges from the dry GBR catchments and the enormous interannual and inter-decadal variability of export patterns, most attempts to model sediment export have
focused on average annual load estimates. Given that human management decisions are generally
required to deal with current situations and projections for the immediate future, it is doubtful if longterm average estimates provide a very meaningful management tool in the dry tropics – the reality that
total inputs into the GBR lagoon from the dry tropics can be negligible over dry decades and enormous in
a flood year is inescapable. Moreover, total exports appear to be dictated by large-scale flood events with
132
a recurrence interval in the order of 8 to 20 years depending on the system and the contaminant flux rates
that occur during such events are so high that minor variations in the intensity, duration and location of
rainfall could lead to export load variations large enough to be equivalent to of low flow discharge.
For example Belperio (1979) estimated the following end of river sediment export values for the
Burdekin River prior to the construction of the Burdekin Dam:
Sediment Load
Tonnes x 106
Washload Bedload
Average Year
Flood Year (1957-58)
Drought Year (1968-69)
24hr Peak Discharge
(3, 600m3/s in 1974)
I1.2
3
0.45
19.7
3.7
0.008
0.001
6.4
1.7
Contaminant Exports to Estuarine Wetlands
To date virtually all research dealing with downstream impacts of catchment runoff has focused on
receiving environments outside of the river mouths (i.e. within the GBR lagoon itself), even though the
boundaries of the Great Barrier Reef World Heritage Area (GBRWHA) are currently defined by the low
water mark and therefore encompass the river estuaries. It is however, patently clear that estuaries are
subject to far greater and more direct anthropogenic pressures than any other marine waters, and that the
risk of water quality-related damage to estuarine ecosystems is substantially higher (in some cases by
orders of magnitude) than it is in other parts of the GBRWHA.
It has been clearly established that in terms of total quantities the overwhelming majority of water,
nutrients, sediment and associated contaminants discharged into the GBR lagoon from dry and wet/dry
catchments are delivered into the sea via the main river channels during large flood events. As discussed
previously, contributions from the relatively small, near-coastal catchments are disproportionally high
due to their close proximity to the sea, but the inland catchment areas of large rivers like the Burdekin
and Fitzroy are so large that they are still the main source of contaminant exports. Furthermore the total
quantities of sediment carried to the marine environment rivers in baseflows are negligible in virtually all
catchments.
It is evident that a very large proportion of the solute and very fine, potentially bio-active, sediments
carried by floods will pass so rapidly through the river estuaries and creeks that they should be expected
to have little, if any, impact on the ecosystem (provided that acutely stressful conditions do not develop).
A significant proportion of this material appears to flocculate and settle before salinity levels in river
plumes increase to about 10 and during floods this generally happens outside the river mouths. (There
may potentially be significant deposition of silt-sized particles in the estuarine parts of the floodplain
during the peak of floods. These are important from a geomorphological and ecological viewpoint but
are much less relevant to water quality). Some of the sediments that flocculate off-shore may, depending
on prevailing conditions, be carried back into the estuary from which they came but most will be
gradually redistributed further north. Unflocculated material, on the other hand, will disperse with the
river plume and can be expected to travel large distances in a short time.
During the falling limb of the hydrograph, stormwaters in the estuary will begin to be displaced by
seawater, and salinities will rise to levels high enough to induce flocculation. This means that there will
be a sudden increase in the probability of bioactive sediments being deposited within the estuary. At the
same time the residence time of stormwater within the estuary increases, providing greater opportunity
for bio-assimilation of dissolved contaminants (including those which desorb from fine sediments). The
potential for contaminants to become involved in ecological processes within the estuary would also be
133
expected to increase as flows fall, and ultimately the relatively small amounts carried into the system
when the river is at baseflow would be expected to exert disproportionate effects on the ecosystem.
Similarly, the contaminants carried into the estuary in runoff from small localised storm events, have the
potential to significantly affect the ecosystem even though the amounts involved represent a negligible
proportion of the total quantities that pass through the system in the longer term.
There is evidence from research conducted in other regions that a significant proportion of the
productivity of estuaries can be attributed to utilisation of fluvial nutrient inputs, so it is likely that
anthropogenic changes to the quantities and/or the timing of delivery of riverine nutrients to estuaries will
have significant effects on the ecosystem. It must also be recognised that unlike coastal waters, estuaries,
especially their upper parts, do not receive the full benefit of the dilution, mixing and re-aeration capacity
potentially available from the sea. During spring tides the tidal wedge can penetrate significant distances
into the estuary and this provides reasonable mixing, dilution and dispersion in the lower estuaries.
However, in the tributaries and intertidal wetlands at the top of the estuary, mixing can be quite poor, and
it is not uncommon to find that the ebb and flow of tidal seawater simply pushes estuary water back and
forth. (During the dry season this can be evidenced by the development of inverse salinity gradients).
Moreover, during extreme neap tide sets (i.e. more than one third of the time in many parts of the dry
tropics) there is very little tidal movement and mixing can be poor even in the lower estuary.
At places and times where mixing is poor, biological processes (such as respiration) can exert large
effects on water quality. This, coupled with low dilution and dispersion capacity, can create conditions
where the ecosystem is vulnerable to acute, episodic water quality impacts. Other factors being equal the
probability of this occurring rises rapidly with increasing distance from the open sea.
The above considerations lead to the general conclusion that individual estuaries can be expected to be
most affected by runoff generated from their own local subcatchment areas and that some of the most
adverse affects could be associated with quite small localised events. This is a significant conclusion
because in many cases the local subcatchments surrounding the estuary are completely dominated by
intensive agricultural and/or urban developments. The small sinusoidal channels and intertidal wetlands
at the top of the estuary are at much higher risk than the larger main channels but because of their
comparatively small size many investigators tend to ignore them. It is important to recognise, however,
that these parts of the system provide specialised aquatic habitats that are poorly represented, sometimes
absent, from the lower estuary. Moreover, the cumulative area of benthic habitats provided by all these
small systems can amount to a very significant total, indicating that they are significant contributors to
the total productivity of the estuarine ecosystem.
From a water quality viewpoint the estuaries of the Burdekin floodplain fall into the following broad
groups based on existing differences in the flow regimes of the systems that feed into them:
1. The right bank of the floodplain (from Yellow Gin Creek to the Elliot River). The Burdekin River
flood plume often extends as far south as Cape Upstart and has reached the Whitsundays on
occasions, (Devlin et al, 2000) so these small southern estuaries may potentially receive some inputs
of fine particulate and dissolved contaminants from the Burdekin River via the sea.
Here streams are largely unsupplemented and ephemeral. Presumably they have always delivered
sediment and nutrients to their small estuaries mainly as pulses during storm events. Landholders
construct temporary bund walls to capture the last of the freshwater flows in several of these systems
but it is doubtful if baseflows were ever sufficiently prolonged for this to significantly affect the
nutrient budget of the estuary. Runoff from horticultural developments on the levee banks of
Molongle Creek and the Elliot River drain away from the main watercourse. This protects the river
from adverse water quality effects but there is potential for agrochemicals to accumulate in the small
coastal receiving streams nearby (eg. RM Creek). Apart from this, anthropogenic impacts seem
likely to be confined to increases in the size of the nutrient and sediment pulses delivered from
grazing lands in the coastal catchment areas during storms. All of the creeks maintain coarse basal
sands all the way to the estuary indicating that fine sediments are transported very efficiently during
floods. Hence small scale stormevents in the coastal catchments are almost certainly the main source
of contaminant loading within the estuaries. Since such events are relatively rare these estuaries are
134
relatively safe from fluvial inputs of anthropogenic contaminants even through some parts of the
catchment are in poor condition.
2. Floodplain creeks and streams that are used to distribute irrigation water (eg. Plantation Ck,
Sheepstation Ck, Warren’s Gully, Iyah Ck, Kalamia Ck, Pink Lily Lagoon).
These once ephemeral streams now support permanent flows of Burdekin Falls Dam BFD water.
This water is turbid but otherwise of good quality when it first enters the streams, but water quality is
currently being spoiled by deoxygenation and other related problems caused mainly by exotic weed
invasions. Levees have been constructed at the top of the estuary in most of these systems in order to
prevent ingress of seawater into the lower reaches of the streams. This has caused the formation of
extensive artificial freshwater wetlands which are now choked with exotic weeds and/or miscreant
monocultures of Typha. These changes could alter the nutrient budget of the estuaries in many ways,
but much closer investigation would be required to determine precise effects.
Perhaps more importantly, the poor quality waters ponded behind the levees are hypoxic and as a
result contain significant concentrations of potential toxicants such as hydrogen sulphide. This water
has the potential to create acutely lethal conditions if suddenly released into the estuary. The risks of
this happening will be greatest during small storm events and/or the initial rising stages of a large
storm event. This is an issue that warrants closer investigation.
3. Floodplain creeks that are used to drain irrigation tailwater and storm waters away from cane farms
(Barrattas Creek and associated lagoons such as Collinson’s and Didjeridoo).
The most closely studied example of this is the Barrattas Creek system. This was once an ephemeral
system which retracted into a series of permanent lagoons during the dry season. The creek is now
perennial and supports high baseflows and frequent swift flows due to supplementation by tailwater
releases. Nutrient concentrations in the tailwater are extremely variable but, on average, approximate
the levels that would be encountered on the falling limb of the hydrograph in a stream that is subject
to moderate grazing pressure. This is significantly higher than would be expected in a similar stream
if it was at baseflow. Regardless of the significance of these concentrations (which is yet to be
determined) artificial flows now constantly supply nutrients to the estuary during the dry season
when natural inputs would have been negligible, and this must have significantly increased dry
season nutrient loading in the estuary.
There is also a high potential for tailwater flows to contain agrochemicals such as pesticides and
heavy metals and for occasional episodic pulses of Biochemical Oxygen Demand (BOD), severely
hypoxic water and/or high ammonia concentrations, each of which could be acutely lethal to marine
organisms under certain circumstances. Since these effects occur mainly as pulses (determined by
the timing of irrigation and rain events relative to cropping operations on individual farms) they are
unlikely to be detected, let alone quantified, by the grab sampling methods currently being employed
in routine monitoring programs. Even occasional brief occurrences of acutely toxic conditions can
have prolonged effects on biological community structure and function. In cases where they recur,
such episodes could be the main factor governing the condition of the ecosystem in the long term.
There is therefore an urgent need to develop and implement more suitable monitoring techniques.
4. The tidal reaches of the Haughton River.
This estuary benefits from connections with the extensive Cromarty Wetlands system to the north and
the large relatively unaltered Barramundi Creek estuary system to the south. Tidal circulation
appears to provide reasonably efficient dilution and dispersion at the confluence of the Haughton
River and Cromarty Creek (during the dry season hypersaline waters indicative of poor mixing are
usually confined to reaches of Cromarty Creek that are well upstream of the junction). In the
aftermath of flood events freshwater discharges from the Cromarty channel provide additional
flushing for periods of several weeks to a few months.
135
It is therefore concluded that storm waters discharged into the estuary from the Haughton River will
be dispersed and diluted more efficiently than they are in most regional estuaries. Hence, the
Haughton estuary is expected to be less vulnerable to water quality problems than most others in the
area. Modifications to baseflow inputs due to the presence of the Haughton Weirs (Giru and Val
Bird) will probably have affected dry season nutrient processing in the tidal reach between the weir
and Cromarty junction, but since natural baseflows were quite low and baseflow modifications are
minor it is unlikely that effects would be measurable further downstream.
Currently there are virtually no interpretable long term ambient water quality data available for the
estuarine wetlands in the dry tropics. A fortnightly grab sampling program was instigated in the
Burdekin River Irrigation Area (BRIA) in 1988. This program, which continued for more than a decade,
originally included two estuarine sites but it became apparent that the data would be uninterpretable due
to natural variations in tidal conditions, etc., so monitoring focused entirely on freshwater sites during
most of the program. The freshwater data provide some indication of the concentrations of nutrients
being carried towards the estuary but due to limited flow rate data it is difficult to confidently determine
flux rates. Moreover, it is impossible to tell how much of this reached the estuary, when it arrived, or
what form it was in when it got there.
Recent research has shown that dissolved oxygen availability is the most important water quality issue in
the freshwater wetlands of the dry tropics. It is possible, perhaps likely, that in many places this problem
extends downstream into the estuaries, the surveys conducted in conjunction with the current study have
demonstrated that the inherent tendencies towards the development of hypoxia that is so evident in
freshwater wetlands are also obvious in estuarine wetlands. There was inadequate rainfall during the
coarse of this study to be able to determine if the severe oxygen sags experienced in freshwater wetlands
during small stormevents are transferred into the estuary, but the potential for adverse effects of this kind
appears to be high, so detailed monitoring of oxygen dynamics should be treated as a priority for future
estuarine investigations. This would require the adoption of specialised monitoring procedures because
the levels of oxygen (and temperature and pH) within the water column can fluctuate massively over very
small scales of time (eg. minutes to hours) and space (centimetres to metres). The spot measurements
currently being collected in regional monitoring programs cannot be interpreted and, because
measurements are inevitably taken during daylight hours when oxygen levels are highest, are usually
misleading.
It should also be noted that there are currently no Australian or Queensland water quality guidelines
available for estuaries. At the moment investigators generally need to employ professional judgement to
determine if the ANZECC (2000) freshwater or marine guidelines are most applicable in any given
situation. However, based on our limited studies in this area to date, we doubt whether either freshwater
or marine water quality criteria are strictly relevant. In contrast, we believe that marine sediment quality
guidelines should generally be applicable in local estuaries and it is expected that reference values for
estuarine sediments will be included in the forthcoming Queensland Water Quality Guidelines.
Due to the episodic nature of toxicant inputs into estuaries, water sampling can be a very inefficient and
expensive monitoring technique, so residue monitoring is generally a preferred alternative. However, a
few naturally occurring toxicants such as ammonia and hydrogen sulphide degrade without leaving
signature residues, so water sampling is the only means by which they can be detected. In this case the
most practical (and affordable) approach is to conduct detailed risk evaluations to propose hypotheses
about the places and times when high concentrations are most likely to occur, and then implement brief
targeted sampling exercises to test these hypotheses. Most toxicants such as metals and pesticides, on the
other hand, will accumulate to some extent in benthic sediments and biota. Sediment and tissue residues
are a reasonably reliable indicator of exposure to these kinds of contaminants.
I1.3
Ambient Water Quality in Freshwater Drainage Systems
The brief intense flow pulses which dominate the hydrograph in most dry and wet/dry drainage systems
have important ramifications for water quality assessment and management. The most important of these
are:
136
•
•
•
•
The potential for impacts on the marine environment depends almost entirely on the quality of
stormwater generated during flood events.
The resource potential of surface waters depends mainly on the quality of water discharged during
the falling limb of large flood events and/or the rising and falling limbs of smaller storm events.
(This is the water which is most likely to be harvested in large quantities or captured in
impoundments; some water may be harvested during the rising stages of large floods but high
suspended sediment loads and bed sediment movements make pumping difficult and these very high
flows are also sustained only briefly).
Estuaries are also likely to be affected mainly by contaminant inputs associated with small events and
the falling limb of flood hydrographs or by baseflows supplemented by irrigation runoff land in a few
cases of industrial or urban effluent discharges.
As a general rule (to which there may be a few important exceptions) instream productivity is mainly
governed by what happens during very small storm events and the very low flows that are present
most of the time.
Water quality research in most of the dry tropics primarily deals with contaminant flux estimates and
relates mainly to the evaluation of downstream, rather than instream, impacts. There have been few
previous attempts to examine or establish links between water quality, ecosystem conditions and
biophysical processes in wetlands in this region of Australia, and even basic water quality data are
unavailable in most cases. Ambient water quality data are available for some major tributary streams and
river reaches, mainly through state agencies (NRM and EPA). These have been collected using standard
routine monitoring techniques which have proved to be effective in many parts of the world and in many
types of water body (eg infrequent random or irregular interval grab samples and spot measurements).
However, our research indicates that there are sound grounds to doubt if such data can be interpreted in
an ecologically meaningful way in most of the dry and wet-dry tropics. (This situation is not a criticism
of the agencies involved, who have correctly followed well established historical conventions; we are
challenging the paradigms upon which these conventions are based).
The reasons for the current deficiencies are numerous but the following factors have a significant bearing:
•
Conventional monitoring practices have been developed mainly in temperate regions of the world
and are primarily based on experience gained from monitoring large, well-mixed water bodies such as
large coastal embayments, lakes and large perennial rivers. Detailed studies of smaller water bodies,
especially in the tropics, have been confined mostly to swiftly flowing well-mixed perennial
mountain streams. Dry tropical rivers are large well-mixed systems for only a few weeks to a few
months a year (at which time instances discharge rates can rival most of the world’s rivers)– most of
the time the systems comprise small shallow streams, a few sluggish, poorly mixed deeper reaches,
numerous small very poorly mixed water holes (both instream and off-channel), permanently
inundated, ground water dependent floodplain wetlands and ephemeral/intermittent wetlands that are
dry most of the time. Systems of this kind are very poorly represented in the water quality literature.
•
Ecologically based water quality monitoring is a relatively new undertaking, especially in tropical
Australia. Many of our water quality scientists gained their experience in resource assessments and
hydrological investigations and have had limited exposure to the principles of aquatic ecology. One
of the most obvious consequences of this is a tendency to focus monitoring and management efforts
on large water bodies and to attempt to collect samples representative of the majority of water volume
contained in an aquatic system (this is usually the appropriate strategy when the requirement is to
quantify contaminant fluxes or assess resource potential; the value of a resource is after-all,
proportional to the volume available for use). In contrast, the productivity of aquatic ecosystems,
especially shallow ones, is usually much more closely linked to the surface area and quality of
benthic habitats than it is to water volume. Aquatic productivity, for instance, is normally measured
in terms of carbon fixation per unit area, not per unit volume. This dictates the adoption of quite
different criteria when evaluating the relative importance of various parts of the system and/or
deciding when or where to focus investigations. For example, a relatively large (by dry tropical
standards) turbid river reach with coarse sandy bottom is likely to support much lower levels of
productivity than a comparatively small off-channel billabong or a series of small permanent water
holes in a nearby tributary. No attempt has yet been made to determine an energy budget for a local
137
river system; however, based on published data the cumulative contribution from numerous small
water bodies to the total area of productive benthic habitats in the system as a whole, is likely to be
substantial. Moreover, such water bodies often provide habitats that are poorly represented or largely
absent in the main river system, so they make a significant contribution to the diversity of the system
as a whole. It follows that the current tendency to focus research and monitoring efforts almost
entirely on the main river channel is inappropriate if general aquatic ecosystem health is being
assessed.
•
•
Monitoring sites that are to be used to quantify fluxes of water and/or contaminants must be
strategically located within the drainage system to ensure that the inputs from various subcatchments
are adequately captured. Additionally, because flow data are needed for flux estimates, the precise
location of the monitoring station will usually be dictated by stream gauging requirements.
Ecological/ambient monitoring sites on the other hand, should ideally be located in areas that best
represent the dominant and/or most valuable aquatic habitats in each area. In large, relatively
homogenous perennial systems these two kinds of sites commonly coincide so it is feasible to utilise
a single monitoring site network for all purposes. However, this is rarely the case in the spatially and
temporally heterogeneous systems of the dry tropics.
Effective water quality monitoring program designs need to be based on a conceptual understanding
of the main biophysical factors and processes that govern water quality and ecological conditions.
ANZECC and ARMCANZ (2000) WQG and ANZECC (2000) Guidelines for Water Quality
Monitoring and Reporting, both promote the use of holistic conceptual models to assist in issue
prioritisation, and the subsequent selection of appropriate indicators and monitoring regimes.
However, few of the examples and case studies contained in these and other reference documents are
applicable in this region, and it is only recently that we have begun to develop sufficient
understanding of local environments to be able to propose viable alternatives.
I1.4
Water Quality and Management in Floodplain Wetlands
I1.4.1 Unsupplemented near-coastal wetlands
On their way to the sea, the contaminants released from highly disturbed coastal subcatchments may be
diluted and dispersed, firstly by influxes of storm waters originating in undisturbed parts of the local
watershed; secondly by riverine discharges arising from the greater catchment area; (which are often
large enough to overwhelm iputs from small coastal catchments) and, ultimately, by the immense
volumes of seawater in the marine environment. At the same time contaminants participate in numerous
biophysical transformations such as settling, oxidation, sorption or desorption, bioassimilation,
decomposition, biochemical or chemical conversion, precipitation or volatilisation. These processes
generally guarantee that the concentrations of contaminants introduced by runoff from intensively
developed subcatchments decrease incrementally with increasing time and distance from the source.
Such processes play a crucial role in determining the susceptibility of different aquatic ecosystems to
impact from various anthropogenic contaminant sources. Freshwater wetland systems, which are the
initial recipients of contaminant inputs in many intensively developed areas, are subject to greater risks of
harm than any other regional ecosystems. These highly vulnerable wetlands include the numerous
tributary streams, coastal distributary creeks, lagoons and swamps that drain subcatchments dominated by
human activity. Most of these discharge into much larger, less vulnerable water bodies that are often
considered to be more ecologically and socio-economically important due to their larger size. However,
many of the smaller wetland systems play a crucial role in maintaining the integrity of regional aquatic
ecosystems – they may, for example, provide breeding and nursery habitats that are virtually absent from
large rivers and/or estuaries. These aquatic ecosystems must currently contend not only with frequent
and prolonged exposure to poorly diluted stormwater and wastewater inputs, but also a unique variety of
other natural and anthropogenic stresses. Most of the water bodies in question are relatively small,
poorly mixed, poorly aerated, and support high levels of natural productivity. They are therefore
inherently prone to the periodic development of stressful water quality conditions. In the pre-European
landscape, wetland habitats were so abundant and diverse that these natural tendencies would not have
presented a problem. The occurrence of “poor” water quality at certain times and places within the
complex wetland mosaic probably played a vital role in the maintenance of high diversity by providing
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habitat for organisms specifically adapted to conditions that are stressful to competitors; and in habitats of
this kind, episodes of catastrophic water quality failure would simply have provided opportunities for
predators and scavengers to feed more easily. Such events were sustainable because there were other
refuge habitats nearby that were more than capable of maintaining regional fish stocks and biodiversity.
The contemporary situation is very different. Most wetlands represent mere fragments of the extensive
habitat mosaics that once occupied most watersheds, especially along the now intensively developed
sections of the coastal plains. The progressive destruction and reclamation of wetlands since European
settlement, and especially the accelerated losses due to sugar industry expansion during the last quarter of
the 20th century, have been well documented. In many subcatchments, wetland habitats that were once
dominant features of the landscape (e.g., Melaleuca (paperbark) swamps) are now either absent or exist
only as small isolated remnants. Most commentators have pointed to the resulting loss of stormwater
detention and contaminant removal capacity as the major adverse impact of these changes. The
associated losses in fishery productivity and regional biodiversity have also been identified as cause for
concern, but the effects on the fundamental ecology of the freshwater wetland ecosystems and the
consequent implications for best management of the remaining habitats have received little attention. For
this reason some important points are often overlooked, as follows:
•
•
•
Most of the remaining wetlands in agricultural areas have been retained because they serve as natural
drains and/or useful water reservoirs, and in most cases they have been altered (by drainage works,
installation of flow regulation structures, tidal barriers, culverts etc) to enhance these capabilities.
Selective retention has thus favoured a few wetland types and these have often been modified in
ways that make them much more similar to one another than they once were. Accordingly there has
been greater loss of habitat diversity than most published figures suggest.
The condition of remnant wetlands is poorly documented, but it is clear that a wetland is not
necessarily capable of performing any of its natural functions simply because it still retains water.
Our research on the Herbert and Burdekin floodplains indicates that, because of loss of riparian
vegetation, aquatic weed invasions and hydrological modifications due to irrigation, the majority of
wetlands are severely degraded. In most cases it is unlikely that they would be able to function
normally even if the quality of water flowing into them was excellent.
Habitat degradation is so widespread that it has been impossible to locate sites that might represent
how these systems would have functioned under natural conditions. Therefore, we have had to resort
to studies examining the effects of different types of degradation to piece together some
understanding of the underlying natural processes.
Habitat loss and degradation, linked with other pressures, such as fishing, have created a situation where
it is extremely doubtful that any wetlands could ever be successfully restored to their natural state.
Moreover, in these heavily altered landscapes it is very likely that the best possible ecological and socioeconomic outcomes would be achieved if remnant wetlands were actively managed to ensure that they
take on new functional roles that better compensate for the overall accrued losses. Such management
measures would require establishment and maintenance of new, essentially artificial ecosystems, capable
of sustaining the ecological values desired by the community, and might involve the construction of new
wetland habitats. Strategies of this kind are already being trialled in several areas, ( in conjunction with
initiatives such as Landcare, NHT, SIIP) but it is too early to judge if their objectives are being
accomplished. Nevertheless, it is clear that the functions and needs of tropical wetland systems are not
sufficiently understood for managers to be able to make informed decisions.
There is a widely held view that management to minimise contaminant export to the marine environment
will automatically be beneficial to freshwater ecosystems (and vice versa). This assumption is rarely
valid; in fact there are many instances where the two objectives can be mutually exclusive. For example,
total retention of agricultural contaminants within the freshwater drainage system would effectively
eliminate exports to the marine environment, but in the process, freshwater habitat values could be totally
destroyed; moreover, the resulting loss of fishery productivity and ecosystem connectivity would almost
certainly impact on marine food webs so the net benefit to the marine environment would be questionable
at best.
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The identification of management measures that are beneficial to all aquatic ecosystems, and the ability
to reach appropriate compromises when such options are not available, require a well developed
understanding of the ecology of the region, and knowledge of the tolerances and vulnerabilities of
individual ecosystems and their components.
Current National Water Quality Guidelines (ANZECC 2000) promote the idea of holistically evaluating
the factors that determine aquatic ecosystem condition in order to ensure that management is directed at
the most important issues. This advice is particularly pertinent to regional wetlands and underpins most
of the research we have carried out in recent years.
The current research project has focused on identifying and examining the most important water-qualityrelated issues that affect wetlands that are the primary recipients of anthropogenic contaminants. The
biophysical factors and anthropogenic influences that make these wetlands particularly vulnerable to
ecotoxicological impacts are also major influences on the relative importance of various water quality
processes. Consequently, the main issues of concern are somewhat different from those that have been
identified for other aquatic ecosystems and especially adjacent marine environments.
I1.4.2 Burdekin floodplain
a)
Water quality influences
Coastal floodplain
Many of the aquatic ecosystems on the coastal floodplain are currently in very poor condition. Factors
such as clearing of riparian vegetation, introduction of exotic weeds, drainage alterations and
hydrological modifications, combine to undermine natural ecological functions and increase the
vulnerability of water bodies to water quality problems. There is unequivocal emerging evidence (see
Perna 2003 for example) that water quality problems exist, but the extent to which they can be attributed
to various potential causes is only now being studied in detail. Nevertheless, there are clear qualitative
indications that the following issues dominate.
Water management and use
Several of the coastal creeks that were once ephemeral have been transformed into perennially flowing
waterways due to constant distribution of irrigation water (eg. Sheep Station and Plantation Creeks) or
tailwater disposal (eg. Barrattas Creek). Theoretically this increase in flow could increase re-aeration,
dispersion and dilution rates and hence, reduce the vulnerability of these systems to certain water quality
problems. In practice three factors act against this:
1. Water usage and demand, and therefore flow rates, often fall to a minimum when rainy
conditions prevail. This results in a sudden increase in vulnerability at the times when water
quality threats are highest.
2. Exotic weeds and Typha monocultures have invaded (or are invading) large sections of most
distributary channels and wetlands. These very efficiently strip oxygen from the water column
so artificial baseflows simply distribute hypoxic water downstream. The risks of other water
quality factors causing damage to these ecosystems is currently very low, simply because they
have already been degraded to the point where few sensitive organisms remain. However, weed
removal and rehabilitation programs are currently underway and initial results indicate rapid
recovery of biodiversity is achievable. There are also clear indications that artificial flows in the
distributaries become a valuable source of oxygen for downstream ecosystems, if weeds are
removed from upstream reaches. The chronic turbidity of BFD irrigation water substantially
alters the ecology of receiving waters and is almost certainly one of the factors contributing to
the invasive success of exotic weeds (which appear to cope with constant turbidity much better
than most native plants). However, it appears that a functional ecosystem, capable of supporting
those elements of natural ecological function that are of most concern to the local community
140
(eg. nursery habitat for barramundi), could be maintained if other problems such as weed
invasions and loss of riparian habitat, are redressed.
3. Virtually all of the waterways on the floodplain other than major tailwater drains like the
Barrattas Creeks have been cut off from the estuary through the construction of weirs, bund
walls and levee banks. Hence watercourses which once entered the estuary through super-tidal
salt flats and intertidal mud flats, which offer little resistance to the passage of stormwaters, now
terminate in permanently inundated freshwater ponds which have become choked with weeds.
These ponds greatly increase the residence time of poor quality first-flush water. Even with the
weeds removed efficient flushing will be delayed significantly while the pond is filling, so the
highest level of protection would only be achieved by removing all flow obstructions. (This
strategy would provide the additional benefits of allowing fish passage and ingress of seawater
which would prevent the re-establishment of most freshwater weeds). It is understood that the
North Burdekin Water Board may be prepared to trial rehabilitation works of this kind in
Plantation Creek.
Note that the concentrations of nutrients in tailwater can at times be considerably higher than they are in
the irrigation water. However, most of the time flow rates in the irrigation channels are substantially
higher than they are in tailwater drains. Hence the total quantity of nutrients passing through the system
may be quite similar in both cases.
Farming and irrigation
•
Laser-levelling of farms, reclaimation of wetland sump areas and drainage improvements, etc.
have undoubtedly increased the rate at which water-borne contaminants enter natural
watercourses and this is likely to exacerbate the effects of first-flush runoff. Some farms use
recycling ponds capable of capturing a small quantity of the initial runoff but the net benefit of
these has not been studied.
•
Recent work by the Sugar CRC, CSIRO Land and Water (CLW) and ACTFR has shown that the
runoff from cane farms can at times contain dangerously high levels of BOD. This BOD appears
to be attributable mainly to sugar (and its degradation products) deposited in fields from cane
juice released during harvesting. In most soils, microbes degrade sugars quickly so the
concentrations in soil can usually be expected to decrease rapidly after harvest. The dynamics of
this have not been examined closely but trials to date show that BOD levels in the tailwaters
generated by the first post-harvest irrigation event can be more than an order of magnitude higher
than the limit of 20mg/L that is normally specified in EPA effluent discharge licences for sewage
treatment plants. Our surveys also show that many receiving water bodies already suffer from
chronic hypoxia and are unlikely to be able to cope with the introduction of any BOD.
