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Montane Meadows in the Sierra Nevada:
A Comparison of Terrestrial and Aquatic Assessment Methods
by
Sarah Elizabeth Purdy B.S. (University of California, Davis) 2005
Thesis
Submitted in partial satisfaction of the requirements for the degree of
Master of Science
in
Ecology
in the OFFICE OF GRADUATE STUDIES
of the University of California
Davis
Approved
2010
Peter B. Moyle Kenneth W. Tate
Valerie Eviner
Committee in Charge
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Sarah Elizabeth Purdy December 2009
Ecology
Montane Meadows in the Sierra Nevada: A comparison of terrestrial and aquatic assessment methods
Abstract
We surveyed montane meadows in the northern Sierra Nevada and southern
Cascades for two field seasons to compare commonly used aquatic and terrestrial-
based assessments of meadow condition. We surveyed 1) fish, 2) reptiles, 3)
amphibians, 4) aquatic macroinvertebrates, 5) stream geomorphology, 6) physical
habitat, and 7) terrestrial vegetation in 79 meadows between the elevations of 1000 and
3000 m. From the results of those surveys we calculated five multi-metric indices based
on methods commonly-used by researchers and land management agencies. The five
indices consisted of 1) fish-only, 2) native fish and amphibians, 3) macroinvertebrates, 4)
physical habitat, and 5) vegetation. We compared the results of the five indices and
found that there were significant differences in the outcomes of the five indices. We
found positive correlations between the vegetation index and the physical habitat index,
the invertebrate index and the physical habitat index, and the two fish-based indices, but
there were significant differences between the indices in both range and means. We
concluded that the five indices provided very different interpretations of the condition in a
given meadow. While the assessment of meadow condition changed based on which
index was used, each provided an assessment of different components important to the
overall condition of a meadow system. Utilizing a multimetric approach that accounts for
both terrestrial and aquatic habitats is the best opportunity to assess meadow condition,
particularly given disproportionate importance of these systems in the Sierra Nevada
landscape. To accept the results of just a single index in the absence of the others is
potentially misleading and costly.
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Introduction
Montane meadows are wetland systems that have disproportionate
importance compared to their area (Kattlemann and Embury 1996, Kondolf et al.
1996). In the Sierra Nevada of California and Nevada, meadows support critical
ecosystem services including biodiversity, flood attenuation, sediment filtration,
water storage, water quality improvement, and carbon sequestration (Potter
1994, Woltemade 2000, Povirk et al. 2001, Hammersmark et al. 2008). In
addition, meadow vegetation has significant direct economic value as forage for
grazing livestock (Torell et al. 1996).
The majority of meadow systems in the Sierra Nevada have suffered
anthropogenic impacts to their soils, hydrologic processes, and biotic integrity
(Ratliff 1985, Knapp and Matthews 1996, Castelli et al. 2000, Blank et al. 2006,
Popp et al. 2008). In particular, streambank erosion and channel incision are
widespread and highly detrimental to meadow function; these erosional
processes can be accelerated by improper livestock grazing, culvert and road
crossing placement, mining, logging, recreational activities, and water diversions.
The impacts of improper management are often exacerbated by episodic natural
events such as drought, fire, and flood (Leege et al. 1981, Belsky 1999, Wemple
et al. 1996, Gucinski et al. 2001).
At severe levels, erosion and channel incision cannot be reversed by
simply removing the disturbance(s). Once critical thresholds of impact have
been reached, the meadows do not recover without active intervention (Ratliff
1985, US Bureau of Land Management 1995, Chambers et al. 2002, Allen-Diaz
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et al. 1999, Micheli and Kirchner 2002, Briske et al. 2008). Incision of a meadow
lowers the local water table and can be viewed as transition to an alternate stable
ecological state (e.g. Briske et al. 2008) from a previous stable state with a high
water table supporting meandering streams and diverse wetland vegetation. This
transition results in a reduction of instream habitat, loss of hydrologic functions,
and changes in community structure in both the aquatic and terrestrial
ecosystems (Zimmer and Bachmann 1978, Hammersmark et al. 2008, Cornwell
and Brown 2008). Without re-elevation of the water table and restoration of
hydrologic connectivity between meadow surface and stream channel, the
meadow remains altered, potentially for centuries, and becomes a terrace
occupied by upland plant communities (Allen-Diaz 1999, Loheide et al. 2009,
Briske et al. 2008). This represents a loss of ecosystem services and economic
value, but is preventable and even reversible if management actions are taken
before such thresholds are crossed. The key is to identify meadows at risk before
this threshold is crossed, so that management actions can be taken.
