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Draft Guidelines for the Selection of Methods to Determine Ecological Flows and Water Levels

Acknowledgements

Executive Summary – Recommendations

1 Introduction

2 Rivers

3 Lakes and Wetlands

4 Groundwaters

5. References of Main Text and Appendices

Appendix 1: Relationship between Total Allocation and Ecological Flow Requirements in Rivers

Appendix 2: Effects of Flow Regime on Stream Ecology

Appendix 3: Methods Used in Ecological Flow Assessments of Rivers

2 Rivers - continued

2.5 Description of individual methods

In Section 2.3 we described the framework for method selection, which was based on the significance of instream values (Section 2.1.1) and degree of hydrological alteration (Section 2.2). In this section we describe all methods and make recommendations for their use within the context of the framework. A stylised representation of the framework (Table 2.4) is given with each method to assist the reader in placing the method within the context of the framework.

2.5.1 Historical flow methods

Framework for use (Table 2.4)

Historical flow methods are based on flow records and are the simplest and easiest to apply. Stalnaker et al (1995) describe this type of method as ‘standard setting’ because they are generally desktop rules-of-thumb methods based on a proportion of a flow statistic to specify a minimum flow. The statistic could be the mean annual low flow, a percentile from the flow duration curve, or an annual minimum with a given exceedance probability (see Historical flow method 1 in Table 2.7). For example, a method might prescribe that abstraction ceases when the natural flow falls below 80% of the MALF. Another method that has been used is to allow the total amount of water taken from the river to vary with the flow, eg, allow abstraction of 10% of the flow at any time (see Historical flow method 2 in Table 2.7).

The aim of historical flow methods is to maintain the flow within the historical flow range, or to avoid the flow regime from deviating largely from the natural flow regime. The underlying assumption is that the ecosystem has adjusted to the flow regime and that a reduction in flow will cause reduction in the biological state (abundance, diversity, etc) proportional to the reduction in flow; or in other words, that the biological response is proportional to flow (Figure 2.3). It is usually also assumed that the natural ecosystem will only be slightly affected as long as the changes in flow are limited and the stream maintains its natural character. It is implicitly assumed that the ecological state cannot improve by changing the natural flow regime.

Note: The biological response is assumed to be proportional to the flow, the wetted perimeter or width, and the weighted usable area – for the historical flow method, the hydraulic method, and the habitat method, respectively.

The most well known historical flow method is the Tennant (1976) method, also known as the Montana method, which specifies that 10% of the average flow is the lower limit for aquatic life and 30% of the average flow provides a satisfactory stream environment. The Tennant method was based on hydraulic data from 11 United States streams (in Montana, Wyoming and Nebraska) and an assessment of the depths and velocities needed for sustaining aquatic life. At 10% of average flow, he found that the average depth was 0.3 m and velocity 0.25 m/s, and considered these lower limits for aquatic life. He found that 30% of average flow or higher provided average depths of 0.45–0.6 m and velocities of 0.45–0.6 m/s and considered these to be in the good to optimum range for aquatic organisms. This is an example of a ‘regional method’, applicable to the region that has the same type of streams as the streams used for developing the method. The Tennant method has been adopted in many different parts of the world, including New Zealand, and in some cases, its recommended minimum flows have been similar to IFIM predictions (eg, Allan 1995; Crowe et al 2004). In New Zealand, Fraser (1978) suggested that the Tennant method could be extended to incorporate seasonal variation by specifying monthly minimum flows as a percentage of monthly mean flows.

Historical flows can also be used to define ‘an ecologically acceptable flow regime’ – for example, Arthington et al’s (1992) ‘holistic method’ that considers the magnitude of low flows, and the timing, duration and frequency of high flows. Such a flow regime would not only sustain biota during extreme droughts, but would also provide high flows and flow variability needed to maintain the diversity of the ecosystem. The building block method (BBM: King et al 2000) is a similar approach. The range of variability approach (RVA) and the associated indicators of hydrological alteration (IHA) identify an appropriate range of variation, usually one standard deviation, in a set of 32 hydrological parameters derived from the ‘natural’ flow record (Richter et al 1997). The holistic, BBM and RVA methods are conservative and maintain the ecosystem by retaining the key elements of the natural flow regime. They are probably most appropriate for river systems where the linkages between ecosystem integrity and flow requirements are poorly understood.

2.5.2 Expert panel

Framework for use (Table 2.4)

Expert panels have been used by some regional councils in New Zealand. These usually comprise interested parties as well as ‘experts’. This method has been used to support or verify other methods of ecological flow assessment, rather than as a method in its own right. The expert panels inspect the stream from the banks and consider the suitability of suggested ecological flow requirements. If the stream is at or close to the suggested ecological flow, it is possible to assess its suitability for many aquatic biota, provided the panel has the relevant experience. However, if the stream is not close to the ecological flow it is very difficult, if not impossible, to envisage hydraulic conditions at other flows.

