APPENDIX 4

PROJECT 4: STREAM ECOLOGY

Managing and harvesting urban stormwater for stream health

LITERATURE REVIEW

SEPTEMBER 2010

Appendix 4: Project 4 – Literature Review, 2010 ii

Executive Summary

Why are waterway ecosystems important?

Humans have always depended on the products and services of healthy river and stream ecosystems, not only for water and waste disposal, but for the important food and fibre sources provided by their plants and animals. Rivers and streams play other important roles, particularly the retention and transformation of materials and pollutants which are generated in their catchments (Alexander, et al., 2000). They can help to protect downstream estuaries and bays from excessive nutrients and are thus vital in protecting important coastal fisheries. Humans also value waterways for their intrinsic and biodiversity values. Freshwaters, despite only making up 0.009% of the world‘s water, contain at least 40% of the world‘s fish species and a quarter of all vertebrate species (Millennium Ecosystem Assessment, 2005). Unfortunately, the rate of extinction in freshwater systems is greater than in any terrestrial ecosystem (Dudgeon, et al., 2006). The degradation of waterways through urbanisation has resulted in the loss of many riverine species, and threatened human wellbeing and livelihoods through phenomena such as algal blooms and toxicant accumulation. The natural form of urban waterways has typically been altered, either directly, or though changes in hydrology, which acts as a ‗master variable‘ driving patterns and processes in waterways (Poff, et al., 1997). So why should we protect and attempt to rehabilitate and restore our urban waterways? The intrinsic and human values described above provide, in themselves, a compelling rationale. Healthy streams, rivers, estuaries and coastal waters are a fundamental prerequisite to achieving sustainable management of our cities; our streams must be maintained in a healthy state for not only this generation, but for future generations (United Nations General Assembly, 1987). However, there are also some simple economic arguments for taking action to prevent stream degradation. Preventing decline through appropriate stormwater management will be considerably less expensive than the cost of restoring already degraded waterways (Rutherfurd, et al., 2000). Fortunately, the integration of stormwater harvesting and other carefully-designed stormwater management provides a great opportunity to protect and restore urban waterways.

Impacts of current stormwater management practice on streams

Current stormwater management practice severely degrades the health of rivers and streams, with consequent impacts on estuaries and coastal waters. In the worst case, small streams are obliterated by piping. Even where streams are retained, current stormwater practice results in changes to hydrology, water quality and geomorphology, leading to inevitable ecological decline. Whilst attempts to improve stormwater management through ―water sensitive urban design‖ have been successful in reducing pollutant loads, current practice has failed to arrest the degradation of streams.

Hydrology Urbanisation profoundly changes the hydrologic cycle, increasing the volume, frequency and flow rate of storm flows. Evapotranspiration is also greatly reduced, decreasing from around 80% of mean annual rainfall in the pre-developed situation to around 15% for impervious areas. The increases in runoff result in increased frequency and magnitude of disturbance to channel substrates and to aquatic ecosystems. Whilst the impacts of urbanisation on stormflows are generally well recognised, the impacts on low-flow hydrology are less well understood, and yet have critical impacts on both the ecological and social value of waterways. The loss of baseflow due to decreased infiltration caused by impervious areas, results in once perennial streams becoming ephemeral, and has consequent impacts on water quality, particularly during summer periods. Many studies have investigated potentially useful indicators for use in assessing the catchment-scale impacts of urbanisation. Such indicators need to have ecological relevance and be able to be measured and modelled. They should be sensitive to the urbanisation gradient, rather than external

Appendix 4: Project 4 – Literature Review, 2010 iii

factors (e.g. catchment size). The indicators found to be most useful relate to frequency and duration of high flow events, flashiness, mean annual runoff volume and duration and frequency of low flow events. At the site-scale, a different set of indicators will be required; these will need to be suitable for application in assessing the performance of stormwater management (and harvesting) systems. Indicators such as the frequency of discharge from the site, the volume of sub-surface (filtered) flow and the mean annual flow volume from the site, have been proposed as being suitable.

Water quality

Like hydrology, changes to water quality are a ubiquitous response to urbanisation. A wide range of pollutants, including toxicants, nutrients oxygen-demanding substances, and of course sediments, are likely to occur in elevated concentrations and loads in catchments subjected to urbanisation. The increased concentrations and loads in waterways results from both increased generation and mobilisation/transport processes (due to the creation of hydraulically efficient drainage networks). A number of studies have shown that pollutant concentrations are well predicted by impervious areas, but importantly, Hatt et al‘s (2004) study showed that the effective imperviousness of a catchment gave the best prediction of a wide range of pollutants, indicating the direct role which stormwater discharge has on receiving waterway water quality. A new approach to the design of urban drainage systems is thus required if we wish to avoid water quality degradation due to urbanisation. The selection of water quality indicators must take into account the nature of the waterway. For large systems, loads targets provide a suitable measure, whilst concentration targets (such as those provided by ANZECC or local SEPPs) provide the most ecologically-relevant target for smaller flowing waterways, which have limited buffering capacity.

Geomorphology

The natural geomorphic function of urban waterways is severely impacted by the increased volume of water generated by the urbanised catchment. The increased flow removes much of the mobile sediment load and preferentially erodes the channel. With conventional urban hydrology there is limited potential to replace (and retain) mobile sediments and hence the natural functioning, namely erosion and deposition in a form of quasi-equilibrium. This includes ubiquitous features of a ‗natural‘ channel such as benches and bars. Unless we reduce those flows responsible for mobile sediment transport returning much of this functioning is not feasible, necessitating the protection of artificially ‗static‘ channels with hard or soft engineering works. In addition, the increased channel capacity of urban streams confounds the hydrologic alterations induced by the urban catchment. Reducing channel capacity and re-engaging the floodplain will reduce the accelerated channel degradation resulting from both the incised channel and altered hydrologic regime. Within the urban environment space is the greatest constraint where ‗internal floodplains‘ may be required to alleviate impacts on the channel while preventing flooding into the development boundary.

Ecology

Given the impacts of urbanisation on hydrology, water quality and geomorphology, it is n surprise that there are major consequent impacts on the ecological values of streams. While the underlying mechanisms are complex, they can be classified into catchment-level impacts (largely hydrology and water quality) and local impacts (largely changes in organic matter supply). The water quality and hydrologic impacts described above cause increases in the frequency of both hydraulic and water quality disturbance, change the structure of in-stream habitat, and directly disturb biota. The majority of ecological impacts are therefore a result of catchment scale processes, whilst local-scale impacts (e.g. riparian vegetation) may play a role in some circumstances. However, catchments with even a small coverage of connected imperviousness are likely to have sufficiently altered hydrology and reduced water quality that even extensive riparian restoration works may show minimal improvement in stream ecological condition. Ecological responses to catchment urbanisation can be measured using a range of indicators. For example, primary producers (e.g. algae) will increase as a result of increased nutrient concentrations (Hatt, et al., 2004; Newall & Walsh, 2005; Resh, et al., 1988; Sonneman, et al., 2001). Ecosystem

Appendix 4: Project 4 – Literature Review, 2010 iv

function indicators (such as nitrogen cycling) have been shown to vary along an urban gradient, as have factors such as organic matter decomposition (Meyer, et al., 2005).

Perhaps the most commonly used indicators have been macroinvertebrates, given their sensitivity to water quality changes and the presence of well-established protocols for comparing impacted and reference populations. Urban streams typically, show an increased abundance of ‗pollution tolerant taxa‘, with reductions in the more ‗pollution sensitive taxa‘ including Ephemeroptera, Plecoptera and Trichoptera (Paul & Meyer, 2001; Suren & McMurtrie, 2005; Walsh, 2004). Urban streams are also typified by organisms with rapid life histories, such as oligochaetes and chironomids. These taxa are highly tolerant to hydrologic disturbance (both in terms of high flows and drying) because they spend a relatively short time as an aquatic larva. Fish have also been used as indicators, with urbanised streams typically having much less diverse fish populations (Arthington, et al., 1983).

Importantly, because a wide range of ecological indicators are highly sensitive to the impacts of stormwater runoff, significant response by their indicators is seen at even very low levels of urbanisation. This makes these indicators very useful in assessing the degree to which urbanisation is impacting on the waterway ecosystem, but, conversely, means that rehabilitation efforts (e.g. through stormwater management and harvesting) will not show an ecological response until the urban-runoff impact has been reduced to very low levels. Whilst stormwater managers may understandably wish to have a more linear (and thus progressive indicator), no studies to date have been able to identify any. Given the role of indicators in identifying the ecological condition of waterways, the search to find a more progressive indicator, while perhaps helpful in terms of garnering political and social support for interventions, will probably provide little ecological insight. The sharply non-linear response of currently indicators accurately identifies the situation; streams are ecologically degraded wherever significant stormwater inputs are permitted to occur.

Potential for stream rehabilitation through stormwater harvesting

Urban stormwater runoff poses an environmental flow problem, but one which is completely the inverse of the ‗normal‘ environmental flow problems we deal with. Most environmental flows problems arise as a result of too much water being extracted from waterways for urban use, with the challenge being to distribute the remainder for the maximum environmental benefit (Arthington, et al., 2010). As has been described above, the reverse is true for urban stormwater; the frequency and volume of stormwater runoff into waterways is far greater than in the pre-developed state. Urban stormwater runoff, delivered through conventional drainage systems is a complex environmental flow problem that can, in large part, be solved through harvesting of stormwater before it reaches aquatic ecosystems. This is an unusual ―win-win‖ situation, in that retention of that water for human uses will have a positive environmental outcome in protecting the waterway. However, it is critical that urban stormwater harvesting systems harvest water before it enters the waterway, rather than harvesting water from the waterways, leaving flow and water quality disturbance upstream unaddressed. Stormwater harvesting can – and should – thus be applied to help restore flow regimes and water quality. To do so requires careful integration of harvesting (which has the potential to reduce flow volume and peak discharge rates) with carefully chosen techniques to restore the low-flows which have been lost due to urbanisation. A wide range of WSUD techniques, such as vegetated infiltration and biofiltration systems, can be used to restore both the evapotranspiration and infiltration fluxes.

Interim recommendations

We propose the following initial set of ecologically-based principles for the design of integrated stormwater management, incorporating both stormwater harvesting and measures focussed at restoring baseflows and water quality:

1. Minimize uncontrolled storm flows to the pre-developed runoff frequency.

2. Restore baseflows, by delivering filtered flows to the stream through treatment measures that ensure flow rates do not exceed pre-urban subsurface flow rates. Appropriate maximum flow rates can be estimated from baseflow separation analysis in reference streams, or by assessment of infiltration capacities of native soils in the catchment.

Appendix 4: Project 4 – Literature Review, 2010 v

3. Filtered flows should aim to meet water quality concentration objectives close to standard objectives for ecosystem protection of freshwaters (ANZECC & ARMCANZ, 2000). For some variables, such as nitrogen, ANZECC objectives might be unattainable in treatment systems with collection pipes, in which case pragmatic acceptance of best attainable concentrations is appropriate. Wherever possible, exfiltration systems, and systems that allow overflow or infiltration flows to drain to pervious land, will help to achieve improved water quality, as well as increasing loss through evapotranspiration.

4. The attainment of the first three objectives will require substantial retention and loss of stormwater runoff, either for indoor use and export to the wastewater stream, or for irrigation and loss to evapotranspiration. Therefore harvesting of a large proportion of stormwater runoff is a central objective for restoration or protection of environmental flows. Modelling and/or analysis of pre-developed flows using relationships derived by Zhang et al (2001) can be used to estimate the amount required to be harvested (Figure S1).

Figure S1. Annual volume of runoff from 1 ha of impervious surface, partitioned into two parts: the volume that needs to be passed through filtration systems to restore lost subsurface flows (grey polygon), and the volume that needs to be retained in the catchment and not delivered to the stream (through evapotranspirational loss or through use and export from the catchment through the wastewater stream (adapted from Zhang et al. (2001).

Knowledge gaps and impediments to overcome

There are a number of important knowledge gaps which remain in relation to the potential impact of stormwater harvesting and water sensitive urban design on urban waterway ecosystems: Hydrological effects, scales and integration; how should stormwater harvesting be integrated with other stormwater management measures to restore all components of the flow regime? At what scales should these systems be located? To what extent do flow regimes needs to be restored in order to see an ecological response? What is the ecologically appropriate flow regime in the context of a channel enlarged by urbanisation? Impacts on water quality regimes; what will be the consequences of harvesting at different scales on the water quality regime in the receiving waters, and what ecological consequences will that have? Will different waterway types have varying tolerance to harvesting-induced water quality changes? Is catchment-scale water quality treatment enough, or is floodplain re-engagement necessary to restore pre-development nutrient retention of urban streams, in order to protect downstream waterways? Geomorphic processes and consequences; what is the effect of urbanisation on sediment budgets and what effect will stormwater harvesting have on these budgets. Do we need to restore the sediment regime? How does the reduction in mobile sediments impact on geomorphic functions of the channels? What level of dynamism is acceptable or feasible for an urban stream, and how can we design our stormwater system to facilitate this ‗desirable‘ level of dynamism. Will channel intervention be necessary, or will restoring pre-development flows be enough to allow channels to

Appendix 4: Project 4 – Literature Review, 2010 vi

self-restore? What role will stormwater detention storage have on erosion potential; how can we design systems to minimise erosion potential whilst maximising water yields? Ecological responses and thresholds; What mitigating factors might explain variations in the level of ecosystem response to urban impacts? How will stream ecosystems respond to stormwater harvesting? How should we monitor and report these responses in ways that are ecologically meaningful?

A vision for urban waterways

In order to underpin the Stream Ecology research project, the research team has developed an initial vision for waterways, focussing on those waterways which are most directly affected by stormwater harvesting; streams. During the course of the research project, we will be developing similar visions for other waterway types, such as lakes and estuaries, and large rivers. However, we hypothesise protection of small streams is a critical prerequisite to protection of larger downstream waterways, because small upland streams have such low buffering capacity. These visions focus on the role that stormwater harvesting and stormwater management can play in protecting and restoring waterways. We consider two cases; those streams currently in good condition and thus potentially able to be protected, and those which are currently highly degraded, for which we hope to use a combination of harvesting and stormwater management to achieve a significant degree of rehabilitation.

A vision for streams with hope of being protected

To use urban stormwater management and harvesting to maintain the stream in its intact

natural state

Driving factor indicators: maintain natural flow volume and variability and pollutant concentrations and loads; natural levels of erosion and sediment deposition, supply and retention of coarse particulate organic matter (CPOM), natural frequency of floodplain engagement, adequate riparian vegetation buffer width and condition Ecological indicators: achieve presence of sensitive species (e.g. high EPT, riparian plants), natural levels of primary and secondary productivity and natural levels of nutrient and sediment retention (with respect to reference conditions).

A vision for streams which are currently highly degraded

To use urban stormwater management and harvesting to re-establish ecological processes

and patterns characteristic of the stream in its natural state

Driving factor indicators: reduce flow volume, magnitude, duration and frequency of stormflows, restore baseflows and reduce pollutant concentrations and loads; retain appropriate levels of mobile sediment, reduce erosion and restore near-natural sediment supply levels, restore supply and retention of CPOM and as much as possible the natural frequency of floodplain engagement, maintain an adequate riparian vegetation buffer.

Ecological indicators: achieve presence of sensitive species (e.g. high EPT, riparian plants), reduced levels of primary and secondary productivity and increased retention of nutrient and sediment (with respect to reference conditions).

We hypothesise that the ‗initial principles‘ guiding the design and operation of stormwater harvesting systems, integrated with other WSUD measures aimed at restoring pre-development water quality and hydrology, will be capable of delivering this vision. The Cities as Catchments Project, and in particular the opportunity to monitor real systems, gives us an ideal framework in which to test this hypothesis.

Appendix 4: Project 4 – Literature Review, 2010 vii

Table of Contents Executive Summary ................................................................................................................................ ii Table of Figures .................................................................................................................................... viii Table of Tables ..................................................................................................................................... viii 1 The ecological function of waterways ............................................................................................. 1 2 Impacts of conventional urban stormwater management ............................................................... 4

2.1 Hydrological cycle ................................................................................................................... 4 2.1.1 Catchment water balance ............................................................................................... 4 2.1.2 Impacts on flow regimes ................................................................................................. 6 2.1.3 Hydrologic indicators ....................................................................................................... 8 2.1.4 Summary ....................................................................................................................... 13

2.2 Water quality ......................................................................................................................... 14 2.2.1 Mechanisms of pollutant generation and transport ....................................................... 14 2.2.2 Effect on pollutant concentrations and loads ................................................................ 16 2.2.3 Pertinent water quality indicators / objectives ............................................................... 16 2.2.4 Summary ....................................................................................................................... 17

2.3 Geomorphology ..................................................................................................................... 18 2.3.1 Impacts of urbanisation on sediment processes ........................................................... 18 2.3.2 Impacts of urbanisation on channel morphology .......................................................... 19 2.3.3 Pertinent geomorphic objectives and indicators ........................................................... 25 2.3.4 Summary ....................................................................................................................... 26

2.4 Ecology .................................................................................................................................. 27 2.4.1 Overview ....................................................................................................................... 27 2.4.2 Drivers of ecological change in urban-impacted streams ............................................. 27 2.4.3 Catchment-scale drivers ............................................................................................... 28 2.4.4 Local-scale drivers ........................................................................................................ 29 2.4.5 Indicators of ecological response .................................................................................. 30 2.4.6 Summary ....................................................................................................................... 35

3 The environmental flows concept.................................................................................................. 36 3.1 Environmental flows; the basics ............................................................................................ 36 3.2 Application of environmental flows concept to urban stormwater ......................................... 37

4 The role of stormwater harvesting and its integration with other WSUD and stream restoration interventions .......................................................................................................................................... 39

4.1 Potential impacts of stormwater harvesting .......................................................................... 39 4.2 Interactions between stormwater harvesting and other stormwater management measures 41

4.2.1 Urban stream restoration; interactions with other stream interventions........................ 42 5 Critical knowledge gaps ................................................................................................................ 44

5.1 Hydrological effects, scales and integration .......................................................................... 44 5.2 Impacts on water quality regimes.......................................................................................... 44 5.3 Geomorphic processes and consequences .......................................................................... 45 5.4 Ecological responses and thresholds ................................................................................... 45 5.5 Selection of indicators at a range of scales .......................................................................... 45

References ............................................................................................................................................ 47

Appendix 4: Project 4 – Literature Review, 2010 viii

Table of Figures Figure 1. Relationship between annual rainfall and evapotranspiration for a range of vegetation types (Source: Zhang, et al., 1999). ................................................................................................................. 5 Figure 2. Typical water balance in natural and 50% impervious urbanised catchment (Base diagram adapted from FISRWG, 1998). ............................................................................................................... 5 Figure 3. Schematic illustration of the pertinent impacts of urbanisation on hydrology at the catchment scale (Marsalek, et al., 2007). ............................................................................................... 6 Figure 4. Schematic illustration of the pertinent impacts of urbanisation on catchment hydrology. ...... 8 Figure 5. Conceptualised change in sediment delivery to streams relative to catchment land-use (after Wolman 1967) ............................................................................................................................. 18 Figure 6. Channel incision and recovery after Schumm et al (1984). a) temporal view, b) spatial view .............................................................................................................................................................. 21 Figure 7. A conceptual model of the impacts of urbanisation on stream ecosystems (after Walsh et al. 2005b). .................................................................................................................................................. 28 Figure 8. Non-linear relationships between a wide range of ecological indicators (D-N) and effective imperviousness. .................................................................................................................................... 35 Figure 9. Conceptual graphs of ecological and human value of water ................................................ 36 Figure 10. Estimated annual runoff coefficients (C) from impervious surfaces (open triangles) from sites across the Melbourne region as a function of mean annual rainfall (R). ...................................... 38 Figure 11. Annual volume of runoff from 1 ha of impervious surface .................................................. 39 Figure 12. Relative changes to flow and water quality indicators from natural conditions for a catchment with no permanent baseflow (i.e. typical small upland catchment) ..................................... 40 Figure 13. Relative changes to flow and water quality indicators from natural conditions for a catchment with no permanent baseflow (i.e. typical small upland catchment) ..................................... 40 Figure 14. Schematic illustration of the pertinent impacts of urbanisation on hydrology at the catchment scale, showing the respective roles of stormwater harvesting and baseflow-restoration techniques such as bioretention and infiltration (adapted from Marsalek, et al., 2007)........................ 42

Table of Tables Table 1. Hydrologic indicators suitable for assessment of urbanisation impacts, as suggested by DeGasperi et al. (2009). ........................................................................................................................ 10 Table 2. Hydrologic indicators used by Fletcher et al. (2007) for evaluation of stormwater harvesting impacts. ................................................................................................................................................. 13 Table 3. Typical urban stormwater pollutants and sources (after: Duncan, 1995; Lawrence & Breen, 2006). .................................................................................................................................................... 15 Table 4. Summary of typical water quality values for runoff from urban, agricultural and forested catchments, urban streams, and secondary treated sewage. .............................................................. 16 Table 5. Example of pollutant load reduction targets (Source: Wong, 2006). ..................................... 17 Table 6. Summary of ecological responses to urbanisation currently reported in the scientific literature. ............................................................................................................................................... 31

Appendix 4: Project 4 – Literature Review, 2010 1

1 The ecological function of waterways

Why are waterways important?