Much research has focused on the releases of sediment and nutrients from farms into the aquatic
environment. Cane farming requires the use of large quantities of nitrogen-based fertilisers and current
indications are that about one third of this is used by the crop and that another third volatilises. The
remainder can potentially be released into the aquatic environment. Accordingly many researchers have
identified releases of nitrogen from canefarms, especially in the dissolved inorganic form, as a major
issue for downstream effects on the GBR. It is generally assumed that there would be similar potential
for impacts on freshwater and tidal wetlands but this is a far more complex issue. Tropical marine
ecosystems, for instance, are generally considered to be nitrogen limited (i.e. the ecosystem would be
expected to exhibit immediate productivity increases in response to nitrogen inputs). The situation in
wetland environments, where other factors clearly limit productivity in various places and/or at various
times, is not as straightforward. Moreover, different plant types and/or species are often simultaneously
subject to different limiting factors (eg. floating weeds vs emergents vs submergents).
The issue of nutrient limitation in shallow wetlands is also complicated by the capacity of many plants to
obtain nutrients from bottom sediments and soil, and in environments dominated by species which do
this, it appears that the major effects of changes in nutrient levels in the water column are to modify the
plant community structure rather than gross productivity levels.
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Regardless of these complications it is clear that dry season nutrient and sediment loads have
substantially increased in all the waterways that receive irrigation water, simply because there is now
water constantly flowing into the system at times when natural flows are absent. The extent to which this
contributes to existing weed problems is not known. It is probable that the modified hydraulic regime
would in itself lead to excessive aquatic weed accumulations. However, the presence of high nutrient
loads has almost certainly to accelerated the rate at which this happens and chronic turbidity levels
definitely give the existing problematic (floating and emergent) species a competitive advantage over
other less invasive submergent native plant species that re-establish when weeds are first removed.
Nutrients and sediments are unlikely to be of concern to wetland environments during the rising stage of
the flood hydrographs because most are non-toxic and will pass so rapidly through the wetlands that they
will have little effect. (Increases in turbidity which would be of concern in other parts of the catchment,
are less of a problem on the floodplain because most of these wetlands have adapted to regular inputs of
turbid BFD water). However, ammonia (derived from fertilisers such as urea, ammonium phosphate and
ammonium nitrate), is toxic at high concentrations and levels which could be acutely toxic are
occasionally detected in cane drains and initial receiving streams. For example the toxicity of ammonia
varies enormously depending on pH, temperature, salinity and dissolved oxygen concentrations, so high
levels do not necessarily indicate that damage is being done (we are currently testing local species under
local conditions to clarify this matter). Fertilisers containing ammonia are converted to comparatively
innocuous nitrate quite rapidly in farm soils, so the risks of harm are only high when rainfall (or
irrigation) occurs soon after fertilisers are applied. It is therefore likely (as is the case for so many of the
potential water quality issues in the Burdekin) that impacts would be manifested in the form of brief,
infrequent, localised but severe exposure episodes.
Other toxicants that are likely to be found in cane-farm runoff include mercury (which is used in organic
forms as a fungicide), copper and zinc (which are applied as trace element fertilisers), cadmium (which is
a contaminant of some phosphate fertilisers) and a range of pesticides (see Table X). The potential for
the release of these contaminants depends on many factors (eg. soil composition and condition, time since
application, method of application, etc.) and, as with ammonia, releases are most likely to be episodic and
difficult to predict. Very little effort has been made to determine if there is evidence of accumulations of,
or exposure to, these toxicants in the wetlands that are most likely to be affected by farm runoff. The
majority of studies that have been conducted focus on the marine environment, which benefits from
enormous dilution capacity and therefore provides very limited insights into conditions in the freshwater
and estuarine wetlands. Hunter et al. (1998) reported on the pesticide monitoring that has been
conducted in the Burdekin floodplain canegrowing areas to date. They summarise results obtained from
samples of water, sediment and biota collected from streams, water holes, irrigation supply channels and
drains of the BHWSS on individual sampling trips conducted in 1990, 1992 and 1993, and samples
collected from a single drain in 1995, 1996 and 1997. Modern pesticides degrade quite rapidly and inputs
are known to be episodic and erratic, hence concentrations would be expected to vary enormously over
time and space. Consequently the total number and spatial coverage of sampling events to date are far
too low to be able to draw any definite conclusions. However, low levels of pesticide residues and
especially the herbicides atrazine and 2,4-D, were detected in several samples on several occasions.
Hunter et al. conclude that the results provide evidence of the presence of a number of pesticide residues
in several drains, streams and lagoons in the BRIA, that the environmental significance of the findings
remain uncertain, but that monitoring should be continued as a precautionary measure.
Mortimer (2000) reported on the concentrations of metals in mud crabs collected from six estuaries along
the Queensland coast. The mean concentration of most metals obtained from analysis of ten crabs
collected “near Ayr” were similar to or lower than the levels reported from other estuaries, but mercury
concentrations were almost twice as high as those reported for the Brisbane River and an order of
magnitude higher than other sites. The author does not comment on this result but to our knowledge
canefarming is the only potential source of such contamination in the Ayr region.
Table 2:
Estimated annual pesticide amounts (kg active ingredient/year) used in the
Burdekin Catchments (Hamilton & Haydon 1996)
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Catchment
b)
Pesticide
Haughton
Burdekin
Atrazine
24300
19300
2,4-D
6900
5500
Ametryn
6200
4900
Glyphosate
5300
4200
Diuron
4100
3300
Management implications
Implications for inland waters and coastal floodplains are treated separately here but deficiencies in
current monitoring techniques are evident in both. During most of the year (i.e. when the
unsupplemented parts of the system are at baseflow or during small flow events) monitoring is especially
difficult because the system is fragmented into effectively independent water bodies within which
localised biophysical factors govern water quality characteristics. Moreover, in many cases water quality
within an individual water body can fluctuate massively over very small scales of space and time.
It is only during the falling limb/tail of the hydrograph that the system ever comes close to matching the
paradigms (of catchment driven water quality gradients) upon which traditional ambient monitoring
programs are largely based. At such times the system generally forms a drainage continuum made up of
a series of reaches defined mainly by confluences of tributaries, within which instream physical and
biological processes exert only subtle and gradual water quality effects. Furthermore, most biological
transformations occur in the water column at this stage of the hydrograph, so contaminants are
bioassimilated into planktonic biomass which can be captured in water samples. Conceptually at such
times the river channel functions as a relatively inert conduit that conveys a biologically active stream of
water. Under these conditions grab samples and spot measurements can usually be interpreted
meaningfully, as long as major discontinuities in mixing and aeration patterns (caused by waterfalls,
rapids, weirs, etc.) are taken into consideration. The risks of acute effects from physical and chemical
stressors are also sufficiently low for the ANZECC and ARMCANZ (2000) recommended approaches for
interpretation to be applicable. Later in the hydrograph benthic biological processes gradually begin to
dominate and the rivers turn into a living conduit that exerts major influences on the quality of water
passing through it and as the system fragments it becomes increasingly heterogeneous making
conventional monitoring techniques entirely ineffective.
The flood hydrograph is of particular interest for resource development because it is this water that is
most likely to be retained in large reservoirs. Current monitoring focuses mainly on determining the
conductivity, salt composition and turbidity of this water. This work provides fairly convincing evidence
that salinity is unlikely to be a significant issue in any major Burdekin catchment (especially when
exploitable flows are present) and it is doubtful if further monitoring of this kind would change that
evaluation. It is equally evident, however, that high turbidity is an important issue for water managers.
The chronic high turbidity levels in the BFD are undoubtedly a significant source of impact on all
downstream aquatic ecosystems. The chronic reduction in water clarity throughout all parts of the
distributary system is clearly visible to all observers and is a potential source of dissatisfaction from the
local community.
The chronic turbidity of the BFD was not anticipated – expert opinion prior to the dam’s construction was
that the suspended sediments would settle in the dam. This poor predictive capability is one of many
examples of the deficiencies of dealing with suspended sediments as a gross property rather than
examining its separate components. In a catchment such as the Burdekin where stormflows carry large
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quantities of sediment, the only means of reliably predicting the quality of waters that will be trapped in
an impoundment is to monitor the composition and settling properties of the sediment that will be carried
into it. Note that simple inexpensive tests capable of conveying very useful information of this sort can
be developed for routine monitoring purposes. It is strongly recommended that managers commission a
small research project to develop such methods for incorporation into routine monitoring programs. The
relatively small number of remaining floodplain wetlands that do not receive inputs of irrigation supply
or tailwater are subject to the same water quality influences as the wetlands on the inland floodplains,
where free-range grazing is the main source of anthropogenic impact.
Coastal floodplain
Most of the wetlands on the developed sections of the floodplain at this time are now of a different type
than they once were, due to the constant artificial throughputs of (mostly turbid) irrigation water.
Aquatic weed management is undoubtedly the priority issue on the floodplain, and until the existing
infestations are brought under control other water quality management measures are unlikely to achieve
many environmental gains. The following recommendations therefore apply mainly to systems where
exotic and excessively invasive weeds have been brought under control (and the community-based NHT
projects which we participated in during the course of this project are attempting to accomplish this).
Note that this does not mean that the water courses should not support a large biomass of native and/or
less invasive aquatic plants.
• The current practice of allowing water flows in the distributary channels (and drains) to fall during
rainy weather (when demands are low) places significant stress on any water body that is accustomed
to higher flows, especially if it supports high levels of submergent biomass. Our work has shown that
the weather only has to become cloudy and still, for oxygen concentrations in open water bodies to
drop to stressful levels if flow rates are low. It would therefore be advisable to adopt a policy of,
wherever possible, providing artificial flows in distributaries and drains during rainy weather until
stormwater flows are sufficiently high to provide dispersion and aeration.
•
•
•
•
The adoption of best practice farming methods including the use of grass filter strips, first-flush
runoff traps, etc. would also be highly recommended. Green cane trash blanketing (GCTB), a
practice which is widely used in other cane growing regions, but not in the Burdekin, is often
promoted as a means of reducing soil and nutrient loss. However, the effect of green cane trash
blanketing on BOD releases has not yet been properly evaluated so it is uncertain if this approach
would be a desirable option in the Burdekin.
It would be worth investigating the feasibility of using irrigation water to supplement stormflows
during small-scale events in order to help flush out weeds (especially hyacinth).
Wherever feasible it would be advisable to re-establish connections to estuaries and remove the weed
infested artificial wetlands that have formed at the end of distributary channels.
There are insufficient data to determine if contemporary releases of toxicants from farms should be
treated as a priority issue. A program to monitor residues in sediments at different times of the year,
throughout the entire floodplain (including the estuaries), should be implemented to evaluate risks.
(Toxicant concentrations in the water column vary far too rapidly for water samples to be useful in
this regard). A less rigorous but more affordable alternative to a broad survey would be to conduct a
risk assessment to determine when and where various toxicants should accumulate and then carry out
a more modest sampling program targeting the highest risk sites. (This should take the times of
chemical applications, etc. into account). More extensive monitoring would only be required if
unacceptable accumulations are detected at the targeted sites.
Prevention of irrigation tailwater releases is often considered to be the best-practice option for
canefarming. This is probably a justifiable stance provided that two prerequisites can be satisfied: 1) that
the receiving water body is in good condition and has never previously been subject to prolonged
tailwater flows, and 2) that the quality of water in recycling dams (which will eventually overflow into
the drainage system when it rains) is acceptable. We do not currently have sufficient data indicative of
the quality of water in recycling ponds to assess if the latter condition is fulfilled. The first prerequisite is
important because several of the wetlands on the floodplain are in such poor and vulnerable condition
that they undoubtedly stand a better chance of maintaining useful ecological functions if they receive the
benefits of re-aeration and mixing, etc. provided by artificial flows of tailwater. This is because they have
144
accumulated such a high biomass that flushing of any kind is preferable to stagnancy. Benefits are likely
to accrue even if the tailwater is moderately contaminated with agrochemicals. Water containing very
high levels of sugar and other BOD causing substances, however, are an exception to this rule. There is
an urgent need to take steps to prevent such water from entering any natural waterway. Other wetlands,
depending on the amount and type of biomass they contain, may indeed benefit from less tailwater.
Basically this is a situation where a horses-for-courses approach would achieve the best net benefit.
Currently, it would be necessary to conduct quite detailed investigations at each individual site to
determine the preferred option. This again illustrates the potential benefit of a typology scheme that
would enable recommendations to be optimised for site types rather than individual sites.
Inland waters
When the hydrograph is falling there is green feed available for livestock (and hence their water
consumption requirements are at a minimum) and there is drinking water available all over the
catchment. Therefore livestock are usually widely dispersed and do not congregate around any particular
water body (or part of a water body) for long. It follows that the main source of ambient water quality
problems in the catchment (i.e. congregations of livestock in riparian zones) is largely absent at the time
of the year when the river is probably best placed to cope with it. Accordingly, the probability of
detecting impacts from grazing at this stage of the hydrograph is low.
Inputs of oxygen demanding organic matter (especially manure), the induced effects of turbidity and
nutrients on natural instream oxygen demand and the consequent development of hypoxia, are the
principal sources water quality problems associated with congregations of livestock in riparian zones.
These effects are seldom manifested over large lengths of river but instead as a scattered array of severe
localised catastrophes in numerous isolated water holes palustrine wetlands and river reaches. Few of
these are large enough to be considered critical in their own right but the cumulative loss of productivity
and diversity from all affected water bodies is potentially very significant.
Dissolved oxygen integrates the effects of most of the biophysical and water quality processes occurring
in the water body, and is therefore a potentially useful indicator. It is also the parameter which is most
likely to limit ecological conditions and therefore is also the main habitat quality determinant. Under base
flow/stagnant conditions, concentrations of contaminants, and particularly oxygen, fluctuate enormously
over even small spatial and temporal scales, making detection and quantification by conventional means
very difficult. Data indicative of average concentrations and conditions are at best, impossible to
interpret meaningfully, and at worst are entirely misleading. Spot measurements of dissolved oxygen are
virtually impossible to interpret unless they are supported by a great deal of biophysical data (allowing
sources of dissolved oxygen variability to be identified) but even then interpretation may prove
impossible. For example, many thousands of spot measurements of oxygen have been collected in the
Burdekin catchment over the years, but no more than a handful of them can be interpreted with any
confidence.
Even if oxygen data indicative of spatial and temporal variations are collected, they can be extremely
difficult to summarise and analyse. We are currently attempting to ameliorate this problem by
developing an indexing system that analyses fluctuations and predicts their ecological significance. It
would not normally be feasible to monitor oxygen fluctuations routinely with the required resolution.
(This requires measurements to be taken at different places in the water column at least several times
between dusk and dawn). We are currently developing models to predict spatial and temporal
distribution patterns in different types of water body, enabling meaningful interpretation with many fewer
measurements and allowing identification of the conditions under which simple spot measurements are
adequate.
Vertical oxic stratification in the water column is a very common and widespread phenomenon during the
hot summer months and this increases in severity when flow rates fall. Any alterations which create
deeper, more turbid waters or increase the oxygen demand of bottom sediments (which is often the
ultimate result of eutrophication) can exacerbate the extent and severity of hypoxia in the bottom waters.
This is clearly a major impact of weir and dam construction and is one of the major ways that
anthropogenic sediment and nutrient inputs impact on local waterways. The formation of a severely
145
hypoxic hypolimnion effectively eliminates aerobic organisms from all of the benthic habitats covered by
that layer of water. In some cases this can lead to the loss of up to 95% of all benthic habitats. It is
noteworthy that water measurements taken at the surface provide no indication of this loss of productive
habitat. Moreover, the surface layer often contains good quality water so macroinvertebrate samples
collected from shallow sections of the water body also yield indications that conditions are good, and the
fact that their total biomass has been decimated is seldom detected.
Nutrient concentrations in the water column are impossible to interpret under low flow conditions unless
biological processes in the water body are understood. It is noteworthy for example that hyper-eutrophic
water bodies can display either extremely high or extremely low nutrient concentrations (and anything inbetween) depending on the balance between benthic and planktonic productivity. (If the system is
dominated by phytoplankton the total nutrient levels reported from water samples will be high because
the phytoplankton themselves are included in the sample, but if benthic plants are dominant they can
assimilate nutrients from the water column very efficiently, and since they are not included in samples,
low nutrient concentrations are reported).
The quantities of contaminants required to cause major problems in a stagnant water hole are negligible
compared to the quantities that flow through the watercourse at other times. Hence, models and
monitoring programs directed at estimating annual or seasonal contaminant loads and fluxes are not
sensitive indicators of potential problems.
Vulnerabilities vary greatly over time and between sites. For example, our research in the Dotswood area
(Burrows and Butler, 2001) showed that Fanning River was inherently much more vulnerable to water
quality problems than Star River and Keelbottom Creek. Early in the research project we failed to
recognise this and pooled all control site data together to use as a reference source, thereby combining
together data from sites with naturally good and naturally poor water quality. These were compared with
data collected from cattle-impacted sites which by coincidence were mostly sites with lower vulnerability
(and therefore with higher natural water quality expectations). Statistical analysis indicated that cattle
were having little impact even though the evidence from other observations was to the contrary. This led
to the development of a prototype classification scheme designed to rank the relative vulnerability of each
site every time a sample was taken, allowing us to ensure that we only compared data representative of
sites with similar natural water quality expectations. By employing this approach, statistical analyses
confirmed that livestock were indeed having significant adverse effects on water quality. This
demonstrates how typological analysis can be used as a tool for selecting appropriate reference sites and
conditions.
Biological indicators are of limited use for detecting the effects of acute episodic water quality
disturbances that occur during the rising stage of the flood hydrograph because once flows increase above
critical levels, mobile species will move away while most sedentary organisms will be flushed away,
along with any evidence of water quality damage that may have occurred leading up to the flow peak.
For example, we have found (Butler and Burrows 2001) that after flood events in the tributaries of the
northern Burdekin, macroinvertebrate populations are so sparse that standard rapid sampling techniques
are incapable of confirming if all of the taxa that should be there are present. The fact that many get
washed away should not be taken to indicate that it doesn’t matter if organisms are killed (for example by
asphyxiation) during events – post-event recoveries are heavily reliant of the survival of the few
individuals that manage to remain in place during these events.
Detection of the adverse effects of concern could only be accomplished by employing relatively
expensive remote data-logging or automated sampling equipment. However, the effectiveness of this
approach will depend heavily on site selection and even the precise location within the water column that
is chosen to take readings. The conventional strategy of deploying data loggers at stream gauging sites is
likely to be particularly ineffective unless it can be demonstrated that the gauging point is representative
of the most valuable parts of the ecosystem both in terms of its re-aeration capacity and its in-stream
biological community. In our experience this is seldom the case. It is also worth noting that probes
deployed at a fixed distance above the bottom will often end up immersed in severely hypoxic bottom
waters when waters rise, thus giving the misleading impression that conditions are worse than they really
146
are. Floating sensors that remain suspended in the near-surface layer at all times are preferred for this
work.
The above factors call for the use of monitoring techniques that are logistically difficult to implement and
far too costly to justify unless there is compelling evidence that the potential risks are high. It will be
important to ensure that limited resources are devoted to studying water bodies that are important enough
and vulnerable enough to justify the effort. The logical means of achieving this is to classify regional
water bodies into types that support similar water quality processes and biological functions, and which
have the same natural tendencies to develop particular water quality problems.
This classification method essentially begins where most ecological and geomorphological classifications
end – existing classification systems identify similarities in the physical and biological structure of water
bodies while ours attempts to identify similarities and differences in the way that these attributes
influence factors such as mixing, water detention time, input sources, re-aeration and contaminant
trapping capacity, etc. which ultimately determine how contaminant inputs affect water quality. In order
to avoid confusion between these different classification methods we have elected to refer to ours as a
typology. Once an appropriate typology has been developed it will be possible to map the occurrence of
water body types over the entire catchment. By linking this to existing and emerging GIS data it will be
possible to determine which are the most valuable, vulnerable and/or threatened water bodies (from both
water quality and ecological perspectives).
The typology is particularly necessary as a basis for determining vulnerability, because vulnerable sites
naturally tend to suffer from water quality problems (that could easily be misinterpreted as being
attributable to anthropogenic impacts). However, it also has many other potentially valuable
management applications. For example, controlling livestock and feral animal access to the riparian zone
(and the water) is the only known means of improving the existing water quality problems in most inland
waters (see Figures x and y). The most effective way to maximise ecological gains from improvements
such as riparian fencing, would be to employ the typology to identify which sites most need to be
protected based on a knowledge of their relative importance and vulnerability.
The potential for brief episodic water quality deterioration during the early rising stages of flow events is
usually not factored into environmental release strategies. However, in cases where local rainfall is
delivering runoff into the river reaches downstream of a partially filled impoundment, there is potential to
employ controlled releases of environmental flows to disperse and aerate locally generated stormwaters
until river water begins to flow over the spillway. (This measure would be difficult to accomplish and
may only be justified if more detailed investigations confirmed both the need for and feasibility of such
action).
147
Figure 4:
Comparison of water quality in fenced and unfenced water holes in the
Dalrymple Shire
Relative deviation from the reference
20
15
TN
TDN
10
Nitrate
Ammonia
5
TP
TDP
0
N=
FRP
7
7
7
7
7
7
Fenced
Figure 5:
7
11 11
11 11 11
11 11
Unfenced
Causes of incidences of severe water quality degradation in water holes in the
Dotswood Station area (Northern Burdekin). Total number of incidents = 36
17%
11%
Instream Livestock
Unexplained
53%
Local Runoff
Low Water Level
19%
148
APPENDIX J
BIBLIOGRAPHY OF WETLAND CLASSIFICATION
149
J1.
ANNOTATED BIBLIOGRAPHY OF KEY WETLAND CLASSIFICATION
REFERENCES
Ainslie, W. B. (1994). Rapid wetland functional assessment: its role and utility in the regulatory arena.
Water, Air, & Soil Pollution 77(3-4): 433-444.
Four basic steps in the regulation of dredge and fill activity are outlined in the US Guidelines of the Clean
Water Act as: 1) evaluation of practicable alternatives; 2) evaluation of relation of discharge to other
environmental standards; 3) assessment of significant degradation to waters of the US; and 4) assessment
of appropriate steps to minimize impacts. Steps 1,3, and 4 require wetland functional assessment The
developing functional assessment procedure, based upon functional indicators which can be recognized in
the field and can form the basis for functional indices, shows potential for being rapid and inexpensive,
scientifically-based and replicable.
Arthington, A. H., and Hegerl, E. J. (1988). The distribution, conservation status and management
problems of Queensland's althalassic and tidal wetlands. In “The Conservation of Australian Wetlands”
(A. J. M. P. S. Lake, Ed.), pp. 59 - 109. Surrey Beatty &Sons Pty Ltd, New South Wales, Australia.
The major definitions of wetlands are discussed. The Ramsar Convention definition is provided but cast
aside for what is considered a more appropriate definition, provided by the Environmental Council of
New Zealand. It is considered that the proliferation of local, regional and state inventories using a variety
of classifications, definitions and colloquial terms is the biggest problem in Australian wetland
classification. The authors provide a brief description of wetland types within Stanton and Morgans
(1977) 12 terrestrial biogeographic regions, allowing a systematic breakdown of wetlands associated with
vegetation and landform patterns. Appendix I lists each of the major wetland aggregations of the
Queensland mainland and gives a brief description of the prominent features and major vegetation
formations. In addition a distribution map of the areas occupied by mangrove species communities is
provided. Threats to wetlands, conservation status of Queensland wetlands and their distribution,
selection of wetlands for conservation, and conservation priorities are discussed.
Augusteijn, M. F. and Warrender, C. E. (1998). Wetland classification using optical and radar data and
neural network classification. International Journal of Remote Sensing 19(8): 1545-1560.
NASA's Airborne Terrestrial Applications Sensor (ATLAS) multi-spectral data and Airborne Imaging
Radar Synthetic Radar (AIRSAR) data were used to investigate the ability of a neural network based
classification technique to delineate upland and forested wetland areas and to distinguish between
different levels of wetness in a forested wetland. The neural network technique separated upland from
wetland spectral signatures and discriminated two out of four different water regimes identified by the
NWI within the wetland area. Both ATLAS and AIRSAR data sources, when used in isolation, could
separate wetland from upland about equally well, but better performance was observed when these data
sources were combined.
Babb, J. S., Cole, C. A., Brooks, R. P. and Rose, A. W. Hydrogeomorphology, watershed geology, and
water quality of wetlands in central Pennsylvania. Journal of the Pennsylvania Academy of Science
71(1): 21-28.
Hydrogeomorphology refers to landscape position and source of water and is a characterising feature of
wetland function. It is hypothesised that wetlands of differing hydrogeomorphology will provide waters
of different quality; this project tested this hypothesis. A subset of reference wetlands were characterized
by their water chemistry (pH, conductivity, alkalinity, ferrous iron, nitrate, ammonia, and phosphate). It
was found that:
• Differences in pH and alkalinity were due more to local bedrock geology than hydrogeomorphic
class,
150
•
•
•
Nitrate concentrations differed by HGM class and nitrate and ammonia were strongly negatively
correlated with the amount of organic matter present,
Phosphorus and ferrous iron were correlated in a complex interaction with soil water, soil organic
matter, and the levels of oxygen present in the soil, and that
Local levels of disturbance did not seem to be related to water quality in the wetlands.
These data show the importance of understanding local geology when planning for a wetland creation
project and indicate the significance of correct landscape placement for wetland creation projects.
Barson, M. M., and Williams, J. E. (1991). “Wetland Inventory - towards a unified approach. (Draft
Discussion Paper for Technical Workshop hosted by the Rural Resources and Australian National Parks
and Wildlife Service, August 5 - 6 1991, Canberra).” Australian National Parks and Wildlife Service,
Canberra.
There is a need for a unified approach in wetland classification and inventory; different regions within
Australia may still utilise classification systems that are appropriate for their region and needs, as long as
a common core data set is adopted. This will ensure that regional variability between similar wetlands
will be minimised. Included are suggestions for a potential core data set, brief reviews of the
international and national literature on wetland classification, and an extensive list of scientific papers
concerned with different classification techniques. The authors suggest that the classification system
developed by Cowardin et al. (1979) is too detailed for use in Australia.
Bernert, J. A., Daggett, S. G., Bierly, K. F., Eilers, J. M., Eilers, B. J. and Blok, E. (1999). Recent
wetlands trends (1981/82-1994) in the Willamette Valley, Oregon, USA. Wetlands 19(3): 545-559.
A two-stage, stratified, systematic sample design was implemented in the Willamette Valley, Oregon,
USA, to quantify wetland and land-use changes from the 1980s to the 1990s. The Stage I sample (n =
711) was drawn from public land survey sections and was stratified by land use and runoff potential. The
Stage II sample (n = 114) re-sampled the Stage I sample stratified by the amount of hydric soils identified
in the Stage I sample. Wetland and upland classes were delineated on large-scale aerial photographs,
digitized into ARC/Info coverages, and compared to quantify land-cover changes. Total loss of wetlands
to uplands during the study period was about 3,800 ha, representing a 2.1 percent wetland loss from the
1980s. The net loss after adjusting for wetland gains was about 2,750 ha. During the study period, 70
percent of the wetland loss was associated with agriculture, six percent was lost to urbanization, and 24
percent was lost to other changes. The loss of wetlands to agriculture and the conversion of wetland types
was consistent with a pronounced climatic component related to below-normal precipitation from 1985 to
1994, although continued installation of tile drains and expansion of irrigated agriculture also may have
contributed to the changes. The loss of wetlands to agriculture raised questions regarding the
effectiveness of current agricultural wetland policy, which appears ill-prepared to protect small wetlands
or to deal with loss of wetlands from intensified use of existing farmland. This study identified a larger
number and area of wetlands compared with national wetland surveys because of the larger scale data
used in this study, the nature of the strata used in the statistical design, and the inclusion of palustrine
farmed wetlands in the land-use classification.
Blackman, J. G., Span, A. V., and Whiteley, L. A. (1992). “Provisional Handbook for the
Classification and Field Assessment of Queensland Wetlands and Deep Water Habitats.” Wetland
Inventory Team, Conservation Strategy Branch, Department of Environment and Heritage, Northern
Regional Centre, Pallarenda.
A provisional system for the classification of wetlands and deep water habitats based on the system
devised in the USA by Cowardin et al.(1979) modified for Queensland conditions. Five major wetland
types are described: marine, estuarine, riverine, lacustrine, and palustrine. The classification system
described is hierarchical, progressing from systems and subsystems, class, subclass, and dominance
types. Modifiers may be applied at the class and subclass level. A Queensland Referencing Number
151
(QRN#) is suggested as a method of identifying individual wetlands by their Brooks Number from the
Australian 1:250 000 topographic map series, sequential number for the wetland assemblage/aggregation
and individual number within an assemblage. Basic concepts and procedures for recording primary field
data for the systematic assessment of wetland habitats throughout Queensland are included.
Boyd, R., Dalrymple, R. and Zaitlin, B. A. (1992). Classification of clastic coastal depositional
environments. Sedimentary Geology 80(3-4): 139-150.
A new classification for clastic coastal environments is proposed. It includes the full range of major
depositional settings including deltas, strand plains, tidal flats, estuaries and lagoons. A ternary process
classification used two axes: 1) the relative power of wave versus tidal processes, and 2) relative fluvial
power (increasing upward), and was used to illustrate the genetic process-response relationships between
major coastal environments. The evolutionary classification combines the concept of two sediment
sources (river and marine) with a relative sea-level parameter to classify embayed as well as linear and
elongate/lobate shorelines.
Brinson, M. M., and Rheinhardt, R. (1996). The role of reference wetlands in functional assessment
and mitigation. Ecological Applications 6, 69 - 76.
This investigation into the use of reference wetland sites on functional assessment and mitigation
includes potential misuses as well as highlighting the advantages. Methods on how to determine
reference sites is provided and the process of functional assessment is described by flow diagrams and a
case study.
Bucher, D. and Saenger, P. (1994). A classification of tropical and subtropical Australian estuaries.
Aquatic Conservation 4(1): 1-19.
571 tropical and subtropical Australian estuaries north from Carnarvon on the west coast around to Coffs
Harbour on the east, were measured for relative areas of mangroves, salt marsh/clay pan, seagrass,
intertidal flats and open water. There were strong relationships between mangrove:salt marsh area and
average annual rainfall, although there are no consistent trends in the absolute area of each wetland type.
Classification of estuaries into three distinct groups is based on rainfall and the dominant wetland type.
Three categories can also be distinguished on wetland proportions. Greater tide range coincides with a
greater area per estuary of all wetland types.
Cole, C. A. and Brooks, R. P. (2000). Patterns of wetland hydrology in the ridge and valley province,
Pennsylvania, USA. Wetlands 20(3): 438-447.
The objective in this project was to examine the hydrology of several wetlands over an extended period
of time to achieve adequate information on moisture regimes and year-round hydrodynamics. Wetlands
and analyses were organised around hydrogeomorphic (HGM) principles.
Results showed that:
•
Ground-water-dominated wetlands (riparian depressions and slopes) were the wettest sites,
•
Surface-water systems (headwater and main-stream floodplain wetlands) were drier,
•
There was little difference between slopes and the floodplain wetlands in the amount of time
water was within the root zone,
•
Riparian depressions were wetter longer,
•
The average duration of water within the root zone was almost a year for riparian depressions and
much less for all other wetland types,
•
Disturbance contributed significantly to hydrologic behaviour, more so than HGM classification.