2.5.3 Generalised habitat models

Framework for use (Table 2.4)

Studies of flow and habitat requirements in more than 60 New Zealand rivers suggest that flow requirements can be generalised for particular species (Jowett 1996). For trout, these generalised relationships vary with fish size and life stage. Trout rivers, even of the same size, vary in the value of the fisheries they support. Moreover, different sizes of trout and life stages have different depth and velocity and related flow requirements. Jowett’s studies indicate that maximum habitat for juvenile trout tends to be provided by flows of 1–2 m³/s, whereas maximum habitat for adult brown trout is provided by flows of 6–15 m³/s. More recently, a larger set of rivers (99) was examined by Lamouroux and Jowett (2005) to show that the shape of the relationships between habitat and flow per unit width was consistent between rivers. The results of this analysis allow us to more closely define the flows that provide maximum habitat for each species, and more importantly, quantify the habitat change that occurs with flow change.

The habitat suitability curves used to generate the generalised habitat models were regarded as the best currently available (2004), but further refinement is possible as more data on habitat use are collected or more refined methods of analysis are developed.

Generalised models can be applied to a specific stream with simple spreadsheet calculations while knowing only the average width at one discharge. This assumes that the hydraulic geometry (relationship between width and flow) is typical of New Zealand rivers, as described in Jowett (1998). However, there are limitations. The generalised models were based on 99 New Zealand streams and rivers and so represent the results of habitat analysis in a river of ‘average’ shape. This assumption breaks down where a river is unusually wide and shallow, as in extensively braided rivers, or where the river is narrow and deep, as in some spring-fed streams. If flow and width measurements are carried out at two flows, WAIORA (Water Allocation Impacts on River Attributes) can be used to apply the generalised relationships with greater certainty. The application and data requirements are described in the WAIORA user guide (Jowett et al 2003).

2.5.4 Fish passage/connectivity models

Framework for use (Table 2.4)

Fish passage can be an issue at low flows when there is inadequate water depth for upstream/ downstream migration of fish species or laterally into wetlands and side braids. Fish migration is often related to spawning and is a seasonal requirement. Survey reaches and cross-section locations are selected that represent potential barriers to fish passage. Hydraulic modelling is carried out to determine the lowest flow that provides adequate passage width through the river. Passage width is the continuous width of river with depth exceeding the minimum passage depth and velocity less than the maximum passage velocity. These models are usually applied to salmonids because these fish have greater depth passage requirements than native fishes.

2.5.5 1D and 2D hydraulic habitat models

1D models
Framework for use (Table 2.4)

2D models
Framework for use (Table 2.4)

In contrast to standard setting methods, hydraulic-habitat methods predict how the river changes incrementally with flow and hence are most suited to evaluating the effects of large-scale changes to the natural flow regime. As such they enable a holistic approach to be taken to the assessment of flow regime requirements, where necessary. Hydraulic-habitat methods are a primary component of the IFIM. Data requirements and hydraulic modelling capability increases with the complexity of the underlying hydraulic models. However, the biological interpretation of results is critical and the process allows other aspects of flow effects on biota (eg, duration of low flows, flow variability and frequency of flushing flows, food supply, and passage restrictions) to be considered, as far as possible with existing knowledge.

At the least, the basic premise of these methods is to maintain instream habitat that is suitable for the biota. The critical values and their associated habitat suitability curves must be appropriate to the stream, particularly its size, and must be related to flow, particularly minimum flows, if hydraulic-habitat models are to produce consistent and sensible results; and to provide a context for connected systems such as riverine wetlands or groundwater systems.

The critical factors and their associated habitat suitability criteria can be perceived in two ways. In most instances, we apply them in a specific sense for providing habitat for the target critical species/life stage and with the added aim of providing for taxa with lower flow requirements. However in some situations, habitat criteria associated with the critical factor can be used in a generic sense to provide instream conditions that, based on experience, are considered appropriate for the ecological function and potential range of instream communities. In this latter situation, the habitat criteria act as general descriptors of instream conditions and stream size; the ‘target species’ is secondary and may in fact not actually be present. Examples of these applications include:

• trout spawning criteria which also provide good depths and velocities for invertebrate habitat (which sustains the fish food base) in small streams

• redfin and common bully habitat criteria that provide good general instream conditions for streams slightly larger than those dominated by diadromous galaxiids.