The development of human civilization has always depended on the products and services provided by healthy river and stream ecosystems. Much of the fresh water on which towns and cities of the world depend is derived from rivers and streams. It is no coincidence that most of the great cities of the world lie on the banks of large rivers (Mumford, 1961). The animals and plants of rivers and streams historically provided important food and fibre sources for human settlements, and this remains true in many parts of the world, such as the great basins of the Mekong and the Amazon. Riverine ecosystems are also important for their retention, transformation and export of materials that are generated by catchment physical and biological processes. The transport of sediments in rivers is a central determinant of landforms (Knighton, 1998) and the productivity of estuaries and coastal waters is strongly determined by the input of nutrients from rivers, particularly during floods (Loneragan & Bunn, 1999). Thus coastal fisheries are dependent on the health of their receiving rivers and streams. However, the health of coastal waters and the productivity of coastal fisheries depend on an appropriate balance of nutrient delivery. Eutrophication resulting from excess nutrient export from degraded rivers is a threat to many coastal waters globally, particularly those downstream of large cities (e.g. Harris, et al., 1996; Palmer, 2004). Small streams and their riparian zones are critical parts of the landscape for the retention of nutrients to prevent excessive export, particularly nitrogen (Alexander, et al., 2000), primarily because a large proportion of the water they carry flows through channel and riparian sediments that are hotspots for critical nutrient transformation and removal processes such as denitrification (McClain, et al., 2003). In addition to protecting downstream waters, this important capacity of healthy streams to retain and transform contaminants also serves to purify the water in which riverine plants and animals live (and which humans extract). Yet, river ecosystems, so central to human well-being, carry only 0.0002% of the world‘s water, and cover less than 0.1% of the world‘s surface (Millennium Ecosystem Assessment, 2005). Given their relatively small coverage of the planet, the diversity of life in rivers is astonishing. Freshwaters (which includes lakes as well as rivers, making up 0.009% of the world‘s water) are home to at least 40% of the world‘s fish species, and a quarter of all vertebrate species. The taxonomy of freshwater biota is woefully understudied compared to terrestrial biota, yet they still account for 6% of all described species (Dudgeon, et al., 2006). While the extraordinary biodiversity of rivers and lakes is not as widely appreciated as that of, say tropical rainforests and coral reefs, the threats and rates of extinction are much greater in freshwaters than in any of the most affected terrestrial ecosystems (Dudgeon, et al., 2006). In part, the increased vulnerability of freshwater biodiversity and freshwater ecosystems in general, is a result of their position in the landscape. At the bottom of valleys, they are the natural destination of the products of land use change. Even in the absence of land use change, the biodiversity of rivers and the ecological processes that support it (and human civilization) depend strongly on physical processes originating in the catchment, with processes in the riparian zone and channel having the most immediate influence.

Human use and perceptions of waterways

Throughout history, humans have focussed on rivers as water resources, and have largely managed them as networks that can be engineered for the storage and distribution of water for human needs (Naiman, et al., 1995). Unfortunately, and somewhat self-defeatingly, the second primary use that humans have put rivers to is as repositories of waste. This narrow, utilitarian approach to rivers has resulted in a grand diminishment of the benefits that rivers provide the world as healthy ecosystems, such as their remarkable biological diversity, and the retention and treatment of nutrients and other contaminants. Poor management of the world's rivers

Appendix 4: Project 4 – Literature Review, 2010 2

has resulted in the loss of many riverine species, with unknown long-term consequences, and in other losses and changes that have more direct and immediate impacts on our social and economic well-being: e.g. loss of native fish stocks, toxic algal blooms and increases in salinity, which threaten urban water supplies and irrigation systems. Historically, the development of cities has followed a pattern of initial dependence on freshwater from the local river, increasing use of the river for disposal of wastes, compromising its use as a water source, followed by the implementation of engineering solutions to source more distant water for the city. With increasing affluence, modern cities have introduced environmental regulation to limit wastewater disposal into rivers, but the use of rivers for the disposal of urban stormwater remains the norm (Walsh, 2000). Improved environmental regulation has resulted in improvement in the aesthetic appearance and reduced health risks associated with primary contact of urban rivers. However, the continued, unregulated discharge of urban stormwater to rivers has largely limited any improvement in urban rivers to aesthetics and reduced human health risk (Walsh, 2000). Urban dwellers value the parklands and open space afforded by urban streams (Lackey, 2001; ResearchWise, 2004). Increasingly, cities are turning to their rivers as attractive open space, but the rivers in most cases are diminished in their ecological structure in function by catchment processes, and local engineering of their banks. The protection or restoration of biodiversity or ecological function in rivers with substantial catchment urbanisation, to a condition close to that of streams with less intense human land use, has yet to be achieved. The renewed and growing enthusiasm of the world‘s urbanites for their rivers and streams is encouraging. However, the perception of today‘s generations of improved urban rivers arises from a starting point of sick polluted rivers of 20-30 years ago. A lowering of expectations of the environment over generations is a common occurrence because the rate of environmental change exceeds generation times (Pauly, 1995). We must therefore be careful to set objectives for environmental improvement based on what is achievable rather than what is perceived by the community as acceptable.

Factors that drive important processes and features of waterways

More than any other ecosystem type, rivers and streams—running waters—are defined by their hydrology. Stream flow has been called the 'master' variable that drives patterns and processes in rivers: certainly geomorphic, ecological and evolutionary processes in rivers are strongly influenced by many aspects of the flow regime (Poff, et al., 1997). The physical setting of a river strongly influences how the flow regime translates into hydraulic conditions experienced by, and available to, the riverine biota (Poff, et al., 2010). The translation of the flow regime into small-scale hydraulic conditions has important implications for ecological processes (Newall & Walsh, 2005). Indeed, Stratzner and Higler (1985) state that ‗physical characteristics of flow (“stream hydraulics”) are the most important environmental factor governing the zonation of stream benthos on a world-wide scale‘ (pg. 127). The abundance and diversity of aquatic biota is closely associated with geomorphic aspects of channel morphology and functioning (Bledsloe, 2002). Physical features such as riffles, pools and bars, and channel characteristics such as width, depth and sinuosity drive the translation of flow regime into hydraulic conditions. For example, greater variation in bed elevation will provide fish and invertebrates with longer pool persistence during drying periods and refuge during higher ‗disturbance‘ flows. Local characteristics such as channel geometry, floodplain height and streambed composition are determinants of the impact of events, such as whether a given flow will create a bed-moving disturbance or an overbank flow (Poff et al. 2010). Natural channels are inherently dynamic in their physical form. Their characteristics vary over time and space with changes in environmental controls and channel morphology being a function of the flow and sediment regime delivered to the channel (Gilvear, 1999; Knighton, 1998). Not only do channel features respond to these inputs, and translate flow into erosive or transportational components, but the channel itself is a deformable boundary. The altered boundary then impacts differently on sediment or velocity.

Appendix 4: Project 4 – Literature Review, 2010 3

The physical template can also greatly influence the social and economic functioning of streams. For example, flooding, loss of land, and channel maintenance requirements are all driven by geomorphic and hydraulic variables. Better understanding the influence man exerts on these variables (e.g. urbanisation and increased stream energy and maintenance) can assist in determining where changes in river management practice will result in social and economic values. Where human activity interacts with stream processes, effective plans for river management and use must take account of some understanding of the principles of river behaviour (Wolman, 1967). The health, biodiversity and human value of urban waterways is also driven by water quality, which has both short-term and long-term effects. Water quality is strongly linked with hydrological changes; increases to runoff coefficients will result in increased mobilisation and transport of contaminants from the catchment into the receiving waters. Aquatic organisms are typically sensitive to elevated levels of toxicants such as heavy metals, as well as the accumulation of excessive nutrients, which can cause eutrophication and the development of (often toxic) algal blooms (Novotny & Olem, 1994; Novotny & Witte, 1997). Over the longer-term, pollutant accumulation in sediments can result in continuing degradation of the ecosystem, even after the pollutant source has been addressed.

A rationale for urban waterway protection

As has been shown above, waterways are valuable natural assets in the urban landscape – both for their intrinsic values and for the wide range of services they provide humans. Sustainable management of our urban streams requires that we maintain them in a healthy state for generations to come (United Nations General Assembly, 1987). However, there are also strong economic arguments for the protection of the existing values of urban waterways. Considerably less investment is required to protect or prevent decline than is required to restore already degraded waterways (Rutherfurd, et al., 2000). In the case where we are uncertain as to the value of ecological health in urban streams, based on current knowledge and values, the precautionary principle should be adopted. This is analogous to environmental flows for rural streams where the dramatically increased understanding in the last decade has brought considerable investment and attention to the issue. In many cases the actions are too late for effective action, particularly actions which are affordable. In the case of urban streams, we suggest that there is strong evidence that stream and river ecosystems that support biota and ecological functions similar to streams with little human impact is possible; that the (generally low) expectations and perceptions of urban dwellers of their streams can be exceeded. Such a future is dependent on new approaches to stormwater management that address the important drivers of degradation, as identified above.

A vision for urban waterways

In order to underpin the Stream Ecology research project, the research team has developed an initial vision for waterways, focussing on those waterways which are most directly affected by stormwater harvesting; streams. During the course of the research project, we will be developing similar visions for other waterway types, such as lakes and estuaries, and large rivers. However, we hypothesise protection of small streams is a critical prerequisite to protection of larger downstream waterways, because small upland streams have such low buffering capacity. These visions focus on the role that stormwater harvesting and stormwater management can play in protecting and restoring waterways. We consider two cases; those streams currently in good condition and thus potentially able to be protected, and those which are currently highly degraded, for which we hope to use a combination of harvesting and stormwater management to achieve a significant degree of rehabilitation. Ideally the only difference between these two visions is time. The ultimate hope would be to return all urban streams to a condition near to their pre-developed state, but clearly this will only be possible over very long timeframes. The second vision thus sets out the goals for the condition of currently degraded streams in the foreseeable future, without precluding future rehabilitation to a close to intact natural state.

Appendix 4: Project 4 – Literature Review, 2010 4

A vision for streams with hope of being protected

To use urban stormwater management and harvesting to maintain the stream in its intact

natural state

Driving factor indicators: maintain natural flow volume and variability and pollutant concentrations and loads; natural levels of erosion and sediment deposition, supply and retention of coarse particulate organic matter (CPOM), natural frequency of floodplain engagement, adequate riparian vegetation buffer width and condition Ecological indicators: achieve presence of sensitive species (e.g. high EPT, riparian plants), natural levels of primary and secondary productivity and natural levels of nutrient and sediment retention (with respect to reference conditions).

A vision for streams which are currently highly degraded

To use urban stormwater management and harvesting to re-establish ecological processes and

patterns characteristic of the stream in its natural state

Driving factor indicators: reduce flow volume, magnitude, duration and frequency of stormflows, restore baseflows and reduce pollutant concentrations and loads; retain appropriate levels of mobile sediment, reduce erosion and restore near-natural sediment supply levels, restore supply and retention of CPOM and as much as possible the natural frequency of floodplain engagement, maintain an adequate riparian vegetation buffer.

Ecological indicators: achieve presence of sensitive species (e.g. high EPT, riparian plants), reduced levels of primary and secondary productivity and increased retention of nutrient and sediment (with respect to reference conditions).

2 Impacts of conventional urban stormwater management

2.1 Hydrological cycle

2.1.1 Catchment water balance

Urbanisation results in major changes to the urban water cycle through:

The creation of impervious areas (decreasing infiltration and evapotranspiration)

The loss of vegetation (decreasing evapotranspiration)

The creation of hydraulically efficient drainage networks, which decrease channel storage and

losses, and which decrease time of concentration (the time taken for flow from all areas of the

catchment to reach the catchment outlet.

In a natural catchment, the vast majority (typically 80-95%) of precipitation which falls on a catchment will be evapotranspired back to the atmosphere (Figure 1, Figure 2), with only a relatively small percentage (typically 5-20%) resulting in streamflow (Argue, 2009; Zhang, et al., 1999). The streamflow coefficient (proportion of catchment rainfall which ends up as streamflow) will depend on factors such as slope, soil and vegetation type and of course climate. Evapotranspiration will occur both in the soil and in the vegetation canopy, as a result of interception, where precipitation is

Appendix 4: Project 4 – Literature Review, 2010 5

intercepted by leaves and evaporated before it can reach the ground. The majority of streamflow in most years will result from subsurface flow (groundwater flow), with a small proportion contributed via direct surface runoff. The subsurface flows may enter waterways through shallow or deep infiltration. For example, throughflow or interflow may result where precipitation is infiltrated only to shallow depths and then makes its way, relatively rapidly, to the receiving waters (Ladson, 2008).

Figure 1. Relationship between annual rainfall and evapotranspiration for a range of vegetation types (Source: Zhang, et al., 1999).

Figure 2. Typical water balance in natural and 50% impervious urbanised catchment (Base diagram adapted from FISRWG, 1998).

In an undisturbed catchment, the frequency with which surface runoff will occur is limited by the ‗sponge effect‘ of the catchment soils (as well as interception by vegetation). For example, Hill et al (1998; 1996) suggest that initial loss values (the amount of rainfall required before runoff is produced) range from around 10 to 50 mm, with a typical value of around 25 mm. Not surprisingly then, natural catchments produce direct surface runoff to streams very rarely. This has both hydrological and water quality implications, since in a natural catchment runoff events capable of mobilising and transporting sediments and pollutants in waterways happen very rarely (Walsh, et al., 2005a; Walsh, et al., 2009). The effects of urbanisation on the urban water balance are well documented, with the key observations being major reductions in evapotranspiration along with major increases in both the frequency and volume of runoff (Fletcher, et al., 2007; Walsh, et al., 2009). Typically a reduction in groundwater level is also observed although this may be obscured by loss of evapotranspiration (due

Appendix 4: Project 4 – Literature Review, 2010 6

to loss of vegetation) and anthropogenic inputs such as water infrastructure leakage (Le Delliou, et al., 2009; Lerner, 1990).

2.1.2 Impacts on flow regimes

Not surprisingly, changes to the catchment water balance due to urbanisation result in major changes to the flow regime of receiving waters. Of these changes, the most obvious – and most reported –is the increase in the frequency and magnitude of peak flow rates and the increase in total runoff volumes (Leopold, 1968). For example, Wong et al. (2000) showed how the moderate urbanisation of a hypothetical catchment could increase the magnitude of the 1 year ARI flow by around ten times. Similarly, Rose and Peters (2001) confirmed that peak storm flows in several urban catchments of the Atlanta area (Georgia, USA) were 30% to 100% greater compared to those in nearby non-urban catchments. This effect is greatest for more frequent storms; as the storm magnitude increases, the difference between a natural and urbanised catchment decreases, with the natural soils becoming saturated and behaving relatively more like impervious areas (DeGasperi, et al., 2009). However, the impacts are far more complex than a simple increase in flow rates, with the following changes also commonly observed:

typically decreased dry weather flows (often referred to as baseflows). The magnitude of this change will depend on the balance between loss of infiltration and loss of evapotranspiration within the catchment;

more frequent surface runoff inputs (with consequences for pollution inputs) as a result of the decreased initial loss (from around 25 mm in a natural catchment to around 1 mm in the urbanised catchment);

decreased lag-time (time of concentration) and recession time;

increased frequency and duration of dry spells (cease-to-flow periods); and

Perennial streams becoming ephemeral (due to the loss of baseflows).

These changes are conceptually represented in Figure 3. Storm events, whilst becoming more frequent, decrease in duration, with shorter lag-time and recession (Burns, et al., 2005). Urban catchments are thus typically referred to as ‗flashy‘. Urban catchments feature a reduced capacity to attenuate runoff because storm runoff is quickly routed to receiving waters, indicated by relatively smaller lag times compared to non-urban catchments (Leopold, 1991).

Figure 3. Schematic illustration of the pertinent impacts of urbanisation on hydrology at the catchment scale (Marsalek, et al., 2007).

In addition to its impacts on stormflows, urbanisation impacts the low flow aspects of the hydrologic regime. DeGasperi et al. (2009) showed that in the Puget Lowland (Washington, USA) urbanisation has likely increased the frequency of days in which streamflow is less than 50% of the annual average streamflow. Konrad (2000) reached a similar conclusion, proposing that urbanisation increases the frequency of streamflows below the annual average streamflow. DeGasperi et al. (2009) suggest that the duration of these low flows is decreased following urbanisation.