It is concluded that knowledge of HGM subclass might serve as a useful surrogate for actual knowledge
of site-specific hydrology. Though the level of uncertainty increases with surface-water systems, HGM
152
subclass allows a large degree of predictability. The applicability of this technique to different latitudes
and longitudes has yet to be tested.
Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T. (1979). Classification of Wetlands and
Deepwater Habitats of the United States. U. S. Department of the Interior, Fish and Wildlife Service,
Washington, D.C.http:// www.npwrc.usgs.gov/resource/1998/classwet/classwet.htm.
A report on what has been considered one of the most versatile classification schemes. Many other
classification systems around the world have been developed using this scheme as their basis. Limits of
wetlands/deepwater habitats are briefly discussed, with the emphasis of the report dedicated to the
description and application of the classification system.
Dale, P. E. R., Chandica, A. L. and Evans, M. (1996). Using image subtraction and classification to
evaluate change in sub-tropical intertidal wetlands. International Journal of Remote Sensing 17(4): 703719.
Remote sensing was used to monitor the impacts on a sub-tropical intertidal wetland before and after it
was physically modified to manage a mosquito breeding problem. Analysis was by subtracting and
classifying digitized images taken before and after modification. Types of change, and the spatial extent
of change, could be determined, including increased inundation as a result of increased tidal flushing,
indicated by reduced spectral values, and increased. mangrove size and spatial extent.
Edgar, G. J., Barret, N. S., Graddon, D. J., and Last, P. R. (2000). The conservation significance of
estuaries: a classification of Tasmanian estuaries using ecological, physical and demographic attributes as
a case study. Biological Conservation 92, 383 - 397.
Although referring to estuaries the concepts presented in this report can be extrapolated to wetlands.
There is considered a need to classify estuaries into assemblages so that it may be determined which of
those assemblages, and which individual estuaries, require protection. In agreement with Barson &
William (1991), this concept considers that different regions within Australia may still utilise systems of
classification specific to that regions needs, provided a core data set be used.
Categorisation of estuaries is based on geomorphological and hydrological attributes, followed by
physical classifications (validated by comparison of biological attributes, such as invertebrate and fish
data sets). Estuaries within each group are then assessed for their level of anthropogenic disturbance.
Methods, including statistical requirements are thoroughly covered.
Everard, M. (1997). Development of a British wetland strategy. Aquatic Conservation: Marine and
Freshwater Ecosystems 7(3): 223-238.
Declining British wetlands have a long history of over-exploitation. The development and acceptance of
a single consistent definition and classification scheme, agreeable to all organizations with interests in
wetlands, will allow for better coordination of wetland policy which is mandatory to avert a continuing
loss of the nation's wetland resource. It will also contribute to the development of a clear national wetland
strategy, in fulfilment of the Ramsar Convention. The establishment of a National Wetland Forum is
essential to build a consensus on key issues, and assist in the development of a national strategy for
wetlands.
153
Fairweather, P. G. (1999). Determining the 'health' of estuaries: priorities for ecological research.
Australian Journal of Ecology 24, 441 - 451.
Another report focusing on estuaries but providing concepts potentially applicable to other wetlands.
'Ecosystem health' and its application to environmental management and assessment is explored, and a
list of potential assessment criteria suggested (including the status of indicator species). Other methods
of assessing ecosystem health, such as the use of 'snap-shot' measurements and eco-assays, are discussed.
Ferren Jr, W. R., Fielder, P. L., and Leidy, R. A. (1996). Wetlands of the Central and Southern
California Coast and Coastal Watersheds: a methodology for their classification and description. U. S.
Environmental Protection Agency.http://lily.mip.berkley.edu/wetlands/titlepag.html.
Rather than simply provide an inventory of the types of wetlands, this report focuses on the proposed
methodology for classifying California wetlands. The report provides a modified Cowardin et al. (1979)
methodology that incorporates keys, water regimes, water chemistry, hydrogeomorphic units (categories,
series, and units), and substrate characteristics, and then places them into a numerical, hierarchical
classification system.
Background information on the study is bolstered by reference to
hydrogeomorphic units, ecosystem functions and socio-economic values.
Following an overview of earlier efforts in wetland classification in California, focus turns towards how
to use the classification methodology, rationale for classification and the presentation of tables of
hierarchical information to be used during the classification of a wetland. Keys to system (estuarine,
marine, palustrine, riverine, and lacustrine), subclass and class are provided.
Ferreira, J. G. (2000). Development of an estuarine quality index based on key physical and
biogeochemical features. Ocean and Coastal Management 43(1): 99-122.
EQUATION - Estuarine QUAlity and condiTION, is an index presented for integrated evaluation of
estuarine quality. It is based on an aggregation of four different components:
• vulnerability,
• water quality,
• sediment quality, and
• trophodynamics.
These four components are combined into a 1 (worse) to 5 (better) overall grade.
Testing of the index was carried out on five different estuaries in the US and Europe, with ecosystems
chosen to study a range of physiography, tidal regimes, organic loading and contamination by persistent
pollutants. Tests were also carried out on 'concept' estuaries and using different scenarios for two of the
chosen estuaries. A range of scores resulted that agreed with other authors, and the methodology is
useful as a decision support system, synthesising the basic descriptors of estuarine quality: physical
aspects, water quality, benthos and higher trophic levels (including socio-economic aspects of fisheries).
Field, D. W., Alexander, C. E., and Broutman, M. (1988). Toward developing an inventory of U. S.
coastal wetlands. Marine Fisheries Review 50, 40 - 46.
The development of the national inventory of coastal wetlands of the USA, compiled from individual
state and county inventories, of which a variety of classification techniques were utilised. From this
variety of classification systems, wetlands could only be easily consolidated into 4 wetland categories
according to vegetation associations. A table of classification methods used by each state/county is
provided. Owing to the variety of state/county based classification methods, and the variation in timing
of the individual inventories, the existing data was unsatisfactory for compiling a national database.
Ongoing efforts for a national database are discussed.
154
Finlayson, C. M., Bailey, B. J., Freeland, W. J., and Fleming, M. R. (1988). Wetlands of the Northern
Territory. In “The Conservation of Australian Wetlands” (A. J. M. P. S. Lake, Ed.), pp. 103 - 126. Surrey
Beatty & Sons Pty Ltd, New South Wales, Australia.
The classification and distribution of Northern Territory wetlands, their conservation status and their
maintenance, is discussed. Drawn largely on the information provided by Paijmans et al. (1985), it does
not strictly adhere with that author's classification system. In general the classification in this case
revolves around the biota of the wetlands of the Northern Territory. A generalised map of the
geographical distribution of the Northern Territory wetlands is provided along with a detailed text
descriptions.
Finlayson, C. M., and von Oertzen, I. (). Wetlands of Australia: northern (tropical) Australia. pp. 195 304.
An overview of the general climatic and geological (eg. drainage pattern) of Australia as an introduction
of the ecological attributes of wetlands. The distribution of wetlands in northern (tropical) Australia and
their classification has primarily been drawn from Piajmans et al. (1985). This is a hierarchical
classification system that assigns different wetlands into classes, each of which have been briefly
described in terms of hydrological and vegetation attributes. Topics include the structural and floristic
classification of wetlands, regional classification systems used in Queensland and the Pilbera, Western
Australia, characteristics and ecological variables associated with seagrass meadows, mangrove swamps,
salt-marshes and flats, freshwater swamps, seasonally inundated floodplains and billabongs, and lakes,
the biological aspects of conservation, and recommendations on the directions that conservation should
take.
Finlayson, C. M., Von Oertzen, I., Jacobs, S. W. L. and Brock, M. A. (1993). Wetlands of Australia.
Wetlands of the world: inventory, ecology and management. Vol. I. D. F. Whigham. Handbook of
Vegetation Science, 15/2, Kluwer: 195-304.
Temperate and tropical Australian wetlands are classified on the basis of physiography. For temperate
Australia, information on the flora and fauna is reviewed on the basis of wetland classification within
each of the major Australian Drainage Divisions, whereas for northern Australia, ecological information
is summarized according to the major wetland types. Biological aspects of conservation issues are
discussed with recommendations on the broad directions that conservation of Australian wetlands should
take: the need 1) to examine wetlands from a total catchment and drainage basin perspective; 2) to
consider conservation as one of several uses for a wetland; 3) to unify the status of reserve systems
between the Australian States; 4) to give equal attention to all developments hazardous to wetlands rather
than focussing on those with a high media profile; and 5) in temperate Australia, the need for more
information on arid zone wetlands to assess their conservation status and value, and an immediate
expansion of current research activities and/or control of threats posed by feral animals and alien weeds
Finlayson, C. M. and Van Der Valk, A. G. (1995). Classification and inventory of the world's wetlands.
Vegetatio 118(1-2): 192.
The objective of this paper is to establish a framework for developing an international classification
system for a global inventory of wetlands. An array of wetland classification and inventories from
around the world are recognised, including those from the Ramsar Convention, Europe, the Boreal mires
in Finland and Scandinavia; India; China; South Africa; Australia; Canada; and the US. A sampling
design with global application is proposed.
155
Flint, R. W. and Younk, J. A. (1983). Estuarine benthos: long-term community structure variations,
Corpus Christi Bay, Texas. Estuaries 6(2): 126-141.
Hierarchical classification procedures applied to channel versus shoal habitats identified benthic
community structure patterns that were associated with a major disturbance (dredging) to the area, as well
as minor, more frequent disturbances associated with large ship traffic and shrimp trawling activities,
illustrating the resilience of benthic communities to disturbance. Species number and total density were
also related to changes in salinity patterns.
Gerla, P. J. (1999). Estimating the ground-water contribution in wetlands using modeling and digital
terrain analysis. Wetlands 19(2): 394-402.
The contribution of ground water to a wetland's water budget can be a major component in classifying a
wetland; to overcome expenses and difficulties associated with such a measure, an estimation technique
is described that combines the use of a digital elevation model (DEM) with transient numerical modeling
and assumes that the water table reflects the general pattern of surface topography. By knowing or
assuming hydraulic conductivity and using the model water-table configuration, an estimate for groundwater flow to and from each discretized grid node can be estimated and mapped, showing the simulated
distribution of recharge and discharge within and surrounding the wetland. Using a 30-m, 1:24,000 scale
DEM grid in combination with data from the U.S. Fish and Wildlife National Wetlands Inventory, the
model predicts the most likely areas of ground-water interaction in and near wetlands and lakes. More
quantitative results can be obtained by applying observed water budget and soil/aquifer parameter data.
Gilvear, D. J. and McInnes, R. J. (1994). Wetland hydrological vulnerability and the use of
classification procedures: a Scottish case study. Journal of Environmental Management 42(4): 403-414.
The relative importance of wetland water balance equation variables (climatology, surface-water and
groundwater) are used to categorise wetlands into 12 types, each class of which can also be assigned a
hydrological vulnerability to a number of man's activities and various types of water pollution. This
classification scheme was successfully tested to assess its usefulness as a first stage in wetland
hydrological vulnerability assessment using Scottish Natural Heritage data.
Gren, I. M., Folke, C., Turner, K. and Bateman, I. (1994). Primary and secondary values of wetland
ecosystems. Environmental & Resource Economics 4(1): 55-74.
A classification of wetlands based on primary (intrinsic development and maintenance) and secondary
(resources and services generated) values is suggested in an attempt to adequately valuate the
multifunctionality of a wetland. Case studies that use different valuation methods demonstrate the
different degrees to which the primary and secondary values can be quantified. It is concluded that only
part of the total wetland value can be captured in monetary terms.
Gwin, S. E., Kentula, M. E., and Shaffer, P. W. (1999). Evaluating the effects of wetland regulation
through hydrogeomorphic classification and landscape profiles. Wetlands 19, 477 - 489.
This classification system has been developed so that changes between naturally occurring wetlands and
mitigated wetlands can be examined. The hydrogeomorphic classes for natural wetlands (depression,
riverine, slope, and lacustrine fringe) are defined using the principles developed by Brinson (1993) and
Smith et al. (1995), as are three hydrogeomorphic classes for the mitigated wetlands.
156
Harper-Arabie, R. M. and Kolm, K. E. A stepwise, integrated hydrogeomorphic approach for the
classification of wetlands and assessment of wetland hydrological function in the Southern Rocky
Mountains of Colorado. Geological Society of America, 1998 annual meeting 30(7): 121.
A step-wise, integrated hydrogeomorphic approach (HGM) for the classification of wetlands in the
southern Rocky Mountains was developed providing a framework for determining wetland hydrologic
function. The approach includes the following analyses:
• surface features, including hydrophytes, peat and mineral soils, surface water, topography, beaver
activity;
• subsurface features, including geomorphologic deposits, geology, and hydrogeology; and
• ground-water system features, including water levels, recharge, and discharge.
Critical variables that could be assessed for determining wetland function, and two wetland classes, slope
and riverine, were identified.
Fifteen hydrologic function equations are proposed for the southern Rocky Mountain wetlands:
• two functions pertain to the atmospheric processes (ATMin and ET);
• seven functions pertaining to the surface water processes (SWin-riverine, SWin-slope, SWoutriverine, SWout-slope, SWstore-dynamic, SWstore-long term, and ED); and
• six functions pertaining to the ground-water processes (GWinterception, GWmovement, GWoutriver, GWout-springs and seeps, GWstorage-dynamic, and GWstorage-long term).
24 variables were defined to complete the hydrologic function assessment process, and each of the
variables was assigned a ranking between 0. 0 and 1. 0 with respect to the reference site.
Hauer, F. R. and Smith, R. D. (1998). The hydrogeomorphic approach to functional assessment of
riparian wetlands: Evaluating impacts and mitigation on river floodplains in the U.S.A. Freshwater
Biology 40(3): 517-530.
Hydrogeomorphic (HGM) classification is applied to riparian wetlands of two alluvial rivers. The HGM
approach to functional assessment of wetlands was developed to appease wetland lost or degraded by
anthropogenic impacts. It is based on:
•
classification of wetlands by geomorphic origin and hydrographic regime,
•
assessment models that associate variables as indicators of function, and
•
comparison to reference wetlands that represent the range of conditions that may be expected in
a particular region.
Riverine wetlands are characterized by formative fluvial processes that occur mainly on flood plains (e.g.
US bottomland hardwood forests that typify the low gradient, fine texture substratum of the south-eastern
coastal plain and the alluvial flood plains that typify the high gradient, coarse texture substratum of
western montane rivers).
Assessment models for fourteen alluvial wetland functions are described, each of which is a composite of
two to seven wetland variables that are independently scored in relation to a reference data set developed
for alluvial rivers in the western U.S.A. Scores are summarized by a 'functional capacity index' (FCI),
which is multiplied by the area of the project site to produce a dimensionless 'functional capacity unit'
(FCU). When HGM is properly used, compensatory mitigation is based on the FCUs lost that must be
returned to the riverine landscape under statutory authority.
157
Heath, R. A. and Grange, K. R. (1982). Classification of New Zealand inlets: short-circuiting
environmental assessment of New Zealand coastal inlets. Proc. river and estuary mixing workshop,
Hamilton, NZ, 1981. J. C. Rutherford. 49, NZ Water & Soil Miscellaneous Publication: 91-92.
The potential for environmental problems associated with inlets can be predicted from the factors such as
residence time of the inlet waters; the tidal excursion of coastal waters; the tidal energy dissipation; the
freshwater inflow; width and depth.
Henderson, F. M., Hart, T. F., Jr., Heaton, B. P. and Portolese, J. E. (1999). Mapping coastal
ecosystems over a steep development gradient using C-CAP protocols. International Journal of Remote
Sensing 20(4): 727-744.
Coastwatch Change Analysis Program (C-CAP) was used to map over 5500 sq. km of Long Island, New
York. This paper describes the adjustments made to the protocol to conform to local environmental
conditions, and the methodology employed to map five upland and four wetland land cover types. The
incorporation of ancillary data sets with TM imagery produced an accuracy of over 90%, remedying the
spectral confusion among land cover resulting from TM alone.
Hills, J. M., Murphy, K. J., Pulford, I. D., and Flowers, T. H. (1994). A method for classifying
European riverine wetland ecosystems using functional vegetation groups. Functional Ecology 8, 242 252.
Plant attributes are used to formulate a model of vegetation in terms of plant functional groups. Easily
measurable plant traits that are associated with the C (competitor), S (stress tolerator), and D (disturbance
tolerator) strategy types are determined and used to classify wetland communities into functional
vegetation types. A brief literature review on the classification of vegetation into functional groups is
accompanied by detailed information on methods to determine which functional traits particular species
use.
Holmes, N. T. H. (1999). Recovery of headwater stream flora following the 1989-1992 groundwater
drought. Hydrological Processes 13(3): 341-354.
Over a three year period in spring, summer and autumn from 1993 to 1995, 118 sites on 24 headwater
reaches of groundwater streams were investigated in an effort to identify the effects of typical low flows
and bed drying on aquatic and wetland plants following an exceptionally long period of poor groundwater
recharge between 1989-92. Determination of the rate, and extent, of recovery of aquatic plants following
drought conditions is crucial for conservation bodies, water regulators and water utilities to help them
make judgements on the degree of impacts groundwater abstractions have/could have on river flows, and
separating these from natural causes. The macrophyte survey data have been used to develop a
classification system for headwater streams fed by groundwater and demonstrate clear insight into the
flora expected in groundwater streams based primarily on flow periodicity and channel form. The
behaviour of individual species helps in environmental assessment of proposed new abstractions, and
allows accurate predictions on which ones might decline, be lost or invade, and for predicting the benefits
of alleviation strategies based on target flows. Species relevant to different scenarios are provided; there
are very distinctive communities associated with streams in the upper reaches of groundwater
catchments, and these can be correlated with different flow and physical habitat characteristics. Some
communities are very stable, and changed little during and after the drought; others are highly sensitive,
and changed dramatically after flows returned to 'normal' at the end of the drought, or when impacting
abstractions were curtailed. In most systems flora returned to predicted stable states within two years of
normal recharge, even at locations where there had been no flow for several years.
158
Huffman, R. T. and Tucker, G. E. (1984) Preliminary guide to the onsite identification and delineation
of the wetlands of Alaska. STAR, 22(19), 1984
The classification system in this guide is adapted from that utilized by the National Wetlands Inventory
Project of the US Fish and Wildlife Service, but frequently departs from this system to describe common
and/or distinct wetland communities or associations.
Imhof, J. G., and FitzGibbon, J. E. (1999). A procedure for developing a riparian/valley classification
system for management. In “Proceedings of the Second Internation Conference on Natural Channel
Systems” (P. Young and A. Boyds, Eds.), March 1 - 4, 1999, Niagra Falls, Canada.
The procedure involves the development of a classification system via the identification of key functions
of riparian zones, determining ecological models, establishing functional principles, and determining the
appropriate scale for the data collection and assessment. It is suggested that any classification system
developed through such a procedure should provide an understanding of the role of a particular valley
form and composition in relation to riparian zone and stream channel form and function. A brief review
of the literature of the various functions within riparian zones is accompanied by a procedure for
developing a classification system for riparian zone management.
Kangas, P. C. (1990). An energy theory of landscape for classifying wetlands. Forested wetlands. A. E.
Lugo. Ecosystems of the World, 15, Elsevier: 15-23.
Arguing that environmental energy sources determine ecosystem characteristics, and that therefore the
spatial expression of energy sources becomes the distinguishing feature of land classification schemes,
such a classification is applied to forested wetlands. Different forms of water distribution are matched
with different wetland forms, distinguishing between background water supplies, centres, zones, strings
and islands, and strips.
Keddy, P. A. (2000). Wetland ecology : principles and conservation. New York, Cambridge University
Press.
A synthesis of the existing field of wetland ecology, using a few central themes:
• basic characteristics of wetlands,
• key environmental factors that produce wetland community types, and
• unifying problems such as assembly rules, restoration and conservation.
The volume draws upon a range of wetland habitats and geographic regions, from Africa, Asia, Europe,
Australia and New Zealand, as well as from North and South America.
Klemas, V. V. (2001). Remote sensing of landscape-level coastal environmental indicators.
Environmental Management 27(1): 47-57.
The efficiency, availability and cost effectiveness of remote sensing (RS) and geographic information
systems (GIS) are allowing researchers and managers to take a broader view of ecological patterns and
processes. Landscape-level environmental indicators (e.g. watershed land cover, riparian buffers,
shoreline and wetland changes) can be detected by Landsat Thematic Mapper (TM) and other remote
sensors to provide quantitative estimates of coastal and estuarine habitat conditions and trends. Landsat 7,
and new satellites, carrying sensors with much finer spatial (1-5 m) and spectral (200 narrow bands)
resolutions has made the monitoring of large coastal areas and estuaries, and the detection of coastal and
wetland condition changes more feasible and accurate. This advancing technology provides effective and
efficient alternative management strategies, particularly when the range of techniques for generating,
organizing, storing, and analysing spatial information are combined with mathematical models.
159
Kolm, K. E. and Emerick, J. C. (1997) An approach for the evaluation of wetland structure, hydrology,
and biogeochemistry for wetland ecosystem management in the Southern Rocky Mountains of Colorado.
Geological Society of America, 1997 annual meeting 29(6): 67
A long-term research effort to develop a multidisciplinary, quantitative approach to evaluate wetland
structure, hydrology, and biogeochemistry for wetland ecosystem management in the southern Rocky
Mountains of Colorado. Wetlands structure encompasses topography, vegetation, soils, and geomorphic
and geologic materials. Wetlands hydrology encompasses the climate, surface water, and ground-water
aspects of wetlands ecosystem process and function. Wetlands biogeochemistry is characterized by the
abiotic and biotic chemical processes that occur in and between the biota, water, and geologic and
geomorphic materials.
This approach integrates:
• Conceptualization of wetland-system structure, hydrology, and biogeochemistry;
• Characterization of wetland ecosystem structure;
• Characterization of wetland system hydrology;
• Characterization of chemical partitioning within the wetland ecosystem;
• Characterization of abiotic and biotic metals behaviour within selected reference wetland
ecosystems; and
• Confirmation of hydrologic and chemical behaviour within a wetland ecosystem using
modelling.
Kudray, G. M. and Gale, M. R. (2000). Evaluation of national wetland and inventory maps in a heavily
forested region in the upper great lakes. Wetlands 20(4): 581-587.
The extent of accuracy of National Wetland Inventory (NWI) maps was reviewed using field data from
148 plots in the Hiawatha National Forest ecological classification and inventory program. NWI maps
were over 90% accurate in identifying uplands and jurisdictional wetlands, with least accuracy (90.7%)
achieved in identifying forested wetlands. The most common error was the NWI classification of
wetlands on a poorly drained upland soil that often occurs in complexes with wetland soils in the region.
Forested wetlands with a cover type similar to adjacent uplands were also a source of error on NWI maps.
Though NWI maps proved highly accurate, improvement could be achieved by the mapping of wetlandupland complexes.
Lent, R. M., Waldron, M. C. and Rader, J. C. (1998). Multivariate classification of small order
watersheds in the Quabbin Reservoir Basin, Massachusetts. Journal of the American Water Resources
Association 34(2): 439-450.
GIS data was used to delineate and define landscape attributes (hydrologic, geologic, and geographic
features) of 83 Massachusetts watersheds. The basin was subdivided into three subbasins based on the
evaluation of chemical constituents collected from water samples over a year (principal component
analysis accounted for about 90 percent of the variance in water chemistry data, with the principal
components defined as a biogeochemical variable related to wetland density, an acid-neutralization
variable, and a road-salt variable related to density of primary roads). ANOVAs determined that all
stream water constituents were significantly different among subbasins. Multiple regression techniques
were used to relate stream water chemistry to landscape attributes. Important differences in landscape
attributes were related to wetlands, slope, and soil type.
Lotspeich, F. B. (1980). Watersheds as the basic ecosystem: this conceptual framework provides a basis
for a natural classification system. Water Resources Bulletin 16, 581 - 586.
Factors that form the basis for any classification system of ecosystems are broken down into 'state' and
'transactional' factors. 'State factors' include climate and geology, whereas 'transactional factors' refer to
such things as soil properties and vegetation characteristics. Rather than a discussion on the development
of a classification system, it emphasises the points that must be verified during the development of such a
system, with particular reference to ecosystem processes.
160
Luternaur, J. L., Atkins, R. J., Moody, A. I., Williams, H. F. L., and Gibson, J. W. (1995). Salt
marshes. In “Geomorphology and Sedimentology of Estuaries” (G. M. E. Perillo, Ed.), pp. 307 - 332.
Elsevier, Amsterdam.
Assessment of the geomorphological and sedimentary processes of estuarine marshes is supplemented by
the status of numerical modelling. Coastal marsh morpho-sedimentology, sediment processes,
colonisation of vascular and non-vascular plants, sedimentation rates, and estuarine marsh dynamics are
discussed with the proposition that numerical models can be approached only after specific questions
relating to these processes be answered.
Ming-Ko, W., Rowsell, R. D. and Clark, R. G. (1993). Hydrological classification of Canadian prairie
wetlands and prediction of wetland inundation in response to climatic variability. Occasional Paper Canadian Wildlife Service 79: 22.
Four groups of wetlands are distinguished with temporal variation in wetland inundation creating a
continuum between groups. Ephemeral, intermittent, semipermanent, and permanent wetlands, are
categorised on their annual proportion of inundation, the variable water source dependent on the local
influences. In late summer, flooding of permanent and some intermittent wetlands can be determined by
fluctuations in the regional water table. Fractions of ephemeral and intermittent wetlands inundated
during spring and early summer were predicted with near certainty using snowfall of current and past
winters
Mucha, A. P. and Costa, M. H. (1999). Macrozoobenthic community structure in two Portuguese
estuaries: Relationship with organic enrichment and nutrient gradients. Acta Oecologica 20(4): 363-376.
A study of the structural and functional aspects of the macrozoobenthic communities and their
relationships with the organic enrichment and nutrient gradients in water and sediment. The study was
carried out in two Coastal ecosystems under strong anthropogenic pressure: the Sado estuary and the
Aveiro Lagoon. Samples were collected at four stations between October 1994 and April 1995. Results
indicated that spatial variability, both horizontal and vertical, was higher than the temporal variability,
due to local hydrodynamism, organic load and granulometric structure of the sediment. Comparing the
community structure at the different stations, the existence of disturbance cases not directly related to
organic enrichment gradients could be observed. The interpretation of the functional role of the
macrozoobenthic communities at the water-sediment interface was based on the trophodynamic group
classification. This approach indicated the sensitivity of different groups to organic enrichment,
confirming the role of burrower subsurface-deposit feeders as opportunists associated with organic
enrichment and, on the other hand, the influence of certain trophodynamic groups in the nutrient transfer
processes, particularly the association between tube-builder omnivores and burrower carnivore and
nitrate content in interstitial water.
NRC. (1995). “Wetlands: characteristics and boundaries (National Research Council).” National
Academy Press, Washington.
A unique definition of wetlands (as opposed to RAMSAR) is presented. Chapters cover various aspects
of wetland ecology including:
•
the nature of wetland ecosystems and their response to alteration,
•
analysis of the properties that delineate wetlands from other ecosystems (e.g. hydrology, soils,
vegetation, and a combination of these),
•
the importance of delineation within regional boundaries, and
•
the use of maps, images, and modelling in the assessment of wetlands
161
Paijmans, K., Galloway, R. W., Faith, D. P., Fleming, P. M., Haantjens, H. A., Heyligers, P. C.,
Kalma, J. D., and Loffler, E. (1985). “Aspects of Australian Wetlands (Division of Water and Land
Resources Technical Paper No. 44).” Commonwealth Scientific and Industrial Research Organisation,
Australia.
Australian wetlands have been classified into 6 major categories: lakes, swamps, land subject to
inundation, channels, tidal flats, and coastal water bodies. Information is provided in the form of rainfall,
surface-runoff, and evaporation maps, and maps covering drainage divisions, landform regions, and
distribution of different wetland types. A hierarchical classification system has been applied with
discussion on the delineation between classes within each wetland category, and their vegetation
characteristics.
Pawlowicz, R. and Farmer, D. M. (1998). Diagnosing vertical mixing in a two-layer exchange flow.
Journal of Geophysical Research C: Oceans 103(13): 30,695-30,711.
The use of temperature as a characterising determinant has been overlooked in most classification
schemes, however a proper understanding of the effects of surface heating can explain observed seaward
changes in the slope of temperature-salinity correlations. This paper demonstrates that along-channel
changes in the slope of T-S correlations are virtually independent of vertical mixing but are directly
related to horizontal layer transport, and that c hanges in the layer salinities can be related to various
ratios of horizontal and vertical transports. A diagnostic determination of Lagrangian transport and
mixing can be made from a combination of these results from standard hydrographic observations of
layer temperature and salinity and an estimate of the surface heat input
Perillo, G. M. E. (1995). Definitions and geomorphological classifications of estuaries. In
“Geomorphology and Sedimentology of Estuaries” (G. M. E. Perillo, Ed.), pp. 17 - 47. Elsevier,
Amsterdam.
Proposes a newly devised morphogenic classification system as distinct from the traditional four:
physiographic, tidal range, evolutionary, and morphological.
Prandle, D. (1985). On salinity regimes and the vertical structure of residual flows in narrow tidal
estuaries. Estuarine, Coastal & Shelf Science 20(5): 615-635.
The level of stratification in narrow estuaries is shown to be related to an algorithm containing elements
of 1) velocity structure and 2) the square of the 'flow ratio', providing a simple classification for estuarine
stratification to indicate the sensitivity of any particular estuary to changing conditions.
Pressey, R. L. (1989a). Wetlands of the Lower Clarence Floodplain, northern coastal New South Wales.
Proceedings of the Linnean Society New South Wales 111, 143 - 155.
A vegetation inventory and description of wetlands of the Lower Clarence floodplain. Wetlands
surveyed were ranked according to their conservation value, which was determined via five attributes.
Pressey, R. L., and Harris, J. H. (1988). Wetlands of New South Wales. In “The Conservation of
Australian Wetlands” (A. J. McComb and P. S. Lake, Eds.), pp. 35 - 57. Surrey Beatty & Sons Pty Ltd,
New South Wales, Australia.
Includes consideration of wetland resources, impacts and threats, present conservation and management,
and future conservation and management. Wetlands are described as being rather homogeneous within
their divisions (coastal, inland, tableland and man-made) in terms of their geomorphology, hydrology,
162
and in particular biological attributes. Recognition is made of the need for a comprehensive definition
and assessment of the NSW's wetlands to produce quantifiable information and map resources.
Ramm, A. E. L. (1990). Application of the community degradation index to South African estuaries.
Water Research 24(3): 383-389.
