26.5 Conclusions
alternative scenarios (e.g., Ratner et aI., 1997).
These analyses require much demographic information that is seldom available except for closely
monitored and carefully enumerated populations
(Rieman and McIntyre, 1993). Where sufficient
data are available, it is prudent to incorporate PV As
for target species that may be particularly endangered or vulnerable. If sufficient data are not available, considering the data needs for PV A (especially through sensitivity analysis using plausible
ranges of parameter estimates) can point out high
priorities for monitoring and assessment, as well as
domains of risk relative to prospective decisions
(Rieman and McIntyre, 1993; Ruggiero et aI., 1994;
Kokko et aI., 1997). For fishes, for which considerable data are sometimes available, the results of
quantitative PV A are often in agreement with credible qualitative assessments performed by independent scientists (Nehlsen et aI., 1991; Ratner et
aI., 1997). However, wherever listed or candidate
species are at issue, managers should consider the
lack of sufficient data to perform PV A as a red flag
indicating that they might be unwittingly jeopardizing these populations through management action or inaction. Wishful thinking and pressing
deadlines notwithstanding, sometimes analysis
simply cannot substitute for field data, and the necessary data must be acquired for a reliable assessment to be completed.
26.5 Conclusions
It is an awkward challenge indeed to find a common organizing principle among the many approaches used to assess patterns in aquatic biota
and linkages between aquatic biota and habitat or
landscape conditions. In keeping with the recent literature, this chapter focuses on issues of scale,
stressing the need to explicitly address the consequences of scale in determining which of the many
available protocols and methods to apply in a given
situation. Table 26.1 contrasts a number of ecological assessment methods and metrics commonly
employed for aquatic biota in terms of the biological scale or level of organization that they address.
Clearly, there is tension between objectives or requirements for assessing patterns in community or
assemblage structure or integrity and assessments
necessary to examine the viability of particular
species, subspecies, metapopulations, and populations. The variation in spatial and temporal scale
of biological systems of concern in Table 26.1 itself spans many orders of magnitude, and the full
consideration of key biophysical linkages and the
399
ecological context of an ecosystem can create additionallevels of complexity and scale. Population
and species viability are not just concerns at the
level of a single lake or stream reach; they also
must be addressed (especially for endangered
species) at the cumulative level of the entire historic and present-day range of the taxon.
Deciding on a standard approach to this multidimensional complexity of scale is daunting,
especially when we consider the diversity of management issues and challenges that ecological assessments are called on to address. Some key questions, if resolved, could potentially simplify the
task of determining appropriate scales for assessment. For example, debate continues over the need
to incorporate a discrete, catchment-level unit in
the hierarchy of aquatic habitat. Some practitioners argue that, from the biological point of view,
there is not sufficient need for such a unit (e.g., The
Nature Conservancy, 1997), while others focus almost exclusively on the catchment as the organizing unit for analysis and synthesis (e.g., Montgomery et aI., 1995). Still others (e.g., Frissell and
Bayles, 1996; Omernik and Bailey, 1997) see it as
a necessary and useful unit if viewed in a regional,
comparative context.
In addition to resolving the question of appropriate scale and geographic scope, the ecological
scope of an environmental assessment must also be
defined. This second overarching question asks if
it is necessary to treat all species or if it will be sufficient to conduct explicit analyses only for selected
target species for which data or a large body of
knowledge already exists. Some species are known
to perform special roles as dominant interactors in
aquatic ecosystems, either as keystones (e.g., predacious fishes) or as ecosystem engineers (e.g.,
beavers). It seems clear that ecological assessments
must pay special attention to these species (Power
et aI., 1996). But can we also identify for aquatic
ecosystems a subset of focal or umbrella species
(Lambeck, 1997) that can be considered to have requirements that, if met, will result in the conservation of numerous other species with similar habitat requirements? This seems the only logical way
to simplify the task of coping with the large diversity of taxa with which we need to be concerned.
Yet little progress has been made in identifying a
basis for aggregating species by habitat requirements or other ecological criteria (but see Townsend and Hildrew, 1994; Poff, 1997). Again, explicit consideration of scale of life-history traits,
habitat templates, and ecological responses will be
necessary to advance this front.