In New Zealand, it has generally been assumed that minimum flows set for salmonids will be adequate to maintain native fish populations. The rationale for this is that trout, because of their large size and drift-feeding requirements, have higher depth and velocity requirements than most native fishes. Many native fishes are most abundant in small streams or on the margins of larger rivers (eg, upland bullies, redfin bullies, inanga). Therefore, habitat for these species is best at low flow and in larger rivers; the margins will still provide some habitat for these native fishes at the higher flows required by salmonids.

Hydraulic modelling is used to predict water depths and velocities at individual points in a section of river over a range of flows. These predictions are then used to show how usable habitat varies with flow. Ecological flow assessments are based on the shape of the curves and the proportional changes engendered by a flow change. Either 1D or 2D modelling can be used, and if done well, there should be little difference between the results. 2D modelling can only be applied to a reach, the length of which is usually up to 1 km, a constraint imposed by survey costs. The reach is usually chosen to represent a longer segment of river.

The difficulties in acquisition of sufficient and accurate bed topography and calibration of 2D models are a practical limitation to their utility, and it should not be assumed that they are better simply because they require more data. A good knowledge of hydraulics is necessary to identify salient features of bed topography, especially in turbid or deep water. Calibration is difficult, subjective, and time-consuming with large files. In contrast, 1D survey methods are straight-forward and calibration procedures are well-developed and reproducible, although empirical. 1D surveys can be carried out over longer sections of river using the habitat mapping method, so that they can include a greater variety of habitats, although not at the same level of detail as a 2D survey. 1D surveys require fewer resources than 2D surveys, and usually produce similar or better accuracies. However, 2D models are better able to extrapolate beyond the calibration range in complex river morphologies, give good graphic representation, and when the modelling is done well, they give better predictions of the direction and distribution of velocities.

It is difficult to calibrate a 2D model so that measured water surface levels are modelled precisely, and any error in water surface level translates to an error in predicted depth and mean cross-section velocity. This becomes particularly critical at low flows, where the definition of upstream and downstream water level controls, such as topography at the head of a braid, or topography of a riffle at the tail of a pool, determines the flow in the braid or the water level in pool. Thus, the accuracy of the topographic model will determine whether water levels are predicted correctly at low flows.

1D models using empirically derived stage-discharge rating curves are easier to calibrate, and predict water surface level more accurately than 2D models, at least within the range of rating curve calibration. Within a reach, a 2D model requires more data points than a 1D model and therefore gives a better measure of the longitudinal variations in depth and velocity. As predicted flows depart from the flow used to calibrate a 1D model, uncertainty in velocity distribution increases. 2D models are likely to predict changes in velocity distribution, particularly eddy and reverse flows, more accurately than 1D models, although in both cases, predicted depths and velocities will be incorrect if water surface levels are not modelled accurately.

The number of cross-sections in a 1D survey depends on the morphological variability within the river. An analysis of a 2D survey showed that similar results could be achieved with a 1D survey with 19 cross-sections (Tarbet and Hardy 1996). In another study, Payne et al (2004) sub-sampled several very large data sets to determine how many cross-sections are required to produce a robust weighted usable area function. They found that 18–20 cross-sections gave results nearly identical to results for 40 to 70 cross-sections per reach.

As with all methods listed, instream habitat surveys and analyses have to be carried out appropriately. The main factors in habitat analyses are that:

1. the physical habitat measured is representative
2. the aquatic organisms may or may not be habitat-limited, but it is conservative to assume that a population is habitat-limited. The preference (habitat suitability) curves should reflect the preferred habitat of the organisms
3. suitability curves should be based on measurements that show how distributions of abundance or presence of organisms vary with physical habitat. If measured abundance or distribution does not vary with physical habitat, habitat methods are not applicable.

Further discussion on hydraulic habitat models is found in Appendix 3.

Regional methods
Framework for use (Table 2.4)

Regional methods are a subset of hydraulic habitat methods and are based on an analysis of ecological flow assessments made by habitat-based methods in a number of rivers. They have already been developed for some areas of New Zealand (eg, Jowett 1993a, b; Wilding 2002). Typically, the ecological flow requirement will be a function of river size. Regional methods are quick and easy to use, once they have been developed, and are almost as biologically defensible as the assessments on which they are based.

Tennant’s (1976) method is a good example of a regional method that combines the best features of historical flow methods and habitat methods, resulting in a biologically defensible method of minimum flow assessment – for the region. Once established, regional methods can be easily applied to rivers within the region using a formula based on the proportion of natural flow, either recorded or estimated. The formula can be as simple as a fixed proportion of flow or can vary the proportion with river size, possibly retaining a higher proportion of the flow in small rivers than in larger rivers, as used in formulae for maintenance of trout and food-producing habitat in Wellington and Taranaki rivers (Jowett 1993a, b). Similar methods could be developed for regions that are hydrologically and morphologically similar, with criteria that apply to trout, native fish, stream insects, or periphyton.