Appendix 4: Project 4 – Literature Review, 2010 7

The influences of urbanisation on dry weather flows (baseflows) are less clear than they are for stormflows, primarily because of the seasonal variation in baseflow. There appears to be some uncertainty in the literature regarding the impact of urbanisation on summer baseflow. Based on the loss of infiltration due to impervious surfaces, urbanisation should decrease summer baseflow, as reported by Kaufman et al. (2009). However, Konrad (2000) and Konrad and Booth (2005) suggest that urbanisation will not decrease the summer baseflow of perennial urban streams where they are supported by regional groundwater systems. Moreover, Rose and Peters (2001) suggest that summer baseflow may be higher in urban catchments because a reduced volume of water is removed from groundwater storage due to decreased evapotranspiration. In addition, a range of exogenous factors can augment the summer baseflow of perennial urban streams and also increase the probability of observing flow during the dry season in naturally ephemeral urban streams. For example, wastewater discharges (Brandes, et al., 2005; Paul & Meyer, 2001) and septic tank effluent (Burns, et al., 2005) can augment the summer baseflow of perennial urban streams, as can leaking water and wastewater pipes (Brandes, et al., 2005; Konrad & Booth, 2002). There does appear to be more consistency regarding the impact of urbanisation on winter baseflow. Konrad and Booth (2005) showed that with increasing urbanisation, winter baseflow decreased, implying that winter baseflow is less influenced by catchment physiographic characteristics than is summer baseflow. Figure 4 provides a schematic summary of the impacts of urbanisation on flow regimes, using real data from two catchments – Olinda Creek (with very low levels of urbanisation) and Brushy Creek (with moderate levels). The impacts of even this relatively low level of imperviousness on peak flows are obvious, as is the reduced storm recession time and the reduction in both summer and winter baseflow.

Appendix 4: Project 4 – Literature Review, 2010 8

Figure 4. Schematic illustration of the pertinent impacts of urbanisation on catchment hydrology. The abbreviation TI is total imperviousness. The data used is the mean daily streamflow for Olinda Creek and Brushy Creek from 1/06/2002 to 31/05/2003. Catchment area (24.3 km

2 and 14.7 km

2 for Olinda Creek and Brushy Creek respectively) was used to

normalize streamflow (where relevant). The rainfall data used was sourced from gauges at Olinda Creek (229690) and Brushy Creek (229249)(adapted from Burns, et al., in prep), with a climate typical of eastern Melbourne climate conditions.

2.1.3 Hydrologic indicators

Hydrologic indicators provide a way of objectively and quantitatively assessing the degree of impact of urbanisation on the flow regime of receiving waters. They also provide a means of assessing or predicting the waterway health consequences of alternative stormwater management approaches. It is important to distinguish, however between hydrologic indicators for receiving waters and those which might be used specifically for management of stormwater within the catchment. In this section we primarily discuss indicators used for quantifying changes to receiving water flow regimes, whilst in Section 3.2 we outline suitable indicators for use in establishing objectives for stormwater management and stormwater harvesting in particular (at scales ranging from allotment to precinct).

Appendix 4: Project 4 – Literature Review, 2010 9

2.1.3.1 Hydrologic indicators for receiving waters

Research on hydrologic indicators for streams affected by urbanisation is relatively rare, perhaps because restoration of the pre-development flow regime (or something approaching it) has been considered too difficult (Baker, et al., 2004), although this perception has begun to change in recent years (Wenger et al., 2009). Appropriate receiving water hydrologic indicators should be:

1. Ecologically relevant (i.e. they should address the mechanisms of disturbance which are shown to be responsible for ecological degradation of receiving waters).

2. Able to be readily measured and modelled, ideally using currently available modelling tools (although it is important not to let limitations in current modelling capability result in selection of indicators which may not be ecologically relevant)

3. Be able to be applied to a range of catchment sizes and types

4. Be relatively intuitive and able to be readily communicated.

5. Be related to impacts of urbanisation on hydrologic regimes (rather than related primarily to hydrologic regime changes due to extraction or agriculture, for example).

Generally, a hydrologic indicator reveals some information relating to a component of the hydrologic regime. The various components of the hydrologic regime include: volume, frequency, duration, rate of change, flashiness, magnitude and timing. Richter et al. (1996) and subsequent studies (Richter, et al., 1998; Richter, et al., 1997) describe 32 different hydrologic indicators. DeGasperi et al. (2009) adopted these hydrologic indicators to evaluate the hydrologic impacts of urbanisation on streams in the Puget Lowland (Washington, USA). They found that only some were significantly correlated to both benthic index of biological integrity scores and measures of urbanisation (Table 1), making them potentially sensitive metrics of the impacts of urbanisation on receiving water health.

Appendix 4: Project 4 – Literature Review, 2010 10

Table 1. Hydrologic indicators suitable for assessment of urbanisation impacts, as suggested by DeGasperi et al. (2009).

Component of

hydrologic regime

Hydrologic indicator

Definition Significance

Frequency

Low pulse count

A low flow pulse is defined as a daily flow observed which is below or equal to a set threshold (estimated based on pre-developed data). The threshold is set at 50 percent of the long-term daily average flow. This indicator is the number of times in each calendar year that distinct low flow pulses occurred. The units of this indicator are count.

The number of low pulse counts increased for a rapidly urbanizating catchment. Moreover, a negative correlation between benthic index of biological integrity scores and low pulse count was observed. Finally, the number of low pulse counts for a forested catchments was around three times as less compared to urbanised catchments.

High pulse count

A high flow pulse as a daily average flow observed which is equal to or greater to a set threshold (estimated based on pre-developed data). The threshold is set at twice the long-term daily average flow. This indicator is the number of distinct high flow pulses encountered. The units of this indicator are count.

The number of high pulse counts was significantly correlated with 1) benthic index of biological integrity scores and 2) measures of urbanisation. It is suggested that this indicator is not correlated with catchment area.

Flow reversals

This indicator is defined as the number of times that the flow rate changed from an increase to a decrease or vice versa during a water year. The change in flow rate must be greater than or equal to 2 percent. The units of this indicator are count.

The number of flow reverses was significantly correlated to measures of urbanisation and catchment area.

Duration

Low pulse duration

In keeping with the definition of low flow pulse, this indicator is the annual average duration of low flow pulses during a calendar year. The units of this indicator are days.

The number of days between low pulses was significantly correlated to measures of urbanisation, although the correlation was statistically weaker compared to that of high pulse duration.

High pulse duration

In keeping with the definition for high flow pulse, this indicator is the annual average duration of high flow pulses during a water year .The units of this indicator are days.

The number of days between high pulses was significantly correlated to measures of urbanisation and channel slope.

Flashiness

TQmean

This indicator is defined as the fraction of time during a water year that the daily average flow rate is greater than the annual average flow rate of that year. The units for this indicator are fraction of year (dimensionless),

This indicator was more correlated with benthic index of biological integrity scores and less with measures of urbanisation. Also, catchment area was correlated with this indicator.

Richards-Baker Flashiness Index

The Richards-Baker Flashiness Index – it is a dimensionless index of flow oscillations relative to total flow based on daily average discharge measured during the water year. This unit is dimensionless.

This indicator was significantly correlated with benthic index of biological integrity scores and measures of urbanisation.

The choice of indicators requires careful consideration of the catchment characteristics. In particular, catchment size will have an important influence, since, for example, larger catchments are likely to have less variable flows. In other words, care should be given to the use of indicators which are related to variables besides those relating to urbanisation (e.g. catchment area, annual rainfall etc). Similarly, a small upland ephemeral (i.e. with streamflow during only part of the year) stream will have very different ecological dependency on flow than a lowland perennial (permanently flowing) stream. Baker et al. (2004) developed the Richards-Baker Flashiness Index (Equation 1). This hydrologic indicator, measures the degree to which streamflow varies throughout a period of interest. For example, the R-B Index for pristine streams is typically close to zero since streamflow tends to gradually rise and fall throughout any given year. In contrast, the R-B Index for urban streams is usually close to one because the capacity of urban catchments to attenuate surface runoff is low. Indeed, Bressler et al. (2009) showed that the R-B Index is significantly correlated with measures of

Appendix 4: Project 4 – Literature Review, 2010 11

urbanisation. However, it is also correlated with catchment size, given that smaller catchments are more flashy than larger catchments. Equation 1 Where qi is the daily streamflow, qi-1 is the daily streamflow on the previous day and n is the number of days throughout the period of interest. Konrad (2000) proposed the hydrologic indicator TQmean which describes the fraction of time during a water year that the average daily flow rate is greater than the annual average daily flow rate of that year. The value of TQmean is negatively correlated with measures of urbanisation. For example, Konrad (2000) showed that for streams in the Puget Lowland (Washington, USA), the difference between the average TQmean for urban streams and that for non-urban streams, was statistically significant. Furthermore, Booth et al. (2004) showed that benthic index of biological integrity (B-IBI) scores were positively correlated with TQmean. For example, streams with TQmean values of around 0.35 or greater, tended to feature a good diversity of macroinvertebrate species. While hydrologic indicators such as the R-B Index and TQmean describe daily fluctuations in streamflow, other indicators relate more to floods. For example, Konrad (2000) proposes the hydrologic indicator CVAMF, which measures the coefficient of variation of annual flood size, and the Txyr indicator, which measures the frequency of time when streamflow is greater than the flood with a magnitude of 1 in X times per year. This indicator was found by Konrad (2000) to be well correlated with aquatic ecosystem health (as measured by the B-IBI score). Kennen et al. (2010) explored the hydro-ecological relationships of 67 streams in north-eastern USA, using 171 hydrologic indicators to characterize the hydrologic regime of their study streams. Similar to Booth et al, they found that only a small number of these critically influenced ecological condition. These most pertinent indicators included:

The mean of all April flow values over the entire record (this criticality of seasonal flows may relating to the timing of important biological events);

Low flood pulse count. Computed as the average number of flow events with flows below a threshold equal to the 25

th percentile value for the entire flow record (number of events/year);

High-flow frequency. Computed as the average number of flow events with flows above a threshold equal to 75 percent exceedence value for the entire flow record (number of events/year); and

The mean of the minima of all April flow values over the entire record.

In a review of hydrologic indicators, Olden and Poff (2003) suggest that a small number of indicators can robustly characterize a streamflow regime, proposing indicators including pulse count, mean annual runoff, changes in flow and flow minima over given periods. Kennen et al. (2008) explored a range of ecologically relevant hydrologic indicators at 856 monitoring sites in New Jersey, United States. They suggested that the indicator NSTORM (average number of storms per year which produce quickflow) was significantly correlated to macroinvertebrate assemblage structure, due to its impact on habitat and channel form. Steuer et al‘s (2010) study of 83 hydrologic condition metrics took a similar approach, relating the metrics to algal, invertebrate and fish communities across the USA. Whilst a number of metrics were found to be relevant, the frequency of high-flow events (events/month, where an event is defined as when flow exceeds 9x median flow) was found to best explain ecological condition. Not surprisingly, Steuer et al also found that the hydrological indicators were themselves explained by catchment urbanisation. They note, however, that their study of the relationship to catchment land use was limited by not having access to connected imperviousness data rather than total imperviousness.

Appendix 4: Project 4 – Literature Review, 2010 12

In summary, there seems to be a consensus in the literature that the changes to hydrologic regimes which impact receiving water ecosystems are best described by a small number of metrics, including:

Frequency and duration of high-flow events

Flashiness: frequency of changes in flow regime from the long-term average

Mean annual runoff volume

Frequency and duration of low flow events

Whilst catchment-scale indicators are necessary to understand the relationship between hydrology

and receiving water condition, the impacts of urbanisation on catchment hydrology will be managed

across the catchment, at a range of scales. For stormwater managers – and for the designers and

operators of stormwater harvesting systems, there is thus a need to understand how these catchment-

scale hydrologic indicators are impacted by changes to stormwater flows at the site-scale (be it

allotment, streetscape or precinct).

2.1.3.2 Hydrologic indicators at the site-scale

To date, there has been relatively little research on the relationship between hydrological indicators at the site-scale in the urban contexts and receiving water hydrology and ecology. Early studies to quantify the link between stormwater management and receiving waters focussed on physical factors of urbanisation, starting with total imperviousness (Booth & Reinelt, 1994; Schueler, 1994). However, the use of total imperviousness leads to the simplistic and potentially counter-productive approach of determining a threshold above which impacts on receiving waters are likely and then using this as a limit on the imperviousness allowed in a catchment (Pew Oceans Commission, 2003; Schueler & Claytor, 1997). Such a result is only likely to cause urban sprawl and fails to recognise that degradation caused by impervious areas is not an inevitable consequence. Later studies showed that it was effective imperviousness (i.e. impervious areas directly connected to waterways via constructed drainage systems) that better explained these impacts (Hatt, et al., 2004; Lee & Heaney, 2003; Sutherland, 1995). Wong et al. (2000) continued this line of enquiry, comparing the consequences of maintaining natural drainage channels with that of lining or piping some or all drainage channels within the catchment. They concluded that 80-90% of the increase in peak flows is explained by the nature of the drainage connection, rather than simply the proportion of the catchment which is impervious. However, whilst effective imperviousness (EI) might be a useful landscape indicator, it remains an oversimplification, as it categorises impervious areas in a binary manner, as either connected or unconnected. A given impervious area may contribute runoff in some storms and not others. So, runoff after a small storm from a particular non-connected surface may not wet its surrounding pervious surfaces sufficiently to initiate overland flow to the nearest waterway or to result in discharge from the nearest sealed, connected drain (Walsh, et al., 2009). However, from a larger storm, runoff may occur either from rainfall intensity greater than the soil infiltration or from saturation of the soil, and the consequent runoff may be sufficient to form a direct hydraulic connection between the unconnected surface and the waterway. In response to the need to develop a more sophisticated approach, several studies (Ladson, et al., 2005; Ladson, et al., 2006; Walsh, et al., 2005a) identified the frequency of runoff as a useful indicator of stormwater hydrologic impacts, reasoning that in a natural catchment runoff frequency is limited by interception, infiltration and large ‗initial losses‘ (Hill, et al., 1998; Hill, et al., 1996). For example, Walsh et al [, 2005 #62} used empirical evidence to propose that the frequency of stormwater discharge explained variation in ecological condition of 16 catchments. In 2009, Walsh et al refined the runoff frequency metric, proposing retention capacity, which measures the capacity of a given stormwater management measure to reduce the runoff frequency back to the pre-urban frequency. Whilst such frequency-based indicators are easy to understand, measure and model, they measures only one aspect of the hydrological disturbance caused by imperviousness, ignoring changes in volume or changes in low flow hydrology. In 2007, Fletcher et al. proposed a wide range of indicators, in a modelling study of the potential impacts of stormwater harvesting on catchment hydrology. They undertook modelling at a sub-

Appendix 4: Project 4 – Literature Review, 2010 13

catchment scale, testing the effect of various stormwater harvesting scenarios on indicators such as: runoff volume and frequency, spell number and length, and peak flow rates (Table 2). Whilst several of these indicators could be applied at the site or sub-catchment scale, no conceptual or mechanistic link between these indicators and the receiving water indicators is made. Indeed, with the exception of frequency of surface runoff, the indicators are essentially catchment-scale indicators applied at a site-scale. Nonetheless, they provide a potentially useful set from which to choose, given that they include similar indicators to those identified as of priority to receiving water hydrology – frequency of high flows, total runoff volume and frequency and duration of low flows.

Table 2. Hydrologic indicators used by Fletcher et al. (2007) for evaluation of stormwater harvesting impacts.

Category Indicator (and abbreviated name) Analysis timestep

Unit

Runoff Total runoff daily ML/yr

Frequency of surface runoff daily times/yr

Flow Spells

Duration (total time of low flows) daily days/yr

Average length of low-flow spells daily days in a row (average/yr)

Number of low-flow events daily events/yr

Duration (total time of high flows) daily days/yr

Average length of high-flow spells daily days in a row (average/yr)

Number of high-flow events daily events/yr

Peak Flow

Q1month hourly m3/sec

Q3month hourly m3/sec

Q1year hourly m3/sec

Q1.5year hourly m3/sec

Q5year hourly m3/sec

Flow Duration Curve

Integral of the flow duration curve hourly Integral of curve

In the most recent developments in this area (Fletcher, et al., in press; Walsh, et al., 2010), a more integrated set of site-scale indicators (all measured relative to the pre-urban state) has been proposed, including:

1. The frequency of runoff discharged directly to waterway (i.e. the frequency of high-flow disturbance);

2. The volume of subsurface (filtered flow) discharged from the site (i.e. the volume of baseflow contribution); and

3. The mean annual flow volume from the site.

Such an integrated suite of indicators forms potentially a useful basis for developing stormwater management objectives aimed at restoring the post-development hydrologic regime as close as possible to the pre-development level.

2.1.4 Summary

Urbanisation profoundly changes the hydrologic cycle, increasing the volume, frequency and flow rate of storm flows, whilst commonly reducing baseflows through reductions in infiltration. Evapotranspiration is also greatly reduced, decreasing from around 80% of mean annual rainfall in the pre-developed situation to around 15% for impervious areas. The increases in runoff result in increased frequency and magnitude of disturbance to channel substrates and to aquatic ecosystems. Many studies have investigated potentially useful indicators for use in assessing the catchment-scale impacts of urbanisation. Such indicators need to have ecological relevance (i.e. be shown to be related to ecological indicators) and be able to be measured and modelled. They should be sensitive to the urbanisation gradient, rather than external factors (e.g. catchment size). The indicators found to be most useful relate to frequency and duration of high flow events, flashiness, mean annual runoff

Appendix 4: Project 4 – Literature Review, 2010 14

volume and duration and frequency of low flow events. At the site-scale, a different set of indicators will be required; these will need to be suitable for application in assessing the performance of stormwater management (and harvesting) systems. Indicators such as the frequency of discharge from the site, the volume of sub-surface (filtered) flow and the mean annual flow volume from the site, have been proposed as being suitable.

2.2 Water quality

This section begins with an overview of the mechanisms driving pollutant generation and transport in urban catchments. It then reviews typical pollutant concentrations and loads in urban runoff, and compares these to non-urban catchments. Finally, pertinent water quality indicators and objectives

are identified and discussed.

2.2.1 Mechanisms of pollutant generation and transport

2.2.1.1 Major stormwater pollutants

Stormwater contains a wide range of pollutants, which have varying impacts on receiving waters. The pollutants that are of primary interest in terms of their potential to cause significant ecological impacts are (Lawrence & Breen, 2006):

Toxicants (heavy metals, hydrocarbons, pesticides, ammonia);

Nutrients (phosphorus, nitrogen, carbon);

Oxygen-demanding substances (organic material, biological oxygen demand, ammonia, hydrocarbons, sulphides); and

Physical contaminants (suspended solids); sediments may cause problems directly (e.g. smothering of habitats) or through their transport of attached pollutants such as heavy metals.

Issues related to these pollutants include impacts on ecosystems due to toxicants in the water column and sediments, impacts on ecosystems due to nuisance plant growth, asphyxiation of respiring organisms due to depletion of oxygen, modified primary production as a result of light attenuation by particles, and smothering of benthic organisms by sedimentation (ANZECC/ARMCANZ, 2000). The principal sources of common stormwater pollutants are summarised in Table 3.