The concepts of spatial refugia and biodiversity
alternative scenarios (e.g., Ratner et aI., 1997).
These analyses require much demographic information that is seldom available except for closely
monitored and carefully enumerated populations
(Rieman and McIntyre, 1993). Where sufficient
data are available, it is prudent to incorporate PV As
for target species that may be particularly endangered or vulnerable. If sufficient data are not available, considering the data needs for PV A (especially through sensitivity analysis using plausible
ranges of parameter estimates) can point out high
priorities for monitoring and assessment, as well as
domains of risk relative to prospective decisions
(Rieman and McIntyre, 1993; Ruggiero et aI., 1994;
Kokko et aI., 1997). For fishes, for which considerable data are sometimes available, the results of
quantitative PV A are often in agreement with credible qualitative assessments performed by independent scientists (Nehlsen et aI., 1991; Ratner et
aI., 1997). However, wherever listed or candidate
species are at issue, managers should consider the
lack of sufficient data to perform PV A as a red flag
indicating that they might be unwittingly jeopardizing these populations through management action or inaction. Wishful thinking and pressing
deadlines notwithstanding, sometimes analysis
simply cannot substitute for field data, and the necessary data must be acquired for a reliable assessment to be completed.
26.5 Conclusions
It is an awkward challenge indeed to find a common organizing principle among the many approaches used to assess patterns in aquatic biota
and linkages between aquatic biota and habitat or
landscape conditions. In keeping with the recent literature, this chapter focuses on issues of scale,
stressing the need to explicitly address the consequences of scale in determining which of the many
available protocols and methods to apply in a given
situation. Table 26.1 contrasts a number of ecological assessment methods and metrics commonly
employed for aquatic biota in terms of the biological scale or level of organization that they address.
Clearly, there is tension between objectives or requirements for assessing patterns in community or
assemblage structure or integrity and assessments
necessary to examine the viability of particular
species, subspecies, metapopulations, and populations. The variation in spatial and temporal scale
of biological systems of concern in Table 26.1 itself spans many orders of magnitude, and the full
consideration of key biophysical linkages and the
399
ecological context of an ecosystem can create additionallevels of complexity and scale. Population
and species viability are not just concerns at the
level of a single lake or stream reach; they also
must be addressed (especially for endangered
species) at the cumulative level of the entire historic and present-day range of the taxon.
Deciding on a standard approach to this multidimensional complexity of scale is daunting,
especially when we consider the diversity of management issues and challenges that ecological assessments are called on to address. Some key questions, if resolved, could potentially simplify the
task of determining appropriate scales for assessment. For example, debate continues over the need
to incorporate a discrete, catchment-level unit in
the hierarchy of aquatic habitat. Some practitioners argue that, from the biological point of view,
there is not sufficient need for such a unit (e.g., The
Nature Conservancy, 1997), while others focus almost exclusively on the catchment as the organizing unit for analysis and synthesis (e.g., Montgomery et aI., 1995). Still others (e.g., Frissell and
Bayles, 1996; Omernik and Bailey, 1997) see it as
a necessary and useful unit if viewed in a regional,
comparative context.
In addition to resolving the question of appropriate scale and geographic scope, the ecological
scope of an environmental assessment must also be
defined. This second overarching question asks if
it is necessary to treat all species or if it will be sufficient to conduct explicit analyses only for selected
target species for which data or a large body of
knowledge already exists. Some species are known
to perform special roles as dominant interactors in
aquatic ecosystems, either as keystones (e.g., predacious fishes) or as ecosystem engineers (e.g.,
beavers). It seems clear that ecological assessments
must pay special attention to these species (Power
et aI., 1996). But can we also identify for aquatic
ecosystems a subset of focal or umbrella species
(Lambeck, 1997) that can be considered to have requirements that, if met, will result in the conservation of numerous other species with similar habitat requirements? This seems the only logical way
to simplify the task of coping with the large diversity of taxa with which we need to be concerned.
Yet little progress has been made in identifying a
basis for aggregating species by habitat requirements or other ecological criteria (but see Townsend and Hildrew, 1994; Poff, 1997). Again, explicit consideration of scale of life-history traits,
habitat templates, and ecological responses will be
necessary to advance this front.
The concepts of spatial refugia and biodiversity