By analysing habitat variation with flow for rivers within a region, it is possible to determine the level of flow as a proportion of median or mean annual low flow that maintains adequate or optimum conditions for various ‘target’ communities. Variation in levels of maintenance could be achieved by assessing requirements for optimum habitat and minimum habitat, as in the Tennant method. Application of the method would involve selecting an appropriate target community and level of maintenance for the river in question and then applying a formula based on flow. The formula may be referenced to an historical flow statistic (eg, MALF: Jowett 1993a, b).

The benefit of regional methods over historical flow methods is that they can have explicit environmental goals, making water management more transparent. Thus, regional methods can be established as biologically defensible, and discussion and consultation can focus on whether the ‘target’ and flow standards of maintenance are appropriate.

The rationale for habitat-based regional methods is primarily that of habitat methods. Within a region, it is possible to develop formulae that predict when hydraulic conditions are optimum or become limiting for a range of aquatic species. For instance, most native fish are small-stream species. Few are found in swift, deep water. In contrast, adult trout are rarely found in water less than about 0.4 m deep. Stream insects are most abundant in shallow swift habitats.

It is also possible to generalise velocity and depth criteria as levels of protection within a region based on a data set from rivers in the region. For instance, average velocities of less than 0.1 m/s might be considered poor, 0.1–0.3 m/s adequate, and 0.3–0.5 m/s good for aquatic organisms such as trout and benthic invertebrates. Similarly, average depths greater than 0.15 m might be considered suitable for native fish, and depths greater than 0.4 m suitable for adult trout.

These methods are potentially useful in that they combine the best features of habitat and flow methods. They are less expensive than habitat methods, yet once developed are cheaper but still likely to result in flow assessments that provide life sustaining flows whilst retaining some degree of the river’s ‘character’.

2.5.6 Periphyton biomass model

Framework for use (Table 2.4)

This is a specialised model that would mainly be applied to ecological flow assessments where the frequency of floods and freshes is altered. The model predicts accumulated periphyton biomass from the time since the last flood and nitrogen and phosphorus concentrations (as mean monthly concentrations measured over at least a year). The model is based on an analysis of periphyton samples collected in a large number of rivers and is described in Biggs (2000). In rivers below impoundments, this relationship can be used to develop the necessary frequency of artificial floods for various growth periods to ensure that peak biomass does not exceed biomass guidelines.

2.5.7 Entrainment, sediment and bank stability models

Framework for use (Table 2.4)

The models are used to assess flushing flow requirements, where flows should move fine sediments, but not the armour layer and channel maintenance flows, where a significant portion of the armour layer is disturbed. Entrainment models are an extension of hydraulic models and would only be applied to ecological flow assessments where the frequency of floods and freshes is altered. These models are used to predict bed shear stress and velocities and hence potential movement of bed sediments at high flows. The output of these models describes how the area of river with shear stresses exceeding critical shear stresses for bed sediment movement varies with flow. For bank stability, shear stresses at the banks are compared to critical shear stresses. Critical shear stresses for non-cohesive sediments are relatively well known (eg, Graf 1971). Critical shear stresses for cohesive sediments are reviewed by Jowett and Elliott (2006).

2.5.8 Suspended sediment model

Framework for use (Table 2.4)

This model, sometimes known as the ‘sticky bed model’ predicts how suspended sediment concentration varies with distance downstream under different flows (Jowett and Milhous 2002). The model can be used to determine downstream changes in suspended sediment concentration (water clarity) that result from water abstraction or diversion. The assumption is that the stream bed is a matrix of gravel and cobbles and that any suspended sediment reaching the water/substrate interface will be trapped within the matrix. This mechanism has been shown to be valid until all voids in the substrate matrix are filled or a surface seal forms. Calibration measurements of flow and suspended sediment concentrations downstream of the abstraction point are advisable but not essential. Currently, river hydraulic habitat simulation (RHYHABSIM) is the only hydraulic model that does this calculation. This type of model would only be applied to relatively large abstractions or diversions; applications and field measurements of suspended sediment concentration indicate that the changes in sediment concentration are small and barely detectable.