Appendix 4: Project 4 – Literature Review, 2010 15

Table 3. Typical urban stormwater pollutants and sources (after: Duncan, 1995; Lawrence & Breen, 2006).

Source

Pollutant

so

lid

s

nu

trie

nts

oxyg

en

dem

an

din

g

meta

ls

oils

Syn

theti

c

org

an

ics

atmospheric deposition

plants & plant debris

soil erosion

cleared land

fertilizers

human waste

animal waste

vehicle fuels & fluids

fuel combustion

vehicle wear

industrial and household chemicals

industrial processes

paints and preservatives

pesticides

stormwater facilities

2.2.1.2 Stormwater pollution mobilisation and transport processes

The quality of stormwater is impacted in urban areas by the increased generation of pollutants and also by the hydrologic and drainage changes which result from urbanisation (Taylor, et al., 2005). This second cause is important because attenuation of some pollutants (primarily particulates and adsorbed pollutants) can occur in rainfall which is filtered through porous soils, but the proportion of porous surfaces is diminished in urban areas. As a result, discharges to receiving waters in urban catchments are dominated by runoff from impervious areas (see Section 2.1), where attenuation is much reduced. The mobilisation and transport processes that influence urban stormwater pollution are complex and highly variable, both in space and time (Duncan, 1995). However, the primary processes include:

wet and dry atmospheric deposition;

interception on vegetation and anthropogenic above-ground structures;

buildup of contaminants on impervious surfaces;

washoff from both pervious and impervious surfaces into formed channels and pipes; and

transport along channels and pipes (Duncan, 1995).

The contribution of these processes to stormwater pollution varies according to the pollutant type. For example, dry and wet atmospheric deposition typically supplies as much nitrogen as is washed off in urban runoff (Duncan, 1995). In contrast, dissolution and washoff from roof materials is a major source of copper (Cu) and zinc (Zn) (Duncan, 1995). It is also important to note that buildup of pollutants on urban surfaces is typically very high compared to washoff loads from any one event, suggesting that in many cases the buildup is not a limiting factor in determining washoff loads (Duncan, 1995).

Appendix 4: Project 4 – Literature Review, 2010 16

Initial attempts to predict urban runoff quality commonly focussed on land use (Carpenter, et al., 1998; Soranno, et al., 1996) as an explanatory variable. The extent of impervious surfaces (imperviousness) has also previously been identified as a catchment feature that influences water quality (Arnold & Gibbons, 1996; May, et al., 1997; McMahon & Cuffney, 2000). However, neither of these approaches captures the underlying mechanisms driving the transport of pollutants to receiving waters. For this reason, it was hypothesised that the proportion of impervious areas directly connected to streams via the stormwater drainage system (drainage connection) is more important than the actual impervious areas themselves (e.g. Wang, et al., 2001). Hatt et al (2004) studied 15 small streams across an urban gradient and demonstrated that effective imperviousness (i.e. the proportion of the catchment made up of directly connected impervious areas) better explained variation in concentrations and loads of a number of stormwater pollutants than did total imperviousness.

2.2.2 Effect on pollutant concentrations and loads

In addition to the hydrological impacts described in Section 2.1, runoff from urban areas is generally of poorer quality than that from natural catchments, and carries higher concentrations and loads of pollutants (Duncan, 2006b; Fletcher, et al., 2005). Table 4 summarises typical stormwater pollutant concentrations and contrasts them with runoff from agricultural and forested catchments, as well treated sewage and urban streams. While there is a great deal of variation in urban runoff quality (Leecaster, et al., 2002), which is attributable to factors such as natural catchment characteristics, land-use and climatic patterns (Novotny & Olem, 1994), previous studies have demonstrated a link between pollutant concentrations and the extent of urbanisation. For example, Hatt et al (Hatt, et al., 2004) reported that concentrations of dissolved organic carbon (DOC), total phosphorus (TP), filterable reactive phosphorus (FRP), ammonium (NH4

+), and electrical conductivity (EC) all increased

with urbanisation (measured by imperviousness and drainage connection). Further, they found that loads of TP, FRP, NH4

+, total suspended solids (TSS), total nitrogen (TN) and oxidised nitrogen (NOx)

all increased with urbanisation. This is consistent with the findings of many others (see for example Carpenter, et al., 1998; Chocat, et al., 2001; Soranno, et al., 1996).

Table 4. Summary of typical water quality values for runoff from urban, agricultural and forested catchments, urban streams, and secondary treated sewage.

Variable (mg/L)

Urban runoff

a

Agricultural runoff

b

Forest runoff

b

Secondary sewage

a

Typical urban stream water

qualitya

TSS 250 (13-1620) 186 79 25 2.5-23

TP 0.6 (0.1-3) 0.54 0.072 8 0.02-1.2

TN 3.5 (0.5-13) 3.9 0.83 35 0.39-4.9

NH4+

0.7 (0.1-2.5) 20 0.002-0.16

NOx 1.5 (0.4-5) 10 0.34-3.2

BOD 15 (7-40) 15 1.0-4.0

Cd 0.002-0.05 0.025 0.002 <0.0005

Cr 0.02 0.01 -

Cu 0.4 0.038 0.03 0.001-0.017

Pb 0.01-2.0 0.025 0.02 <0.002-0.024

Zn 0.01-5.0 0.2 0.1 0.009-0.14 a (Lawrence & Breen, 2006); only for cities with separate stormwater and sanitary sewer systems.

b (Duncan, 2006a; Duncan, 1999)

2.2.3 Pertinent water quality indicators / objectives

Stormwater quality targets may take two principal forms; (i) concentration targets or (ii) load targets. In Australia, loads are most commonly used in setting water quality targets, such as those proposed in Chapter 1 of Australian Runoff Quality (Table 5).

Appendix 4: Project 4 – Literature Review, 2010 17

Table 5. Example of pollutant load reduction targets (Source: Wong, 2006).

Pollutant Stormwater Treatment Objective

Suspended Solids (SS) 80% retention of the average annual load

Total Phosphorus (TP) 45% retention of the average annual load

Total Nitrogen (TN) 45% retention of the average annual load

Litter Retention of litter greater than 50 mm for flows up to 25% of the 1-year ARI peak flow

Such targets are widely applied in Queensland, Victoria, and New South Wales. Loads targets can be readily reported against using currently available modelling tools such as MUSIC (Gold Coast City Council, 2006) and have been considered as useful given their integrative nature. However, whilst loads-based treatment targets are suitable for lentic waterways such as lakes, estuaries and bays which have a high buffering capacity due to their large volumes, they are not appropriate for smaller flowing (lotic) waterways (Taylor, et al., 2005; Walsh, et al., 2004). Indeed lotic systems such as small streams are far more sensitive to temporary spikes in stormwater pollution (Roesner, 1999; Walsh, et al., 2005a), thus concentration-based treatment targets would be more appropriate for these receiving waters. In Australia many concentration targets exist, such as the national ANZECC guidelines for freshwater and marine waters (ANZECC/ARMCANZ, 2000). Similar concentration targets exist for most Australian States and Territories. They are generally derived from an assessment of receiving water requirements and are typically risk-based, in that they specify maximum permissible probabilities of a given threshold concentration (e.g. median concentration of TSS not to exceed 20 mg/L). Such guidelines use trigger levels, below which there is a low risk of harm to the ecosystem. When trigger levels are exceeded, further investigation to understand ecosystem response is required. It is possible that concentration targets may not all be achievable using current stormwater treatment technologies (Hatt, et al., 2009; Walsh, et al., 2010). In this case, a compromise ―best practical treatment‖ target may need to be considered as an interim measure only. It is thus critical that appropriate indicators and targets be chosen for the particular receiving water of interest, with concentration targets used for lotic receiving waters and loads targets used for large, well-buffered systems. This also means that where stormwater is to be infiltrated, it should be required to meet water quality objectives to ensure that groundwater is not polluted. As a starting point, the same targets as applied to surface receiving waters could be used. While imperviousness is an important source of pollutants, the main driver of pollutant transport in urban catchments is the drainage system, which efficiently delivers runoff and pollutants to receiving waters.

2.2.4 Summary

Like hydrology, changes to water quality are a ubiquitous response to urbanisation. A wide range of pollutants, including toxicants, nutrients oxygen-demanding substances, and of course sediments, are likely to occur in elevated concentrations and loads in catchments subjected to urbanisation. The increased concentrations and loads in waterways results from both increased generation (through land use activities) and mobilisation/transport processes (due to the creation of hydraulically efficient drainage networks). A number of studies have shown that pollutant concentrations are well predicted by impervious areas, but importantly, Hatt et al‘s (2004) study showed that the effective imperviousness of a catchment gave the best prediction of a wide range of pollutants, indicating the direct role which stormwater discharge has on receiving waterway water quality. A new approach to the design of urban drainage systems is thus required if we wish to avoid water quality degradation due to urbanisation. The selection of water quality indicators must take into account the nature of the waterway. For large systems, loads targets provide a suitable measure, whilst concentration targets (such as those

Appendix 4: Project 4 – Literature Review, 2010 18

provided by ANZECC or local SEPPs) provide the most ecologically-relevant target for smaller flowing waterways, which have limited buffering capacity.

2.3 Geomorphology

2.3.1 Impacts of urbanisation on sediment processes

The focus of urban impacts on stream channels has firmly been on water quality and hydrologic alteration and the role of changes in sediment yield. The impact this has on channel degradation or ecological functioning has been largely ignored. Alterations to hydrology of streams in urban catchments, and of the catchment landscape, are accompanied by altered sediment dynamics (Wolman, 1967). Sediment dynamics, which incorporate the spatial and temporal presence and variation in constituents of the bed and bank materials, is an ecologically important aspect of stream health. It is well established that reduced sediment supply to a stream increases the energy available to degrade the channel for a given flow, or alternatively stated, stream energy increases as sediment supply decreases (Schumm, 1977). Sediment budgets in the urban environment are relatively poorly understood, with the majority of research in this field focused on forested and rural catchments (Nelson & Booth, 2002). Sediment yield from an urban catchment is generally considered to decrease, although these conclusions refer to coarse-grained sediments (bed load) more commonly than fine-grained suspended sediment (Bledsloe, 2002; Gurnell, et al., 2007). Walsh et al. (2005b) highlighted the inconsistency across studies in suspended sediment yield response following urbanisation. The prevailing model of sediment dynamics related to land use is that of Wolman (1967), who conceptualised the changes in sediment delivery to streams relative to catchment land-use change (Figure 5). He found catchment-cover change from forest to crops resulted in significant increases in sediment load to receiving streams. The most dramatic change in sediment dynamics resulting from land-use change results from construction impacts of a developing catchment (3 to 5 times higher; see Keller, 1962), although this is likely to overestimate the contemporary contribution of suspended sediment to the stream system during construction due to improvements in construction practice. Most pertinently Wolman (1967) suggested that sediment yields post-construction phase can be lower than for the intact catchment, though, the findings for the established urban phase were the least compelling, based predominantly on images of drain sedimentation.

Figure 5. Conceptualised change in sediment delivery to streams relative to catchment land-use (after Wolman 1967). Note the uncertainty regarding the urban phase (circled).

Gurnell et al. (2007) suggested that reduction in sediment supply in urban environments is a result of erosion-resistant sealing of catchment surfaces. This theory, when comparing urban to the former agricultural land use, is intuitively convincing considering the reduced areal extent of exposed soils and reduced potential for sediment liberation, in comparison to the increased frequency and volumes of runoff.

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The erosion of riverbanks commonly increases post-urbanisation and this can be considered an important source of sediment for urban rivers (Gurnell, et al., 2007; Trimble, 1997; Wolman, 1967). Trimble (1997) estimated that channel erosion provides about two-thirds of the total sediment yield from an urban catchment. However, widespread reinforcement of river banks in the laterally constrained urban setting often reduces this sediment source (Gurnell, et al., 2007). In the Issaquah Creek catchment, Washington, Nelson and Booth (2002) found development to almost double sediment production even though relatively little sediment was liberated directly from the urban areas. The increase was primarily attributed to sediment production resulting from discharge-induced channel erosion (20% of the total sediment budget). However, the approach used to determine pre- and post-urbanisation channel capacity, namely an empirically derived relationship between channel capacity and the 2 year ARI, is simplistic (as acknowledged by the author). Other approaches are discussed in this section. The increased sediment transport capacity resulting from conventional approaches to managing stormwater (and its subsequent increased runoff volume) has been well demonstrated (Bledsloe, 2002; Grove & Ladson, 2006; Pomeroy, et al., 2008). Of the three phases of sediment movement in the urban environment; supply (source), transport (and erosion) and deposition, the latter is the most poorly understood. This, no doubt, has much to do with the absence of sediment deposition to inspire these studies and the prevalence, and immediately relevant concerns, surrounding erosion. Issues result from both decreased sediment loads (e.g. accelerated clearwater erosion) and increased sediment loads (e.g. smothering of substrates, Wood & Armitage, 1997). In addition, better understanding the dynamics of sediment deposition in urban stream is particularly pressing particularly when compared with rural streams, because fine-grained sediments play an important role in the storage and transport of contaminants. There are differing findings on the effects of urbanisation on sediment calibre but in general a progression towards finer sediments (‗sediment fining‘) is observed. Booth and Jackson (1997) suggested that this fining with urbanisation is explained by the dominance of overland flow driving increased sediment transport. This is supported by Gurnell et al. (2007), whereas Pizzuto et al. (2000) found that gravel-bed urban streams, when compared with rural streams, were lacking the finer particles ranging in size from sands to pebbles which they suggested had been selectively removed, resulting in coarsening. This may suggest a bimodal distribution following the adjustment of streams to urbanisation: increased clays/silts, decreased sands and gravels, and a dominance of coarser material (cobbles, boulders) where present. Interestingly, overall increases in sediment size have been observed due to intervention activities (Grable & Harden, 2006), which is discussed further in the following section. Irrespective of sediment size it appears the mobilisation effectiveness of stormwater flows from urban catchments cannot be discounted. Booth and Henshaw (2001) suggested that no unconsolidated sediments in urban stream channels are immune from disturbance given adequate upstream urbanisation.

2.3.2 Impacts of urbanisation on channel morphology

It is well established that river channels adjust to the flow and sediment regimes received (Leopold & Maddock, 1953; Leopold, et al., 1964; Wolman & Miller, 1960). The adjustment is via lateral migration or incision, or a combination of both (Bledsloe, 2002). Hydrologic changes associated with urban catchments tend to lead to increased rates of lateral migration where a channel is not constrained (Nelson, et al., 2006). Indeed, Elliott et al. (2010) found urbanisation to increase the erosion potential index of a stream by a factor of between 1.6 and 9.3. In theory, as a stream adjusts over time to the hydrologic and sedimentologic inputs it should be capable of reaching a relative stable phase without intervention. In reality, many stream responses are non linear and, irrespective of the stability of the inputs with time, the channel remains dynamic (Rhoads, et al., 2008). Stream states have been referred to as ‗dynamic equilibrium‘; moving between phases of erosion and deposition (Hack, 1960); or ‗quasi-equilibrium‘: whereby channels undergo adjustment to the range of discharges and sediment loads experienced and continue to change

Appendix 4: Project 4 – Literature Review, 2010 20

capacity or shape only slightly and slowly over periods of time (Leopold & Maddock, 1953). Some terms such as ‗dynamically stable‘ (Rhoads, et al., 2008) are truly oxymoronic. A ‗metastable‘ state has been used to define riverine ecosystems (Thorp, et al., 2006) and this may be an appropriate way to consider geomorphic functioning of urban streams. Metastability is a long lived unstable or transient state. In essence, the stream can exist in a number of ‗stable‘ states and hover between those states for long periods of time depending on the conditions, changing readily between a more stable or less stable condition. While semantics can often become tiresome there is a need to define, at least conceptually, the desires for dynamism when considering river management in a physically constrained environment. In an urban setting attainment of any form of ‗equilibrium‘ or ‗stable‘ state is most challenging for two reasons: a lack of adjustment room, a.k.a. riparian ‗buffer‘; and the magnitude of channel adjustment required to cater for the flow volume and hydrologic regime delivered to the channels through stormwater systems. Whether urban stream channels will ever achieve a quasi-equilibrium appears to still be an open question (Grable & Harden, 2006). Notwithstanding, Booth and Henshaw (2001) found that the age of the upstream development was closely related to the rate of channel change. Streams are most stable downstream of more established developments. They suggested the ‗reason for this influence was enigmatic‘, yet, it seems quite plausible that (as they further suggest): a) the longer the period of adjustment the closer to a state of quasi-equilibrium, b) erodible sediments are likely to have been removed and low-erosivity sediments remain, and c) undergo a stabilising influence from cementation of suspended sediments and regeneration of vegetation.

2.3.2.1 Changes in channel capacity

Wolman (1967) suggested that enlargement of urban channels is likely where an increase in discharge accelerates erosion, and a decrease in available sediment is unable to ‗keep pace‘ with the erosion as it might under a normal sediment supply regime. Indeed, urbanisation is commonly associated with channel enlargement, channel widening and channel deepening (Booth & Henshaw, 2001; Booth & Jackson, 1997; Doll, et al., 2002; Grable & Harden, 2006; Gregory, et al., 2002; Grove & Ladson, 2006). Grable and Harden‘s (2006) data clearly identified the dominance of erosion over deposition throughout an urban catchment. Yet, there is also wide variability in these responses (Gurnell, et al., 2007) and in cases little to no change has been observed (e.g. Nelson, et al., 2006). As highlighted by Booth and Henshaw (2001) studies have also identified channel width reductions following urbanisation, but these are rare. Booth and Henshaw (2001) suggested there is a potential for research bias through focusing on degrading channels. Attention is most commonly, and not surprisingly, paid to particularly dramatic channel changes during urbanisation of a catchment. Geomorphologists are inherently interested in measuring change. Channel capacity has been found to increase by a factor of at least 2: Pizzuto et al. (2000) found an increase of 2.3 times; Gregory et al. (2002) 2-2.5 times; and MacRae (1997) 4.2 times. These increases in capacity are not uniformly distributed by bed and bank. Pizzuto et al. (2000) found capacity increase solely resulted from widening, attributed, in part, to armouring of bed sediments (a common geomorphic response to the removal of the finer-grained sediments). Channel incision is a well known geomorphic response to either increased flow, decreased sediment load, an oversteep channel gradient, or decreased calibre of sediment inputs (Lane, 1955). Incising channels tend to degrade vertically, lowering the bed, prior to a phase of lateral degradation, eroding the channel banks and widening the channel. This process can be cyclical. The classic representation of channel incision is provided by Schumm (1984), as shown in Figure 6. Where flow and sediment regimes are not conducive to recovery, the return of an inset channel (Type V) is unlikely.