We propose that current techniques to assess the condition of both the
terrestrial and aquatic components of montane meadows do not provide
adequate information to: 1) determine key factors impacting meadow condition;
nor 2) determine how close the meadow is to crossing the threshold to a
different, less desirable, state (e.g., Belsky et al. 1999, Allen-Diaz 1991, Auble et
al. 1994, Chambers et al. 2004, Blank et al. 2006). In particular, terrestrial and
aquatic components are rarely assessed together, although the two components
are highly interdependent. We suggest that meadow condition assessments
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which integrate local terrestrial and aquatic conditions are needed to evaluate
montane meadow status. Such evaluations can provide the basis for future
monitoring and can help to determine how to balance ecological benefits with the
economic benefits associated with various land management practices. At
present there are three commonly-used types of assessments for meadows and
their associated stream systems: vegetation surveys, qualitative habitat
assessments such as Proper Functioning Condition and the EPA Rapid Habitat
Assessment, and indices of biotic integrity (IBIs). These assessment tools were
not developed specifically for meadow evaluation; instead they have been
typically used for rangeland assessments, high gradient stream assessments, or
fish surveys.
Vegetation Surveys
Vegetation surveys have been
the standard method used to evaluate meadow condition by most natural
resource management agencies, where a meadow in good condition is one that
has herbaceous vegetation composition which benefits seasonal grazing by
livestock. These surveys use metrics such as plant species composition,
vegetative cover, plant rooting depth, community type, and seral status to
determine meadow condition (i.e., Ratliff 1985, 1993, Weixelman et al. 1997,
Winward 2000). They provide quantitative data, usually through the use of
transects and quadrats, and allow for accurate re-measurement to determine
trends through time. However, these methods require a high degree of plant
taxonomic expertise to perform. However, the heterogeneous nature of meadow
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systems makes it difficult to extrapolate the conditions found in transects and
quadrats to the larger surroundings. Currently, the predominant Forest Service
range assessment method in the Sierra Nevada is the vegetation survey method
developed by Weixelman et al. (2003), based in part on the methods of Winward
(2000).
Qualitative Habitat Assessments
Two qualitative assessment techniques have been commonly applied to
meadows: Proper Functioning Condition (PFC) assessment for the
terrestrial/hydrological portions and Rapid Habitat Assessment (RHA) for the
aquatic/riparian habitat portions. The PFC assessment was developed jointly by
Bureau of Land Management (BLM), USDA Forest Service, and Natural
Resource Conservation Service (NRCS) and focuses assessment on 17 metrics
such as hydrologic connectivity, balance of sediment deposition and erosion, and
vegetation composition required to stabilize deposited sediment. The impetus for
developing PFC was the need for an assessment method that was rapid,
required minimal expertise, and distinguished the range of conditions
encountered in the field from pristine to highly impacted (Prichard et al. 1994,
1996, 1998; Mitchell and Tippy 1993). Similarly, Barbour et al. (1999) developed
RHA protocols as a part of their larger Rapid Bioassessment Protocol (RBP) for
small (wadeable) streams. This ten-metric index focuses predominantly on
instream and streambed components such as available habitat for invertebrates
and fishes; siltation and erosion, bank stability, riparian width, meander ratios,
flow regimes, and access to the floodplain (Appendix 1). The RHA provides a
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numerical basis to visually determine the condition of the stream habitat. It uses
few direct measures of the habitat components but provides guidelines for
categorizing each metric into four broad condition categories (poor, marginal,
sub-optimal, optimal) to ease interpretation. While these two assessment
protocols are more integrative in their approaches and easier to perform than
vegetation survey, they are qualitative and have been criticized for lack of
sensitivity to change and inability to accurately monitor trends over time, and
excessive observer variability (e.g., Coles-Ritchie et al. 2004).