The categorisation of 62 estuarine and lagoonal systems situated on the Natal coast in South Africa was
based on a community degradation index (CDI) of fish assemblages. The systems were first classified
into six major groups based upon eight physical-hydrologic parameters. CDI values were calculated for
each system by comparing a reference faunal list with species lists resulting from biological surveys
conducted between 1981-1986, resulting in values ranging from 0.2 (un-degraded) to 8.2 (severely
degraded).
Rio, J. N. R. and Lozano-Garcia, D. F. (2000). Spatial filtering of radar data (RADARSAT) for
wetlands (brackish marshes) classification. Remote Sensing of Environment 73(2): 143-151.
Mapping of important wetlands for migratory birds is essential for management and conservation; the
object of this report was to determine the most effective spatial filter treatment applied to a two-angle
RADARSAT data set to reduce speckle effects and improve marsh classification performance of
Tamaulipas wetlands. Eighteen spatial filter treatments were applied to the data based on four different
algorithms: Maximum a Posteriori, Lee-Sigma, Local Region, and Median. Classification of the filtered
images resulted in five general classes, which were statistically compared. The Lee-Sigma algorithm
with three iterations and square window sizes of three, five, and seven pixels, respectively, produced the
best overall results with the 8-m pixel RADARSAT data.
Rochford, D. J. (1959). Classification of Australian estuarine systems. Archivio Di Oceanografia e
Limnologia 11, 171 - 177.
Meteorological, morphometric, hydrological and circulatory data are seen as 'standard' requisites for the
estuarine classification system. An idealised estuary is considered to consist of four zones, (marine, tidal,
gradient and freshwater) distinguished by hydrological (pH, salinity, temperature, dissolved 02 and
inorganic nutrients), circulation, sediment and biological properties. This system displays weaknesses in
its limited use to within Australia (due to terminology difficulties especially with respect to salinity and
temperature), and its utilisation of average conditions only. A concluding discussion by an American
based scientist provides a brief review of the paper and describes the pertinence of this system in the
USA.
Rosgen, D. L. (1994). A classification of natural rivers. Catena 22, 169 - 199.
A classification system for rivers in which morphological characteristics are organised into stream types.
Seven major stream types are described in which six additional types are delineated. A stream
hierarchical inventory system utilising the stream classification system is presented, along with specific
examples.
Roy, P. S. (1984). New South Wales estuaries: their origin and evolution. Coastal geomorphology in
Australia. B. G. Thom, Academic Press: 99-121.
A model of evolutionary change is developed from a classification of estuaries and stages in estuary
evolution accounting for both the physical changes and the accompanying biological changes.
163
Semeniuk, C. A. (1987). Wetlands of the Darling System - a geomorphic approach to habitat
classification. Journal of the Royal Society of Western Australia 69, 95 - 112.
Wetlands range through many inter-gradational categories (in particular between riverine, estuarine and
marine systems). There are advantages and disadvantages of attributes such as vegetation, water quality,
and internal and external morphology in the classification of wetlands, all of which are discussed
supported by a list of references. This classification system is primarily targeted towards 'land-based'
wetlands of the Darling System.
Semeniuk, C. A., Semeniuk, V., Cresswell, I. D. and Marchant, N. G. (1990). Wetlands of the Darling
System, southwestern Australia: a descriptive classification using vegetation pattern and form. Journal Royal Society of Western Australia 72(4): 109-121.
A classification is proposed for wetlands of the Darling System based on the scale of wetland vegetation
copmlexes; extent of vegetation cover over the wetland; internal organization of vegetation in plan;
vegetation structure; and details of the floristic/structural components.
Shaffer, P. W., Kentula, M. E. and Gwin, S. E. (1999). Characterization of wetland hydrology using
hydrogeomorphic classification. Wetlands 19(3): 490-504.
Water levels were monitored in 45 wetlands for three years to characterize the hydrology of wetlands in
the vicinity of Portland, Oregon, USA and classified wetlands by hydrogeomorphic (HGM) class to
determine whether hydrologic regimes differed in wetlands in different HGM classes. Hydrologic
regimes were also compared in naturally occurring wetlands (NOWs) and mitigation wetlands (MWs)
and in wetlands with/without a human-made water-retention structure, to determine whether and how
human modifications are changing the hydrology of wetlands. No relationship was found between
hydrologic attributes and land use, soil association, or wetland area, nor significant differences related to
presence of a water-retention structure and to wetland type (NOW or MW). Water levels were higher
with less temporal variability and more extensive inundation (as % wetland area) in MWs and in
wetlands modified to include a retention structure. HGM class was very effective for characterizing
wetland hydrology, with significant differences among HGM classes for water level and for extent and
duration of inundation. Slope wetlands demonstrated the lowest water levels and inundation period;
riverine wetlands experienced intermediate conditions, while depressions had the highest water levels and
greatest extent and duration of inundation. Geomorphic setting and wetland structure are both shown to
be important in defining wetland hydrology and support the use of HGM for wetland classification.
Sharifi, M. (1990). Assessment of surface water quality by an index system in Anzali Basin. The
hydrological basis for water resources management. Proc., symposium, Beijing, 1990. U. Shamir.
Publication, 197, Iahs: 163-171.
A water quality index was developed in order to integrate the composite influence of various physical,
chemical and biological parameters measured by the Bureau of Reclamation of Anzali wetland in 18
different places in the area and some water courses in the basin over 5 years. This index depends upon a
hierarchy of weights of parameters which were selected on the basis of a policy of land use in the area as
well as considering nature and the amount of the polluted material and pattern of Caspian Sea water
penetration into the Anzali wetland. Eventually a five group classification system was developed to
examine the possibility of improving the general management of water. The water quality index system,
which was used for the first time in Iran, also served as a satisfactory means of unambiguous
communication between experts and the public.
164
Tiner, R. W. (1999). Wetland indicators : a guide to wetland indentification, delineation, classification,
and mapping. Boca Raton, Fla., Lewis Publishers.
US wetland indicators are presented that illustrate the current state of scientific wetland recognition and
mapping. Plants, soils, and other signs of wetland hydrology in the soil, or on the surface of wetlands are
the focus towards understanding the current concept of wetland and methods for identifying, describing,
classifying, and delineating wetlands.
Tomczak, M. (1996). Definition of Estuaries; empirical estuary classification.
http://www.es.flinders.edu.au/~matton/ShelfCoast/notes/chapter11.htm.
A discussion on different estuarine classification schemes and their theoretical foundations. The first
classification scheme is solely based on topographical features (proposed by Pritchard, 1952). Types of
estuaries include 'positive' estuaries (salt-wedge, highly stratified, slightly stratified, and vertically
mixed), 'inverse' or 'negative' estuaries and salt plug estuaries. Each type of estuary is discussed in
relation to the rate and location of fresh water volume and salt water volume mixing, and classified
within a 2-dimensional framework (vertical and lateral, from estuary mouth to upstream). Additional of a
third dimension (lateral, from bank to bank), imposes classification modification due to the Coriolis force
which concentrates flow on the left of the estuary (looking in the direction of flow). A final type of
estuary, the intermediate estuary, includes hydrodynamic characteristics such as the temporary
disappearance of the thermohaline (at the interchange from estuary to ocean embayment).
Töyrä, J., Pietroniro, A. and Martz, L. W. (2001). Multisensor hydrologic assessment of a freshwater
wetland. Remote Sensing of Environment 75(2): 162-173.
Synthetic aperture radar (SAR) and visible/infrared (VIR) satellite imagery for mapping the extent of
standing water in the Peace-Athabasca Delta are assessed. SAR images contain data about the geometric
and electrical characteristics of the objects, while VIR images contain information about the reflectivity
of objects. Radar pulses can also penetrate vegetation to some degree depending on the wavelength and
vegetation thickness.
Generally technologies that complement eachother can be used in combination to enhance any output;
thus it is hypothesized that flood mapping will be more efficient when Radarsat SAR and SPOT
multispectral imagery are applied together. This hypothesis was tested with results indicating that flood
mapping in both spring and summer conditions has significantly higher accuracy when Radarsat and
SPOT imagery are used in combination, rather than separately. Radarsat imagery is more accurate when
acquired at low incidence angles.
Unknown. (1999a). Classification, vegetation analysis, and inventory of coastal wetlands. U. S.
Environment Protection Agency & Environment Canada.
http://www.epa.gov/grtlakes/solec...ers/coastabia/classification.html.
A discussion on the classification of the Great Lakes coastal wetlands based on 'site-types' (riverine,
lacustrine etc.), geomorphic type, and vegetation. This classification system has been developed from the
Michigan natural Features Inventory. A table of the classification scheme, plates of most of the wetland
types, and a description of nine identified groups based on both abiotic and vegetation analysis, are
provided.
165
Van Soest, F., Williams, A., Parkinson, R. and Kent, M. (2001). Methodology development for
wetland restoration planning based on topography and soil hydrology. IAHS-AISH Publication (268):
271-278.
This paper presents a novel approach for the identification of suitable wetland restoration sites at a
regional scale, using readily available soil and topographic data. The study focuses on Rhôs pasture, a
floristically diverse wet grassland community, which has been under threat from agricultural
improvement in the past and more recently from neglect. The HOST classification was used to describe
soil hydrological properties and topography was expressed using the In(a/ta&nbeta;) topographic index.
Distributions of soil hydrology and the topographic index under Rhôs pasture sites were compared to the
distributions of the total area and significance of the differences was assessed with chi-squared tests.
Results were integrated into a decision support system, which predicted 72% of the existing sites. The
method provides a useful tool for restoration planning at the decision-making scale.
Wells, E. D. and Zoltai, S. (1985). The Canadian system of wetland classification and its application to
circumboreal wetlands. Aquilo Series Botanica 21: 45-52.
The Canadian Wetland Classification System consists of 4 levels: 1) wetland classes (eg bog, fen, marsh,
swamp); wetland forms (eg domed bogs, slope fen, stream marsh); 3) vegetation (plant communities,
associations); and 4) specialized needs of particular disciplines. Its development and structure, its
application to the classification of peatlands in E Newfoundland and its relationship with wetland
classification systems in northern Europe are discussed.
Wells, J. T. (1995). Tide dominated estuaries and tidal rivers. In “Geomorphology and sedimentology of
Estuaries” (G. M. E. Perillo, Ed.). Elsevier, Amsterdam.
'How are tidal-dominated estuaries classified?' is addressed with a summary of the important processes
and attributes of 'tidal-dominated' estuaries. Physical processes such as the effects of tide on sediment
dynamics, estuarine morphology, and sedimentology are reviewed accompanied by a list of examples,
with description/discussion, of this type of estuary from around the world.
Whigham, D. F., Dykyjova, D. and Hejny, S. (1993). Wetlands of the world: inventory, ecology and
management. Volume I. Handbook of Vegetation Science, 15/2, Kluwer: 768.
Distributions of wetlands, description of wetland types, wetland classification systems, ecological
characteristics, wetland use and conservation and recommendations on what should be done for each type
of wetland are covered. The intention of this work is to provide an overview of the world's wetlands and
an introduction to the literature on their distribution, biota, management and ecology.
Whigham, D. F., Rains, M. C., Mason, J. A., Kahn, H., Ruhlman, M. B., Nutter, W. L., Lee, L. C.,
Brinson, M. M. and Rheinhardt, R. D. (1999). Hydrogeomorphic (HGM) assessment - a test of user
consistency. Wetlands 19(3): 560-569.
User consistency in the application of hydrogeomorphic (HGM) functional assessment models is assessed
by comparing two teams of individuals trained in the HGM methodology. Results from the analysis of
44 riverine wetlands on the Coastal Plain of Delaware, Maryland, and Virginia, USA over a three-week
period demonstrated a high degree of agreement between the two assessment teams for both Variable
Subindices and Functional Capacity Index Scores, indicating that the assessment models were robust and
results were repeatable. It was found that it is important to use only those variables whose measurements
are repeatable, otherwise HGM functional capacity scores are detrimentally affected, especially functions
that are modelled by only a few variables.
166
Whittecar, G. R. and Daniels, W. L. (2000). Use of hydrogeomorphic concepts to design created
wetlands in southeastern Virginia. Geomorphology 31(1-4): 355-371.
To overcome detrimental affects of constructed wetlands, sufficient study of the geomorphic and
hydrologic processes active at the mitigation site, and greater understanding of the geomorphic processes
that created natural wetlands in that area, need to be incorporated into the construction planning. HGM
classification for the assessment of wetland functions requires examination of surface and subsurface
processes, but refinement of concepts to include geomorphic evolution would vastly improve wetland
construction design and management.
Winter, T. C. and Llamas, M. R. (1993). Hydrogeology of wetlands. Journal of Hydrology 141(1-4):
269.
The Hydrogeology of Wetlands Symposium, 28th International Geological Congress in Washington, DC,
in July 1989A was held to assemble papers describing hydrogeologic studies of wetlands representative
of different geographic regions, wetland types, and study approaches. Study locations ranged
geographically from wetlands in the Arctic to the Subtropics and covered wetland types including coastal,
riverine, depressional glacial terrane, and dunal depressions. Different study approaches included
regional syntheses, analyses of groundwater flow systems, wetland-river interaction, and geomorphologyvegetation interaction.
Zhang, S. Q., Zhang, S. K. and Zhang, J. Y. (2000). A study on wetland classification model of remote
sensing in the Sangjiang Plain. Chinese Geographical Science 10(1): 68-73.
The identification and discriminate isolation of around 20 categories of wetlands in the Sanjiang Plain of
China is reliant upon remote sensing. However, TM Landsat imagery is inefficient where there is
wetland similarity or illegibility of the wetland spectrum, and to compensate spectrum enhancement,
pseudo colour composites, and algebra enhancement of TM images have been applied. Distracting noise
e.g. from atmosphere transportation, make even these improvements deficient. It is considered that
geographical analysis, based on physical features of wetlands reflecting the diversification of spectrum
information of wetlands, which include the spatial-temporal characteristics of the wetlands distribution,
the landscape differences of wetlands from season to season, the growing environment and the vertical
structure of wetlands vegetation and so on, must be included. Artificial alterations to spatial structure of
wetlands are also suggested as important symbols of wetlands identification from remote sensing images.
Zogg, G. P. and Barnes, B. V. (1995). Ecological classification and analysis of wetland ecosystems,
northern Lower Michigan, USA. Canadian Journal of Forest Research 25(11): 1865-1875.
Ecosystem types are identified on the basis of simultaneous integration of physiography, climate,
hydrology, soil, and vegetation. Wetland types were better discriminated using ground-flora vegetation,
hydrology, or soil characteristics, rather than by canonical variates analysis of overstory composition
data.
167
J2.
LIST OF WETLAND CLASSIFICATION REFERENCES
Ainslie, W. B. (1994). Rapid wetland functional assessment: its role and utility in the regulatory arena.
Water, Air, & Soil Pollution 77(3-4): 433-444.
Anonymous. National wetland classification system. U. S. Geological Survey Professional Paper: 318.
Arthington, A. H., and Hegerl, E. J. (1988). The distribution, conservation status and management
problems of Queensland's althalassic and tidal wetlands. In The Conservation of Australian Wetlands (A.
J. M. P. S. Lake, Ed.), pp. 59 - 109. Surrey Beatty &Sons Pty Ltd, New South Wales, Australia.
Augusteijn, M. F. and Warrender, C. E. (1998). Wetland classification using optical and radar data and
neural network classification. International Journal of Remote Sensing 19(8): 1545-1560.
Augustinus, P. G. E. F. (1995). Geomorphology and sedimentology of estuaries. In Geomorphology and
Sedimentology of Estuarie” (G. M. E. Perillo, Ed.), pp. 333 - 357. Elsevier, Amsterdam.
Babb, J. S., Cole, C. A., Brooks, R. P. and Rose, A. W. Hydrogeomorphology, watershed geology, and
water quality of wetlands in central Pennsylvania. Journal of the Pennsylvania Academy of Science 71(1):
21-28.
Baker, J. M., and Wolff, W. J. (1987). Biological Surveys of Estuaries and Coasts - estuarine and
brackish-water sciences association handbook. Cambridge University Press, Cambridge.
Barson, M. M., and Williams, J. E. (1991). “Wetland Inventory - towards a unified approach. (Draft
Discussion Paper for Technical Workshop hosted by the Rural Resources and Australian National Parks
and Wildlife Service, August 5 - 6 1991, Canberra).” Australian National Parks and Wildlife Service,
Canberra.
Bayly, I. A. E. (1974). Australian estuaries. Ecological Society of Australia: Proceedings 8, 41 - 66.
Bayly, I. A. E. (1980). Estuaries and coastal lakes. An ecological basis for water resource management.
W. D. Williams. Canberra, Australian National University Press: 391-398.
Beadle, N. C. W. (1981). The Vegetation of Australia. (Chapter 23, Part 4). Gustav Fuscher Verlag, New
York.
Beeftink, W. G. (1977). Salt-marshes. The coastline - a contribution to our understanding of its ecology
and physiography in relation to land-use and management and the pressures to which it is subject. R. S.
K. Barnes. London, John Wiley and Sons: 93-122.
Beer, T. (1983). Australian estuaries and estuarine modelling. Search 14, 136 - 140.
Bell, F. C., and Vorst, P. C. (1981). Geomorphic Parameters of Representative Basins and Their
Hydrologic Significance - Technical Paper No. 58. Department of National Development and Energy,
Australian Water Resources Council, Australian Government Publishing Service, Canberra.
Bernert, J. A., Daggett, S. G., Bierly, K. F., Eilers, J. M., Eilers, B. J. and Blok, E. (1999). Recent
wetlands trends (1981/82-1994) in the Willamette Valley, Oregon, USA. Wetlands 19(3): 545-559.
Bird, E. C. F. (1984). Coasts: an introduction to coastal geomorphology. Australian National University
Press, Canberra.
Bird, E. C. F., and Barson, M. M. (1977). Measurement of physiographic changes on mangrove-fringed
estuaries and coastlines. Marine Research in Indonesia 18, 73 - 80.
168
Blaber, S. J. M. (2000). Estuary health and value. National Land and Water Resources Audit, Australia Audit themes and projects, Natural Heritage Trust.
Blackman, J. G., Gardiner, S. J., and Morgan, M. G. (1995). Framework for biogeographic inventory,
assessment, planning and management of wetland systems: the Queensland approach. In “Wetlands
Research in the Wet-Dry Tropics of Australia. Workshop, Jabiru, N. T. 22 - 24 March 1995” (C. M.
Finlayson, Ed.), pp. 114 - 122. Supervising Scientist, Barton.
Blackman, J. G., Span, A. V., and Whiteley, L. A. (1992). “Provisional Handbook for the Classification
and Field Assessment of Queensland Wetlands and Deep Water Habitats.” Wetland Inventory Team,
Conservation Strategy Branch, Department of Environment and Heritage, Northern Regional Centre,
Pallarenda.
Blanton, J. O., and Gross, T. F. (). LU-CES State of Knowledge Report on Physical Oceanography. South
Atlantic Bight, Land Use - Coastal Ecosystem Study. /www.baruch.sc.edu/luces2/blanton.html.
Bokuniewicz, H. (1995). Sedimentary systems of coastal-plain estuaries. In Geomorphology and
Sedimentology of Estuaries (G. M. E. Perillo, Ed.), pp. 49 - 67. Elsevier, Amsterdam.
Boon, P. J., Calow, P., and Petts, G. E. (1992). River Conservation and Management. John Wiley &
Sons, Chichester.
Bowden, K. F. (1968). Circulation and Diffusion. In Estuaries (G. H. Lauff, Ed.), pp. 15 - 36. American
Association for the Advancement of Science, Washington, DC.
Boyd, R., Dalrymple, R. and Zaitlin, B. A. (1992). Classification of clastic coastal depositional
environments. Sedimentary Geology 80(3-4): 139-150.
Bradford, A. L. and Watt, W. E. Hydrogeological and other controls on the distribution of plant
communities at Minesing Swamp, Ontario, Canada. Geological Society of America, 1998 annual meeting
30(7): 119.
Brinson, M. M., and Rheinhardt, R. (1996). The role of reference wetlands in functional assessment and
mitigation. Ecological Applications 6, 69 - 76.
Brooks, K. N. (1997). Hydrology and the management of watersheds. Ames, Iowa, Iowa State University
Press.
Bucher, D., and Saenger, P. (1991). An inventory of Australian estuaries and enclosed marine waters: an
overview of results. Australian Geographical Studies 29, 370 - 381.
Bucher, D. and Saenger, P. (1994). A classification of tropical and subtropical Australian estuaries.
Aquatic Conservation 4(1): 1-19.
Bulger, A. J., Monaco, M. E. and McCormick-Ray, M. G. (1993). Biologically-based estuarine salinity
zones derived from a multivariate analysis. Estuaries 16(2): 311-322.
Bunt, J. S., and Williams, W. T. (1980). Studies in the analysis of data from Australian tidal forests
('Mangroves'). I. Vegetational sequences and their graphic representation. Australian Journal of Ecology
5, 385 - 390.
Bunt, J. S., Williams, W. T., and Clay, H. J. (1982a). River water salinity and the distribution of
mangrove species along several rivers in north Queensland. Australian Journal of Botany 30, 401 - 412.
Bunt, J. S., Williams, W. T., and Duke, N. C. (1982b). Mangrove distribution in north-east Australia.
Journal of Biogeography 9, 111 - 120.
169
Burban, P. Y., Lick, W., and Lick, J. (1989). The flocculation of fine-grained sediments in estuarine
waters. Journal of Geophysical Research 94, 8323 - 833.
Bureau of Soils and International Rice Research Institute (1985). Wetland soils: characterization,
classification, and utilization: proceedings of a workshop held 26 March to 5 April 1984. Laguna,
Philippines, International Rice Research Institute. United States. Soil Management Support Services,
Philippines.
Burton, J. D., and Liss, P. S. (1976). Estuarine Chemistry. Academic Press, London.
Callaway, R. M., Jones, S., Ferren, J. W. R., and Parikh, A. (1990). Ecology of a mediterranean-climate
estuarine wetland at Carpinteria, California: plant distributions and soil salinity in the upper marsh.
Canadian Journal of Botany 68, 1139 - 1146.
Carne, R. J. (1991). Landform Vegetation Relationships in the Minnamurra Estuary, N.S.W. Monograph
Series No. 6. Department of Geography and Oceanography, University College, Australian Defence
Academy, Canberra.
Carter, V. National wetland-classification system. U. S. Geological Survey Professional Paper: 233.
Carter, V. P. Wetland hydrology, water quality, and associated functions. National water summary on
wetland resources: 35-48.
Cole, C. A. and Brooks, R. P. (2000). Patterns of wetland hydrology in the ridge and valley province,
Pennsylvania, USA. Wetlands 20(3): 438-447.
Colombo, G. (1977). Lagoons. The coastline - a contribution to our understanding of its ecology and
physiography in relation to land-use and management and the pressures to which it is subject. R. S. K.
Barnes. London, John Wiley and Sons: 63-82.
Corre, J. J. (1985). Environmental structures and variation in coastal vegetation of the Golfe du Lion
(France). Vegetatio 61, 15 - 22.
Cowardin, L. M., Carter, V., Golet, F. C., and LaRoe, E. T. (1979). Classification of Wetlands and
Deepwater Habitats of the United States. U. S. Department of the Interior, Fish and Wildlife Service,
Washington, D.C. http:// www.npwrc.usgs.gov/resource/1998/classwet/classwet.htm.
Cowardin, L. M., and Golet, F. C. (1995). Us Fish and Wildlife Service 1979 wetland classification: a
review. Vegetatio 118, 139 - 152.
Cowardin, L. M. Wetland classification in the United States. Journal of Forestry 76(10): 666-668.
Dale, P. E. R., Chandica, A. L. and Evans, M. (1996). Using image subtraction and classification to
evaluate change in sub- tropical intertidal wetlands. International Journal of Remote Sensing 17(4): 703719.
Day, J. H. (1981). The nature, origin and classification of estuaries. Estuarine ecology, with particular
reference to southern Africa. J. H. Day, A.A. Balkema: 1-6.
DCM (2000). Coastal wetlands: get to know them. Division of coastal management, North Carolina.
http://dcm2.enr.state.nc.us/Wetlands/wetlandsknow.htm.
Digby, M. J., Saenger, P., Whelan, M. B., McConchie, D., Eyre, B., Holmes, N., and Bucher, D. (1999).
A Physical Classification of Australian Estuaries. Centre for Coastal Management, Southern Cross
University (Land and Water Resources Research and Development Corporation), Lismore, NSW.
170
Driscoll, P. V. (1995). Wetland Definition and Fauna Assessment of Cape York Peninsula. Cape York
Peninsula Land Use Strategy, Office of the Coordinator General of Queensland, Brisbane, Department of
Environment, Sport and Territories, Canberra, and Queensland Department of Environment and Heritage,
Brisbane.
Dyer, K. R. (1995). Sediment transport processes in estuaries. In “Geomorphology and Sedimentology of
Estuaries” (G. M. E. Perrillo, Ed.), pp. 423 - 449. Elseviewr, Amsterdam.
Duttmann, R. and Mosimann, T. Die oekologische Bewertung und dynamische Modellierung von
Teilfunktionen und -prozessen des Landschaftshaushaltes; Anwendungen und Perspektiven eines
geooekologischen Informationssystems in der Praxis. Petermanns Geographische Mitteilungen 138(1): 317.
Dyer, K. R. and Huntley, D. A. (1999). The origin, classification and modelling of sand banks and ridges.
Continental Shelf Research 19(10): 1285-1330.
Edgar, G. J., Barret, N. S., Graddon, D. J., and Last, P. R. (2000). The conservation significance of
estuaries: a classification of Tasmanian estuaries using ecological, physical and demographic attributes as
a case study. Biological Conservation 92, 383 - 397.
Engelen, G. B. and Kloosterman, F. H. (1996). Hydrological systems analysis : methods and applications.
Boston, Kluwer.
Everard, M. (1997). Development of a British wetland strategy. Aquatic Conservation: Marine and
Freshwater Ecosystems 7(3): 223-238.
Eyre, B. (1993). Nutrients in the sediments of a tropical north-eastern Australian estuary, catchment and
near-shore coastal zone. Australian Journal of Marine and Freshwater Research 44, 845 - 866.
Eyre, B. (1995). A first-order nutrient budget for the tropical Moresby Estuary and catchment, North
Queensland, Australia. Journal of Coastal Research 11, 717 - 732.
Eyre, B. (1999). Methodology for Evaluating the Impact of Sewage Effluent Loads on the Clarence
Estuary. Centre for Coastal Management for DPWS.
Eyre, B. (). A regional evaluation of nutrient transformation and phytoplankton growth at the land-sea
interface of nine rivers dominated sub-tropical East Australian estuaries using modified mixing diagrams.
Marine Ecology Progress Series .
Eyre, B., and Balls, P. (1999). A comparative study of nutrient behaviour along the salinity gradient of
tropical and temperate estuaries. Estuaries 22, 313 - 326.
Fairweather, P. G. (1999). Determining the 'health' of estuaries: priorities for ecological research.
Australian Journal of Ecology 24, 441 - 451.
Ferreira, J. G. (2000). Development of an estuarine quality index based on key physical and
biogeochemical features. Ocean and Coastal Management 43(1): 99-122.
Ferren Jr, W. R., Fielder, P. L., and Leidy, R. A. (1996). Wetlands of the Central and Southern California
Coast and Coastal Watersheds: a methodology for their classification and description. U. S.
Environmental Protection Agency. http://lily.mip.berkley.edu/wetlands/titlepag.html.
Field, D. W., Alexander, C. E., and Broutman, M. (1988). Toward developing an inventory of U. S.
coastal wetlands. Marine Fisheries Review 50, 40 - 46.
171
Finlayson, C. M., Bailey, B. J., Freeland, W. J., and Fleming, M. R. (1988). Wetlands of the Northern
Territory. In “The Conservation of Australian Wetlands” (A. J. M. P. S. Lake, Ed.), pp. 103 - 126. Surrey
Beatty & Sons Pty Ltd, New South Wales, Australia.
Finlayson, C. M., Von Oertzen, I., Jacobs, S. W. L. and Brock, M. A. (1993). Wetlands of Australia.
Wetlands of the world: inventory, ecology and management. Vol. I. D. F. Whigham. Handbook of
Vegetation Science, 15/2, Kluwer: 195-304.
Finlayson, C. M., and van der Valk, A. G. (1995). Wetland classification and inventory: a summary.
Vegetatio 118, 185 - 192.
Finlayson, C. M., and von Oertzen, I. (). Wetlands of Australia: northern (tropical) Australia. pp. 195 304.
Flint, R. W. and Younk, J. A. (1983). Estuarine benthos: long-term community structure variations,
Corpus Christi Bay, Texas Estuaries 6(2): 126-141.
Frissel, C. A., Liss, W. J., Warren, C. E., and Hurley, M. D. (1986). A hierarchical framework for stream
habitat classification: viewing streams in a watershed context. Environmental Management 10, 199 - 214.
Garofalo, D. (1980). The influence of wetland vegetation on tidal stream channel migration and
morphology. Estuaries 3, 258 - 270.
Gebhardt, K., Leonard, S., Staidl, G. and Prichard, D. "Riparian area management; riparian and wetland
classification review." Technical Reference - U. S. Department of the Interior, Bureau of Land
Management Service Center: 56.
Gebhardt, K. A. (1989). "Ground-water consideration in riparian/wetland classification."
Gerla, P. J. (1999). Estimating the ground-water contribution in wetlands using modeling and digital
terrain analysis. Wetlands 19(2): 394-402.
Gerritson, J., Carlson, R. E., Dycus, D. L., Faulkner, C., Gibson, G. R., Harcum, J., and Markowitz, S. A.
(). Lake and Reservoir Bio-assessment and Bio-criteria: technical guidance document. U. S.
Environmental
Protection
Agency,
Office
of
Water,
Washington,
DC
/www.epa.gov.owow/monitoring/tech/lakes.htm.
Ghassemi, F., Jakeman, A. J., Nix, H. A. and Australian National University. Centre for Resource and
Environmental Studies. (1995). Salinisation of land and water resources: human causes, extent,
management, and case studies. Sydney, N.S.W.,, NSW University Press.
Gilvear, D. J. and McInnes, R. J. (1994). Wetland hydrological vulnerability and the use of classification
procedures: a Scottish case study. Journal of Environmental Management 42(4): 403-414.
Goda, T. Scope and limit in the application of ecological models to environmental management; I-IV.
Ecological Modelling 31(1-4): 363.
Gowing, D. J. G. and Youngs, E. G. The effect of the hydrology of a Thames flood meadow on its
vegetation pattern. Floodplain rivers; hydrological processes and ecological significance 8: 69-80.
Gren, I. M., Folke, C., Turner, K. and Bateman, I. (1994). Primary and secondary values of wetland
ecosystems. Environmental & Resource Economics 4(1): 55-74.
Grime, J. P. (1974). Vegetation classification by reference to strategies. Nature 250, 26 - 30.
Guilcher, A. (1968). Origin of sediments in estuaries. In “Estuaries” (G. H. Lauff, Ed.), pp. 149 - 157.
American Association for the Advancement of Science, Washington DC.