Appendix 4: Project 4 – Literature Review, 2010 21

Figure 6. Channel incision and recovery after Schumm et al (1984). a) temporal view, b) spatial view

Susceptibility of streams to urbanisation varies. Characteristics such as slope, substrate, riparian vegetation cover and presence of bedrock or man-made hydraulic controls influence the robustness of channel morphology (Booth & Jackson, 1997). Steep channel slopes in themselves are not are determinant of the susceptibility of urban channels to incision, but Booth and Henshaw (2001) suggested that steeper gradients may increase the magnitude of change: channels with slopes > 4 percent exhibited the largest changes (> 0.3 m/yr). They suggested that susceptibility is particularly dependent on substrate geology.

2.3.2.2 Modes of channel widening

The three commonly cited modes of bank erosion are appropriate to the investigation of urban waterway degradation: Fluvial scour removes individual sediment particles or aggregates by water flow. This occurs when the force applied to the bank by flowing water exceeds the resistance of the bank surface to these forces (Abernethy and Rutherfurd, 1999). The removal of bank material is therefore closely related to near-bank velocity conditions and in particular to the velocity gradient and turbulence close to the bank, which determines the magnitude of hydraulic shear (Knighton, 1998). The process of scour and fill has been stated as contributing to the adjustment of size and shape in natural channels (Leopold et al, 1964), though, with increased flow, decreased sediment supply and incised channels containing flow the dominance of scour over fill is likely to be significant. Mass failure or slumping occurs when large segments of the bank break off through the process of erosion. This process is generally triggered when a critical stability condition is exceeded, either by reduction of the internal strength of the bank, due to sub-areal preparation, or a change of river geometry, commonly through fluvial scour (Abernethy and Rutherfurd, 1999). Therefore the susceptibility of river banks to mass failure depends on their geometry, structure and material properties (Knighton, 1998). The collapsed blocks produced by mass failures may break on impact and be removed or they may remain intact to be eroded by hydraulic action, sometimes protecting the lower bank from further erosion (Knighton, 1998). A number of factors influence erosion by mass failure including: bank and material composition, climate, subsurface conditions, channel geometry and bioperturbation. In particular, flow characteristics, such as rates of fall, play a considerable role in the potential for mass failure. Sub-aerial preparation (drying and desiccation) occurs when bank areas are exposed to air (above the waterline). This includes piping, rain splash, rill erosion, stock trampling and desiccation (Abernethy and Rutherfurd, 1999). Cycles of wetting and drying are especially important as they cause swelling and shrinkage of the soil, leading to the development of fissures and tension cracks which promote failure (Knighton, 1998). Desiccation causes extremely dry and cracking bank material which is highly erodible (Abernethy and Rutherfurd, 1999).

Appendix 4: Project 4 – Literature Review, 2010 22

The proportional role of each of these erosion mechanisms in urban bank failure is poorly understood. Most importantly, the relationship between bank erosion mechanisms and components of the urban stormwater flow regime has not been specifically demonstrated. MacRae and Rowney (1992) summarised a number of process-form investigations to suggest fluvial scour at the bank toe followed by mass failure was most common. Indeed, the mechanism of fluvial scour is most commonly used to determine erosion potential indices (Elliot, et al., 2010), as discussed in Section 2.3.3.

2.3.2.3 Loss of channel diversity

Urban development has been found to reduce channel morphology to a uniform or simplified form (Bernhardt & Palmer, 2007; Booth & Henshaw, 2001; Booth & Jackson, 1997; Gurnell, et al., 2007). A simplification of channel morphology, such as the loss of bars and benches, is associated with the overall widening and deepening (Booth & Henshaw, 2001). The cause of which can be attributed to the loss of hydraulic diversity and reduction in mobile sediments as transport capacity exceeds yield. At a finer scale reduction in mobile sediments impacts on substrate diversity and hyporheic exchange, which plays an important role in the chemical and biological functions of streams (Ryan & Boufadel, 2007). The physical structure of the stream can affect how a flow translates to an important ecological event such as a disturbance (Poff et al. 2010). Booth and Henshaw (2001) highlighted the impact of morphologic simplification on biological and aesthetic values and demonstrated that ‗channel instability does correlate with low quality habitat’ (p. 19). The role of wood and other roughness elements in maintaining the diversity and stability of a channel through the provision of roughness cannot be underestimated. The removal of wood from an otherwise stable stream may be the tipping point for catastrophic channel change (Booth & Henshaw, 2001). While channel stability is often the aim of the river engineer it generally competes with the desires of the river manager or geomorphologist sympathetic with ecological functioning. The attainment of quasi-equilibrium may not in fact be the ideal state for an urban stream where some semblance of ecological functioning is desirable. Booth and Henshaw (2001) highlighted that the re-equilibration of a channel following major incision does not necessarily coincide with the return of improved habitat quality.

2.3.2.4 Stream modification

Increased runoff and potentially decreased sediment from urban environments should, in theory, result in greater channel erosion and an increase in channel width (Wolman, 1967), though, many urban channels are not permitted to proceed through these phases. A common response in the urban environment by stream managers or engineers is to control channel adjustment through the use of ‗hard engineering‘ such as rock and concrete bank lining and bed control. Planform change is often the result of interventions such as meander cutoffs or channel straightening (Nelson, et al., 2006; Rhoads, et al., 2008). Historically, obvious benefits to this approach have been highlighted including: increasing hydraulic efficiency (hydraulic conveyance), reducing erosion and reducing maintenance (Wolman, 1967). This, however, often only transfers the problem, and channel straightening inevitably leads to increased hydraulic gradient and an erosional response (Rhoads, et al., 2008) The concept of an ‗equilibrium‘ stream implies stability in elevation, gradient and channel form (Wolman, 1967). This concept might be appealing from an engineering viewpoint but ignores many of the aspirational goals for a healthy stream ecosystem. Modes of adjustment, including the presence of erosion and substrate sediment dynamics, can be important elements of a healthy stream (Florsheim, et al., 2008; Wolman, 1967). The ecological and biological environment, adjusting to the input variables, is unlikely to be dynamic if stream morphology is moribund. Stabilised urban channels have been found to exhibit demonstrably lower habitat values (Gurnell, et al., 2007). The role of engineering protection in urban streams is often considerably more influential than hydrologic or sedimentologic change. The geomorphic response of a stream to changed flow and sediment inputs is often moot if protection prevents this adjustment. The vexing problem is that to

Appendix 4: Project 4 – Literature Review, 2010 23

satisfy many of the concerns associated with urban streams, such as flooding and effects on infrastructure, stream protection is often required. Until some of the geomorphically effective flows are reduced increasing knowledge of stream channel adjustment is of little use. Physical interventions in urban streams may be responsible, in part, for the poor correlations between stormwater regimes and channel degradation. Numerous authors highlight the role of grade controls and within-channel protection in increasing the complex response of streams to urbanisation (Booth & Henshaw, 2001; Grable & Harden, 2006; Gurnell, et al., 2007). In particular Grable and Harden (2006) describe the concept of ‗decoupling‘ for an urban stream in Tennessee, USA. Artificial structures such as bedrock controls and culverts allow for flow and fine-grained sediments to pass between reaches but not bedload or channel adjustment ‗information‘, upstream or downstream. This information includes knick points, ‗sediment waves‘, or channel adjustments. The role of decoupling in channel degradation is supported by Jordan et al. (2009) who found that structures within the channel disrupt the bed load sediment continuity resulting in long-term channel instability. In understanding the relative role of urbanisation in shaping urban channels the modifying role of preparatory land uses cannot be discounted. Channel incision is often ongoing, and related to urban stormwater as a driver, but these adjustments may be minor in comparison to channel degradation prior to urbanisation. Catchment clearing, and clearing of the riparian zone, encourages channel incision through altered streamflows and reduced structural stability via vegetation. Incision processes depend on land use and the hydrogeology of the channel in question. For poorly drained riparian lands physical interventions such as channel dredging confines flow, decreases infiltration and reduces flow attenuation to the trunk channel. Direct intervention in the channel with the express intention of drainage ‗efficiency‘ leads to an enlarged and straightened stream. Increased containment of flow leads to higher stream powers greatly exacerbating the already increased energy from stormwater runoff. For well drained land, commonly associated with steeper terrain, stream modification is most commonly related to flow detention through the construction of online storages. This often leads to an alternating cut and fill channels. The potential for abundance and diversity of bed and bank habitats in urban streams, and the relationship of these features to mobile channels, supports efforts to focus on urban runoff to streams (the cause) rather than reactively treating the stream (the symptom).

2.3.2.5 Stormwater as a driver of channel change

Flow alterations due to stormwater, as described in Section 2.1, increase the quantity and rates of runoff. With increased ‗flashiness‘ the frequency of events exceeding the threshold for erosion is greater. The hydrologic challenge most commonly pursued is to decrease the duration of events which exceed an erosive threshold. Theoretically, power functions relating the overall discharge in a river channel to river dimensions (Leopold & Maddock, 1953) clearly suggest that for increased discharge (up to 10 times the volume) there will be a corresponding increase in channel width and depth. EarthTech (2006) highlighted the increase in energy expenditure (not excess energy) in an urban stream, Little Stringybark Creek, based on standard urban stormwater design. Peak total energy expenditure was found to be bimodal: greatest during low flows (0.1 m3/s) and medium frequency events (1 in 4 yr ARI approx.) and generally 4 to 5 times greater than for natural conditions. WSUD was found to return total energy expenditure to levels similar to a condition closer to that of a forested/cleared catchment. For ease of extrapolation urban stream degradation is most commonly related to a surrogate measure of flow hydrology: Total Imperviousness Area (TIA) or more appropriately Effective Impervious Area (EIA). Based on an empirical study nearby Seattle, Washington, Booth and Jackson (1997) found that 6 percent EIA for a watershed was a threshold beyond which the receiving channel was significantly wider. Moving beyond a static consideration of the impacts of effective impervious area Booth and Henshaw (2001) analysed an 11 year data set to identify EIA relative to the rate of channel change. They found that the rate of channel change is poorly correlated with EIA. Some sites with up to 40 percent EIA

Appendix 4: Project 4 – Literature Review, 2010 24

experienced moderate to minor change while another with as little as 3 percent experienced very large change. This was attributed to the dominance of the vagaries of local conditions over hydrologic processes. The important role of local geomorphic conditions for sensitivity to erosion is often surmised (Bledsloe, 2002; Pomeroy, et al., 2008). The relative sensitivity of streams to urbanisation propagates the uncertainty surrounding response. Not surprisingly, streams in western Washington‘s granular hillslope deposits considerably greater change occurred compared with those in cohesive silt-clay deposits (Booth & Henshaw, 2001). In addition to local conditions, the poor correlation between rate of channel change and EIA may represent stabilisation of channels over time, or the inadequacy of EIA as a surrogate measure of the impact of urbanisation on channel morphology. Channel capacity and erosion of urban channels are most commonly related to the 1 in 2yr ARI event (Bledsloe, 2002; MacRae, 1996; Nelson & Booth, 2002). This appears to be a relic of the early findings of Wolman and Miller (1960). MacRae and Rowney (1992) found channel erosion, based on scour potential, following urbanisation to be better associated with moderate flow events: sub-bankfull flows with recurrence intervals between the 1 in 0.5 to 1 in 1.5 year ARI. The greatest sediment transport potential was found by MacRae and Rowney (1992) to occur at moderate depths: 0.5 to 0.85 of bankfull depth. The greatest increase in scour from pre to post-urbanisation corresponded to flow events less than 0.7 bankfull depth. They concluded that urbanisation shifts channel forming dominance from bankfull flows to more frequent smaller events.

2.3.2.6 Geomorphic thresholds and current stormwater management approaches

Efforts to reduce the hydrologic impact of urbanisation are most commonly reliant on detention basins and flow control ponds connected to stormwater pipes (Booth & Henshaw, 2001; Elliot, et al., 2010). These approaches often do not address the increased volume and are rarely effective in reducing the duration of geomorphically effective flows resulting from an urban catchment (Elliot, et al., 2010; Pomeroy, et al., 2008); indeed, whilst reducing the peak discharge rate, they may actually result in an increase in the duration of elevated flows. For example, Booth and Henshaw (2001) investigated the Timberline tributaries, King County, Washington, downstream of intensive urban development and the tributary with a detention basin exhibited the most drastic channel change. Clearly, the role of detention basins in prolonging geomorphically effective flows requires further investigation. Urban flow attenuation is commonly, particularly in the United States, threshold based. The ‗duration standard‘ for detention basin design is aimed at maintaining the aggregate of post-development flows at or below a threshold for sediment mobility (Booth & Jackson, 1997; Pomeroy, et al., 2008). Common practice is to relate channel erosion to hydrologic thresholds for sediment mobility to determine ‗geomorphically effective‘ or ‗geomorphically detrimental‘ flows, e.g. one half of the 2 year pre-development flow: Booth and Jackson 1997, using excess energy expenditure or erosion potential (EP) (Bledsloe, 2002; Booth, 1990; Elliot, et al., 2010; Grove & Ladson, 2006; Pomeroy, et al., 2008; Tilleard & Blackham, 2010). The erosion potential index is continuously simulated to assess ‗work done‘ on the channel above the critical shear stress (Bledsloe, 2002; Pomeroy, et al., 2008). Although, critical shear stress values are rarely field verified (e.g. Grove & Ladson, 2006; Pomeroy, et al., 2008; Tilleard & Blackham, 2010) often relying on the flume study results of Chow (1959). Furthermore, the concept that stream power can be used to determine urban stream stability has been refuted by Stacey and Rutherfurd (2007) based on a study of over a 1000 sites in Fairfax County, Virginia, USA. They found that substrate type provided a statistically greater influence on stability , but could not define stability (Stacey & Rutherfurd, 2007). Nevertheless, deterministic studies of excess energy or EP provide indications of the impacts of conventional stormwater relative to WSUD techniques. An analysis of EPI for a hypothetical development on a ‗geomorphically sensitive‘ stream in Victoria found that urban development led to a 30 percent increase in EP compared with an increase of 10 percent when WSUD was employed (Tilleard & Blackham, 2010). As noted this suggests the stream is still likely to erode with WSUD implemented. However, this approach presumably also allows the dimensions of the hypothetically stable channel to be determined. A ‗peak standard‘, or ‗peak shaving‘ approach, whereby peak flows pre and post are matched, has been found to be considerably less effective in reducing elevated sediment transport rates because

Appendix 4: Project 4 – Literature Review, 2010 25

flows may still exceed the threshold for movement over longer durations (Bledsloe, 2002; Booth & Jackson, 1997).

2.3.3 Pertinent geomorphic objectives and indicators

2.3.3.1 Geomorphic objectives

Available habitat is often seen as a limiting factor for urban stream health and physical integrity is seen as the fundamental scale on which to base river rehabilitation (Findlay & Taylor, 2006). The desire of the geomorphologist considering urban streams is to allow for the greatest amount of ‗natural‘ (stream controlled) adjustment within the constraints of the urban environment. This includes providing ample buffers (Coleman, et al., 2005) and realising erosion as a positive attribute of healthy streams (Florsheim, et al., 2008). Florsheim et al. (2008) concludes that bank erosion is integral to the functioning of river ecosystems. These notions require a paradigm shift in current thinking on urban stream management most often based on a goal for stream stability. Some of the challenges in implementing, and determining the geomorphic success of appropriate rehabilitation, are captured by Rhoads et al. (2008):

Investment in initiatives to improve environmental attributes of the streams are deemed worthwhile as long as the resulting project does not increase risk for channel instability (erosion and sedimentation problems) or for flooding. (p. 225)

The four key contributions which fluvial geomorphology can make in this regard are highlighted by Gilvear (1999, p. 229):

1. to promote recognition of lateral, vertical, and downstream connectivity in the fluvial system and the inter-relationships between river planform, profile, and cross-section;

2. to stress the importance of understanding fluvial history and chronology over a range of time scales, and recognizing the significance of both palaeo and active landforms and deposits as indicators of levels of landscape stability;

3. to highlight the sensitivity of geomorphic systems to environmental disturbances and change, especially when close to geomorphic thresholds, and the dynamics of the natural systems; and

4. to demonstrate the importance of landforms and processes in controlling and defining fluvial biotopes and to thus promote ecologically acceptable engineering.

The challenge is to appropriately describe the geomorphic response of urban streams to altered stormwater and sediment inputs, reducing uncertainty and quantifying risk. As identified by Rhoads et al. (2008):

Uncertainty itself derives largely from a lack of information on how the streams might respond to naturalization. Alleviation of this uncertainty involves the implementation of various evaluation methods to generate information on potential or actual geomorphological performance. (p. 225)

A dilemma facing the geomorphologist rehabilitating an urban stream is the appropriate level of intervention. The geomorphic form and function of streams cannot be ‗recolonised‘ at the same rate nor the same extent, as the biological condition can under ideal circumstances. Physical form can either be reinstated through intervention, e.g. physically recreating pools, bars and riffles using machinery, or, with appropriate hydrologic and sediment inputs e.g. allowing for self adjustment. While the former is an immediate response, the period of response for the latter, even with appropriate inputs, could be measured in decades rather than years (the latter comparable to biota recovery times with favourable conditions). This raises the question should we intervene to expedite the recovery process? According to Rhoads et al. (1999) rehabilitation programs should focus on creation or naturalisation in order to improve the health and value of a system. Newson (2002) suggested intervention was often necessary: ‗assisted natural recovery‘.

Appendix 4: Project 4 – Literature Review, 2010 26

2.3.3.2 Managing perceptions of dynamic streams

Findlay and Taylor (2006) highlighted that the primary concern of the community is recreational, aesthetic and civil aspects (e.g. flood mitigation) of urban streams. The provision of a ‗naturally‘ functioning system, or ecological health, rarely enters the conscience of the general populous. The urban geomorphologist‘s role is a particularly challenging endeavour to convince engineers and the community of the benefits of a dynamically functioning channel. In particular, geomorphic dynamism can co-exist with flooding and maintenance concerns if appropriately understood and implemented and erosion is a natural and ecologically beneficial process (Florsheim, et al., 2008). The success of appropriate rehabilitation ‗will be difficult unless public expectations are tailored to an appropriate understanding of stream dynamics‘ (Rhoads, et al., 2008 p. 227). The steps to stakeholder confidence include highlighting the values of geomorphically functioning streams and improved understanding of channel response in urban environments so that outcomes and associated risks can be practically assessed.