2.5.9 Seston flux model

Framework for use (Table 2.4)

The seston flux model is the suspended sediment model described above, applied to the special case of seston (plankton) flowing from a lake. Lake outlets commonly support high densities of filter-feeding benthic invertebrates and seston is their main food source. Flow affects the distance that seston is carried downstream and effective length of ‘lake outlet’. As with the suspended sediment model, field measurements of seston concentration at a number of points below the lake outlet should be taken to calibrate the model.

2.5.10 Inundation models

Framework for use (Table 2.4)

Inundation models are hydraulic models, usually 2D, that show the extent of inundation of wetlands or similar riparian zones for different flows. The model would only be applied to ecological flow assessments where the frequency of floods and freshes is altered. The model can be used to determine critical flows for wetland maintenance or could be used for flood hazard mapping for damage avoidance.

2.5.11 Fish bioenergetics models

Framework for use (Table 2.4)

These models have been developed for brown trout in New Zealand and the United Kingdom and for rainbow trout in the United States, but have not been widely applied. They are an extension of habitat models in that they assess the habitat suitability for drift-feeding fish taking into consideration factors such as drift-food availability, swimming ability, foraging behaviour, and metabolic processes. Both bioenergetic models and drift-feeding habitat suitability models predict trout feeding locations, with the bioenergetic model providing an alternative to the empirical habitat suitability observations. Comparison of New Zealand brown trout bioenergetic and habitat suitability models shows excellent agreement.

The advantage of bioenergetic models is that they predict biological meaningful metrics such as net rate of energy intake, growth potential, and carrying capacity for trout, in graphical outputs that are easily understood by stakeholders. This is a significant advance over weighted useable area (WUA) – flow relationships predicted by traditional 1D or 2D habitat modelling based on empirical habitat suitability criteria. However, at present they can be applied only at limited spatial scale (eg, over a riffle, pool sequence), and are expensive. The bioenergetic model developed at the Cawthron Institute operates on the output of 2D or 1D representative reach hydraulic models (Hayes et al 2003; Kelly et al 2005). Another similar bioenergetics based model developed in the United Kingdom operates on the output of a 3D hydraulic model (eg, Booker et al 2004).

2.5.12 Dissolved oxygen models

Framework for use (Table 2.4)

Dissolved oxygen concentrations can fall below acceptable levels in low-gradient streams that contain macrophytes or decomposing organic matter on the bed; in these slow-flowing streams the flow required to maintain an adequate dissolved oxygen concentration is an important ecological consideration. Three important parameters, as well as habitat and water temperature, are required to calculate flow effects on dissolved oxygen concentration. These are:

• daily community respiration rate (the average rate of oxygen consumption by aquatic plants and micro-organisms)

• production/respiration ratio (ratio of the daily rates of photosynthetic production of oxygen to daily oxygen respiration by plants and micro-organisms)

• reaeration coefficient (the coefficient that describes the rate at which oxygen is exchanged between the atmosphere and the stream).

The WAIORA DO model applies to streams with a reasonably homogenous distribution of aquatic plants (which can include algae) in a reach. At present, we do not have a model that applies to streams where low concentrations of dissolved oxygen are caused by anaerobic decomposition of organic matter. Model calibration using field measurements of dissolved oxygen, stream flow, width, depth, water temperature, and climatic conditions is essential, because modelling parameters are difficult to estimate.

2.5.13 Temperature models

Framework for use (Table 2.4)

Water temperatures may affect aquatic systems in many ways ranging from acute lethal effects to chronic stresses. When the flow of a river is reduced, it becomes more responsive to solar radiation because it is shallower and flowing more slowly. Water temperature modelling is soundly based on physical principles and a calibrated water temperature model is capable of predicting flow effects accurately. The water temperature models used to assess flow effects are usually one-dimensional heat transport models that predict water temperatures from the abstraction point as a function of stream distance downstream and environmental heat flux. In general terms, the heat is gained or lost from a parcel of water as it passes through a stream segment. This is accomplished by simulating the various heat flux processes that determine that temperature change. These physical processes include convection, conduction, evaporation, as well as heat to or from the air (long-wave radiation), direct solar radiation (short-wave), and radiation back from the water.

The temperature of water in a river is influenced more by climate than by river flow. The effect of water temperatures on aquatic biota is difficult to determine, largely because lethal water temperatures only occur for short periods when climatic conditions are extreme and flows low. Because of the dependence of water temperature on climate, it is difficult to use water temperatures to set ecological flow requirements, although it is possible to quantify the temperature changes that occur.

2.5.14 Groundwater models

Framework for use (Table 2.4)

Groundwater abstractions can influence the surface flows in a river, depending on the hydro-geology and the distance of the groundwater take from the river. Similarly, changes in river water levels can influence groundwater levels. Groundwater models are discussed in the Groundwater section of this report.

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