172
Gwin, S. E., Kentula, M. E., and Shaffer, P. W. (1999). Evaluating the effects of wetland regulation
through hydrogeomorphic classification and landscape profiles. Wetlands 19, 477 - 489.
Hansen, D. V., and Rattray, J. M. (1966). New dimensions in estuary classification. Limnology and
Oceanography XI, 319 - 326.
Hansen, P. L., Pfister, R. D., Boggs, K., Cook, B. J., Joy, J. and Hinckley, D. K. Classification and
management of Montana's riparian and wetland sites. MFCES Miscellaneous Publication 54: 646.
Harding, W. R. (1994). Water quality trends and the influence of salinity in a highly regulated estuary
near Cape Town, South Africa. South African Journal of Science 90(4): 240-246.
Harper-Arabie, R. M. and Kolm, K. E. (1998) A stepwise, integrated hydrogeomorphic approach for the
classification of wetlands and assessment of wetland hydrological function in the Southern Rocky
Mountains of Colorado. Geological Society of America, 1998 annual meeting 30(7): 121.
Harris, R. R. (1988). Associations between stream valley geomorphology and riparian vegetation as a
basis for landscape analysis in the Eastern Sierra Nevada, California, U. S. A. Environmental
Management 12, 219 - 228.
Hart, B. S. (1995). Delta front estuaries. In “Geomorphology and Sedimentology of Estuaries” (G. M. E.
Perillo, Ed.). Elsevier, Amsterdam.
Biology 40(3): 517-530.
Hayden, B. P., Ray, G. C., and Dolan, R. (1984). Classification of coastal and marine environments.
Environmental Conservation 11, 199 - 207.
Biology 40(3): 517-530.
Heath, R. A. and Grange, K. R. (1982). Classification of New Zealand inlets: short-circuiting
environmental assessment of New Zealand coastal inlets. Proc. river and estuary mixing workshop,
Hamilton, NZ, 1981. J. C. Rutherford. 49, NZ Water & Soil Miscellaneous Publication: 91-92.
Henderson, F. M., Hart, T. F., Jr., Heaton, B. P. and Portolese, J. E. (1999). Mapping coastal ecosystems
over a steep development gradient using C-CAP protocols. International Journal of Remote Sensing
20(4): 727-7
Hill, A. L., Semeniuk, C. A., Semeniuk, V., and Del Marco, A. (1996). Wetlands of the Swan Coastal
Plain - wetland mapping, classification and evaluation, main report., Vol. 2a. Water and Rivers
Commision & Department of Environmental Protection, W. A., Perth.
Hills, J. M., Murphy, K. J., Pulford, I. D., and Flowers, T. H. (1994). A method for classifying European
riverine wetland ecosystems using functional vegetation groups. Functional Ecology 8, 242 - 252.
Holmes, N. T. H. (1999). Recovery of headwater stream flora following the 1989-1992 groundwater
drought. Hydrological Processes 13(3): 341-354.
Howard, J. D., and Frey, R. W. (). Characteristic physical and biogenic sedimentary structures in Georgia
estuaries..
173
Huffman, R. T. and Tucker, G. E. "Preliminary guide to the onsite identification and delineation of the
wetlands of Alaska."
Hunter, K. A., and Liss, P. S. (1982). Organic matter and the surface charge of susoended particles in
estuarine waters. Limnology and Oceanography 27, 322 - 335.
Huthchinson, M. F., and Dowling, T. I. (1991). A continental hydrological assessment of a new gridbased digital elevation model of Australia. Hydrological Processes 5, 45 - 58.
Imhof, J. G., Fitzgibbon, J., and Annable, W. K. (1996). A heirarchical evaluation system for the
characterizing watershed ecosystems for fish habitat. Canadian Journal of Fisheries and Aquatic Science
53, 312 - 326.
Imhof, J. G., and FitzGibbon, J. E. (1999). A procedure for developing a riparian/valley classification
system for management. In “Proceedings of the Second Internation Conference on Natural Channel
Systems” (P. Young and A. Boyds, Eds.), March 1 - 4, 1999, Niagra Falls, Canada.
Institute, I. R. R. (1984). Wetland Soils: Characterization, Classification, and Utilization. Laguna,
Phillapines: Internation Rice Research Institute.
Jay, D. A. (). Selected Publications -Abstracts- ("Interaction of Fluctuating River Flow with a Barotropic
Tide: a demonstration of wavelet tidal analysis methods"; "An Ecological Perspective on Estuarine
Classification"; "A Comparison of Methods for Analysis of Tidal Records Containing Multi-Scale NonTidal Background Energy"; "A Review of Recent Developments in Estuarine Scalar Flux Estimation";
"Downstream Effects of Water Withdrawel in a Small, High-Gradient Basin: erosion and deposition on
the Skokomish River delta"; "Formation of the Columbia River Plume: hydraulic control in
action").http://www.ccalmr.ogi.edu/~djay/pubs_long.htm.
Jennings, J. N., and Bird, E. C. F. (1968). Regional geomorphological characteristics of some Australian
estuaries. In “Estuaries” (G. H. Lauff, Ed.), pp. 121 - 128. American Association for the Advancement of
Sciences, Washington DC.
Jorgensen, S. E. Physical and chemical unit processes of ecological modelling. Application of ecological
modelling in environmental management, Part A 4A: 107-129.
Jorgensen, S. E. Application of ecological modelling in environmental management, Part A.
Developments in Environmental Modelling 4A: 731.
Jorgensen, S. E. (1994). Fundamentals of ecological modelling. Amsterdam ; London, Elsevier.
Jorgensen, S. E. (2001). Thermodynamics and ecological modelling. Environmental and ecological
modelling: 373.
Jorgensen, S. E. and Gromiec, M. J. (1989). Mathematical submodels in water quality systems.
Amsterdam ; New YorkNew York, NY, USA, Elsevier ;Distributors for the U.S. and Canada Elsevier
Science Pub. Co.
Jorgensen, S. E., Jorgensen, L. A., Mejer, H. F. and Kamp-Nielsen, L. A water quality model for the
Upper Nile lake system. Analysis of ecological systems; State-of-the-Art in ecological modelling 5: 631639.
Kamp-Nielsen, L. Sediment-water exchange models. Application of ecological modelling in
environmental management, Part A 4A: 387-420.
Kamp-Nielsen, L., Jorgensen, L. A. and Jorgensen, S. E. (1981). "A sediment-water exchange model for
lakes in the Upper Nile Basin."
174
Kangas, P. C. (1990). An energy theory of landscape for classifying wetlands. Forested wetlands. A. E.
Lugo. Ecosystems of the World, 15, Elsevier: 15-23.
Keddy, P., Fraser, L. H. and Wisheu, I. C. (1998). A comparative approach to examine competitive
response of 48 wetland plant species. Journal of Vegetation Science 9(6): 777-786.
Keddy, P. A. (2000). Wetland ecology : principles and conservation. New York, Cambridge University
Press.
Kennish, M. J. (1986). “Ecology of Estuaries. Volume 1: physical and chemical aspects.” CRC Press Inc.,
Florida.
Klemas, V. V. (2001). Remote sensing of landscape-level coastal environmental indicators.
Knox, G. A. (1986). Estuarine Ecosystems: A Systems Approach. CRC Press Inc., Florida.
Kolm, K. E. and Emerick, J. C. (1997) An approach for the evaluation of wetland structure, hydrology,
and biogeochemistry for wetland ecosystem management in the Southern Rocky Mountains of Colorado.
Geological Society of America, 1997 annual meeting 29(6): 67.
Kudray, G. M. and Gale, M. R. (2000). Evaluation of national wetland and inventory maps in a heavily
forested region in the upper great lakes. Wetlands 20(4): 581-587.
Lassey, K. R. Conceptually simple mathematical models of filtration (and exchange) processes.
Ecological Modelling 18(1): 1-26.
Lent, R. M., Waldron, M. C. and Rader, J. C. (1998). Multivariate classification of small order
watersheds in the Quabbin Reservoir Basin, Massachusetts. Journal of the American Water Resources
Association 34(2): 439-450.
Leonard, S., Staidl, G., Fogg, J., Gebhardt, K., Hagenbuck, W. and Prichard, D. "Riparian area
management; procedures for ecological site inventory; with special reference to riparian-wetland sites."
Technical Reference - U. S. Department of the Interior, Bureau of Land Management Service Center:
135.
Lotspeich, F. B. (1980). Watersheds as the basic ecosystem: this conceptual framework provides a basis
for a natural classification system. Water Resources Bulletin 16, 581 - 586.
Lucero, A. R. and James Cook University of North Queensland. Dept. of Earth Sciences. (1996).
Hydrogeochemical studies of an aquifer below a tropical wetland Ingham, Queensland. Townsville, Qld.
Luczkovich, J. J. Estuarine Ecology (LectureNotes). /www.drjoe.biology.ecu.edu/estuary/estuary.htm.
Lugo, A. E., and Snedaker, S. C. (1974). The ecology of mangroves. 39 - 64.
Lukacs, G. P. (1995). Wetlands of the Lower Burdekin Region, north Queensland. In Wetland Research
in the Wet-Dry Tropics of Australia. Workshop, Jabiru, N. T. 22 - 24 March 1995 (C. M. Finlayson, Ed.),
pp. 86 - 94. Supervising Scientist, Barton.
Lung, W.-S. and Light, R. N. Modelling copper removal in wetland ecosystems. Ecological Modelling
93(1-3): 89-100.
Lunn, R. J., Adams, R., Mackay, R. and Dunn, S. M. Development and application of a nitrogen
modelling system for large catchments. Journal of Hydrology 174(3-4): 285-304.
175
Luternaur, J. L., Atkins, R. J., Moody, A. I., Williams, H. F. L., and Gibson, J. W. (1995). Salt marshes.
In Geomorphology and Sedimentology of Estuaries (G. M. E. Perillo, Ed.), pp. 307 - 332. Elsevier,
Amsterdam.
Lyon, J. G. (1993). Practical Handbook for Wetland Identification and Delineation. Lewis Publishers,
London.
McComb, A. J., and Lake, P. S. (1988). The Conservation of Australian Wetlands. Surrey Beatty & Sons
Pty Ltd in association with World Wildlife Fund Australia, Chipping Norton, N. S. W.
McComb, A. J., and Lake, P. S. (1990). Australian Wetlands. Angus & Robertson Publishers Australia,
North Ryde, N. S. W.
McLay, C. L. An inventory of the status and origin of New Zealand estuarine systems. Proceedings New Zealand Ecological Society 23: 8-26.
McLusky, D. S. (1981). The Estuarine Ecosystem Blackie, Glasgow.
McVoy, C. W., Kersebaum, K. C. e., Diekkrueger, B. e. and Soendgerath, D. e. Modelling of geobiosphere processes; Validation of agroecosystem models. Ecological Modelling 81(1-3): 300.
Megonigal, J. P., Patrick, W. H., Jr. and Faulkner, S. P. Wetland identification in seasonally flooded
forest soils; soil morphology and redox dynamics. Soil Science Society of America Journal 57(1): 140149.
Michaelis, W. and GKSS-Forschungszentrum Geesthacht. (1990). Estuarine water quality management
monitoring, modelling, and research. Berlin ; New York, Springer-Verlag.
Ming-Ko, W., Rowsell, R. D. and Clark, R. G. (1993). Hydrological classification of Canadian prairie
wetlands and prediction of wetland inundation in response to climatic variability. Occasional Paper Canadian Wildlife Service 79: 22.
Mitsch, W. J., and Gosselink, J. G. (1986). Wetlands. Van Nostrand Reinhold Company, New York.
Mitsch, W. J., Straskraba, M. and Jorgensen, S. E. (1988). Wetland modelling. Amsterdam, Elsevier.
Morrison, M., Bennett, J., Blamey, R. K. and University of New South Wales. School of Economics and
Management. (1998). Valuing improved wetland quality using choice modelling. Canberra, School of
Economics and Management University College University of New South Wales.
Mucha, A. P. and Costa, M. H. (1999). Macrozoobenthic community structure in two Portuguese
estuaries: Relationship with organic enrichment and nutrient gradients. Acta Oecologica 20(4): 363-376.
Nelson-Smith, A. (1977). Estuaries. The coastline - a contribution to our understanding of its ecology and
physiography in relation to land-use and management and the pressures to which it is subject. R. S. K.
Barnes. London, John Wiley and Sons: 123-146.
NRC. (1995). Wetlands: characteristics and boundaries (National Research Council). National Academy
Press, Washington.
Oberrecht, K. (1997). Classification of Estuaries.http://www.harborside.com/~ssnerr/class1.htm.
Okruszko, H. Klassifizierung von Feuchtgebieten auf der Grundlage der in ihnen entstehenden
hydromorphen Bodensubstrate. Telma 10: 17-24.
Osborne, L. L., and Kovacic, D. A. (1993). Riparian vegetated buffer strips in water-quality restoration
and stream management. Freshwater Biology 29, 243 - 258.
176
Paijmans, K., Galloway, R. W., Faith, D. P., Fleming, P. M., Haantjens, H. A., Heyligers, P. C., Kalma, J.
D., and Loffler, E. (1985). “Aspects of Australian Wetlands (Division of Water and Land Resources
Technical Paper No. 44).” Commonwealth Scientific and Industrial Research Organisation, Australia.
Patrick, R., and Palavage, D. M. (1994). The value of species as indicators of water quality. Proceedings
of the Academy of Natural Sciences of Philadelphia 145, 55 - 92.
Pawlowicz, R. and Farmer, D. M. (1998). Diagnosing vertical mixing in a two-layer exchange flow.
Journal of Geophysical Research C: Oceans 103(13): 30,695-30,711.
Perillo, G. M. E. (1995). Geomorphology and sedimentology of estuaries. Developments in
Sedimentology, 53, Elsevier: 471.
Perillo, G. M. E., Cintia Piccolo, M., Del Carmen Scapini, M. and Orfila, J. (1989). Hydrography and
circulation of the Chubut River estuary (Argentina). Estuaries 12(3): 186-194.
Perillo, G. M. E. (1995). Definitions and geomorphological classifications of estuaries. In
Geomorphology and Sedimentology of Estuaries (G. M. E. Perillo, Ed.), pp. 17 - 47. Elsevier,
Amsterdam.
Perry, J. A. and Vanderklein, E. (1996). Water quality : management of a natural resource. Cambridge,
Mass., Blackwell science.
Prandle, D. (1985). On salinity regimes and the vertical structure of residual flows in narrow tidal
estuaries. Estuarine, Coastal & Shelf Science 20(5): 615-635.
Prescott, K. L. and Tsanis, I. K. Mass balance modelling and wetland restoration. Ecological Engineering
9(1-2): 1-18.
Pressey, R. L. (1989a). Wetlands of the Lower Clarence Floodplain, northern coastal New South Wales.
Pressey, R. L. (1989b). Wetlands of the Lower Maccleay Floodplain, northern coastal New South Wales.
Pressey, R. L., and Adam, P. (1995). A review of wetland inventory and classification in Australia.
Vegetatio 118, 81 - 101.
Pressey, R. L., and Harris, J. H. (1988). Wetlands of New South Wales. In “The Conservation of
Australian Wetlands” (A. J. McComb and P. S. Lake, Eds.), pp. 35 - 57. Surrey Beatty & Sons Pty Ltd,
New South Wales, Australia.
Pritchard, D. W. (1968). What is an estuary: physical viewpoint. In “Estuaries” (G. H. Lauff, Ed.), pp. 3 5. American Association for the Advancement of Science, Washington, DC.
Prichard, D., Clemmer, P., Gorges, M., Meyer, G. and Shumac, K. Riparian area management; using
aerial photographs to assess proper functioning condition of riparian-wetland areas. Technical
Reference - U. S. Department of the Interior, Bureau of Land Management Service Center: 41.
Rabalais, S. C., Pulich, J. W. M., Rabalais, N. N., felder, D. L., Tinnin, R. K., and Kalke, R. D. (1989). A
biological and physiological characterization of the Rio Carrizal estuary, Tamaulipas, Mexico.
Contributions in Marine Science 31, 25 - 37.
Rabenhorst, M. C. (2001). "Soils of tidal and fringing wetlands."
177
Rajar, R. and Cetina, M. Hydrodynamic and water quality modelling; an experience.
Modelling 101(2-3): 195-207.
Ecological
Rajar, R., Cetina, M. and Sirca, A. Hydrodynamic and water quality modelling; case studies. Ecological
Modelling 101(2-3): 209-228.
Ramm, A. E. L. (1990). Application of the community degradation index to South African estuaries.
Water Research 24(3): 383-389.
Richardson, J. L., Arndt, J. L. and Montgomery, J. A. (2001). "Hydrology of wetland and related soils."
Richardson, J. L. and Brinson, M. M. (2001). "Wetland soils and the hydrogeomorphic classification of
wetlands."
Richardson, J. L. and Vepraskas, M. J. e. (2001). "Wetland soils; genesis, hydrology, landscapes, and
classification." 417.
Rio, J. N. R. and Lozano-Garcia, D. F. (2000). Spatial filtering of radar data (RADARSAT) for wetlands
(brackish marshes) classification. Remote Sensing of Environment 73(2): 143-151.
Rochford, D. J. (1959). Classification of Australian estuarine systems. Archivio Di Oceanografia e
Limnologia 11, 171 - 177.
Roggeri, H. (1995). “Tropical Freshwater Wetlands: a guide to current knowledge and sustainable
management.” Kluwer Academic Publishers, London.
Rosgen, D. L. (1994). A classification of natural rivers. Catena 22, 169 - 199.
Rosgen, D. L. (1988). "A stream classification system."
Roulet, N. T. and McKenzie, J. M. Role of groundwater in determining the pattern of peatlands in the
Hudson Bay Lowland. Geological Society of America, 1998 annual meeting 30(7): 119.
Roy, P. S. (1984). New South Wales estuaries: their origin and evolution. Coastal geomorphology in
Australia. B. G. Thom, Academic Press: 99-121.
Saenger, P. (1999). The Status of Australian Estuaries and Enclosed Marine Waters - Technical Annex I.
State of the Environment Report for Australia. Environment Australia, Department of Environment,
Canberra. http://www.environment.gov.au/marine/publications/somer/somer_annex1/som_ann5.html.
Saenger, P., and Robson, J. (1977). Structural analysis of mangrove communities on the Central
Queensland coastline. Marine Research in Indonesia 18, 101 - 118.
Salathé, T. (1993). Towards integrated management of coastal wetlands of Mediterranean type. The
Ramsar Library - Towards the wise use of wetlands. T. J. Davis: //www.ramsar.org/lib_wise_4.htm.
Sattler, P., and Williams, R. (1999). “The Conservation Status of Queensland's Bioregional Ecosystems.”
Environmental Protection Agency, Brisbane.
Schum, M. and Varnell, J. A simulation model of radionuclide migration through coastal ecosystems.
Analysis of ecological systems; State-of-the-Art in ecological modelling 5: 969-978.
Scott, C. F. (1993). Canonical parameters for estuary classification. Estuarine, Coastal and Shelf Science
36, 529 - 540.
Scott, D. A., and Jones, T. A. (1995). Classification and inventory of wetlands: a global overview.
Vegetatio 118, 3 - 16.
178
Semeniuk, C. A. (1987). Wetlands of the Darling System - a geomorphic approach to habitat
classification. Journal of the Royal Society of Western Australia 69, 95 - 112.
Semeniuk, C. A., and Semeniuk, V. (1995a). Geomorphic approach to classifying wetlands in tropical
north Australia. In “Wetland Research in the Wet-Dry Tropics of Australia. Workshop, Jabiru, N. T. 22 24 March 1995” (C. M. Finlayson, Ed.), pp. 123 - 128. Supervising Scientist, barton.
Semeniuk, C. A., and Semeniuk, V. (1995b). A geomorphic approach to global classification for inland
wetalnds. Vegetatio 118, 103 - 124.
Semeniuk, C. A., Semeniuk, V., Cresswell, I. D. and Marchant, N. G. (1990). Wetlands of the Darling
System, southwestern Australia: a descriptive classification using vegetation pattern and form. Journal Royal Society of Western Australia 72(4): 109-121.
Shaffer, P. W., Kentula, M. E. and Gwin, S. E. (1999). Characterization of wetland hydrology using
hydrogeomorphic classification. Wetlands 19(3): 490-504.
Sharifi, M. (1990). Assessment of surface water quality by an index system in Anzali Basin. The
hydrological basis for water resources management. Proc., symposium, Beijing, 1990. U. Shamir.
Publication, 197, Iahs: 163-171.
Shubel, J. R. (1971). The Estuarine Environment: estuaries and estuarine sedimentation (Short Course
Lecture Notes). American Geological Institute, Washington DC.
Smith, I. T. J., and Duke, N. C. (1987). Physical determinants of inter-estuary variation in mangrove
species richness around the tropical coastline of Australia. Journal of Biogeography 14, 9 - 19.
Smith, M., and Storey, A. (1999). Evaluation/Design and Implementation of Baseline Monitoring
(DIBM3) - South-East Queensland Regional Water Quality Management Strategy. Streamtec Pty Ltd
Ecological Consultants, Nedlands, W. A.
Somlyody, L. Water-quality modelling; a comparison of transport-oriented and ecology-oriented
approaches. Ecological Modelling 17(3-4): 183-207.
Stephens, K. (2001). Wetland plants of Queensland, CSIRO.
Stevenson, R. J., Bothwell, M. L., and Lowe, R. L. (1996) Algal ecology: freshwater benthic ecosystem,.
Academic Press.
Stewart, W. D. P. (1972). Estuarine and brackish waters - an introduction. The estuarine environment,
Estuarine and Brackish-water Biological Association, Applied Science Publishers.
Summers, J. K. and Engle, V. D. (1993). Evaluation of sampling strategies to characterize dissolved
oxygen conditions in northern Gulf of Mexico estuaries. Environmental Monitoring & Assessment 24(3):
219-229.
Supervising Scientist for the Alligator Rivers Region (Australia) and Land and Water Resources
Research and Development Corporation (Australia) (1995). Wetland research in the wet-dry tropics of
Australia : workshop, Jabiru, NT, 22-24 March 1995. Barton, A.C.T., Supervising Scientist.
Svendsen, H. (1986). Mixing and exchange processes in estuaries, fjords and shelf waters. The role of
freshwater outflow in coastal marine ecosystems. Proc. Bodo, 1985. S. Skreslet. NATO ASI, Series G, 7,
Springer-Verlag: 13-45.
Thackway, R., and Cresswell, I. D. (1995). An Interim Biogeographic Regionalisation for Australia.
Reserves Systems Unit, Australian Nature Conservation Agency, Canberra, ACT.
179
Thackway, R., and Cresswell, I. D. (1998). Interim Marine and Coastal Regionalisation for Australia: an
ecosystem-based classification for marine and coastal environments. IMCRA Technical Group,
Environment Australia, Department of Environment, Canberra.
Thom, B. G. (). Mangrove ecology and deltaic geomorphology: Tabasco, Mexico. Unknown, 301 - 343.
Tiner, R. W. (1999). “Wetland Indicators: a guide to wetland identification, delineation, classification,
and mapping.” Lewis Publishers, Washington DC.
Tiner, R. W. Wetland definitions and classifications in the United States. National water summary on
Tiner, R. W. (1999). Wetland indicators : a guide to wetland indentification, delineation, classification,
and mapping. Boca Raton, Fla., Lewis Publishers.
Tomczak, M. (1996). Definition of Estuaries; empirical estuary classification.
http://www.es.flinders.edu.au/~matton/ShelfCoast/notes/chapter11.htm.
Tomczak, M. (1997). Estuary classifications based on fluid dynamics principles.
http://www.es.flinders.edu.au/~mattom/ShelfCoast/notes/chapter12.htm.
Töyrä, J., Pietroniro, A. and Martz, L. W. (2001). Multisensor hydrologic assessment of a freshwater
wetland. Remote Sensing of Environment 75(2): 162-173.
Trudgill, S. T., Webb, B. and Walling, D. E. (1999). Water quality : processes and policy. Chichester,
Wiley.
Tubbs, C. R. (1977). Muddy foreshores. The coastline - a contribution to our understanding of its ecology
and physiography in relation to land-use and management and the pressures to which it is subject. R. S.
K. Barnes. London, John Wiley and Sons: 83-92.
UNEP (2000). Coastal wetlands: mangroves, marshes and seagrass beds. United Nations Environment
Programme. http://www.unep.ch/seas/main/hwetl.htm.
Unknown. (1997). Estuaries in Oregon.
http://www.inforain.org/epb/chpt1/ch1s2.htm
http://www.inforian.org/epb/chpt1/ch1s3.htm
http://www.inforain.org/epb/chpt3/ch3s3.htm.
Unknown. (1999a). Classification, vegetation analysis, and inventory of coastal wetlands. U. S.
Environment Protection Agency & Environment Canada.
http://www.epa.gov/grtlakes/solec...ers/coastabia/classification.html.
Unknown. (1999b). National Estuarine Inventory.http:neonet.nlr.nl/ceos-idn/campaigns/NEI.htm.
Unknown. (2000). Classification of Maine Waters. Department of Environmental Protection, Maine, U.
S. Http://janus.state,me.us/dep/blwq/class.htm.
Unknown. (-a). The Estuarine Ecosystem: ecology of the intertidal and subtidal areas. National Oceanic
and Atmospheric Administration, U. S.
http://www.csc.noaa.gov/otter/htmls/ecosys/ecology/estuary.htm.
180
Unknown. (-b). South Carolina Coastal Region. National Oceanic and Atmospheric Administration, U.
S.http://www.csc.noaa.gov/otter/htmls/ecosys/physical/ph_coast.htm.
USFWS (1992). Wetlands and deepwater habitats: hierarchical structure - systems and subsystems. Water
Words Dictionary - Appendix D-2. Washington, US Department of the Interior, Fish and Wildlife
Service.
van der Zee, S. E. A. T. M., Weeks, J. M., O'Hare, S. and Rattner, B. (1998). Contaminant displacement
modelling in soil and groundwater"
Van Soest, F., Williams, A., Parkinson, R. and Kent, M. (2001). Methodology development for wetland
restoration planning based on topography and soil hydrology IAHS-AISH Publication(268): 271-278.
Vardavas, I. M. A simple groundwater recharge depletion model for the tropical Magela Creek
catchment. Ecological Modelling 68(3-4): 147-159.
Vepraskas, M. J. and Richardson, J. L. (2001). Wetland soils : genesis, hydrology, landscapes, and
classification. Boca Raton, Fla., Lewis Publishers.
Walkey, M., Swingland, I. R., Russell, S. and British Council. (1999). Integrated protected area
management. Boston, Kluwer Academic Publishers.
Warner, B. G. and Rubec, C. D. A. e. (1997). "The Canadian wetland classification system." 68.
Wells, J. T. (1995). Tide dominated estuaries and tidal rivers. In “Geomorphology and sedimentology of
Estuaries” (G. M. E. Perillo, Ed.). Elsevier, Amsterdam.
Wells, E. D. and Zoltai, S. (1985). The Canadian system of wetland classification and its application to
circumboreal wetlands. Aquilo Series Botanica 21: 45-52.
Wentworth, T. R., Johnson, G. P., and Kologiski, R. L. (1988). Designation of wetlands by weighted
averages of vegetation data: a preliminary evaluation. Water Resources Bulletin 24, 389 - 396.
Whigham, D. F., Rains, M. C., Mason, J. A., Kahn, H., Ruhlman, M. B., Nutter, W. L., Lee, L. C.,
Brinson, M. M. and Rheinhardt, R. D. (1999). Hydrogeomorphic (HGM) assessment - a test of user
consistency. Wetlands 19(3): 560-569.
Whigham, D. F., Dykyjova, D., and Hejny, S. (1993). “Wetlands of the World I: inventory, ecology, and
management.” Kluwer Academic Publishers, Dordrecht.
Whitehead, P. G., O'Connell, P. E. and Institute of Hydrology (Great Britain) (1984). Water quality
modelling, forecasting and control: proceedings of an international workshop at the Institute. London,
Institute of Hydrology.
Whittecar, G. R. and Daniels, W. L. (2000). Use of hydrogeomorphic concepts to design created wetlands
in southeastern Virginia. Geomorphology 31(1-4): 355-371.
Wilen, B. O., Carter, V. P. and Jones, J. R. Wetland mapping and inventory. National water summary on
Wilson, J. G. (1988). “The Biology of Estuarine Management.” Croom Helm, London.
Wootton, L. (). Estuaries and Coastal Regions.http://www.georgian.edu/~wootton/estuaries.html.
Zedler, J. B. (1977). Salt marsh community structure in the Tijuana estuary, California. Estuarine and
Coastal Marine Science 5, 39 - 53.
181
Zhang, S. Q., Zhang, S. K. and Zhang, J. Y. (2000). A study on wetland classification model of remote
sensing in the Sangjiang Plain. Chinese Geographical Science 10(1): 68-73.
Zogg, G. P. and Barnes, B. V. (1995). Ecological classification and analysis of wetland ecosystems,
northern Lower Michigan, USA. Canadian Journal of Forest Research 25(11): 1865-1875.
182
APPENDIX K
BIBLIOGRAPHY OF WATER QUALITY
183
ACTFR (1997) Baseline stormwater monitoring. Australian Centre for Tropical Freshwater Research,
Townsville. Report No. 97/06.
ACTFR (1999) Halifax Bay water and sediment quality study 1999. Australian Centre for Tropical
Freshwater Research, Townsville. ACTFR Report, 99/33.
ACTFR (2002a) Water quality in the Townsville/Burdekiin dry tropics region. Australian Centre for
Tropical Freshwater Research, Townsville. National Heritage Trust Project No. 2002153,
ACTFR Report No. 02/12.
ACTFR (2002b) Environmental impact assessment of Paluma Town weir supply. Augumentation
proposal: flora, fauna and water quality. Australian Centre for Tropical Freshwater
Research, Townsville. ACTFR Report, No. 02/14. 29.
Alexander J., Fielding, C. R. and Podock, G. D. (1999) Flood behaviour of the Burdekin River, Tropical
North Queensland, Australia. In: S. B. Marriot & Alexander, J. (Eds.), Floodplains:
Interdisciplinary Approaches. Geological Society, London, Special Publications 163:2740.
Alexander J., Fielding, C. R., Wakefield, S. J., George, M.T. and Cottnam, C. F. (in press) Fluvial
geochemistry through a short duration, tropical cyclone induced discharge event in the
Burdekin River and Hann Creek, North Queensland, Australia. Aquatic Geochemistry, in
press.
Anon (1993a) The Condition of River Catchments in Queensland, a Broad Overview of Catchment
Management Issues. QDPI, Brisbane.
Anon (1993b) Water Quality Data Report Queensland, January 1988 to December 1991. Environment
Data Report No. 4. (Qld. Dept. Env. and Heritage, Brisbane.)
Anon (1994) Queensland Water Quality Atlas. (DPI, Brisbane)
Anon (1998) A scooping study of water infrastructure options and related issues in the Burdekin River
Catchment. Engineering Services, Dept. of Natural Resources, Queensland.