2.3.3.3 Geomorphic indicators: monitoring and geomorphic metrics susceptible to change

One of the challenges for the geomorphologist is the paucity of data collected on physical aspects of stream channels, particularly in comparison with discharge and water quality data availability (Grable & Harden, 2006). There appears little need for monitoring protocols specifically tailored to urban streams as inputs (sediment and water) and degrees of freedom (width, depth, planform etc.) are comparable to non-urban streams. The most considerable differentiator is the scale of modification to inputs. In this regard the rates of change should be greater. The constraints on degrees of freedom, such as rock protection, have obvious implications for channel stability but, if suitably sized by the engineer, there is little potential for channel change in these locations. In these locations ‗environmental‘ monitoring is a fruitless exercise merely demonstrating the river engineer‘s ability to design for a rare event. A number of approaches are available to the geomorphologist monitoring geomorphic change in urban streams over timescales which are short term (<1 year), medium term (<= 5 years) and long term (> 5 years). For example, short to medium term change may be detected through: deposition mapping and suspended sediment concentrations; and medium to long term changes may be detected through: repeat cross sections and planform alignment mapping. Where cross sectional or alignment comparisons are to be made the importance of permanent benchmarks and accurate survey techniques cannot be over emphasised. Channel changes are often barely discernible from potential operator or equipment errors. Monitoring geomorphic change must be spatially targeted to those sites in which change is likely within management timeframes. Timeframes for geomorphic adjustment, even within urban streams, are particularly greater than that experienced for ecological attributes. Geomorphic change is also highly dominated by thresholds (Booth & Jackson, 1997; Schumm, 1977) and as such a linear response is rare.

2.3.4 Summary

In summary, while impervious cover and channel degradation are often correlated, altered stormwater inflows are cited as the driver of physical form degradation in urban streams (e.g. Grable & Harden, 2006; Walsh, et al., 2005b). In order to maintain (or reduce) the channel capacity of urban streams, and increase the potential to reach a quasi-equilibrium, reductions are required in the overall stormwater volumes to the channel. Reductions are also required in the frequency and duration of scouring flows which have been found to be more prevalent in urban stormwater regimes (MacRae & Rowney, 1992). Returning some of the physical components of physical habitat and a healthy stream, such as benches, bars, and spatially variable and temporally mobile substrates, is not feasible unless the components of the stormwater flow regime are amenable. Most importantly, the focus of the river manager should move beyond reactively patching the symptom (bank and bed erosion) rather than pursuing a solution to the problem (an increase in geomorphically effective flows).

Appendix 4: Project 4 – Literature Review, 2010 27

Developing predictable relationships between urban hydrology and channel form and function is an extremely vexing problem. There is a paucity of medium to long term (5 to >10 year) datasets. Local conditions such as geology, protection works and riparian vegetation have been found to greatly confound geomorphic responses to urban hydrology. The ability to distil stormwater flow regimes to the ‗problem‘ components in geomorphic terms remains limited. To identify the role urban stormwater management can play in restoring the geomorphic functioning of streams, and hence river health, we need to consider the alterations to the hydrologic and sediment regimes in concert with changes to channel morphology. In particular, sediment regimes in urban environments and the impact on stream morphology are poorly understood. In addition, we must be mindful of the explicit impact of urbanisation and disentangle relic impacts from prior land disturbances such as rural development, the impacts of which may be realised for some time. Physical intervention (e.g. channel widening and bed and bank protection) during the rural development and urban phases plays a major role in the geomorphic and hydraulic functioning of a channel. In addition to the physical sciences there are further social and institutional challenges faced by the geomorphologist. If ecological health of urban streams is a priority then there are essentially three steps in the process for geomorphic recovery of an urban stream: 1) - Convincing river managers, engineers and applied geomorphologist, who have historically been focused on channel ‗stabilisation‘ in urban settings, that ‗naturally‘ functioning stream morphology leads to ecological health; 2) - Determining the level of intervention, and the riparian land required to return geomorphic functioning to the urban stream, or in the case of mildly impacted streams whether ‗assisted natural recovery‘ (Newson, 2002) will suffice; and 3) - Determining the feasibility of returning the flow and sedimentologic regime required to facilitate geomorphic functioning within the constraints of the urban environment. It is important to accept that ‗natural‘ geomorphic functioning may not be desired by the community or managers in an urban setting.

2.4 Ecology

2.4.1 Overview

For the last decade, scientists have been studying the effects of urbanisation on aquatic ecosystems (Paul & Meyer, 2001; Suren & McMurtrie, 2005; Taylor, et al., 2004; Wenger, et al., 2009). There is now a well-described suite of impacts of urbanisation on stream and river ecosystems, and an emerging deep understanding of the mechanistic bases for those impacts. Studies of the effects of urbanisation have moved from being correlational studies between land use and ecological structure (e.g. Sonneman, et al., 2001) or function (e.g. Imberger, et al., 2008), to more sophisticated spatial analysis of ecological response to elements of urban land use in relation to flow paths, that has identified urban stormwater runoff delivered through conventional drainage systems as the most plausible driver of urban related degradation of stream ecosystems (Wenger, et al., 2009). Increasingly, having determined some of the cause and effect relationships underlying urban impacts on streams, attention is moving to how best mitigate those impacts. It is timely therefore to summarise the described impacts of urbanisation on stream biota within an organised framework that identifies how best to mitigate these effects. In keeping with this broader review, we will focus on stormwater runoff as a primary mechanism of urban degradation.

2.4.2 Drivers of ecological change in urban-impacted streams

Broadly, the effects of urbanisation can be divided into drivers operating at two scales (Figure 7). The water quality (Section 2.2) and hydrologic changes (Section 2.3) already outlined are largely a product of runoff generated from urban (primarily impervious) surfaces across the catchment, delivered to the stream through conventional stormwater drainage. The increased frequency, magnitude and reduced quality of storm flows and reduced dry-weather flows cause major changes to the structure of in-stream habitat, and direct effects of disturbance on biota. The poor quality of stormflows results from increased concentrations of a range of contaminants including heavy metals, hydrocarbons, excessive nutrients and organic matter. At local scales the loss of riparian vegetation acts to alter light and temperature regimes in streams and to change or reduce organic matter supply (largely as leaf and wood litter) to stream ecosystems.

Appendix 4: Project 4 – Literature Review, 2010 28

We consider local-scale drivers to be independent of upstream catchment condition, and driven only by local riparian condition. Local-scale responses in terms of channel structure, sediment composition and biota are influenced by catchment scale drivers, local drivers and the interaction between the two (Figure 7).

Figure 7. A conceptual model of the impacts of urbanisation on stream ecosystems (after Walsh et al. 2005b). The primary driver of stream degradation in conventionally drained urban catchments (A) is most likely urban stormwater runoff that is delivered to streams through stormwater drains and pipes. These drainage systems efficiently transport pollutants generated in urban catchments to the stream every time there is enough rain to elicit impervious runoff, resulting in more frequent and larger storm flows with increased concentrations of nutrients and toxicants. A secondary driver often associated with urbanisation is the loss of riparian forests, which compounds the effects of stormwater runoff on sediment and organic matter dynamics and growth of autotrophs (algae and bacteria) in-stream. These changes compound the direct influence of stormwater runoff on the animals inhabiting the stream. If impervious surfaces are permitted to drain to pervious land (B), or to stormwater retention measures, the effects of stormwater can be attenuated before reaching the stream. Where the effects of stormwater runoff are adequately reduced, the local scale effects of riparian condition and function are likely to become a more dominant driver of in-stream ecological structure and function. Upward and downward arrows denote increased and decreased effects, respectively, and delta denotes a change of unspecified direction.

2.4.3 Catchment-scale drivers

Catchment-scale drivers transmit impacts to stream biota even if the area immediately adjacent to the stream channel has intact natural vegetation. A characteristic of streams is that they integrate across impacts on catchments and reflect the overall condition of a watershed (Hynes, 1975). Catchment-scale effects are likely to be magnified in conventionally drained urban catchments, where stormwater drainage networks connect the most distant parts of the catchment to the stream through hydraulically efficient stormwater drainage pipes. The catchment-scale effects of urban stormwater runoff are transmitted as the previously described changes to storm flows, dry-weather flows and their quality. Stormwater runoff can also effect organic matter supply, through leaf material entering streams via stormwater drains distributed throughout the

Appendix 4: Project 4 – Literature Review, 2010 29

catchment (Miller & Boulton, 2005). Gross pollutant traps or other stormwater treatment devices may intercept at least part of this load. Effects of hydrologic change have complex consequences for stream ecosystems but can be broadly grouped into the effects of disturbance (both increased ‗flashiness‘ of flows and increased drying), impacts on channel form and sediment composition, and reductions in retentiveness of organic matter. Direct effects of disturbance on aquatic biota have been well studied in non-urbanised settings (e.g. Engelhardt & Kadlec, 2001; Stanley, et al., 2010; Townsend, et al., 1998). These impacts are as a result of a number of factors including increased abrasion from particles entrained in the flow, disturbance of the stream bed and direct wash-out of biota (both sloughing of algae and loss of animals downstream). Drying of streams also impacts biota through dehydration and through a requirement for recolonisation or emergence from resting stages when flow returns (e.g. Lake, 2003). Complex interactions between stream channel form and drying determine the degree of impact of drying; as deeper areas may form pools that act as refugia (Lake, 2003). Altered hydrology may also alter in-stream habitat structure. Accumulations of wood and leaves (snags) in streams are important sites of biodiversity and are often highly productive (Bilby & Likens, 1980). Accumulations of organic matter can also contribute to alterations in stream channel form through the generation of eddies and backwater areas that affect sediment accumulation (Bilby, 1984). Changes in the stream hydrologic regime flushes out accumulations of wood and leaf litter, resulting in reductions in the standing crop of these materials and further simplification of urban stream channels (Roy, et al., 2005). Accumulation of toxins in sediments (derived from polluted stormwater runoff from the catchment) has been posited as a possible driver of reductions in in-stream biodiversity, based on mesocosm experiments using wetland sediments (e.g. Anson, et al., 2008). However, a recent study comparing biota sampled from the streams along an urban gradient concluded, and biota colonizing sediments from the same sites in controlled mesocosms, found no evidence that differences in sediment quality explained the observed declines in biodiversity with increasing catchment-scale urban impact (O'Brien, et al., 2010). The majority of early studies on the effects of urbanisation were based on inference of catchment-scale effects of urbanisation to correlational studies of urban density (usually measured as total imperviousness). More recently studies have sought to infer the most important local-scale mechanisms caused by urbanisation. Most recently considerable advances have been made in seeking to separate the different aspects of catchment urbanisation that are the most plausible drivers of stream ecosystem change. A recent assessment of the most important hydrologic flow paths for transmitting urban impacts to streams, identified the stormwater drainage network as the most plausible pathway by which stream biota are degraded (Walsh & Kunapo, 2009). The identification of stormwater runoff directed through a piped drainage network has been identified as potentially the primary driver of environmental harm to stream ecosystems (Walsh, et al., 2005b).

2.4.4 Local-scale drivers

Loss of native riparian vegetation due to land clearance is often associated with increased urbanisation (e.g. Suren & McMurtrie, 2005). Riparian vegetation is intimately linked with in-stream conditions and biota via a number of different pathways (e.g. Baxter, et al., 2005) Alteration in the biological processes of streams has been suggested to be directly linked to the extent and cover of the adjacent riparian zone (Power & Dietrich, 2002). Riparian vegetation can reduce local stream temperatures through shading of the stream bed, and reduce local light availability. Local inputs of organic matter are important both in the provision of habitat and resources for consumption (Bilby & Likens, 1980). High rates of local supply may maintain these functions even when retention of material is relatively low. Where riparian vegetation along streams is intact it may also function to filter surface and shallow sub-surface flows of water before they enter the stream channel (Gregory, et al., 1991; Lowrance, 1998). However, the interception and filtering of water quality to streams are likely to be limited in reaches in which stormwater runoff is directed directly through pipes to the stream (Walsh, et al., 2007), or where stream incision has resulted in hydrologic isolation of the stream channel from the floodplain (Groffman, et al., 2003).

Appendix 4: Project 4 – Literature Review, 2010 30

Loss of riparian vegetation and consequent reduction of terrestrial inputs including large woody debris (LWD) have been shown to reduce habitat and food availability, affect stream temperature and disrupt sediment, nutrient and toxin concentrations due to increased levels of surface run off (Edwards & Huryn, 1995; Meyer, et al., 2005). Changes in leaf litter type resulting from exotic riparian vegetation common along urban streams (Miller & Boulton, 2005) can also alter ecosystems structure and function such as leaf breakdown rates (Imberger, et al., 2008). Yet few studies have unambiguously demonstrated a strong effect of riparian forest on stream ecosystem structure and function. Some studies have posited that riparian forest might ameliorate the effects of catchment urbanisation, but these have been based on limited data (Morley & Karr, 2002) or have not clearly accounted for spatial correlations between catchment urbanisation and riparian forest cover (Moore & Palmer, 2005; Urban, et al., 2006). However, correlational studies that have explicitly sought to separate the influences of catchment-scale and riparian effects have consistently found that catchment-scale urban density has a stronger effect on biotic assemblages than riparian forest cover (Roy, et al., 2005; Thompson & Parkinson, in press; Walsh, et al., 2007). Similarly, an experimental manipulation of light climate in streams across an urban gradient found that the effect of light on algal biomass accrual and assemblage composition was relatively small compared to differences in catchment urban density (Catford, et al., 2007).

2.4.5 Indicators of ecological response

The responses to the two broad classes of drivers are generally measured at local scales as in-stream ecological indicators, such as the dominant primary producers (algae, macrophytes and emergent plants, collectively referred to as autotrophs), the diversity of aquatic biota and the presence of iconic taxa (usually amphibians, fish and species such as platypus). Measures of ecosystem function (fluxes of energy and materials) are increasingly being used as ecological indicators, although these have only rarely been applied in an urban context. Reviewing the literature on the ecological indicators of change in urban streams reveals a consistent set of patterns and relationships to key drivers of change (Table 6).

Appendix 4: Project 4 – Literature Review, 2010 31

Table 6. Summary of ecological responses to urbanisation currently reported in the scientific literature. The nature of the effect on the response (direction, magnitude) is shown, together with the putative driver and the scale at which the driver operates (where this is known and the data from the literature is consistent). A key reference is shown for each response as an entry point into the literature rather than including a detailed list of references.

Ecological Response General trend in response to

urbanisation

Driver and Scale

(catchment = C, local = L) Reference

Primary producers

Algae

Aquatic plants

Increased biomass

Shift to eutrophic species

Increased biomass

Increased invasive species

Water quality (C) + Light (L)

Water quality (C) + Disturbance (C)

Water quality (C) + Light (L)

Water quality (C)

Taylor et al. (2004)

Newall and Walsh (2005)

Suren (2009)

Suren (2009)

Ecosystem function

Litter decomposition

Nitrogen cycling

Increased

Altered

Litter supply (L) + Water Quality (C)

Abrasion (C)

Riparian condition (L)

Imberger et al. (2008)

Meyer et al. (2005)

Grimm et al. (2005)

Macroinvertebrates

Reduced biodiversity

Dominance of pollution tolerant taxa

Dominance of taxa with short life cycles

Reduction in shredder species

Changes in food web dynamics

Reduced substrate heterogeneity (C)

Impacts of multiple stressors (C+L)

Changed sediment composition (C)

Riparian impacts on adults (L)

Water quality (C)

Physical disturbance (C)

Drying (C)

Reduced/changed litter supply (L)

Reduced litter retention (C)

Impacts of multiple stressors (C+L)

Blakely and Harding (2005)

Walsh et al. (2005)

Roy et al. (2003)

Samways and Steytler (1996)

Roy et al. (2003)

Walsh et al. (2005)

Miller and Boulton (2005)

Miller and Boulton (2005)

Thompson and Parkinson (in press)

Faeth et al. (2005)

Fish

Increased invasive species

Changed community composition

Loss of sensitive species

Multiple

Altered hydrology (C)

Impacts of multiple stressors (C)

Arthington et al. (1983)

Roy et al. (2005b)

Danger and Walsh (2008)

Amphibians Reduced biodiversity Water quality (C) Barrett et al. (2010)

Mammals Reduced occurrence of resident platypus Sediment toxicity and water quality (C)

Impacts of multiple stressors (C)

Serena and Pettigrove (2005)

Danger and Walsh (2008)

Appendix 4: Project 4 – Literature Review, 2010 32

2.4.5.1 Primary producers

Urbanisation of streams has broadly consistent effects on primary producers, although the mechanisms underlying those effects are often complex. Inputs of nitrogen and phosphorus in urban streams, coupled with increased light reaching the stream channel due to clearance of riparian vegetation promotes the growth of both benthic algae and aquatic macrophytes (Roy, et al., 2005). However hydrologic disturbance acts as a strong filter in determining the composition of diatom biofilms in particular, which can be scoured by flow alone or high flows in combination with fine sediments (Hatt, et al., 2004; Newall & Walsh, 2005; Resh, et al., 1988; Sonneman, et al., 2001). Where fine sediment inputs from upstream are large, and bed slopes are low, the establishment of macrophytes is commonly observed, although these are often predominantly tolerant (and often cosmopolitan or invasive) species (Suren & McMurtrie, 2005).

2.4.5.2 Ecosystem functions

Ecosystem functions are the fluxes of energy and material through food webs. Urbanised streams often have high primary productivity due to the availability of light and nutrients, but standing crops of biomass are strongly influenced by hydrologic disturbance (Hatt, et al., 2004). Productivity of consumers (primarily invertebrates) can also be high. In terms of energy, urban streams are thus typified by high productivity and high turnover of biomass. A number of studies in recent years have investigated the effects of urbanisation on the processing of organic matters. Decomposition of leaf litter and woody inputs into streams is dependent on a number of environmental conditions which are altered by urbanisation. Studies on leaf breakdown rates have suggested that that leaf litter breaks down significantly faster in urban streams, due to physical fragmentation associated with increased stormwater inflows (Meyer, et al., 2005), or through increased microbial breakdown resulting from high nutrient availability (Imberger, et al., 2008). Changes in the type of litter provided can also alter function; reduced decomposition rates have been found in urban streams where the predominant litter is from exotic species (Miller & Boulton, 2005). However the rates of organic matter processing are strongly influenced by retention within the stream channel. Hydrologic impacts can greatly reduce the availability of organic matter in urban streams due to washout from simplified channels during high flows. Broad scale changes in nitrogen cycling have been shown along urban gradients (Grimm, et al., 2005). Nitrogen is supplied to urban systems in a variety of sources throughout the catchment, including aerial deposition from exhaust fumes, delivered to streams through stormwater drainage systems. Processing rates are generally high, although the disconnection of floodplains from streams (Section 2.3) can result in drying of soils and a reduced capacity to process nitrogen (Goldman, et al., 1995; Groffman, et al., 2003). The ways in which in-stream nutrient limitation is influenced by urbanisation have been surprisingly little studied (but see Chessman, et al., 1992).