Anon (1999) 1998 fish kill summary - central and north Qld cane areas. CRC Sugar. Developed from
Fish Kill Investigation Report prepared by Vern Veitch and made available to CRC Sugar
for internal use. 2pp.
Anon (2001) Queensland water quality guidelines, Draft for comment. Qld. Environmental Protection
Agency, March 2001.
ANZECC (1992) Australian Water Quality Guidelines for Fresh and Marine Waters. Australian and New
Zealand Environment and Conservation Council.
ANZECC and ARMCANZ (2000a) Australian and New Zealand Guidelines for Fresh and Marine Water
Quality. Australian and New Zealand Environment and Conservation Council (ANZECC)
and Agriculture and Resource Management Council of Australia and New Zealand
(ARMCANZ)
ANZECC and ARMCANZ (2000b) Australian Guidelines for Water Quality Monitoring and Reporting.
Australian and New Zealand Environment and Conservation Council (ANZECC) and
Agriculture and Resource Management Council of Australia and New Zealand
(ARMCANZ)
APVMA (2003) Pubcris registration products database. National Registration Authority for Agricultural
and Veterinary Chemicals. www.nra.gov.au/pubcris/.
184
APVMA
(online)
Australian
Pesticides
http://www.apvma.gov.au/index.html.
and
Veterinary
Medicines
Authority.
Arakel, A. V., Hill, C. M., Piorewicz, J. and Connor, T. B. (1989) Hydro-sedimentology of the Johnstone
River estuary. Hydrobiologia 176-177: 51-60.
Arruda, J. A., Marzolf, G. R. and Faulk, R. T. (1983) The role of suspended sediments in the nutrition of
zooplankton in turbid reservoirs. Ecology 64:1225-1235.
Arthington, A. H., Marshall, J. C., Rayment, G. E., Hunter, H. and Bunn, S. E. (1996) Potential impact
of sugarcane production on riparian and freshwater environments. International
Symposium 'Sugar 2000' - Intensive Sugarcane production: Meeting the challenge beyond
2000, Brisbane, Australia, CAB International.
Australian Govt. (2001) The National Action Plan for Salinity and Water Quality. Agriculture, Fisheries
and Forestry - Australia (AFFA), Environment Australia (EA) and the Council of
Australian
Governments.
http://www.napswq.gov.au/downloads/pdf/action-plannov2001.pdf.
Ayling, A. M. and Ayling, A. L. (1998) The effects of the Daintree River flood plume on Snapper Island
coral reefs. A report to the Great Barrier Reef Marine Park Authority: Research Publication
No. 53, ISBN 0 642 23049 8.
Aylmore, L. A. G. (1998) Soil factors in the transport of pesticides from cotton farms: overview paper.
Mnimising the impact of pesticides on the riverine environment: key findings from
research with the cotton industry - 1998 conference. N. J. Schofield and V. E. Edge.
Canberra, LWRRDC. LWRRDC Occassional Paper 23/98: 58-61.
Bacher, G. J., Chapman, J. C., Lim, R. P., Napier, G. M., Nowak, B., Sunderam, R. M. and Thompson, G.
B. (1992) The impact of agriculture on inland temperate waters: field validation.
Australian Biologist 5(4): 196-202.
Baker J. (2003) A report on the Study of Land-Sourced Pollutants and their impacts on water quality in
and adjacent to the Great Barrier Reef. Science Panel report to the Intergovernment
Steering Committee, GBR Water Quality Action Plan, Premier’s Dept.
Baker, J. and Mickelson, S. (1994) Application Technology And Best Management Practices For
Minimizing Herbicide Runoff. Weed Technology(4): 862-869.
Barry, G. A. (1997) Total heavy metal status of horticultural soils in Queensland. Queensland
Department of Natural Resources, Brisbane. Final Report to Horticultural Resource and
Development Corporatio on Project VG 404.
Barry, G. A. (2000) Recycling of wastes. Sustainable nutrient management in sugarcane production short
course - Course Manual. R. C. Bruce. Townsville, CRC for Sustainable Sugar Production:
73Barry, G. A. and Rayment, G. E. (1992) Heavy metals in biota, sediments and soils of Raine Island.
Department of Natural Resources, Queensland. Unpublished final report to the Raine
Island Corporation.
Barry, G. A., Rayment, G. E., Jeffrey, A. J. and Price, A. M. (2001) Changes in cane soil properties from
applications of sugarmill by-products. Proceedings of the 2001 Conference of the
Australian Society of Sugar Cane Technologists held at Mackay, Queensland, Australia,
1st-4th May 2001. D. M. Hogarth, PK Editorial Services Pty Ltd, Brisbane, Australia: 185191.
185
Barry, G., Bloesch, P., Gardner, EA and Rayment, GE (1998) Re-Use of Sewerage Biosolids on
Canelands. Proc. Aust. Soc. Sugar Cane Technol. 20: 69-75.
Baskaran, S., Budd, K. L., Larsen, R. M. and Bauld, J. (2002) A groundwater quality assessment of the
Lower Pioneer Catchment, Qld. Bureau of Rural Sciences, Department of Agriculture,
Fisheries and Forestry - Australia, Canberra. 62.
Batley, G. E., Apte, S. C. and Stauber, J. L. (1999) Acceptability of aquatic toxicity data for the
derivation of water quality guidelines for metals. Marine and Freshwater Research
50(729-738)
Bauld, J., Leach, L. L. and Sandstrom, M. W. (1995) Impact of land use on groundwater quality in the
Burdekin River Irrigation Area In In Downstream effects of land use. H. Hunter, Eyles, AG
and Rayment GE, (Eds) Department of Natural Resources.
Bauld, J., Leach, L. L. and Sandstrom, M. W. (1996) Impact of land use on groundwater quality in the
Burdekin River Delta and the Burdekin River Irrigation Area : Abstract. In Downstream
effects of land use. H. Hunter, Eyles, AG and Rayment GE, (Eds) Department of Natural
Resources, Queensland: 195.
Bell, M. J., Garside, A. L., Moody, P. W., Pankhurst, C., Halpin, N. V. and Berthelsen, J. (2002) Nutrient
dynamics and root health in sugarcane soils. Proceedings of the 2002 Conference of the
Australian Society of Sugar Cane Technologists held at Cairns, Queensland, Australia, 29
April - 2 May 2002. D. M. Hogarth, PK Editorial Services Pty Ltd, Brisbane, Australia:
92-98.
Belpario, A. P. (1988) Terrigenous and carbonate sedimentation in the Great Barrier Reef province.
Carbonate-clastic transitions, developments in sedimentology. L. J. Doyle and H. H.
Roberts, Elsevier. 42: 143-174.
Belperio, A. P. (1979) The combined use of washload and bed material load rating curves for calculation
of total load: an example from the Burdekin River, Australia. Catena 6:317-329.
Belperio, A.P. (1983) Terrigenous sedimentation in the central Great Barrier Reef lagoon: a model from
the Burdekin region. BMR Journal of Australian Geology and Geophysics 8: 179-190
Belperio, A.P. (1988) Terrigenous and carbonate sedimentation in the Great barrier reef province. In:
Doyle, L.J., Roberts, H.H. (Eds), Carbonate-clastic Transitions, Developments in
Sedimentology 42. Elsevier, pp. 143-174
Bevenue, A., Hylin, J. W., Kawano, Y. and Kelley, T. W. (1972) Organochlorine pesticide residues in
water, sediment, algae and fish. Pesticides Monitoring Journal 6(1): 56-64.
Biggs, B. J. F. and R. J. Davies-Colley (1990) Optical properties of Lake Coleridge: the impacts of turbid
inflows. New Zealand Journal of Marine and Freshwater Research 24: 441-451.
Bishop, K. A. (1980) Fish kills in relation to physical and chemical changes in Magela Creek (East
Alligator River System, Northern Territory) at the beginning of the tropical wet season.
Australian Zoology 20(3): 485-500.
Bloesch, P. M., Rayment, G. E. and Pulsford, J. S. (1997) Regional total phosphorus budgets for sugar
production in Queensland. Proceedings of Australian Society of Sugar Cane Technologists
1997: 213-220.
Bohl, H. P., Bonnett, G. D., Fanninng, D. J., Rayment, G. E. and Davidson, A. B. (2002) Biological
oxygen demand and sugars in irrigation water runoff from sugarcane fields. Proceedings of
the Australian Society of Sugar Cane Technologists 24: 297 - 303.
186
Bohl, H. P., Fanning, D. J., Bonnett, G. D. and Rayment, G. E. (2001) Preliminary results on biological
oxygen demand and sugars in runoff from sugarcane land. GEOTROP 2001: 4th
International Conference on Environmental Chemistry and Geochemistry in the Tropics,
Townsville, May 7-11.
Bohl, H. P., Mitchell, D. C., Penny, R. S. and Roth, C. H. (2000) Nitrogen losses by subsurface flow from
sugarcane on subsurface soil in the Australian Wet Tropics. Proceedings of the Australian
Society of Sugar Cane Technologists 22: 302-307.
Bohl, H. P., Roth, C. H., Tetzlaff, D. and Timmer, J. (2001) Estimation of groundwater recharge and
nitrogen leaching under sugarcane in Ripple Creek catchment, Lower Herbert. Proceedings
of the Australian Society of Sugar Cane Technologists 23: 84-89.
Boudou, A. and Ribeyre, F. (1989) Fundamental concepts in aquatic ecotoxicology. Aquatic
ecotoxicolgy: fundamental concepts and methodologies. A. Boudou and F. Ribeyre.
Forida, CRC Prss: 35-76.
Bourg, A. C. M. (1989) Adsorption of trace inorganic and organic contaminants by solid particulate
matter. Aquatic ecotoxicolgy: fundamental concepts and methodologies. A. Boudou and F.
Ribeyre. Forida, CRC Prss: 107-124.
Brady, G. B. A., Johns, R. B. and Smith, J. D. (1994) Trace metal geochemical association in sediments
from the Cairns Region of the Great Barrier Reef, Australia. Marine Pollution Bulletin 28:
230 - 234.
Bramley, R. a. J., AKL (1996) Land use impacts on nutrient loading in the Herbert River. In Downstream
effects of land use. H. Hunter, Eyles, AG. and Rayment, GE, Department of Natural
Resources, Queensland.
Bramley, R. G. V. and Muller, D. E. (1999) Water quality in the lower Herbert River - the CSIRO
dataset. CSIRO Land and Water
Bramley, R. G. V. and Wood, A. W. (1996) More precise P fertiliser recommendations to minimise
potential downstream impacts of the sugar industry. In Downstream effects of land use. H.
M. Hunter, A. G. Eyles and G. E. Rayment, Department of Natural Resources: 269-272.
Bramley, R. G. V., Edis, R. B., White, R. E., and Wood, A. W. (1998) Environmentally sound
phosphorous management for sugarcane soils. Sugar Research and Development
Corporation. SRDC Project No.CSS3S.
Bramley, R. G. V., Ellis, N., Nable, R. O. and Garside, A. L. (1996) Changes in soil chemical properties
under long-term sugar cane monoculture and their possible role in sugar yield decline.
Australian Journal of Soil Research 34: 967-984.
Bramley, R. G. V., Johnson, A. K. L. and Smettem, D. R. J. Quantifying the effects of different rural land
use on water quality in the catchment of the Herbert river - preliminary results.
Bramley, R. G. V., Johnson, A. K. L. and Wood, A. W. Nutrient loss in the Herbert River catchment and
strategies for its reduction.
Bramley, R. G. V., Smettem, K. R. J. and Johnson, A. K. L. (1994) Measurement and prediction of nonpoint source nutrient losses from a tropical catchment: preliminary results from the Herbert
river study. Soils '94. Proceedings of the Third Triennial Western Australian Soil Science
Conference, Broadwater Resort, Busselton, Western Australia, Australian Society of Soil
Science Inc. (WA Branch)
187
Brodie, J. (1991) Urban impact on the Great Barrier Reef. Land use patterns and nutrient loading of the
Great Barrier Reef Region. Proceedings of the workshop held at the James Cook
University of North Queensland, 17-18 November, 1990. D. Yellowlees. Townsville,
James Cook University: 18-26.
Brodie, J. E. (1995) Nutrients and eutrophication in the Australian marine environment. In: L.P. Zann
and Sutton, D. C. (Eds.), State of the Marine Environment Report for Australia. Technical
Annex No. 2, pp.1-29. GBRMPA, Townsville, Australia.
Brodie, J. E., Hicks, W. S., Richards, G. N. and Thomas, F. G. (1984) Residues related to agricultural
chemicals in the groundwaters of the Burdekin River delta, North Queensland.
Environmental Pollution (Series B) 8: 187-215.
Brodie, J., Christie, C., Devlin, M., Haynes, D., Morris, S., Ramsay, M., Waterhouse, J. and Yorkston, H.
(2001) Catchment management and the Great Barrier Reef. Water Science and Technology
43(9): 203-211.
Brodie, J., Hicks, WS., Richards, GN. and Thomas, FG (1984) Residues related to agricultural chemicals
in the groundwaters of the Burdekin River delta, North Queensland. Environmantal
Pollution (Series B) 8: 187-215.
Brodie, J., Steven, A. and Baer, M. (1997) The extent of the river plumes associated with cyclone Sadie
rainfall. In: A.D.L. Steven (Ed.), Cyclone Sadie flood plumes in the Great Barrier Reef
lagoon: composition and consequences. GBRMPA Workshop Series 22: 27-33.
GBRMPA, Townsville, Qld.
Brooks, K. N., Ffolliott, P. F., Gregersen, H. M. and DeBano, L. F. (1997) Hydrology and the
management of watersheds. Ames, Iowa, Iowa State University Press.
Brown, T. E., Morley, A. W., Sanderson, N. T. and Tait, R. D. (1983) Report of a large fish kill resulting
from natural acid water conditions in Australia. Journal of Fisheries Biology 22: 335-350.
Bruce, R. C. and Rayment, G. E. (1982) Analytical methods and interpretations used by the Agricultural
Chemistry Branch for soil and land use surveys. QDPI Bulletin QB 82004.
Brune, G. M. (1953) Trap efficiency of reservoirs. Trans. AGU 34(3):407-418.
Brunskill, G. J. (2001a) Summary of information about change in river sediment deposition in the Great
Barrier Reef lagoon and continental shelf of North Queensland. BRIG Forum: Water
Quality in the Burdekin Catchment.
Brunskill, G. J. (2001b) Summary of continental shelf research for the Herbert River sector of the Great
Barrier Reef Lagoon, North Queensland. Herbert River Integration Study 1st HRIS
technical workshop, October 2000. B. Musso: 14.
Bryan, G. W. (1976) Some aspects of heavy metal tolerance in aquatic organisms. Effects of pollutants
on aquatic organisms. A. P. M. Lockwood. Cambridge, Cambridge University Press: 7-34.
Bryce, S. (1999) A sediment transport model for mesotidal estuaries of north Queensland. North
Queensland Sedimentologists' 1st Annual Workshop - Extended Abstracts. A. D. Heap.
Townsville, James Cook University: 92-95.
Bryce, S., Larcombe, P. and Ridd, P. V. (1996) Sediment transport in the Normamby River estuary,
Northern Great Barrier Reef shelf, Australia. Great Barrier Reef: Terrigenous Sediment
Flux and Human Impacts. P. Larcombe, K. Woolfe and R. Purdon. Townsville, Earth
Science Department, James Cook University.
188
BSES
(2000)
Biopesticide breakthrough for
http://www.bses.org.au/news/Biopesticide.
Australian
sugar
industry.
BSES.
Bunn, S. E., Davies, P. M. and Kellaway, D. M. (1997) Contributions of sugar cane and invasive pasture
grass to the aquatic food web of a tropical lowland stream. Marine and Freshwater
Research 48:173-179.
Bunn, S. E., Davies, P. M. and Mosisch, T. D. (1999) Ecosystem measures of river health and their
response to riparian and catchment degradation. Freshwater Biology 41: 333-345.
Burdekin Project Assessment Committee (1978) Report on Burdekin River Project. Brisbane. 191pp.
Burdon-Jones, C. and Denton, G. (1984a) Metals in marine organisms from the Great Barrier Reef
Province, Part 1. Baseline Survey. Final report to the Australian Marine Science and
Technologies Grants Committee. 162.
Burdon-Jones, C. and Denton, G. (1984b) Metals in marine organisms from the Great Barrier Reef
Province, Part 2. Regional and seasonal variations. James Cook University, Townsville.
162.
Burdon-Jones, C., Denton, G. R. W., Jones, G. B. and Mcphie, K. A. (1975) Long-term sub-lethal effects
of metals on marine organisms. Part 1. Baseline survey. James Cook University,
Townsville. Progress report to Water Quality Council of Queensland. 105.
Burkholder, J. M., Jr, H. B. G. and Hobbs, C. W. (1995) Fish kills linked to a toxic ambush-predator
dinoflagellate: distribution and environmental conditions. Marine Ecology Progress Series
124: 43-61.
Burkholder, J. M., Mallin, M. A. and Jr., H. B. G. (1999) Fish kills, bottom-water hypoxia, and the toxic
Pfiesteria complex in the Neuse River and estuary. Marine Ecology Progress Series 179:
301-310.
Burrows, D. (2001) Changes to, and management of, water quality in the Burdekin River Catchment.
BRIG Forum: Water Quality in the Burdekin Catchment, pp.46-50.
Burrows, D. W., Butler, B. M. and Pearson, R. G. (2000) Aquatic ecology of the Townsville Field
Training Area – LWRRDC/ CSIRO Sustainable Use of Military Lands Program. ACTFR
Report No. 00/08.
Burrows, D. W., Kutt, A. S. and Kemp, J. E. (1998) Burdekin catchment study – a desktop study of
environmental issues associated with dam and irrigation area development. ACTFR Report
No. 98/19.
Butler, B. and Crees, M. (1997) Baseline stormwater monitoring. Australian Centre for Tropical
Freshwater Research, Townsville. ACTFR Report No. 97/06.
Butler, B., Birtles, A., Pearson, R. and Jones, K. (1996) Ecotourism, water quality and Wet Tropics
streams. Australian Centre for Tropical Freshwater Research, Townsville. ACTFR Report
No. 96/11.
Butler, G. (1977) Algae and pesticides. Residue Reviews 66: 19-62.
Calcino, D. V. (1994) Australian Sugarcane Nutrition Manual. Indooroopilly, Australia, BSES/SRDC.
Calcino, D., Kingston, G. and Haysom, M. (2000) Nutrition of the plant. Manual of Canegrowing. M.
Hogarth and P. Allsopp. Brisbane, Bureau of Sugar Experimental Stations: 153 - 194.
189
Camilleri, C., Markich, S., van Dam, R. and Pfeifle V. (1996) Toxicity of the herbicide tebuthiuron to
Australian tropical freshwater organisms. Supervising Scientist, Canberra, Australia.
Report 131.
Campbell, P. G. C. and Tessier, A. (1989) Geochemistry and bioavailability of trace metals in sediments.
Aquatic ecotoxicolgy: fundamental concepts and methodologies. A. Boudou and F.
Ribeyre. Forida, CRC Prss: 125-150.
Canegrowers (1999) Australian sugar industry handbook 1999, Canegrowers.
Canegrowers Sustainable cane growing in Queensland - Code of Practice. Canegrowers, Department of
Environment, CRC for Sustainable Sugar Production, BSES, Brisbane. 27.
Capelin, M. A. and Prove, B. G. (1983) Soil conservation problems of the humid coastal tropics of North
Queensland. Proceedings of the Australian Society of Sugarcane Technologists. B. T.
Egan. Brisbane, Watson Ferguson and Company: 87-93.
Carlin, G., Shaw, RJ., Brebber, LJ., Simpson, BW and Geritz AF (1996) A relocatable stream water
quality monitoring system. In Downstream effects of land use. H. Hunter, Eyles AG and
Rayment, GE, Department of Natural Resources, Queensland.
Carter, R. M., Johnson, D. P. and Hooper, K. G. (1993) Episodic post-glacial sea-level rise and
sedimentary evolution of a tropical embayment (Cleveland Bay, Great Barrier Reef shelf,
Australia) Australian Journal of Earth Sciences 40: 229-255.
Cavanagh, J., Burns, KA., Brunskill, GJ., Coventry, RJ., Ryan DAJ. and Ahokas, JT (1999) Induction of
hepatic cytochrome P-450 1A in Pikey Bream (Acanthopagrus berda) collected from
agricultural and urban catchments in far north Queensland.
Cavanagh, J.E., Burns, K.A., Brunkill, G.J. ans Coventry, R.J. (1999) Organochlorine pesticide residues
in soils and sediments of the Herbert and Burdekin River regions, north Queensland –
implications for contamination of the Great Barrier reef. Marine Pollution Bulletin 39:
367-375
Chapman, L. (1996) Australian sugar industry by-products recycle plant nutrients. In Downstream effects
of land use. H. Hunter, Eyles, AG and Rayment, GE, Department of Natural Resources:
257-260.
Chapman, L. S., Haysin, M. B. C. and Chardon, C. W. (1981) Checking the fertility of Queenslands sugar
land. Proceedings of the Australian Society of Sugar Cane Technologists 3: 325-332.
Chapman, L. S., Haysom, M. B. C. and Saffigna, P. G. (1992) N cycling in cane fields from 15N labelled
trash and residual fertiliser. Proceedings of the Australian Society of Sugar Cane
Technologists 14: 84-89.
Chapman, P. M. (1995) Ecotoxicolgy and pollution - key issues. Marine Pollution Bulletin 31(4-12):
167-177.
Chessman, B. C. (1995) Rapid assessment of rivers using macroinvertebrates: a procedure based on
habitat-specific sampling, family identification and a biotic index. Australian Journal of
Ecology 20: 122-129.
Chessman, B. C. (2003) New sensitivity grades for Australian river macroinvertebrates. Marine and
Freshwater Research 54: 95-103.
Chessman, B. C., Growns, J. E. and Kotlash, A. R. (1997) Objective derivation of macroinvertebrate
family sensitivity grade numbers for the SIGNAL biotic index: application to the Hunter
River system, New South Wales. Marine and Freshwater Research 48: 159-172.
190
Choy, S. a. T., P (1996) River bioassessment and monitoring using benthic macroinvertebrate community
structure: The monitoring river health initative. H. E. Hunter, AG and Rayment, GE,
Department of Natural Resources, Queensland.
Christiansen, I. (2000) Environmental management. Manual of Canegrowing. M. Hogarth and P. Allsopp.
Brisbane, Bureau of Sugar Experimental Stations: 27-69.
Chudek, E., Joo, M., Horn, A. and McLaren, S. (1998) Stratification behaviour of Queensland dams.
Water Monitoring Group, Dept. of Natural Resources, Queensland.
Clarke, G. M., Arthington, A.H and Pusey, B.J (1995) Fluctuating asymmetry of chironomid larvae as an
indicator of pesticide contamination in freshwater environments. Chironomids: From genes
to ecosystems. P. Cranston. Melbourne, Victoria, CSIRO Publications: 101-109.
Clarke, R. D. (1998) Water quality of the Burdekin River Catchment and proposed monitoring network.
Dept. of Natural Resources and Mining, April 1998.
Clarke, R. D. (2002) Trends in total nitrogen loads in streams discharging to the Great Barrier Reef.
Department of Natural Resources and Mines, Queensland.
Clayton, P. D., and Pearson, R.G. (1996) Instream effects of dunder application to canefields in the
Sarina region, Queensland: Continued monitoring. Australian Centre for Freshwater
Research, James Cook University, Townsville, Townsville. Final Report, 96/09.
Cogle, A. L., Langford, P. A., Kistle, S. E., Ryan, T. J., McDougall, A. E., Russell, D. J. and Best, E.
(2000) Natural resources of the Barron River catchment 2. Water quality, land use and
land management interactions. QI 00033 (Queensland Department of Primary Industries),
Brisbane.
Congdon, R. A. (1991) Effects of irrigation discharge on the Barrattas wetlands. Australian Centre for
Tropical Freshwater Research, James Cook University, Townsville, Townsville. ACTFR
Report, 91/06.
Congdon, R. A. and Lukacs, G. P. (1995) Limnology and classification of tropical floodplain wetlands,
with particular reference to the effects of irrigation drainage. LWRRDC R&D Project,
JCU/ ACTFR Report No. 95/12.
Congdon, R. A. and Lukacs, G. P. (1996) Water quality aspects of irrigation runoff from the Burdekin
River Irrigation Area. In Downstream effects of land use. H. E. Hunter, AG and Rayment,
GE, Department of Natural Resources, Queensland.
Connell, D. W. (1999) Introduction to ecotoxicology. Carlton, Victoria, Blackwell Science.
Connell, D. W. and Hawker, D. W. (1992) Pollution in Tropical Aquatic Systems. Boca Raton, Fla., CRC
Press.
Connell, D. W., Miller, G. J., Mortimer, M. R., Shaw, G. R. and Anderson, S. M. (1996) Persistent
organic pollutants in the southern hemisphere: executive summary. Department of
Environment, Sport and Territories, Environment Protection Agency, Environment
Standards
Branch,
Brisbane.
Consultancy
report.
www.chem.unep.ch/pops/indxhtms/manexp14.html.
Connelly, R. D., Kennedy, I. R., Silburn, D. M., Simpson, B. and Freebairn, I. R. (1998) Exploring farm
design and management options with modelling. Mnimising the impact of pesticides on the
riverine environment: key findings from research with the cotton industry - 1998
191
conference. N. J. Schofield and V. E. Edge. Canberra, LWRRDC. LWRRDC Occassional
Paper 23/98: 94-98.
Cook, F. J., Gardner, E. A. and Carlin, G. (1999) Acid drainage from canelands: quantities exported and
mechanism of production. Proceedings of Austalian Society of Sugar Cane Technologists
21: 246-258.
Cooper, B. and Riley, G. T. (1996) Storm transport of pollutants from dryland agriculture. In: H. M.
Hunter, Eyles, A. G. and Rayment, G. E. (Eds.), Downstream effects of land use. pp 173175, 1996. Dept. of Natural Resources, Queensland, Australia.
Cooperative Research Centre for Sustainable Sugar Production, Aitkenvale, Queensland. CSIRO Land
and Water Technical Report, 16/99.
Crossland, C. (1998) Sugar Production - Effects on the marine environment? Sugar Environment Forum Sustainable canegrowing into the next century, Mackay, CANEGROWERS.
Crossland, C.J., Done, T.J. and Brunskill, G.J. (1996) Potential impacts of sugarcane production on the
marine environment. In: B.A. Keating and J.R. Wilson (Eds), Intensive Sugarcane
Production, Meeting the Challenges Beyond 2000, pp. 423-436. (CAB, International.)
Davies, P. E. (1994) Effects of pesticides on freshwater invertebrate fauna. Land and Water Resources
Research, Canberra.
Davies, P. E. (1998) Session commentary: understanding the impact of pesticides in the riverine
environment. Mnimising the impact of pesticides on the riverine environment: key findings
from research with the cotton industry - 1998 conference. N. J. Schofield and V. E. Edge.
Davies, P. E., Cook, L. S. J. and Goenarso, D. (1994) Sublethal responses to pesticides of several species
of Australian freshwater fish and crustaceans and rainbow trout. Environmental Toxicology
and Chemistry 13(8): 1341-1354.
Davies-Colley, R. J. (1988) Measuring water clarity with a black disk. Limnology and Oceanography
33(4, Part 1): 616-623.
Dawson, H. (1997) Land use practices for wet tropical floodplains. Queensland Department of Natural
Resources, Brisbane. 42pp.
Dean, T. L. and Richardson, J. (1999) Response of seven species of native freshwater fish and a shrimp to
low levels of dissolved oxygen. New Zealand Journal of Marine and Freshwater Research
33: 99-106.
DEH & DNR (1999) Testing the Waters, a Report on the Quality of Queensland Waters. QDEH and
QDNR, Dept. of Natural Resources, Brisbane.
Denton, G. R. W. and Burdon-Jones, C. (1986) Trace metals in algae from the Great Barrier Reef.
Marine Pollution Bulletin 17(3): 98-107.
Department of Primary Industries (1995) An overview of water resources and other related issues – upper
Burdekin region. QDPI (Water Resources), Brisbane, Queensland.
Devlin, M. J., Lourey, M. J., and Sweatman, H. (1997) Water quality. In: Sweatman, H. (Ed.), LongTerm Monitoring of the Great Barrier Reef Status Report 2. Australian Institute of Marine
Science, Townsville.
192
Devlin, M., Waterhouse, J. and Brodie, J. (1999) Properties of flood plumes in the GBR lagoon. In: Heap,
A.D. (Ed.), North Queensland Sedimentologists’ 1st Annual Workshop: Extended
Abstracts. School of Earth Sciences, James Cook University, Townsville, pp. 32-36.
Devlin, M., Waterhouse, J., Christie, C., Haynes, D. and Brodie, J. (in prep.) Long term chlorophyll-a
monitoring in the Great Barrier Reef Lagoon: Status Report 2, 1993-1999.
Devlin, M., Waterhouse, J., Taylor, J. and Brodie, J. (2000) Flood plumes in the Great Barrier Reef:
Spatial and temporal patterns in composition and distribution. Research Publication No.
xx. Great Barrier Reef Marine Park Authority, Townsville.
Devlin, M., Waterhouse, J., Taylor, J. and Brodie, J. (2001) Flood plumes in the Great Barrier Reef:
spatial and temporal patterns in composition and distribution. Great Barrier Reef Marine
Park Authority, Townsville. GBRMPA Research Publication No. 68.
Doherty, G. B., Coomans, D. and Brunskill, G. J. (2000) Modelling natural and enhanced trace metal
concentrations in sediments of Cleveland Bay, Australia. Marine and Freshwater Research
51: 739 - 747.
Droppo, I. G. (2001) Rethinking what constitutes suspended sediment. Hydrological Processes 15:15511564.
Duke, N. C., Roelfsema, C., Tracey, D. and Godson, L. (2001) Preliminary investigation into dieback of
mangroves in the Mackay regioin: intial assessment and possible causes. Marine Botany,
The University of Queensland - Report to the Queensland Fisheries Service, Northern
Region and the Community of Mackay Region, Brisbane. 80.
EPA (1999a) Draft Queensland water quality guidelines Version 1.01. Scientific Asssessment Section.
Queensland Department of Environment.
EPA (1999b) Water quality sampling manual for use in testing for compliance with the Environmental
Protection Act 1994. Queensland Environmental Protection Agency, Brisbane.
Eyre, B. a. D., P (1996) A preliminary assessment of suspended sediment and nutrient concentrations in
three far north Queensland catchments. In Downstream effects of land use. H. E. Hunter,
AG and Rayment, GE, Department of Natural Resources, Queensland.
Eyre, B. and McConchie, D. (1993) Implications of sedimentological studies for environmental pollution
assessment and management: examples from fluvial systems in North Queensland and
Western Australia. Sedimentary Geology 85: 235-252.
Eyre, B. D. (1993) Nutrients in the sediments of a tropical North-eastern Australian estuary, catchment
and nearshore coastal zone. Australian Journal of Marine and Freshwater Research 44:
845-866.
Fabricius, K. E. and Wolanski, E. (2000) Rapid smothering of coral reef organisms by muddy marine
snow. Estuarine, Coastal and Shelf Science 50: 115-120.
Faithful, F. W. and Griffiths, D. J. (2000) Turbid flow through a tropical reservoir (Lake Dalrymple,
Queensland, Australia): responses to a summer storm event. Lakes and Reservoirs:
Research and Management 2000 5:231-247.
Faithful, J. W. (1996) The fate of phosphorus in wetlands: A review. Australian Centre for Tropical
Freshwater Research, Townsville, Townsville. ACTFR Reort No. 96/15.
193
Faithful, J. W. and Griffiths, D. J. (2000) Turbid flow through a tropical reservoir (Lake Dalrymple,
Queensland, Australia): responses to a summer storm event. Lakes and Reservoirs:
Research and Management 2000 5: 231-247.
Ferris, J. M. and R. Christian (1991) Aquatic primary production in relation to microalgal responses to
changing light. Aquatic Sciences 53(2/3): 187-217.
Fielding, C. R. and Alexander, J. (1996) Sedimentology of the upper Burdekin River of North
Queensland, Australia – an example of a tropical, variable discharge river. Terra Nova
8:447-457.
Fielding, C. R., Alexander, J. and Newman-Sutherland, E. (1997) Presentation of in-situ, aborescent
vegetation and fluvial bar construction in the Burdekin river of North Queensland,
Australia. Paleogeography, Paleoclimatology, Paleoecology 135:123-144.
Finlayson, B. a. S., M (1996) Soil, nutrient and pesticide movements from different land use practices,
and subsequent transport by rivers and streams. In Downstream effects of land use. H.
Hunter, Eyles, AG and Rayment GE, Department of Natural Resources, Queensland.
Finlayson, B. and Silburn, M. (1996) Soil, nutrient and pesticide movements from different land use
practices and subsequent transport by rivers and streams. In: H. M. Hunter, Eyles, A. G.
and Rayment, G. E. (Eds.), In Downstream effects of land use. pp 129-140, 1996. Dept. of
Natural Resources, Queensland, Australia.
Furnas, M. (1992) Behaviour of nutrients in tropical aquatic ecosystems. Pollution in tropical aquatic
systems. D. a. H. Connell, DW, CRC Press, Florida: 29-65.
Furnas, M. (2003) Catchments and corals: terrestrial runoff to the Great Barrier Reef. Townsville,
Australian Institute of Marine Science.
Furnas, M. and Mitchell, A. (2001) Runoff of terrestrial sediment and nutrients into the Great Barrier
Reef World Heritage Area. In: E. Wolanski (Ed.), Oceanographic Processes of Coral
Reefs, CRC Press 4:37-51.
Furnas, M. J. & Brodie, J. (1996) Current status of nutrient levels & other water quality parameters in the
Great Barrier Reef. In: Hunter, H, Eyles, A. & Rayment, G. (Eds.), In Downstream effects
of land use. Department of Natural Resources, Brisbane, Australia. pp. 9-23.
Furnas, M. J. and Mitchell, A. W. (1996) Biological oceanography of the Great Barrier Reef. The Great
Barrier Reef - science, use and management . Proceedings of a national conference,
Townsville 25-29 November 1996. Townsville, Reef CRC
Furnas, M., Mitchell, A. W. and Skuza, M. (1995) Nitrogen and phosphorus budgets for the Central
Great Barrier Reef. Great Barrier Reef Marine Park Authority, Townsville. GBRMPA
Research Publication, 36.
Gardner, E. a. B., GA (1996) Rural and urban waste products: Their generation, composition and re-use
on land. Downstream effects on land use. H. Hunter, Eyles, AG and Rayment, GE,
Department of Natural Resources: 243-251.
Gardner, T., Morton, D., Matthews, P., Bruce, A., Cook, F. and Clouston, S. (2000) Report on the
cropping cycle (milestone no.6) to Department of Communication and Information, Local
Government and Planning on the filter project. CIRM, Brisbane.
Garman, D. E. J. and Sutherland, P. D. (1983) Pesticides, fertilizers, heavy metals and salinity - an
assessment of the non-point source problems in Australia. Workshop on non-point sources
of pollution in Australia. Proceedings, Monash University, Melbourne, Australian
Government Publishing Service.
194
Garside, A., Magarey, R. and Braunack, M. (2000) Soil health. Manual of Canegrowing. M. Hogarth and
P. Allsopp. Brisbane, Bureau of Sugar Experimental Stations: 141-151.
Gaus, C., Brunskill, G. J., Weber, R., Papke, O. and M³ller, J. F. (2001) Historical PCDD inputs and their
source implications from dated sediment cores in Queensland (Australia) Environmental
Science and Technology 35(23): 4597-4603.
GBRWQAP (2001) Great Barrier Reef Water Quality Action Plan. Water Quality and Coastal
Development,
Great
Barrier
Reef
Marine
Park
Authority.
http://www.gbrmpa.gov.au/corp_site/key_issues/water_quality/action_plan/index.html.
GHD (1996) Canegrowers Environmental Audit of the Queensland Canegrowing Industry. Gutteridge
Haskins and Davey Pty Ltd, Brisbane. Environmental Audit. 199.
Gilmour, D. A. (1977) Effect of rainforest logging and clearing on water yield and quality in a high
rainfall zone of north east Queensland. Hydrology Symposium, Brisbane, Institute of
Engineers: Brisbane.
Gippel, C. J. (1995) Potential of turbidity monitoring for measuring the transport of suspended solids in
streams. Hydrological Processes 9: 83-97.
Great
Water
Quality
Action
Plan.
Barrier
Reef
Ministerial
Council
(2001)
http://www.gbrmpa.gov.au/corp_site/key_issues/water_quality/action_plan/pdf/gbr_water_
quality_action_plan)
Griffiths, D. J. and Faithful, J. W. (1996) Effects of the sediment load of a tropical north Australian river
on water column characteristics in the receiving impoundment. Arch. Hydrobiol. Suppl.
113 Large Rivers 10:147-157.
Grinter, S., Thomson, C. and Choy, S. (2000) An assessment of the ecological condition of the Burdekin
River using macroinvertebrates as bioindicators. Freshwater Biological Monitoring Report
No. 21, March 2000 Draft. Dept. of Natural Resources, Queensland.
Hamilton, D. & Haydon, G. (1996) Pesticides and fertilisers in the Queensland sugar industry -Estimates
of usage and likely environmental fate. Final report of project ICM-R&D.9401 to
Department of Primary Industries, Queensland.
Hamilton, D. and Haydon, G. (1997) Pesticides and Fertilisers in the New south Wales Sugar Industry Estimates of Usage and Likely Environmental Fate. Volume 1: Pesticides. Queensland
Department of Natural Resources and CRC for Sustainable Sugar Production, Brisbane.
Final Report.
Handlinger, J. (1996) Effects of algal blooms on fish. Fish kills: causes and investigation in Australia. D.
O'Sullivan. Lindisfarne, Tasmania, Aquaculture Sourcebook: 21-29.
Harding, J. S., Young, R. G., Hayes, J. W., shearer, D. A. and Stark, J. D. (1999) Changes in agricultural
intensity and river health along a river continuum. Freshwater Biology 42: 345-357.
Hargreaves, P. A., Simpson, B. W., Ruddle, L. J. and Packett, R. (1999) Persistence and fate of pesticides
in sugarcane soils. Proceedings of Australian Sugar Cane Technologists 21: 287-293.
Hart, B. T. (1982) Australian water criteria for heavy metals. Australian Water Resources, Australian
Government Publishing Service, Canberra. Technical Paper, 7.
Hart, B. T., Ed. (1986) Water quality management: freshwater ecotoxicology in Australia. Melbourne,
Water Studies Centre, Chisholm Institute of Technology.
195
Hart, B. T., Maher, B. and Lawrence, I. (1999) New generation water quality guidelines for ecosystem
protection. Freshwater Biology 41: 347-359.
Haynes, D. and Johnson, J. E. (2000) Organochlorine, heavy metal and polyaromatic hydrocarbon
pollutant concentrations in the Great Barrier Reef (Australia) Environment: a review.
Marine Pollution Bulletin 41(7-12): 267-278.
Haynes, D., Muller, J. and Carter, S. (2000) Pesticide and herbicide residues in sediments and seagrasses
from the Great Barrier Reef World Heritage Area and Queensland coast. Marine Pollution
Bulletin 41(7-12): 279-287.
Haynes, D., Ralph, P., Müller, J. F., Prange, J., Michalek-Wagner, K. and Waterhouse, J. (2000c) The
occurrence and impact of herbicides in the Great Barrier Reef, Australia. Reef Rersarch,
June-December 2000, pp.3-5.
Haynes, D., Ralph, P., Prange, J. and Dennison, B. (2000) The impact of the herbicide diuron on
photosynthesis in three species of tropical seagrass. Marine Pollution Bulletin 41:288-293
Heady, G. a. G., NR (1996) Streamlining An environmentally sustainable drainage network for the Swan
coastal plain (Peel-Harvey catchment, Western Australia) In Downstream effects of land
use. H. Hunter, Eyles, AG and Rayment, GE, Department of Natural Resources,
Queensland: 265-267.
Heinemann, H. G. (1981) A new sediment trap efficiency curve for small reservoirs. Water Resources
Bulletin (AWRA) 17(5):825-830.
Horne, A. J. and C. R. Goldman (1994) Limnology. 2nd Ed. New York, McGraw-Hill International.
Hunter, H. (1993) Fluxes of nitrogen, phosphorus and suspended solids in the Johnstone River system
April 1991-March 1993. Department of Primary Industries, Indooroopilly. Agricultural
Chemistry unpublished report, ACU93.19.
Hunter, H. (1997) From land to river to reef lagoon: land use impacts on water quality in the Johnstone
River catchment. The State of Queensland, Department of Natural Resources, Brisbane.
10pp.
Hunter, H. and Armour, J. D. (2001) Offsite movement of nutrients: contrasting issues at three
Australian study sites. Offsite movement of agrochemicals in troopical sugarcane
production, Extension Workshop, Bundaberg, Australia. 8-9 May 2001, QNRM, ACIAR,
MSIRI.
Hunter, H. M. (1992) Agricultural Contaminants in Aquatic Environments: A Review. Queensland
Department of Primary Industries, Brisbane. 53pp.
Hunter, H. M., Hargreaves, P. A. and Rayment, G. E. (1998) Pesticide residues in surface waters of the
Burdekin River Irrigation Area. Department of Natural Resources, Brisbane. Draft.
Hunter, H. M., Sologinkin, S. J., Choy, S. C., Hooper, A. R., Allen, W. S., Raymond, M. A. A. and
Peeters, J. (2001) Water management in the Johnstone Basin. Queensland Government
Publication, State of Queensland Department of Natural Resources and Mines, Brisbane.
105.
Hunter, H. M., Walton, R. S. and Russell, D. J. (1996) Contemporary water quality in the Johnstone
River Catchment. In Downstream effects of land use. H. M. Hunter, A. G. Eyles and G. E.
Rayment. Brisbane, Department of Natural Resources, Queensland, Australia: 339-345.
196
Hunter, K. A., Kim, J. P. and Reid, M. R. (1999) Factors influencing the inorganic speciation of trace
metal cations in fresh waters. Marine and Freshwater Research 50: 367-372.
Johnson, A. K. L. (1995) Risk perceptions and nutrient management responses in the Australian sugar
industry: preliminary results from the Herbert River district. Proceedings of Australian
Society of Sugar Cane Technologists(March): 7pp.
Johnson, A. K. L. (1996) Quantifying current and historical inputs of fertiliser N and P to rural lands in
the Herbert River catchment. In Downstream effects of land use. H. M. Hunter, A. G. Eyles
and G. E. Rayment. Brisbane, Queensland, Department of Natural Resources: 273-275.
Johnson, A. K. L. and Ebert, S. P. (2000) Quantifying inputs of pesticides to the Great Barrier Reef
Marine Park - a case study in the Herbert River Catchment of North-East Queensland.
Marine Pollution Bulletin 41(7-12): 302-309.
Johnson, A. K. L., Ebert, S. P. and Murray, A. E. (1997) Spatial and temporal distribution of wetland and
riparian zones in catchments adjacent to the Great Barrier Reef Marine Park. Proceedings,
Great Barrier Reef Marine Park Authority Cross Sectoral Workshop on Wetlands and
Water Quality, Babinda, Queensland, 25-26 September 1997.
Johnson, A. K. L., McDonald, G. T., Shrubsole, D. A. and Walker, D. H. (1998) Natural resource use and
management in the Australian sugar industry: current practice and opportunities for
improved policy, planning and management. Australian Journal of Environmental
Management 5: 97-108.
Johnson, A., Walker, D. and Wood, A. (1998) Spatial analysis of the impact of land use on nutrient loads
in the Herbert River catchment. CRC for Sustainable Sugar Production, Townsville. Sugar
research development corporation final report - project CSC13S. 90pp.
Jones, G. B. (1992) The efffect of Trichodesmium blooms on water quality in the Great Barrier Reef
Lagoon. Marine Pelagic Cyanobacteria: Trichodesmium and other Diazotrophs. Carpenter,
Capone and Reuter. Dordrecht, Kluwer Academic Press: 273-287.
Jones, G. B., Mercurio, P. and Olivier, F. (2000) Zinc in fish, crabs, oysters, and mangrove flora and
fauna from Cleveland Bay. Marine Pollution Bulletin 41(7-12): 345-352.
Jones, R. J., Muller, J. and Schreiber, U. (in prep.) The effects of the herbicides diuron and atrazine on
corals of the Great Barrier Reef (Australia)
Jury, W. A., Winer, A. M., Spencer, W. F. and Focht, D. D. (1987) Transport and transformations of
organic chemicals in the soil-air-water ecosystem. Reviews of Environmental
Contamination and Toxicology 99: 119.
Keating, B. A., Bauld, J., Hillier, J., Ellis, R., Weier, K. L., Sunners, F., and Connell, D. (1996) Leaching
of nutrients and pesticides to Queensland groundwaters. In Downstream effects of land use.
H. M. Hunter, Eyles, A. G., and Rayment, G. E. Brisbane, Department of Natural
Resources.
Keating, B. A., Bauld, J., Hillier, J., Ellis, R., Weier, K.L., Sunners, F. and Connell, D. (1996) Leaching
of nutrients and pesticides to Queensland ground waters. H. M. Hunter, Eyles, A. G. and
Rayment, G. E. (Eds.), In Downstream effects of land use. pp 151-163, 1996. Dept. of
Natural Resources, Queensland, Australia.
Keating, B. A., Vallis, I., Hughes, M. and Ridge, D. R. (1993) Strategic directions for nitrogen research a review from the south. Proceedings of the Australian Society of Sugar Cane
197
King, B., McAllister, F. and Done, T. (2000) Modelling the impacts of the Burdekin, Herbert, Tully and
Johnstone River plumes on the Central Great Barrier Reef. Final Report to the CRC Reef
Research Centre from Asia-Pacific Applied Science Associates and the Australian Institute
of Marine Science, Townsville.
King, B., McAllister, F., Wolanski, E., Done, T. and Spagnol, S. (2001) River plume dynamics in the
central Great Barrier Reef. pp. 145-159 In: Wolanski, E. (Ed), Oceanographic Processes of
Coral Reefs: Physical and Biological Links in the Great Barrier Reef. CRC Press, London.
Kinsey, D. W. (1991) Water quality and its effect on reef ecology. Land use patterns and nutrient loading
of the Great Barrier Reef Region. Proceedings of the workshop held at the James Cook
University of North Queensland, 17-18 November, 1990. D. Yellowlees. Townsville,
James Cook University: 192-196.
Kirk, J. T. O. (1981) Monte Carlo study of the nature of the underwater light field in, and the
relationships between optical properties of, turbid yellow waters. Australian Journal of
Marine and Freshwater Research 32(517-532)
Kirk, J. T. O. (1985) Effects of suspensoids (turbidity) on penetration of solar radiaition in aquatic
ecosystems. Hydrobiologia 125: 195-208.
Kirk, J. T. O. (1989) The upwelling light stream in natural waters. Limnology and Oceanography 34(8):
1410-1425.
Kirk, J. T. O. (1994a) Light and Photosynthesis in Aquatic Ecosystems. 2nd Ed. Cambridge, University
Press.
Kirk, J. T. O. (1994b) Characteristics of the light field in highly turbid waters: a Monte Carlo study.
Limnology and Oceanography 39(3): 702-706.
Kirk, J. T. O. and P. A. Tyler (1986) The spectral absorption and scattering properties of dissolved and
particulate components in relation to the underwater light field of some tropical Australian
freshwaters. Freshwater Biology 16: 573-583.
Klumpp, D. W. and Westernhagen, H. v. (1995) Biological effects of pollutants in Australian tropical
coastal waters: Embryonic malformations and chromosomal abberrations in developing
fish. Marine Pollution Bulletin 30(2): 158-168.
Kookana, R. S., Baskaran, S. and Naidu, R. (1998) Pesticide fate and behaviour in Australian soils in
relation to contamination and management of soil and water: a review. Australian Journal
of Soil Research 36: 715-764.
Kookana, R. S., Correll, R. L. and Simpson, B. (1998) Assessing risk at catchment or regional level
through a Pesticide Impact Ranking Index (PIRI) Minimising the impact of pesticides on
the riverine environment: key findings from research with the cotton industry - 1998
conference. N. J. Schofield and V. E. Edge. Canberra, LWRRDC. LWRRDC Occassional
Paper 23/98: 81-88.
Kookana, R. S., Kerekes, A., Oliver, D., Salama, R. B. and Byrne, J. D. (2000) Sorption of selected
pesticides and their metabolites in soil profiles of the Swan Coastal Plain. Agrochemical
Pollution of Water Resources, ACIAR Proceedings No. 104, Hat Yai, Thailand, Australian
Centre for International Agricultural Research.
Korth, W., Thomas, M., Foster, S., Napier, G., McCorkelle, G and Bowmer, KH (1996) Downstream
reduction in toxicity and pesticides in irrigated rice and maize drainage water: Abstract. In
Downstream effects of land use. H. Hunter, Eyles, AG and Rayment, GE, Department of
Natural Resources, Queensland.
198
Larcombe, P. and Taylor, J. (1997) Port of Mourilyan water and sediment quality monitoring baseline
surveys, January 1995-February 1996. James Cook University and the Ports Coorporation
of Queensland, Brisbane. EcoPorts Monograph Series, No. 7.
Larcombe, P. and Woolfe, K.J. (1999) Increased sediment supply to the Great Barrier Reef will not
increase sediment accumulation at most reefs. Coral Reefs 18: 163-169
Larcombe, P., Ridd, P.V. , Pritz, A. and Wilson, B. (1995) Factors controlling suspended sediment on
inner shelf reefs, Townsville, Australia. Coral Reefs 14: 163-171.
Larcombe, P., Woolfe, K. and Purdon, R.G. (Eds.) (1996) Great Barrier Reef: Terrigenous Sediment Flux
and Human Impacts. 2nd Edition, CRC Reef Research Centre, Townsville, 174 pp.
Lind, O. T. (1986) The effect of non-algal turbidity on the relationship of secchi depth to chlorophyll a.
Hydrobiologia 140(1): 27-35.
Lind, O. T. and Davalcos-Lind, L. (1991) Association of turbidity and organic carbon with bacterial
abundance and cell size in a large, turbid, tropical lake. Limnol. Oceanogr.36:1200-1208.
Lough, J. M., and Barnes, D. J. (1996) Centuries-long records of coral growth: a means to assess human
impacts on the GBR. In: P. Larcombe, Woolfe, K. and Purdon, R. (Eds.), Great Barrier
Reef: Terrigenous Sediment Flux and Human Impacts. James Cook University of North
Queensland.
Mueller, J. F., duquesne, S., Ng, J., Shaw, G. R., Krrishnamohan, K., Manonmanii, K., Hodge, M. and
Eaglesham, G. K. (2000) Pesticides in sediments from Queensland irrigation channels and
drains. Marine Pollution Bulletin 41(7-12): 294-301.
Mueller, J. F., Sutton, M., Wermuth, U. D., McLachlan, M. S., Will, S., Hawker, D. and Connell, D. W.
(1996) Polychlorinated dibenzodioxins and polychlorinated dibenzo-furans in topsoils
from northern Queensland, with a history of different trash management practices.
Sugarcane: research towards efficient and sustainable production. J. R. Wilson, D. M.
Hogarth, J. A. Campbell and A. L. Garside. Brisbane, CSIRO Division of Tropical Crops
and Pastures: 273-274.
Mueller, J. F., Sutton, M., Wermuth, U. D., McLachlan, M. S., Will, S., Hawker, D. W. and Connell, D.
W. (1996) Burning sugar cane - a potential for PAHs and PCDD/Fs. Intersect 96.
Mander, Kull, A. and Kuusemets, V. (2000) Nutrient flows and land use change in a rural catchment: a
modelling approach. Landscape Ecology 15: 187-199.
Mann, R. (1996) Phosphorous removal by constructed wetlands. In Downstream effects of land use. H.
Hunter, Eyles, AG and Rayment, GE, Department of Natural Resources, Queensland: 305307.
Markich, S. and Camilleri, C. (1997) Investigation of metal toxicity to tropical biota: recommendations
for revision of the Australian Water Quality Guidelines. Supervising Scientist, Canberra.
Supervising Scientist Report, 127. 87.
Marsh, H. (1989) Mass stranding of dugongs by a tropical cyclone in northern Australia. Marine
Mammal Science 5: 78-84.
Martinick Bosch and Sell (2002) Mt Gordon operations. Water management strategy and procedures for
containment, recycling and release. Martinick Bosch and Sell, Perth. Revised draft.
199
Max Winders & Associates (1998) Preliminary review of the 1998 fish kills, North Queensland. Max
Winders & Associates Pty Ltd Consulting Engineers and Environmental Scientists.
Prepared for Canegrowers. 24pp.
McAllister, F., King, B. and Done, T. (2000a) Burdekin River plume modelling.
McAllister, F., King, B. and Done, T. (2000b) The database of predicted river plume distributions (19661998) in the Central Great Barrier Reef Marine Park – CD ROM. Report to the CRC Reef
Research Centre from Asia-Pacific Applied Science Associates and the Australian Institute
of Marine Science, Townsville.
McCulloch, M. (2002) Coral record of Burdekin River run-off and water quality guidelines for the inner
Great Barrier Reef. Research School of Earth Sciences, ANU, Canberra.
McGregor, G. B. (2000) Algal management planning for state water project storages. Freshwater
Biological Monitoring Report March 2000. Dept. of Natural Resources, Queensland.
McKee, L. J. and Eyre, B. D. (2000) Nitrogen and phosphorus budgets for the sub-tropical Richmond
River catchment, Australia. Biogeochemistry 50(3): 207-239.
McLaughlin, M., Tiller, K.G., Naidu, R and Stevens, DP (1996) Review: the behaviour and
environmental impact of contaminants in fertilizers. Australian Journal of Soil Research
34: 1-54.
McMahon, G., Lawrence, P. and O'Grady, T. (2000) Weed control in sugarcane. Manual of
Canegrowing. M. Hogarth and P. Allsopp. Brisbane, Bureau of Sugar Experimental
Stations: 241-261.
McMahon, P. a. H., BT (1996) Water quality guidelines for the protection of aquatic ecosystems in
Australia. In Downstream effects of land use. H. E. Hunter, AG and Rayment, GE,
McNeil, V. H., Churchill, R. C., Joo, M., Everding, K. L., Clarke, R. D., Bloedel, L. and Horn, A. M.
(2000) Evaluation of the surface water quality of the Haughton-Barratta wetland system.
Water Monitoring Group, Dept. of Natural Resources, Queensland.
Mitchell, A. W. and Furnas, M. J. (1994) River inputs of N and P to the central Great Barrier Reef (16? 19?S) Proceedings of Australian Society of Sugar Cane Technologists.
Mitchell, A. W. and Furnas, M. J. (1996a) Differences in suspended sediment flux to the GBR shelf
between wet and dry catchments of North Queensland. In: P. Larcombe, Woolfe, K. and
Purdon, R. (Eds.), Great Barrier Reef: Terrigenous Sediment Flux and Human Impacts.
James Cook University of North Queensland.
Mitchell, A. W. and Furnas, M. J. (1996b) Terrestrial inputs of nutrients and suspended sediments to the
GBR lagoon. The Great Barrier Reef - science, use and management . Proceedings of a
national conference, Townsville 25-29 November 1996. Townsville, Reef CRC
Mitchell, A. W., Bramley, R. G. V. and Johnson, A. K. L. (1997) Export of nutrients and suspended
sediment during a cyclone-mediated flood event in the Herbert River catchment, Australia.
Marine and Freshwater Research 48: 79-88.
Mitchell, A. W., Reghenzani, J. R., Hunter, H. M. and Bramley, R. G. V. (1996) Water quality and
nutrient fluxes from river systems draining to the GBR Marine Park. H. M. Hunter, Eyles,
A. G. and Rayment, G. E. (Eds.), In Downstream effects of land use. pp. 23-33, 1996.
Dept. of Natural Resources, Queensland, Australia.
200
Mitchell, A., Rasmussen, C., Blake, S., Congdon, R., Reghenzani, J., Saffigna, P., and Sturmey, H.
(1991) Nutrient concentrations and fluxes in North Queensland coastal rivers and streams.
In: D. Yellowlees (Ed.), Land Use Patterns and Nutrient Loading of the Great Barrier Reef
Region. pp.108-161. (Sir George Fisher Centre for Tropical Marine Studies, James Cook
University, Townsville, North Queensland.)
Mitchell, A., Reghenzani, JR. Hunter, HM. and Bramley, RGV (1996) Water quality and nutrient fluxes
from river systems draining to the Great Barrier Reef Marine Park. Downstream effects of
landuse. H. Hunter, Eyles, AG. and Rayment, GE, Department of Natural Resources,
Queensland.
Mitchell, A.W. and Furnas, M.J. (1994) River inputs of N and P to the Central Great Barrier Reef. Proc.
Aust. Soc. Sugar Cane Technol. 16:147-153.
Mitchell, A.W. and Furnas, M.J. (1997) Terrestrial inputs of nutrients and suspended sediments to the
GBR lagoon. In: The Great Barrier Reef - Science, Use and Management. Proceedings of a
National Conference, Townsville 25-29 November 1996, 1:59-71. (Reef CRC, GBRMPA
and James Cook University, Townsville.)
Moody, P. a. C., L (1990) Phosphorus and Nitrogen in the soils of north Queensland. Land use patterns
and nutrient loading of the Great Barrier Reef region, James Cook University, Townsville,
James Cook University of North Queensland.
Moody, P. a. P., C (1996) Phosphorous dynamics of sediments from the Johnstone Rivers. In
Downstream effects of land use. H. Hunter, Eyles, AG and Rayment, GE, Department of
Natural Resources: 359-365.
Moody, P. W. and Aitken, R. L. (1997) Soil acidification under tropical agricultural systems. 1. Rates of
acidification and contributing factors. Australian Journal of Soil Research 35: 163-173.
Moody, P., Prove, BG. and Reghenzani, J (1997) DAQ3S Nutrient Balances and Transport from
Agricultural and Rainforest Lands: A case studt in the Johnstone River Catchment. Sugar
Research and Development Corporation.
Moody, P., Reghenzani, JR., Armour, JD., Prove, BG. and McShane, TJ (1996) Nutrient balances and
transport at farm scale - Johnstone river catchment. In Downstream effects of land use. H.
Hunter, Eyles, AG and Rayment, GE, Department of Natural Resources, Queensland: 347351.
Mortimer, M. (1998) Organochlorine pesticide residues in Brisbane waterways. Moreton Bay and
Catchment. I. R. Tibbetts, N. J. Hall and W. C. Dennison. Brisbane, Shcool of Marine
Science, The University of Queensland: 241-246.
Mortimer, M. R. Pesticide trace metal concentrations in Queensland estuarine crabs.
Mortimer, M.R. (2000) Pesticide and trace metal concentrations in Queensland estuarine crabs. Marine
Pollution Bulletin 41: 359-366.
Moss, A. and Bennett, J. (1991) Focus on some North Queensland water quality issues. Water December:
16-19.
Moss, A. and Costanzo, S. (1998) Levels of heavy metals in the sediments of Queensland rivers, estuaries
and coastal waters. Department of Environment, Queensland, Brisbane. Environmental
technical report No. 20 - ISSN 1037-4671. 20.
Moss, A. J., Rayment, G. E., Reilly, N. and Best, E. K. (1993) Sediment and nutrient exports from
Queensland coastal environments. Department of Environment, Queensland, Brisbane.
Environment Technical Report No. 5.
201
Moss, A., Bennett, J., Poplawski, W., Shaw, R. and Moller, G. (1996) Land use factors affecting the
condition of rivers, estuaries and bays in southern Queensland. In Downstream effects of
land use. H. Hunter, Eyles, AG. and Rayment, GE, Department of Natural Resources,
Queensland.
Moss, A.J., Rayment, G.E., Reilly, N. and Best, E.K. (1992) A preliminary assessment of sediment and
nutrient exports from Queensland coastal catchments. Qld Department of Environment &
Heritage Technical Report No. 4, Brisbane. 33pp.
Müller, J.F., Duquesne, S., Ng, J., Shaw, G.R., Krrishnamohan, K., Manonmannii, K., Hodge, M. and
Eaglesham, G.K. (2000) Pesticides in sediments from Queensland irrigation channels and
drains. Marine Pollution Bulletin 41: 294-301.
Muschal, M. and Cooper, B. (1998) Regional level monitoring of pesticides and their behaviour in rivers.
Mnimising the impact of pesticides on the riverine environment: key findings from
research with the cotton industry - 1998 conference. N. J. Schofield and V. E. Edge.
Neil, D. T. (1996) Spatial and temporal variation of fluvial sediments in Rockingham Bay, North
Queensland. CRC Reef Research Centre, James Cook University, Townsville. 60-66.
Neil, D.T. and Yu, B. (1996) Fluviatile sediment yield to the Great Barrier Reef lagoon: spatial patterns
and the effect of land use. pp.281-286 H.M. Hunter, A.G.Eyles & G.E.Rayment (Eds.), In
Downstream effects of land use. Department of Natural Resources, Brisbane, Australia.
Neil, D.T. and Yu, B. (1996) Simple climate-driven models for estimating sediment input to the GBR
lagoon. In: P. Larcombe, Woolfe, K. and Purdon, R. (Eds.), Great Barrier Reef:
Terrigenous Sediment Flux and Human Impacts. James Cook University of North
Queensland.
Neilson, A. (1994) Organic chemicals in the aquatic environment - Distribution, persistence and toxicity,
Lewis Publishers, CRC Press, Florida.
Nelson, P. N., Noble, A. D., Bramley, R. G. V. and Schroeder, B. L. (2000) Why do soils behave as they
do? Sustainable nutrient management in sugarcane production short course. R. C. Bruce.