2.4.5.3 Macroinvertebrates

Macroinvertebrate communities have been widely studied in urban streams for over a decade. The impacts of urbanisation on patterns of diversity (the number of taxa) and composition (the types of taxa and their numerical abundance) are well understood, although the nature of the cause and effect relationships between urbanisation as a stressor and changes in invertebrate communities are not always consistent or clear. Studies from a range of urban streams have found that urban streams have lower diversity than similar non-urbanised streams (e.g. Blakely & Harding, 2005; Paul & Meyer, 2001; Walsh, et al., 2001). Reduced diversity has been attributed to a number of mechanisms associated with urban stormwater runoff, including direct toxic effects on some biota (Iannuzzi, et al., 2004), reduced habitat heterogeneity in urbanised streams, particularly in substrate (Roy, et al., 2003), reduced availability of resources (Moore & Palmer, 2005) or more generally by the complex multiple stressors delivered by urban stormwater runoff (Walsh, et al., 2001). The altered composition of stream macroinvertebrates in urban streams provides some clear evidence of the likely drivers for reduced diversity in these systems. Organic pollution is a major factor altering invertebrate community structure and diversity. Urban streams typically, show an increased abundance of ‗pollution tolerant taxa‘, with reductions in the more ‗pollution sensitive taxa‘

Appendix 4: Project 4 – Literature Review, 2010 33

including Ephemeroptera, Plecoptera and Trichoptera (Paul & Meyer, 2001; Suren & McMurtrie, 2005; Walsh, 2004). These changes are likely to be driven in part by tolerances to anoxia. Urban streams are also typified by organisms with rapid life histories, which are correlated with small body size and with taxonomic groups such as oligochaetes and chironomids. These taxa are highly tolerant to hydrologic disturbance (both in terms of high flows and drying) because they spend a relatively short time as an aquatic larva. Walsh et al., (2001) investigated the effects of differing levels of urbanisation around metropolitan Melbourne, Australia on macroinvertebrate communities. These authors hypothesized that urban streams were impacted by increased levels of effective imperviousness in the catchment, leading to low water quality and flashy hydrology, resulting in communities with high abundances of pollutant tolerant taxa, an hypothesis that was supported by subsequent studies that separated the effects of general urban density from stormwater drainage connection (Walsh, 2004; Walsh & Kunapo, 2009). A number of studies have posited the existence of thresholds of imperviousness before an effect on macroinvertebrate assemblage composition is detected (e.g. Moore & Palmer, 2005; Morse, et al., 2003). All such studies have used total imperviousness (TI) as their measure of urban density, and have demonstrated a wide range of assemblage response at low levels of TI, which could be explained by the wider variation in drainage connection at low levels of TI (Hatt, et al., 2004; Roy & Shuster, 2009). Some studies, such as Wang et al. (2001), have claimed to use connected imperviousness, but have in fact only adjusted total imperviousness by a fixed proportion, where direct estimates of connected imperviousness (CI) have shown that CI and TI are not consistently related, particularly in less densely developed catchments (Roy & Shuster, 2009). To date the only studies that have directly estimated CI by spatial analysis have demonstrated a strong negative decline in macroinvertebrate diversity (and other ecological indicators) with increasing CI, with no evidence of a threshold before an effect is observed (Walsh, et al., 2005a; Walsh & Kunapo, 2009). This suggests that the hydrologic and water quality disturbance of urban stormwater runoff has strong negative impacts on stream ecosystems at very low levels of conventionally drained catchment urbanisation. Disturbance is an important structuring force in stream communities, with major influences on body form, life histories and community composition (Resh, et al., 1988; Townsend, et al., 1997). Changes in invertebrate communities may also be driven by changes in the nature and volume of organic matter in urban streams. The absence of accumulations of leaves and wood in urban streams reduces the availability of both habitat and food resources for groups such as shredders (aquatic invertebrates which feed on leaves and wood), although most species of shredders also happen to be species with low tolerances to physical disturbance and pollution. Changes in invertebrate communities can also occur as a result of urban-induced changes from native to exotic riparian trees (Miller & Boulton, 2005). There are also other, more complex interactions between riparian vegetation and aquatic invertebrate communities. Riparian vegetation can also enhance dispersal and reproductive success in the adult stages of aquatic macroinvertebrates (Samways & Steytler, 1996), although urban infrastructure has been shown to be a significant barrier to dispersal of flying aquatic insects (Blakely, et al., 2006). The broader food web consequences of urbanisation are the product of interactions between multiple stressors and their direct and indirect effects via all parts of the aquatic food web. In general these have been little studied, although Faeth et al. (2005) found that compared to nearby control streams the importance of predation by birds was greater in urban streams, and that resource fluctuations were lower, with highly complex food web impacts on a large number of taxa.

2.4.5.4 Fish and other vertebrates

The impacts of aquatic environments on aquatic vertebrates have been little studied compared to the impacts on macroinvertebrates. However many of the patterns apparent in invertebrates are also true for fish and amphibians. Urbanised streams typically have a less diverse fish fauna than similar streams not subject to urbanisation (Arthington, et al., 1983). These changes in fish communities have been attributed to a combined effect of increased hydrological disturbance and a reduction in-stream habitat heterogeneity (Roy, et al., 2005).

Appendix 4: Project 4 – Literature Review, 2010 34

Danger and Walsh(2008) demonstrated that several fish species such as river blackfish (Gadopsis marmoratus) responded to CI in a similar pattern to other ecological indicators, being absent from streams of the Melbourne region with more than 0.5% CI. Amphibians show similar patterns, although they are notably intolerant of many pollutants (particularly heavy metals), and are generally more intolerant of urban water quality than fish (Barrett, et al., 2010). Invasive fish are generally highly tolerant of disturbance, and can be favoured over native species in urbanised streams. In temperate areas invasive species such as the eastern mosquitofish (Gambusia holbrooki), and carp (Cyprinus carpio) are commonly present, and appear to outcompete native fish in urban settings. An exception is brown trout, which are absent from the degraded urban streams of Melbourne (Danger & Walsh, 2008). Danger and Walsh also noted a weak positive correlation between the native Galaxias maculatus and CI. They posited that this species, which is found less commonly in rural streams that contain trout, could use the sub-optimal conditions of metropolitan streams as a refuge from competitive exclusion by trout.

2.4.5.5 Selection of pertinent indicators

Ecological indicators for urban waterways are typically correlated with indicators of hydrology (Section 2.1), water quality (Section 2.2), and geomorphology (Section 2.3). However, the selection of pertinent ecological health indicators will also be quite locally context-specific. For example, use of platypus or fish as an indicator will only be useful if the waterway of concern is a known or potential habitat for these organisms. Fortunately, a number of indicators are readily measured and have been identified as being highly sensitive to impacts of urbanisation (making them effective indicators). For example, macroinvertebrates are commonly used, including the widely applied Ephemeroptra, Plecoptera, Trichoptera (EPT) index EPT (Roy, et al., 2003). In Australia both the AusRIVAS and SIGNAL indices have been used, with SIGNAL considered to be more effective across urban gradients (Walsh, 2006). As identified in Table 6, some fish can also be effective indicators, as can algal biomass and composition. Whilst functional indicators (e.g. ratio of photosynthesis:respiration, nitrogen cycling, litter decomposition) have been suggested as potentially useful, no study to date has yet identified a suitable indicator across the urban gradient. There is a difficult paradox in the effectiveness of indicators for assessing the impacts of urbanisation – and conversely – for predicting the impacts of stormwater management aimed at redressing these impacts. Given their sensitivity to the impacts of urban stormwater runoff, all ecological indicators currently in use show a very steep decline with even very low levels of stormwater runoff (as measured by directly connected imperviousness). Whilst this makes the indicators useful in assessing stormwater impacts, it means that the reverse trajectory (along the rehabilitation pathway) will not see a substantial ecological response until impacts have been brought back to these very low levels (Figure 8). Whilst stormwater managers may understandably wish to have a more linear (and thus progressive indicator), no studies to date have been able to identify any. Given the role of indicators in identifying the ecological condition of waterways, the search to find a more progressive indicator, while perhaps helpful in terms of garnering political and social support for interventions, will probably provide little ecological insight. The sharply non-linear response of currently indicators accurately identifies the situation; streams are ecologically degraded wherever significant stormwater inputs are permitted to occur.

Appendix 4: Project 4 – Literature Review, 2010 35

Figure 8. Non-linear relationships between a wide range of ecological indicators (D-N) and effective imperviousness. Note that water quality parameters (A-C) follow the same relationship (Source: Walsh, et al., 2005a).

2.4.6 Summary

This review shows the major impacts that urbanisation can have on the ecological values of streams. While the underlying mechanisms are complex, they can be classified into catchment level impacts (largely hydrology and water quality) and local impacts (largely changes in organic matter supply). The majority of ecological impacts are as a result of catchment scale processes (Table 6). Our review therefore clearly indicates a disjunct between the scale of the main impacts (predominantly catchment-scale urban stormwater runoff) and the scale of most restoration activities (predominantly local scale riparian replantings). Riparian restoration will result in some improvement in ecological condition (Moore & Palmer, 2005; Thompson & Parkinson, in press) however the amount of improvement will be contingent on the condition of the catchment (Hatt, et al., 2004; Roy, et al., 2006). Catchments with even a small coverage of connected imperviousness are likely to have sufficiently altered hydrology and reduced water quality that even extensive riparian restoration works may show minimal improvement in stream ecological condition. One of the major mechanisms underlying this is likely to be the interaction between organic matter supply (which is restored by riparian restoration) and organic matter retention (a function of hydrology). Riparian restoration activities are most likely to see major increases in stream condition where the overall catchment condition is highest. Similarly, riparian restoration activities will be most effective if they also seek to address both catchment scale and local scale impacts of urbanisation. Where riparian replantings can be paired with in-catchment retention of stormwater to reduce total runoff volume, reduce the frequency of uncontrolled stormwater flows, and restore the quality and flow regime of filtered sub-surface flows, ecological improvement is likely to be greater.

Appendix 4: Project 4 – Literature Review, 2010 36

3 The environmental flows concept

3.1 Environmental flows; the basics

We face a major challenge to provide the world's growing human population with reliable and affordable water supplies, while protecting the ecological integrity of freshwater ecosystems (Millennium Ecosystem Assessment, 2005). The concept of environmental flows has developed to meet this challenge, with the aim of identifying the critical elements of flow regimes that should be retained or restored, following extraction of water from river ecosystems (Arthington, et al., 2010). Environmental flow assessment and provision has to date focussed mainly on mitigation of the effects of extractive uses of water, but increasingly, changes to the flow regime resulting from land-use changes, are being considered (Kennen, et al., 2008; Poff, et al., 2010). Most environmental flow problems arise from water being extracted for human use: the challenge for environmental flow researchers and practitioners is how to distribute the remainder for maximum environmental benefit (Arthington, et al., 2010). The starting point in considerations of the effects of water extraction from aquatic ecosystems is that there is a monotonic decline in the condition of aquatic ecosystems with increasing extraction of water from them (Figure 9). This focus on extraction of water from aquatic ecosystems, compounded with the tendency for water resource managers to prefer centralized systems, leads to a tendency of urban water managers to first consider extraction from urban rivers and drains when identifying urban stormwater harvesting projects (Knights & McAuley, 2009; Newton & Ewert, 2009). Such a conception of stormwater harvesting has led to a misconception that urban stormwater runoff has some environmental flow benefit (Victorian Government: Department of Sustainability and Environment, 2006). As is clear from this review, in fact the reverse is true. Urban stormwater runoff, delivered through conventional drainage systems is a complex environmental flow problem that can, in large part, be solved through harvesting of stormwater before it reaches aquatic ecosystems.

Figure 9. Conceptual graphs of ecological and human value of water. I. adapted from Gleick and Palaniappan (2010),

assumes that any extraction from aquatic ecosystems has some negative ecological impact, predicting a monotonic decline of increasing gradient with greater extraction. The benefits accrued by the human population rise linearly with the volume extracted. Beyond peak ecological water (P: Gleick & Palaniappan, 2010), any increase in human benefit is outweighed by reduced ecological benefit. II. illustrates substantially different trends in ecological and human cost and benefit with increasing retention and use of stormwater before it reaches aquatic ecosystems. No stormwater use (A) results in ecological degradation of receiving waters. It also presents greater costs in microclimate control and drainage than if stormwater was used sustainably. Using a volume of stormwater equivalent to the volume lost to ET in pre-urban state (B), if coupled with infiltration systems to restore lost sub-surface flows, provides maximum environmental benefit, by maximising performance of the infiltration systems. Using all available stormwater runoff (C) has an environmental cost by reducing subsurface flow delivery to stream. However, in many urban settings this loss can be compensated by increased infiltration in non-treed open spaces, or by leakage of water supply systems.

Appendix 4: Project 4 – Literature Review, 2010 37

3.2 Application of environmental flows concept to urban stormwater

A primary step in assessing and implementing environmental flows is to estimate how ecologically relevant components of the flow regime are altered by a human activity or construction (Poff, et al., 2010). Typically, ecologically relevant indicators of hydrologic alteration are selected by identifying those that are correlated with changes in ecological indicators (e.g. Kennen, et al., 2008; Kennen, et al., 2010). Kennen et al. (2008) found change in the following indicators, linked to urban land use, explained variation in macroinvertebrate assemblage composition across a disturbance gradient in New Jersey, USA,: ratio of 25% exceedance flow to 75% exceedance flow, the mean number of storms producing quick flow, which was in turn highly correlated with annual change in low-pulse durations and frequency of high-flow pulses; minimum storm size required to initiate quickflow; overland flow generated when precipitation rates exceed infiltration rates of the soils. Similarly, Roy et al. (2005) found urban-driven increases in the frequency of autumn and summer storm events and rates of the rising and falling limb of the hydrograph, and reduced duration of autumn low flows were related to changes in fish assemblages of streams in Atlanta, Georgia. The general conclusion from these works, consistent with the review of hydrologic indicators in Section 2.1.3 is that the primary hydrologic changes driving ecological degradation centre around the increased frequency, magnitude (and worsened quality; see Section 2.2.3) of storm flow events, and the reduced duration of low flows. Several of the studies aiming to identify the change in hydrologic patterns driving ecological response not focussed on or even identified the increase in total volume caused by urban stormwater runoff. While this change might not be a direct driver of ecological response it casts the problem of stormwater runoff as a different class of environmental flow problem. As described in Section 2.1.1, urban runoff volumes are increased due to the loss of evapotranspiration which occurs with urbanisation. As the proportion of rainfall that becomes streamflow in vegetated catchments, and the proportion of rainfall that runs off impervious surfaces are well known, it is a simple task to estimate the volume of new, excess water that is generated by impervious surfaces (Figure 10 and Figure 11). In most cities, at least 60-90% of impervious runoff is water that would never have reached the stream in the pre-urban catchment. While it is not necessarily a problem of direct ecological relevance, the excess volume of stormwater aggravates the challenge of retaining and treating stormwater adequately to provide filtered flows that could mimic lost subsurface-fed dry-weather flows. The literature reviewed in this document suggests that there are four the critical elements of the hydrograph that are likely to be the primary drivers of urban stream degradation, from which we propose objectives that effective stormwater management should target.

1. Minimize uncontrolled storm flows. The increased frequency of hydraulic and pollutant disturbance from stormwater drainage flows has been identified as a primary driver of ecological degradation in streams (see Section 2.1.3). While a natural forested stream might receive one or two substantial floods a year that are associated with increased hydraulic disturbance and delivery of increased contaminant concentrations, streams receiving urban runoff receive such flows every time it rains enough to generate impervious runoff. A primary objective of stormwater management is thus to reduce the frequency of piped flows as close to the pre-urban level as possible (Walsh et al., 2009). Such objectives have recently been mandated for federal projects in the US (US Environmental Protection Agency, 2009), expressed as a requirement to retain the 95th percentile rain event on site.

2. Infiltration flows must be delivered to the stream through treatment measures that ensure flow rates do not exceed pre-urban subsurface flow rates. This objective aims to restore lost dry-weather flows. Appropriate maximum flow rates can be estimated from baseflow separation analysis in reference streams, or by assessment of infiltration capacities of native soils in the catchment (or a combination of these two techniques).

Appendix 4: Project 4 – Literature Review, 2010 38

3. Infiltration flows should aim to meet water quality concentration objectives close to standard objectives for ecosystem protection of freshwaters (ANZECC & ARMCANZ, 2000). For some variables, such as nitrogen, ANZECC objectives might be unattainable in treatment systems with collection pipes, in which case pragmatic acceptance of best attainable concentrations is appropriate. Wherever possible, exfiltration systems, and systems that allow overflow or infiltration flows to drain to pervious land, will help to achieve improved water quality, as well as increasing loss through evapotranspiration.

4. In almost all locations, the attainment of the first three objectives will require substantial retention and loss of stormwater runoff, either for indoor use and export to the wastewater stream, or for irrigation and loss to evapotranspiration. Therefore harvesting of a large proportion of stormwater runoff is a central objective for restoration or protection of environmental flows. Our knowledge of the wide difference between natural and impervious runoff coefficients (Figure 10) allow clear guidelines for the volumes of stormwater that should be kept out of receiving waters altogether (Figure 11). The predicted annual streamflow coefficients for grassland and forest catchments derived by Zhang et al. (2001) serve as useful bounds for the appropriate volume of runoff that should be allowed to reach the stream, primarily through infiltration systems. Using the curves of Figure 11., in a region with an average rainfall of 800 mm/y, of the 8 ML/y that would fall on a 1-ha roof, ideally 1.5–2 ML/y should be allowed to reach the stream through filtration, infiltration and natural topographic flow paths. Around 4–5.5 ML/y should be harvested and retained in the catchment.

Figure 10. Estimated annual runoff coefficients (C) from impervious surfaces (open triangles) from sites across the

Melbourne region as a function of mean annual rainfall (R). Impervious runoff was estimated from daily rainfall data at each of 11 sites (3–45 years of data), assuming an initial loss of 1 mm/d. Regression line: C = 0.234 + 0.203*log10(R). R

2 = 0.90. Annual runoff coefficients from 12 streams with forested (closed circles), grassland (open circles) or mixed

forested and grassland catchments (grey circles) across the Melbourne region as a function of mean annual rainfall. The lines surrounding these stream points are the relationship between streamflow derived by Zhang et al. (2001) for grassland (dashed curve) and forested catchments (dotted curve) of the world.

Appendix 4: Project 4 – Literature Review, 2010 39

Figure 11. Annual volume of runoff from 1 ha of impervious surface (from the relationship between impervious runoff

coefficient and annual rainfall shown in Figure 10), partitioned into two parts: the volume that needs to be passed through filtration systems to restore lost subsurface flows (grey polygon), and the volume that needs to be retained in the catchment and not delivered to the stream (through evapotranspirational loss or through use and export from the catchment through the wastewater stream). For each part, a range is indicated between situations in which the target streamflow is predicted by the grassland curve (more stream flow, less retention in catchment) or by the forest curve (less streamflow, more retention in the catchment) of Zhang et al. (2001; Figure 1).

It could be argued that the natural flow paradigm may not be entirely appropriate for urban streams, because their physical form is generally different compared to their natural condition (Chin & Gregory, 2005). However, it is certain that the flow regime must be restored back towards the pre-development regime, if ecologically successful stream restoration (sensu Palmer et al 2008) is to be possible.