Mercure Inn, Townsville, Cooperative Research Centre for Sustainable Sugar Production:
27-42.
NHMRC (2001) Atrazine (revised 2001) Fact sheets - pesticides Fact Sheet No. 103: 3.
NHMRC and ARMCANZ (1996) Australian drinking water guidelines. National Health and Medical
Research Council, Agriculture and Resource Mangement Council of Australia and New
Zealand, ACT.
NLWRA
(2001)
National Land and
http://www.nlwra.gov.au/.
Water
Resource
Audit.
National
Heritage
Trust.
Noble, R. and Collins, C. (2000) Physical and Chemical Water Quality of the Dawson River. River health
in the Fitzroy Catchment - Community Ownership, January 1997 - December 1999. R. M.
Noble. Brisbane, Queensland Department of Natural Resources: 25-39.
Noble, R. M., Duivenvoorden, L. J., Rummenie, S. K., Long, P. E. and Fabbro, L. D. (1997) In
Downstream effects of land use in the Fitzroy Catchment. Queensland Department of
Natural Resources, Brisbane. Summary report. 97.
202
NOHSC (1994) National Occupational Health and Safety Commission - List of designated hazardous
substances [NOHSC: 10005 (1994)]. Canberra, Australian Government Publishing
Service.
Nowak, B. (1991) Toxicity of organic chemicals. 128-134.
Nowak, B. (1993) Deformities and associated sublethal effects in fish exposed to sewage-borne
contaminants literature review. Centre for Environmental Toxicology. EPA 93/72. 41.
Nowak, B. F. (1997) Biological significance of xenobiotic residues in fish. TEN 4(5): 148-154.
Nowak, B. F., Deavin, J. G., Sarjito and Munday, B. L. (1992) Scanning electron microscopy in aquatic
toxicology. Journal of Computer-Assisted Microscopy 4(3): 241-246.
Nowak, B., Clarke, G., Pettigrove, V., Hughes, J. and Madden, C. Biological indicators and
biomonitoring of water quality in agricultural areas. 83-91.
NRA (1997) NRA Requirements and Guidelines for registering agricultural & veterinary chemicals.
National
Registration
Authority,
ACT.
http://www.nra.gov.au/guidelines/subpage_guidelines.shtml.
NRA (1997) Limnology Survey Final Report, Dry Season 1997. Natural Rersource Assessments,
Brisbane.
NRA (2001) Western Metals Mt Gordon Operations 2001 Dry Season Limnology. Natural Resource
Assessments Environmental Consultants. Report for Western Metals Mt Gordon
Operations.
OECD (1986) Water pollution by fertilizers and pesticides, Organisation for economic co-operation and
development.
Olafson, R. W. (1978) Effect of agricultural activity on levels of organochlorine pesticides in hard corals,
fish and molluscs from the Great Barrier Reef. Marine Environmental Research 1: 87-107.
Olszowy, H., Torr, P. and Imray, P. (1995) Trace element concentration in soils from rural and urban
areas of Australia. Contaminated Sites Monograph Series No. 4. South Australian Health
Commission, Adelaide.
Orpin, A. R. and Ridd, P. V. (1996) Sediment distribution and transport mechanisms, Burdekin region,
central Great Barrier Reef lagoon. In: P. Larcombe, Woolfe, K. and Purdon, R. (Eds.),
Great Barrier Reef: Terrigenous Sediment Flux and Human Impacts. James Cook
University of North Queensland.
Pailles, C. M., D.M., Arakel, A.V. and Saenger, P (1993) The distribution of phosphate in sediments of
the Johnstone Rivers catchment-estuary system, North Queensland, Australia. Sedimentary
Geology 85: 253-269.
Pailles, C., and Moody, P. W. (1996) Assessment of the downstream impact of sediment bound
phosphorous from southern canelands. Queensland Department of Natural Resources.
SIRP Project 3, Final Report.
Patra, R., Chapman, J. C., Lim, R. P. and Gehrke, P. C. (1996) Effects of temperature on the toxicity of
endosulfan, chlorpyrifos and phenol to Australian Ceriodaphnia dubia. Poster abstract
0112. Proceedings InterSect' 96, International symposium on environmental chemistry and
toxicology, Sydney, 14-18 July 1996.
Pearson, R. G. and Connolly, N. M. (2000) Nutrient enhancement, food quality and community dynamics
in a tropical rainforest stream. Freshwater Biology 43: 31-42.
203
Pearson, R. G. and Penridge, L. K. (1987) The effects of pollution by organic sugar mill effluent on the
macro-invertebrates of a stream in tropical Queensland, Australia. Journal of
Environmental Management 24: 205-215.
Pearson, R. G. and Penridge, L. K. (1992) An ecological survey of selected rivers in Queensland - with
particular reference to the effects of sugar mill effluents. James Cook University,
Townsville.
Pearson, R. G., and Connolly, N. M. (1998) Macroinvertebrates as monitors of river health in the tropics.
Australian Centre for Tropical Freshwater Research, James Cook University, Townsville,
Townsville. ACTFR Final Report, 98/20.
Pearson, R. G., M. Crossland, B. Butler and S. Manwaring (2003) Effects of cane-field drainage on the
ecology of tropical waterways. Report prepared for SRDC ACTFR Report No. 03/04.
Pearson, R.G. and Penridge, L.K. (1987) The effects of pollution by organic sugar mill effluent on the
Journal of
macro-invertebrates of a stream in tropical Queensland, Australia.
Environmental Management 24:205-215.
Pearson, R.G. and Penridge, L.K. (1992) An ecological survey of selected rivers in Queensland, with
particular reference to the effects of sugar mill effluents. p.150. Australian Centre for
Tropical Freshwater Research/ Dept. of Zoology Report No. 92/02. (JCU, Townsville)
Peters, E. C., Gassman, N. J., Firman, J. C., Richmond, R. H., and Power, E. A. (1997) Ecotoxicology of
tropical marine ecosystems. Environmental Toxicology and Chemistry 16(1): 12-40.
Powell, B. and Ahern, C. R. (1999) QASSMAC Acid sulfate soils management strategy for Queensland.
QASSMAC and Queensland Department of Natural Resources, Indooroopilly,
Queensland.
Powell, B. and Ahern, C. R. (1999) Nature, origin and distribution of acid sulfate soils: issues for
Queensland. Acid sulfate soils and their management in coastal Queensland: forum and
technical papers, Brisbane, Queensland Department of Natural Resources.
Power, M. E. (1997) Perspectives of the scientific community on the status of ecological risk assessment.
Prange, J. A. and Dennison, W. C. (2000) Physiological responses of five seagrass species to trace
metals. Marine Pollution Bulletin 41(7-12): 327-336.
Pringle, A. W. (1991) Fluvial sediment supply to the north-east Queensland coast, Australia. Australian
Geographical Studies 29: 114-138.
Prosser, I. P., Moran, C. J., Lu, H., Pickup, G., Gallant, J., Young, W. J., Hughes, A. O. and Rustomji, P.
(2002) Sediment and nutrient budgets for rivers draining to the Great Barrier Reef. CSIRO
Land and Water Technical Report xx (in press)
Prosser, I. P., Moran, C. J., Lu, H., Scott, A., Rustomji, P., Stevenson, J., Priestley, G., Roth, C. H. and
Post, D. (2002) Regional patterns of erosion and sediment transport in the Burdekin River
Catchment. Final report prepared for MLA by CSIRO Land and Water.
Prove, B. G. (1988) Soil erosion research in cane fields on the wet tropical coast of north east
Queensland. Workshop on nutrients in the Great Barrier Reef region. C. L. Baldwin.
Townsville, GBRMPA: 30-32.
Prove, B. G., Moody, P. W. and Reghenzani, J. R. (1997) DAQ3S Nutrient balances and transport from
agricultural and rainforested lands. Brisbane. 1997 Final Report to SRDC.
204
Prove, B., Doogan, V. J. and Truong, P. N. V. (1995) Nature and magnitude of soil erosion in sugarcane
land on the wet tropical coast of north-eastern Queensland. Australian Journal of
Experimental Agriculture 35: 641-649.
Pulsford, J. S. (1991) Historical inputs of fertiliser nutrients on to agricultural lands of coastal north
Queensland. Land use patterns and nurient loading of the Great Barrier Reef region, James
Cook University, Townsville, James Cook University of North Queensland.
Pulsford, J. S. (1996) Historical nutrient usage in coastal Queensland river catchments adjacent to the
Great Barrier Reef Marine Park. Great Barrier Reef Marine Park Authority, Townsville.
Research Publication, 40. 63pp.
Pulsford, J. S. (1998) Compilation of historic N and P fertilizer usage for sugar production for all of
coastal Queensland on a catchment basis. June 1998, unpublished data report.
QDPI (1999) Registration and reviews of agricultural and veterinary chemicals, Jenny Bibo. 2000.
QDPI (2000) Application of agricultural and veterinary chemicals. Animal and Plant Health Service,
Queensland
Department
of
Primary
Industries.
http://www.dpi.gov.au/aphs/agvet/application.htm.
Queensland Water Quality Task Force (1992) Freshwater algal blooms in Queensland. Brisbane, Qld.
Volume I Main Report. 104pp.
Qureshi, M. E., Wegener, M. K. and Mason, F. M. (2001) Economics of sugar mill by-products as a
source of nutrients in the Australian sugar industry in Mackay. Proceedings of the 2001
Conference of the Australian Society of Sugar Cane Technologists held at Mackay,
Queensland, Australia, 1st-4th May 2001. D. M. Hogarth, PK Editorial Services Pty Ltd,
Brisbane, Australia: 192-198.
Radcliffe, J. C. (2002) Pesticide use in Australia. Australian Academy of Technological Sciences and
Engineering, Victoria. 310
Ragnarsdottir, K. V. (2000) Environmental fate and toxicology of organophosphate pesticides. Journal of
the Geological Society 157(4): 859-876.
Raisin, G. a. M., DS (1996) Diffuse pollution and the use of wetlands for ameliorating water quality in
the Australian context. In Downstream effects of land use. H. Hunter, Eyles, AG and
Rayment, GE, Department of Natural Resources, Queensland.
Ralph, P. (2000) Herbicide toxicity of Halophila ovalis using chlorophyll a fluorescence. Aquatic Botany
66: 141-152.
Rao, P. S. C., Mansell, R. S., Baldwin, L. B. and Laurent, M. F. (1998) Pesticides and their behaviour in
soil and water. Institute of Food and Agricultural Sciences
Rasiah, V. and Armour, J. (1999) Potentially transportable nitrate-N in fluctuatingground water under
sugarcane in the Wet Tropics. Department of Natural Resources, Centre for Wet Tropics
Agriculture, Sth. Johnstone. unpublished report.
Rayment, G. E. (1990) Alienation of nutrient/soil from agricultural lands. Land use patterns and nutrient
loading of the Great Barrier Reef region, James cook University, Townsville, James Cook
Rayment, G. E. (1998) Background and possible scenarios for fish mortalities (1998) in waterways
draining canelands. CRC for Sustainable Sugar Production, Brisbane. Draft briefing paper.
2pp.
205
Rayment, G. E. (1999) Water quality pressures and status in sugar catchments. Cooperative Research
Centre for Sustainable Sugar Production. Exposure Draft. 19.
Rayment, G. E. (2002a) Land use and surface water quality in sugar catchments. In: K. Bristow (Ed.),
CRC Water Issues Paper. 10pp. (CRC Sugar, in preparation)
Rayment, G. E. (2002b) Water quality pressures and status in sugar catchments, Exposure Draft +. CRC
for Sustainable Sugar Production, Brisbane. 27.
Rayment, G. E. (2002c) Water quality pressures and status in Australian sugar catchments - a science
based review and assessment. CRC for Sustainable Sugar Production. Powerpoint
presentation.
Rayment, G. E. and Bohl, H. (2002) Dissolved oxygen in waterways of sugar catchments. In: R. C.
Bruce (Ed.), Managing Soils, Nutrients and the Environment for Sustainable Sugarcane
Production, Course Manual. pp. 15-23. (CRC Sugar, Townsville)
Rayment, G. E. and Higginson, F. R. (1992) Electrical conductivity/total soluble salts. Australian
Laboratory Handbook of Soil and Water Chemical Methods. Melbourne, Inkata Press:
217-223.
Rayment, G. E. and Neil, D. T. (1996) Sources of material in river discharge. The Great Barrier Reef science, use and management. Proceedings of a National Conference, Townsville 25-29
November 1996, Townsville, Reef CRC, GBRMPA, JCU.
Rayment, G. E., Jeffrey, A, J., and Barry, G. A. (1998) Heavy metals in New South Wales canelands.
Proceedings of Australian Society of Sugar Cane Technologists 20: 63-68.
Rayment, G. E., Jeffrey, A. J., Barry, G. A. and Baker, D. E. (1998) The environmental and productive
significance of nutrients in Queensland canelands. National Soils Conference,
Environmental Benefits of Soil Management, Conference Proceedings, Brisbane,
Australian Society of Soil Science.
Rayment, G. E., Jeffrey, A. J., Barry, G. A. and Chapman, L. (1997) Heavy metals in southern and
central Queensland canelands. Proceedings of Australian Society of Sugar Cane
Rayment, G. E., Moss, A., Mortimer, M., and Haynes, D. (1997) Monitoring of Pesticides in
Queensland's Aquatic Environment and in the Great Barrier Reef Marine Park. CRC for
Sustainable Sugar Production.
Rayment, G., Kuhn, G. and Pulsford, JS (1996) Fertiliser options and improved nutrient management
practices. In Downstream effects of land use. H. Hunter, Eyles, AG and Rayment, GE,
Rayment, G.E. and Neil, D.T. (1996) Sources of material in river discharge. pp.42-58 In: The Great
Barrier Reef: Science, Use and Management: Proceedings 1: Invited Papers. Great Barrier
Reef Marine Park Authority, Townsville.
Raymont, J.E.G (1980) Plankton and Productivity in the Oceans. Vol. 1 Phytoplankton. 2nd Ed.
Pergamon Press, Oxford
Reghenzani, J. (2001) Background information on productivity and environmental implications of surface
drainage in Ripple Creek catchment. Recommendations to enhance environmental
performance of drainage works proposed in the Lower Herbert Water Management
Scheme. C. H. Roth, D. Burrows, B. Butler, J. Reghenzani and D. Post. Townsville,
206
CSIRO Land and Water, Australian Centre for Tropical Freshwater Research, BSES.
Appendix D: 66 - 74.
Reghenzani, J. R., Armour, J. D., Prove, B. J., Moody, P. W. and McShane, T. (1996) Nutrient balances
for sugarcane plant and first ratoon crops in the wet tropics. Sugarcane: Research Towards
Efficient and Sustainable Production. J. R. Wilson, D. M. Hogarth, J. A. Campbell and A.
L. Garside. Brisbane, CSIRO Division of Tropical Crops and Pastures: 275-277.
Reichelt, A. J. and Jones, G. B. (1993) Characterization and fate of suspended sediments associated with
dredging activities in cleveland Bay. Townsville. Unpublished report to the Great Barrier
Reef Marine Park Authority.
Reichelt, A. J. and Jones, G. B. (1994) Trace metals as tracers of dredging activity in Cleveland Bay field and laboratory studies. Australian Journal of Marine and Freshwater Research 45:
1237 - 1257.
Reichelt-Brushett, A. J. and Harrison, P. L. (1999) The effect of copper, zinc and cadmium on
fertilization success of gametes from scleractinian reef corals. Marine Pollution Bulletin
38: 182-187.
Reichelt-Brushett, A. J. and Harrison, P. L. (2000) The effect of copper on the settlement success of
larvae from the scleractinian coral Acropora tenuis. Marine Pollution Bulletin 41: 385-391.
Robertson, F. A. and Thorburn, P. J. Trash management - consequences for soil carbon and nitrogen.
Rodi, W., Pavlovic, R. N. and Srivatsa, S. K. (1981) Prediction of flow and pollutant spreading in rivers.
Transport models for inland and coastal waters. Proceedings of a symposium on predictive
ability. H. B. Fischer. New York, Academic Press: 63-111.
Roth, C. H., Burrows, D., Butler, B., Reghenzani, J. and Post, D. (2001) Recommendations to enhance
environmental performance of drainage works proposed in the Lower Herbert Water
Management Scheme. CSIRO Land and Water, Australian Centre for Tropical Freshwater
Research, BSES, Townsville. Consultancy report to EPA and NRM.
Russell, D. J. and Hales, P. W. (1993) Downstream effects of agricultural chemicals on aquatic biota The Johnstone River, a case study, 15th Federal Convention of the Australian Water and
Wastewater Association, Gold Coast, Australia.
Russell, D. J., Hales, P. W. and Helmke, S. A. (1996) Stream habitat and fish resources in the Russell and
Mulgrave Rivers catchment. Northern Fisheries Centre, DPI, Cairns. 52.
Russell, D. J., Hales, P. W. and Moss, A. (1996) Pesticide residues in aquatic biota from North-East
Queensland coastal streams. Proceedings of the Royal Society of Queensland 106(1): 2330.
Sallaway, M. M. (1979) Soil erosion studies in the Mackay district. Proceedings of the Australian Society
of Sugarcane Technologists. B. T. Egan. Brisbane, Watson Ferguson and Company: 125132.
Sallaway, M. M. (1980) Effect on soil movement of continued cultivation into the wet season.
Proceedings of the Australian Society of Sugarcane Technologists. B. T. Egan. Brisbane,
Watson Ferguson and Company: 51-54.
Sallaway, M., Koon, D. A. and Cheong, R. N. (2001) Processes driving agrochemical movement Workshop Paper 4. Extension Workshop - Offsite movement of agrochemicals in tropical
sugarcane production, Bundaberg, Australia, 8-9 May 2001, MSIRI, ACIAR, Qld Dept. of
Natural Resources and Mines.
207
Sammut, J., Calliinan, R. B. and Dove, M. (1999) A brief review of the aquatic impacts of acid sulfate
soils. Acid sulfate soils and their management in coastal Queensland: forum and technical
papers, Brisbane, Queensland Department of Natural Resources.
Samson, P. R., Allsop, P. G., Brodie, A. J., Chandler, K. J., and Robertson, L. N. (1998) Pesticides under
pressure - options for sustainable pest management. Proc. Aust. Soc. Sugar Cane Technol.
20: 90-96.
Samson, P. R., Milner, R. J. and Bullard, G. K. (1999) Development of Metarhizium-based biopesticides
for sugarcane pest management - current progress and future prospects. Proceedings of
Australian Society of Sugar Cane Technologists 21: 156-163.
Schnoor, J. L. (1996) Environmental modelling - fate and transport of pollutants in water, air, and soil.
New York, John-Wiley and Sons.
Schofield, N. a. E., VE (eds) (1998) Minimising the impact of pesticides on the riverine environment: key
findings from research with the cotton industry, Canberra, Australia, Land and Water
Resources Research and Development Corporation.
Semple, P. and Williams, R. L. (1998) Metals in sediment data summary – Queensland, 1970 to 1992.
Environment Technical Report No.10. Dept. of Environment and Heritage, Qld
Government, Brisbane.
Shannon, B. J. (1987) Burdekin Falls Dam. Water Power and Dam Construction. December 1987,
pp.35-42.
Simpson, B. W. (2001) Pesticides - Australia. Extension Workshop - Offsite movement of agrochemicals
in tropical sugarcane production. Bundaberg, Australia, 8-9 May 2001, MSIRI, ACIAR,
Qld Dept. of Natural Resources and Mines: 61.
Simpson, B. W., Calcino, D. and Haydon, G. (2001) Pesticide and fertiliser use in the Australian sugar
industry - Workshop Paper 3.2. Extension Workshop - Offsite movement of agrochemicals
in tropical sugarcane production, Bundaberg, Australia, 8-9 May 2001, MSIRI, ACIAR,
Qld Dept. of Natural Resources and Mines.
Simpson, B. W., Calcino, D. and Pillay, K. P. (2001) Managing environmental risks - Workshop Paper 6.
Extension Workshop - Offsite movement of agrochemicals in tropical sugarcane
production, Bundaberg, Australia, 8-9 May 2001, MSIRI, ACIAR, Qld Dept. of Natural
Resources and Mines.
Simpson, B. W., Fraser, G., Armour, J., Hargreaves, P. and Ruddle, L. (2001) Pesticide studies Australia - Workshop Paper 5.4. Extension Workshop - Offsite movement of
agrochemicals in tropical sugarcane production, Bundaberg, Australia, 8-9 May 2001,
MSIRI, ACIAR, Qld Dept. of Natural Resources and Mines.
Simpson, B. W., Hargreaves, P. A. and Ruddle, L. J. (2002) Responsible pesticide use. Manageing soils,
nutrients and the environment for sustainable sugar production. R. C. Bruce. Townsville,
Cooperative Research Centre for Sustainable Sugar Production: 137-141.
Simpson, B. W., Kwong, R. N. K., Bholah, A., Cheong, R. N., Sallaway, M., Pillay, K. P., Calcino, D.,
Haydon, G., Koon, D. A., Hunter, H. M., Armour, J. D., Umrit, G., Soobadar, A., Fraser,
G., Hargreaves, P. and Ruddle, L. (2001) Offsite movement of agrochemicals in tropical
sugarcane production. Extension Workshop, Bundaberg, Australia, 8-9 May 2001. Papers
and Presentations., Bundaberg, Australia, Australian Centre for International Agricultural
Research, Mauritius Sugar Industry Research Institute, Queensland Department of Natural
Resources and Mines.
208
Simpson, B., Ruddle, L., Ham, G., Green, B. and Noble, A. (2002) Pesticides - issues and options for
growers. CRC for Sustainable Sugar Production. Factsheet. 2.
Skei, J., Hylland, K., Schaanning, M. T., Berge, J. A., Gunnarsson, J. S., Sk÷ld, M. and Eriksen, D. Ï.
(1996) Interactions between eutrophication and contaminants. I. Principles, experimental
design and synthesis. Marine Pollution Bulletin 33 (1-6): 64-70.
Steven, A., Devlin, M., Brodie, J., Baer, M. & Lourey, M. (1996) Spatial influence and composition of
river plumes in the central Great Barrier Reef. Hunter, H. M., Eyles, A. G. & Rayment, G.
E. (Eds.), In Downstream effects of land use. Department of Natural Resources,
Queensland, Australia, pp.85-92.
Stiegler, J. H., Criswell, J. T. and Smolen, M. D. Pesticides in ground water. Division of Agricultural
Sciences and Natural Resources
Tiller, K. G. (1983) Micronutrients. Soils: an Australian Viewpoint. London, Division of Soils, CSIRO,
Melbourne and Academic Press: 365-387.
Trott, L. A. and Alongi, D. M. (1999) Variability in surface water chemistry and phytoplankton biomass
in two tropical, tidally dominated mangrove creeks. Marine and Freshwater Research 50:
451-457.
Truong, P. (online) Pesticide run-off control using artificial wetlands. Department of Natural Resources,
Queensland. http://www.vetiver.org/AUS_wetland.htm.
Truong, P., Mason, F., Waters, D. and Moody, P. (2000) Application of vetiver grass technology in offsite pollution control. Second International Vetiver Conference, Thaliand, Queensland
Department of Natural Resources.
Turner, L. A. (1996) Abstract: Current water quality of the Condamine-Balonne river system. In
Downstream effects of land use. H. M. Hunter, A. G. Eyles and G. E. Rayment. Brisbane,
Department of Natural Resources, Queensland, Australia: 119-120.
Ullman, W.R. and Sandstrom, M.W. (1987) Dissolved nutrient fluxes from the nearshore sediments of
Bowling Green Bay, Central Great Barrier Reef lagoon (Australia) Estuarine, Coastal &
Shelf Science 24: 289-303.
Vallis, I. and Keating, B. A. (1994) Uptake and loss of fertiliser and soil nitrogen in sugarcane crops.
Proceedings of Australian Society of Sugar Cane Technologists 16: 105-113.
Vandenbrink, P. J., Vandonk, E., Glystra, R., Crum, S. J. H. and Brock, T. C. M. (1995) Effects of
chronic low concentrations of the pesticides chlorpyrifos and atrazine in indoor freshwater
microcosms. Chemosphere 31(5): 3181-3200.
Veitch, V. (1999) Fish kills: a report on water quality related fish kills on the North Queensland Coast
between Sarina and Cardwell from August 1997 to December 1998. Sunfish, Margate,
Queensland.
Walker, G. S. and Brunskill, G. J. (1996) Detection of anthropogenic and natural mercury in sediments
of the Great Barrier Reef lagoon. In: P. Larcombe, Woolfe, K. and Purdon, R. (Eds.),
Great Barrier Reef: Terrigenous Sediment Flux and Human Impacts. James Cook
University of North Queensland. Townsville
Walker, J. a. R., DJ (1996) Indicators of Catchment Health : A Technical Perspective, CSIRO Australia.
209
Walker, T. A. (1982) Use of a secchi disc to measure attenuation of underwater light for photosynthesis.
Journal of Applied Ecology 19: 539-544.
Walmsley, R. D., Butty, M., van der Piepen, H. and D. Grobler (1980) Light penetration and the
interrelationships between optical parameters in a turbid sub-tropical impoundment.
Hydrobiologia 70: 145-157.
Wangersky, P. J. (1994) Sampling and analysis of particulate and dissolved matter. The biology of
particles in aquatic systems. R. S. Wotton. USA, CRC Press: 7-44.
Ward, G. M., Ward, A. K., Dahm, C. N. and Aumen, N. G. (1994) Origin and formation of organic and
inorganic particles in aquatic systems. The biology of particles in aquatic systems. R. S.
Wotton. USA, CRC Press: 45-74.
Warne, M. S. (1998) Critical Review of methods to derive water quality guidelines for toxicants and a
proposal for a new framework. Supervising Scientist, Canberra Australia. Report 135.
Wasson, R. J. (1997) Run-off from the land to rivers and the sea. In: Proceedings of the CRC Reef
GBRMPA Great Barrier Reef Conference. James Cook University, Townsville.
Watzin, M. C. and McIntosh, A. W. (1999) Aquatic ecosystems in agricultural landscapes: a review of
ecological indicators and achievable ecological outcomes. Journal of Soil and Water
Conservation 54(4): 636-650.
Wedemeyer, G. A., Meyer, F. P. and Smith, L. (1976) Environmental stress and fish diseases, TFH
Publications.
Weier, K. L. (1999a) Workshop report Mackay, Proserpine groundwater study, CSIRO Tropical
Agriculture, Brisbane. 17 August 1999.
Weier, K. L. (1999b) Workshop report Burdekin groundwater study, CSIRO Tropical Agriculture,
Brisbane. 19 August 1999.
Weier, K. L. (1999c) Workshop report Ingham, Tully, Innisfail, Gordonvale, Atherton, Mareeba and
Mossman groundwater study, CSIRO Tropical Agriculture, Brisbane. 21 August 1999.
Wetzel, R. G. (2001) Limnology – Lake and River Ecosystems. 3rd Ed. Philadelphia, Saunders College
Publishing.
White, I., Brodie, J. and Mitchell, C. (in prep) Pioneer River Catchment event based water quality
sampling.
White, R. E. and Kookana, R. S. (1998) Measuring nutrient and pesticide movement in soils: benefits for
catchment management. Australian Journal of Experimental Agriculture 38: 725-743.
WHO (1996) Health criteria and other supporting information. Guidelines for drinking water quality.
Geneva, World Health Organisation. 2.
Wijngaarden, R. P. A. v. and Brock, T. C. M. (1999) Population and community responses in pesticidestressed freshwater ecosystems. Human and environmental exposure to xenobiotics. A. A.
M. DelRe, C. Brown, E. Capriet al: 571-589.
Williams D. McB. (2001) Impacts of Teresstrial Runoff on the GBRWHA. Report to CRC Reef,
Townsville QLD. (www.reef.crc.org.au)
Williams, D. M. (2001) Impacts of terrestrial run-off on the Great Barrier Reef World Heritage Area.
CRC Reef Research Centre, Australian Institute of Marine Science, Townsville. 100.
210
Wilson, C., argent, R., Bunn, S. E., Davies, P. M., Grayson, R. B., Hairsine, P. and Rutherford, I. (1996)
Toward an integrated model of the effects of riparian revegetation on stream values in
large catchments. H. M. Hunter, Eyles, A. G. and Rayment, G. E. (Eds.), In Downstream
effects of land use. pp 205-212, 1996. Dept. of Natural Resources, Queensland, Australia.
Wolanski, E. and Jones, M. (1981) Physical properties of Great Barrier Reef lagoon waters near
Townsville. I. Effects of Burdekin River floods. Aust. J. Mar. Freshwater Res., 32:305319.
Wolanski, E. and van Senden, D. (1983) Mixing of Burdekin River flood waters in the Great Barrier
Reef. Australian Journal of Marine and Freshwater Research 34:49-63.
Wood, A. W. (1985) Soil degradation and management under intensive sugarcane cultivation in north
Queensland. Soil Use and Management 1: 120-124.
Wood, A. W. and Kingston, G. (2000) Nutrient use efficiency. Sustainable nutrient management in
sugarcane production short course. R. C. Bruce. Townsville, Cooperative Research Centre
for Sustainable Sugar Production. Course Manual 21-22 March, 2000: 97-102.
Woolfe, K. J. and Larcombe, P. (1998) Terrigenous sediment accumulation as a regional control upon the
distribution of reef carbonates. Reefs and carbonate platforms in the Pacific and Indian
Oceans. G. F. Camoin and P. J. Davies, IAS Special Publication. 25: 295-310.
Woolfe, K.J., Larcombe, P., Orpin, A.R., Purdon, R.G., Michaelson, P., McIntyre, C.M. and Amjad, N.
(1998) Controls upon inner-shelf sedimentation, Cape York Peninsula, in the region of
12ºS. Australian Journal of Earth Sciences 45: 611-622.
Woolfe, K.J., Larcombe, P., Ridd, P. V., Orpin, A.R., Bryce, S. and McIntyre, C.M. (1996) A brief field
guide to the Burdekin delta and Cocoa Creek. In: P. Larcombe, Woolfe, K. and Purdon, R.
(Eds.), Great Barrier Reef: Terrigenous Sediment Flux and Human Impacts. James Cook
Wotton, R. S. (1994) The classificationof particulate and dissolved matter. The biology of particles in
aquatic systems. R. S. Wotton. USA, CRC Press: 1-6.
Yellowlees, D. e. (1990) Land use patterns and nutrient loading of the Great Barrier Reef region.
Yen, J.-H., Hsiao, F.-L. and Wang, Y.-S. (1997) Assessment of the insecticide carbofuran's potential to
contaminate groundwater through soils in the subtropics. Ecotoxicology and
Environmental Safety 38: 260-265.
Young, W. J., Prosser, I. P. and Hughes, A. D. (2001) Modelling nutrient loads in large-scale river
networks for the National Land and Water Resources Audit. CSIRO Land and Water
Technical Report 12/01.
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