4 The role of stormwater harvesting and its integration with other WSUD and stream restoration interventions

4.1 Potential impacts of stormwater harvesting

The idea of stormwater harvesting being used as a strategy to reduce environmental impacts on receiving waters is relatively new and the studies available to date are based on conceptual or modelling analysis, rather than monitoring of actual systems in place. In 2007, Fletcher et al (2007) undertook a hypothetical modelling study, examining the impact of various stormwater harvesting scenarios on a range of hydrologic and water quality indicators in both Melbourne and Brisbane. They found, for both low density (Figure 12) and high density (Figure 13) urban development, that urban stormwater harvesting was capable of bringing both flow and water quality back towards pre-development levels. Over-extraction of water causing a depletion of natural flow levels resulted only from very intensive application of harvesting. However, Knights and McAuley (2009) have observed, in an assessment of stormwater harvesting projects in Sydney, that many are designed to harvest low flows, effectively allowing high flows to bypass. This is often done because it requires smaller storage and diversion infrastructure. However, such an approach has the complete opposite effect of that which receiving waters need; it results in further reduction of baseflows (already normally depleted by urbanisation) without effectively reducing peaks (the increasing magnitude and frequency of which is primary mechanism of receiving water degradation).

Appendix 4: Project 4 – Literature Review, 2010 40

Figure 12. Relative changes to flow and water quality indicators from natural conditions for a catchment with no permanent baseflow (i.e. typical small upland catchment). The post-development ratio of the pre-developed level (which is denoted by the dotted line) is shown for (a) urban development at 14% imperviousness, and (b) after application of stormwater harvesting. The peak flow indicators for 1 and 3 month ARI have not been calculated, because at an hourly timestep for a catchment with no baseflow, they are zero for the pre-development case.

Figure 13. Relative changes to flow and water quality indicators from natural conditions for a catchment with no permanent baseflow (i.e. typical small upland catchment). The post-development ratio of the pre-developed level (which is denoted by the dotted line) is shown for (a) urban development at 70% imperviousness, and (b) after application of stormwater harvesting. The peak flow indicators for 1 and 3 month ARI have not been calculated, because at an hourly timestep for a catchment with no baseflow, they are zero for the pre-development case.

Other studies have demonstrated the benefits of stormwater harvesting in terms of reducing flow peaks downstream. For example, even a distributed network of rainwater tanks, specially designed to provide permanent freeboard by having a ‗trickle-outlet‘ at a pre-determined point below the top of the

0

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20.6 21.0 20.2

Appendix 4: Project 4 – Literature Review, 2010 41

tank, has been shown to significantly reduce peak flows (Hardy et al., 2004). More detailed modelling showed reductions of around 40-50% in the 3 month ARI peak flow, dropping to around 5-10% for the 100 year ARI event. Gerolin et al (2010)also found that rainwater could reduce overflows in the stormwater network in the UK However, Burns et al (2010) used a joint-probability approach to modelling the impacts of allotment-scale rainwater harvesting alone (accounting simultaneously for probability distributions of rainfall and rainfall excess, and found that flood reduction benefits diminished with increasing storm magnitude. They concluded that integrated management of stormwater at allotment, streetscape and precinct scales is required to significant reduce flood peaks. Petrucci et al (2010) et al found similar results, finding rainwater harvesting reduced the frequency of combined sewer overflows, but was not sufficient to prevent overflows during larger storm events. Clearly, the impact of harvesting on peak flows will depend on the catchment context, the nature of the harvesting (extent of area captured, available storage and extent and temporal distribution of demand, etc), and further research is needed to give better guidance on likely outcomes. By its very nature, the harvesting of stormwater will affect downstream water quality. It will reduce pollutant loads to receiving waterways, simply as a function of the water taken away (depending on the use and disposal of that harvested water). For example, if urban stormwater has a typical concentration of around 2.5 mg/L TN (Duncan, 2006; Taylor et al., 2005; Brombach et al., 2005), each ML of water harvested will reduce the N load to the receiving water by 2.5 kg. For example, Hatt et al.‘s (2006a) review found significant reductions in N loads to receiving water, and quantified the equivalent cost of this reduction, using well-established costs for constructing stormwater treatment wetlands. The influence of stormwater harvesting on pollutant concentrations, however, is more complex. If only roofwater, with its lower concentrations of particulate pollutants (Duncan, 2006), was harvested, the concentration at the receiving water may increase. In the case of harvesting of overall stormwater runoff, this is unlikely to be the case. However, no known monitoring data yet exist on the impact of stormwater harvesting on downstream pollutant concentrations. In summary, the potential for stormwater harvesting to assist in mitigating impacts of urban runoff seems to be justified by the (few) studies undertaken to date. However, clear consensus on the performance and effectiveness has not yet developed, meaning that guidance on system design to optimise environmental outcomes is not yet available. Perhaps most importantly, the studies to state focus exclusively on the ability of stormwater harvesting to reduce stormwater volumes and peak flows, along with pollutant loads. Given the nature of stormwater harvesting systems, their ability to restore other parts of the flow regime, such as subsurface flow and low flows, is likely to be limited and will require careful integration with other stormwater management measures such as retention and infiltration systems.

4.2 Interactions between stormwater harvesting and other stormwater management measures

Stormwater harvesting cannot be seen as a complete stormwater management strategy in itself, but as part of an integrated strategy to restore, as close as possible, the water quality and flow regime of the receiving waterway towards its pre-urban state (Figure 14). Stormwater harvesting can be used to reduce overall runoff volumes, frequency and peak flow rates, whilst infiltration and bioretention techniques can be used to restore infiltration and evapotranspiration (Department of Environment and Conservation, 2006; Shuster, et al., 2007). For example, it may be beneficial to include a component of ‗active detention‘ storage in harvesting storages, allowing water to slowly trickle out into downstream infiltration-based systems, thus maximising the efficiency of those infiltration systems, and providing storage to help reduce peak flows from subsequent storms. Unfortunately, research on the interaction of stormwater harvesting and stormwater retention and infiltration techniques is virtually non-existent, leaving designers and practitioners with a lack of guidance and tools to help them integrate technologies of each type.

Appendix 4: Project 4 – Literature Review, 2010 42

Figure 14. Schematic illustration of the pertinent impacts of urbanisation on hydrology at the catchment scale, showing the respective roles of stormwater harvesting and baseflow-restoration techniques such as bioretention and infiltration (adapted from Marsalek, et al., 2007)

Unfortunately, however, little guidance exists on how to integrate stormwater harvesting with other stormwater management techniques. Even the modelling of such systems is limited (Elliott, et al., 2010; Lee, et al., 2008), particularly given the problems of likely dispersed nature of such techniques throughout the urban catchment. In addition to the inability of current tools to consider interactions between harvesting and other WSUD systems such as infiltration and bioretention, there is a deficiency in the ability to represent the influence of scales and spatial arrangement of systems and their interactions on flow regimes at a given point within the catchment. This is a critical prerequisite to optimising the extent and location of source control techniques in relation to larger-scale end-of-pipe systems. For example, the widespread use of rainwater tanks connected to biofiltration systems at the allotment scale might be effective in restoring flow regimes at the catchment scale. One attempt was made recently by Endreny and Collins (2009) who assessed the impact of distributed, clustered, and single (centralized) infiltration systems with the use of MODFLOW (a finite element model). Their results showed that centralised infiltration systems resulted (perhaps not surprisingly) in more pronounced groundwater mounding than did homogenously distributed at-source systems. Hardy et al. (2004) developed a model to assess the effect of distributed rainwater storage (UrbanCycle), but did not include consideration of interactions with low-flow hydrology. Significant advances in modelling of urban hydrology has been developed by Rodriguez and others (Rodriguez, et al., 2008; Rodriguez, et al., 2005) in the form of the distributed hydrological model, URBS. However, we are yet to see these research tools being used in regular practice to determine the optimal arrangement of systems throughout a catchment (based on an assessment of impacts on receiving water hydrology). Despite this lack of empirical data, it would seem reasonable to hypothesise that restoring hydrological water balances at more local scales would consequently help to restore catchment-scale flow regimes towards their natural condition. It is hoped that the various scales at which the Cities as Catchments project will undertake its research and demonstration will help to test this hypothesis.

4.2.1 Urban stream restoration; interactions with other stream interventions

Restoration targets in urban streams have generally been intended to a) mitigate catchment level water quality and hydrology impacts, b) to address local scale removal of riparian restoration or c) to restore in-stream habitat complexity. Constructed stormwater treatment wetlands have been widely applied to mitigate the hydrological and water quality effects of urban runoff (Kadlec & Brix, 1995), and rely on aquatic plants and sediment processes to remove contaminants, in addition to retaining water to broaden hydrologic peaks (Heaney, et al., 1999). Studies of efficiency of nutrient removal have reported up to 75% load reduction for nitrogen, and in the vicinity of 40- 90% removal for phosphorus (Mungasavalli & Viraraghavan, 2006). However treatment performance can be extremely variable, depending on wetland size and design, retention time, season, stormwater volumes and pollutant concentrations (Breen & Dellarco, 1992). It has been suggested that wetlands may have a detrimental effect on downstream ecosystems due to release of contaminated water for extended periods after storm events (Hatt, et al., 2004; Helfield & Diamond, 1997).

Appendix 4: Project 4 – Literature Review, 2010 43

The second major target for restoration of urban streams has involved the replanting or riparian zones with native vegetation. These planting are intended to restore organic matter supply to streams and also to reduce in-stream temperatures and algal growth through shading. Riparian zones are critical areas of landscape to maintain integrity within urban streams (Roy, et al., 2005), and re-establishing inputs of native riparian detritus is an essential element of successful restoration of ecosystem function (Wallace, et al., 1997). Riparian vegetation has also been shown to stabilise stream banks and reduce sediment run off in urban settings (Roy, et al., 2006). However, the efficacy of riparian planting to restore urban streams has been challenged, as in many cases stormwater systems pass under riparian buffers and continue to impact directly on the stream (Newall & Walsh, 2005; Roy, et al., 2006). Most other attempts to improve ecological condition of streams in urbanised catchments have focused on reach-scale enhancement of physical habitat(Brown, 2000). Unfortunately, ecological effects of such habitat enhancement often are not assessed. In almost all cases where assessments were done, changes in biotic composition were small, with only a few taxa colonizing new habitat (Davis, et al., 2003; Larson, et al., 2008; Purcell, et al., 2002; Sudduth & Meyer, 2006; Suren & McMurtrie, 2005; Walsh & Breen, 2001). A few studies have shown some limited success is possible in certain circumstances. Charbonneau and Resh (1992) reported significant improvements in the composition of urban stream animal assemblages following restoration, but that restoration involved both reach-scale habitat improvement and catchment-scale actions such as the removal of sewage pollution from the creek. Larned et al. (2006) successfully displaced an exotic submerged macrophyte from weed beds with a native macrophyte species in a Christchurch stream, although this change in macrophyte species composition had no effect on the degraded macroinvertebrate assemblage in the stream, and subsequent attempts to replicate this success in other streams were less successful, and not universally appreciated by residents (Suren, 2009). Groffman et al. (2005) showed that retention structures in severely degraded urban streams acted as hot spots for nutrient processes. However, such small-scale studies of nutrient retention need to be assessed against the efficiency of in-catchment nutrient retention, and such a comparison has not yet been made. Furthermore retention structures are often unsustainable in the highly modified flow regime of urban streams (Booth, 2005; Frissell & Nawa, 1992). The weight of evidence suggests that urban stream ecosystems are generally limited in their ecological capacity by the impacts of urban stormwater runoff (Paul & Meyer, 2001; Walsh et al., 2005b). Attempts at restoration of stream ecosystems by in-stream or riparian habitat improvement are, therefore, likely to fail because they do not match the scale of the restoration action to that of the constraining impact (Hobbs & Norton, 1996; Lewis et al., 1996). This situation is more strongly true for urban (than rural) catchments because links between the catchment and the stream are more pronounced. The paradigm in urban stream management to date has been to consider riparian replanting activities and broader catchment scale management such as provision of wetlands, entirely separately. However impacts on stream ecology are a result of a complex interaction between catchment level effects on water quality and hydrology, and local effects on organic matter supply. There is strong evidence for an effect of riparian vegetation on stream macroinvertebrates, but this is contingent on the degree of urbanisation in the upstream catchment. The actual mechanism for the interaction between catchment and local scale impacts appears to be complex, but is likely to be mediated by organic matter supply and retention, and the degree to which urbanisation has altered other habitat variables such as sediment composition (Roy, et al., 2006). In most cases riparian replantings are relatively small relative to the area of the catchment, and local scale restoration is simply not adequate to mitigate for the majority of impacts on streams (Newall & Walsh, 2005; Suren & McMurtrie, 2005). Our review of the major impacts of urbanisation (Table 6) clearly indicates that the impacts of urbanisation on streams are a product of catchment level and local scale impacts. The majority of ecological components of interest are primarily affected by hydrologic and water quality impacts occurring at a catchment scale as a result of stormwater. Although local riparian restoration can restore organic matter supply, the retention and availability of that organic matter will be greatly

Appendix 4: Project 4 – Literature Review, 2010 44

affected by hydrology. In systems with heavily altered hydrology and poor water quality the effects of local riparian restoration are likely to be very weak, if indeed any can be detected at all. Attempts at restoration of stream ecosystems by in-stream or riparian habitat improvement are, therefore, likely to fail because they do not match the scale of the restoration action to that of the constraining impact (Hobbs & Norton, 1996; Lewis, et al., 1996). This situation is more strongly true for urban (than rural) catchments because links between the catchment and the stream are more pronounced.

5 Critical knowledge gaps

5.1 Hydrological effects, scales and integration

Whilst the impacts of urbanisation on catchment water balance and consequently on flow regimes are now relatively well understood, there a major gaps in our understanding of how to optimise stormwater harvesting and management strategies for eco-hydrological outcomes. The gaps in knowledge are both theoretical and include:

1. What are the impacts of focussing stormwater management on relatively small frequent storms, in terms of peak discharges of major events? What will be the ecological impacts of such changes?

2. What will be the hydrological consequences of applying stormwater harvesting alone, without focussing on complementary baseflow-restoration techniques? Will such approaches improve ecologically-relevant flow indicators or is a fully integrated approach necessary?

3. If stormwater harvesting can be used to reduce flow peaks and volumes, how should it be integrated with other stormwater management techniques (e.g. infiltration) to restore all aspects of the urban water cycle? For example, what proportion of flows passed through an infiltration system or biofiltration system contributes to streamflow, and what proportion is evapotranspired? How can the combined effects of such systems be modelled at the catchment scale, so that optimal combination and spatial arrangement of measures can be implemented? Are there optimal scales at which one or both of stormwater harvesting and complementary baseflow-restoration techniques should be applied?

4. How much is enough? To what extent do the various hydrologic indicators need to be restored towards their pre-development level, in order to see a tangible positive ecological response? To what extent might such tolerances vary between ecosystem types?

5. What is the ecologically appropriate flow regime in the context of an enlarged channel (see also 5.3)? What will be the ecological consequences of restoring pre-developed flow rates in an enlarged channel?

5.2 Impacts on water quality regimes

Superficially, the impacts of stormwater harvesting on receiving water quality might seem straightforward, given that the concentrations of pollutants in harvested water can be predicted. However, as described in Section 5.1, there will be effects due to the scales and arrangement of application.

1. For example, a focus on rainwater (roofwater) harvesting might result in an increase in the concentration of some pollutants, as other impervious surface runoff contributes a greater proportion to streamflows. What impact will this have on important water quality indicators? What consequences will this have for aquatic ecosystems?

2. How would the importance of such phenomena vary between receiving water types (small streams vs. large lentic waterways, for example)?

3. Is flow restoration and water quality treatment enough or is floodplain re-engagement necessary to restore pre-development nutrient retention of urban streams, and thus protect downstream receiving waters and bays?

Appendix 4: Project 4 – Literature Review, 2010 45

5.3 Geomorphic processes and consequences

Of all the impacts of urbanisation on receiving waters, perhaps the least well understood is those relating to geomorphology. However, this is a vital ‗missing link‘ in determining the feasibility of protection and/or restoration of urban streams, particularly given the likelihood for urban stream channels to have substantially enlarged relative to their pre-developed state, thus altering the ecological consequences that a given flow regime will have on the aquatic ecosystem. A number of important questions remain:

1. What is the effect of urbanisation on sediment budgets, for a range of development intensities? What impact might stormwater harvesting have on these sediment budgets? What impact might other stormwater management measures have? How should stormwater harvesting and management be designed to restore the pre-development sediment supply as much as possible? Should we restore sediment regime at the same time as restoring flow regimes?

2. How does the reduction of mobile sediments (deposition) in urban channels (fine and coarse-grained) impact on ‗natural‘ geomorphic functioning of urban channels.

3. What is the impact of increased runoff volumes on channel morphology? How much flow is too much? Are there clear thresholds? If so, what impact do (a) traditional stormwater management (b) current WSUD and (c) stormwater harvesting have?

4. What are the acceptable (and indeed desirable) levels of ‗dynamism‘ in the urban environment, considering the needs of the aquatic ecosystem and the needs of society?

5. To intervene or not to intervene?

6. Should we design the flow-regime to match the channel, or the channel to match the flow regime?

7. What role will stormwater detention storage have on erosion potential? Might it result in an increased erosion potential index through prolongation of above-threshold flows? If so, can stormwater harvesting be used to resolve this problem? If so, what implications does this have for the optimal scales of application of harvesting?

5.4 Ecological responses and thresholds

The ecological responses to urban gradients have only received attention in relatively recent times. Despite significant gains in our understanding of the mechanisms explaining these responses, a number of uncertainties remain:

1. What mitigating factors might (at catchment-scale) explain variations in the level of stream ecosystem response to urban impacts?

2. How will stream ecosystems (in both greenfields and retrofit situations) respond to stormwater harvesting? How should we monitor and report these responses in ways that are ecologically meaningful?

5.5 Selection of indicators at a range of scales

In every section of this review – those related to hydrology, water quality, geomorphology and ecology, significant questions relating to the selection of indicators are as yet unaddressed:

1. Can the proposed hydrologic indicators (see Section 3.2) at the project/site scale be lumped up at the catchment-scale? What is the impact of scale and arrangement on these indicators? Can catchment-scale water balance objectives be readily determined, and if so, do we have the analysis/modelling capability to allow stormwater managers to assess them?

2. What water quality indicators should be applied at what scales / receiving water types? If concentration targets are the most appropriate for flowing waterways, (i) can they be met by existing technologies and (ii) can they be readily assessed and modelled? What range of other water quality indicators should be used for other receiving waters, such as estuaries, wetlands and large rivers?

3. Are the ecological targets which have been commonly used to date for assessing the impacts of urbanisation suitable for assessing the impacts of harvesting? Are there any other suitably sensitive indicators which should be considered?

Appendix 4: Project 4 – Literature Review, 2010 46

4. What are the appropriate indicators to use for assessing geomorphic regimes, and how should the natural dynamism of stream geomorphology be taken into account?

Appendix 4: Project 4 – Literature Review, 2010 47

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ANZECC/ARMCANZ. (2000). Australian and New Zealand Guidelines for Fresh and Marine Water Quality, Volume 1, The Guidelines (Chapters 1-7).

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