Indices of Biotic Integrity
Karr (1981, 1986, Karr and Chu 1997) developed the concept of the Index
of Biotic Integrity (IBI) as a means of determining the condition of fish populations
in Midwestern rivers. The premise of this method is that the biological community
responds to anthropogenic stressors in a predictable fashion. The metrics used
for assessment are diversity, abundance, life history, sensitivity, and other factors
that are responding to changes in habitat quality which are in turn responding to
stressors. Barbour et al. (1999) used the IBI concepts in developing the Rapid
Bioassessment Protocols used by the Environmental Protection Agency (EPA).
The approach incorporates fish, aquatic macroinvertebrates, periphyton, and a
qualitative habitat assessment similar to RHA assessment.
Purpose
Our study aimed to compare five rapid assessment methods to determine
the condition of montane meadows in the Sierra Nevada. We compared
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methods that were either previously developed or which we developed
specifically for montane meadows by modifying established methods. We
developed three original IBIs based methods (sensu Karr 1981, Harrington and
Born 2000) to quantify condition of 1) fishes, 2) native fishes and amphibians,
and 3) aquatic macroinvertebrates as indicators of aquatic and riparian condition.
We employed a modified version of the Weixelman (year) approach to determine
vegetation condition, and used the EPA Rapid Bioassessment Protocol (Barbour
et al. 1999) habitat assessment to determine stream channel and overall
meadow habitat condition.
This study addressed two questions. First, were all five of the measures of
meadow condition we examined in agreement? Secondly, if not, what were the
differences among the methods? Our hypothesis was that given the inherent
complexity and variability of meadow ecosystems, it is unlikely that a single-focus
approach to assessment adequately captures the condition of the meadow and
its components. Rather, a multi-functional approach is necessary to get the best
information on the true status of the meadow (Karr 2005, 2006; Pellant et al.
2005). However, the constraints imposed on monitoring by time, budget, and
expertise require utilization of an assessment approach that most efficiently
captures meadow condition. This paper shows how some commonly used rapid
assessments, modified for montane meadow systems, can provide quite different
results when used independently, but provide useful assessments when used
together.
Methods
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Study Site Selection and Study Design
Over two field seasons (June through September of 2006 and 2007), we
assessed 79 meadows in the northern Sierra Nevada and southern Cascade,
ranging from Sierra County in the south to Modoc County in the north, visiting
each site a single time. We surveyed only meadows associated with a stream
(flowing near baseflow at time of assessment), and that had previously been
surveyed by USDA Forest Service vegetation crews utilizing the protocol
developed by Weixelman et al. (2003). We surveyed a broad assortment of
meadows over a large geographic range in order to capture the variability
present in montane meadow systems. Site selection was focused primarily in
Plumas, Lassen, and Modoc Counties with some sites in Sierra and Nevada
Counties. We eliminated sites that did not have flowing water. We chose
meadows between the elevations of 1000 and 3000 m, which were less than 5
km from a vehicle access point. Meadow type was determined by utilizing the
key to Region 5 meadow types developed by Weixelman et al. (2003). This
dichotomous key uses depth to water table, elevation, and plant community to
categorize meadow into one of seven types. Our survey area predominantly
consisted of mesic or hydric montane or subalpine meadow types, although
some had converted to more xeric communities due to stream channel incision
and lowered water table.
Fish Survey
We sampled a minimum of one 50 m stream reach within each meadow,
placing blocknets at each end of the reach to prevent entrance or egress of
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fishes from the sampling area. We chose reaches that were both accessible and
representative of the meadow. In large meadows (> 1km in length), we sampled
two 50 m reaches to account for habitat heterogeneity. Basic fish sampling
procedures followed those of Moyle et al. (2002). We conducted single pass
backpack electrofishing surveys using Smith-Root type 12 backpack electrofisher
and systematically sampled all habitat within the stream reach from the lower
blocknet to the upper blocknet. Stunned fish were captured by two to three
people using dip nets. The fish were kept alive in buckets or live wells until they
were identified to species (using Moyle 2002), measured (standard length, mm),
and weighed (volumetric displacement); then returned alive to the water near
where they were caught.
Amphibian and Reptile Survey
We surveyed day-active amphibians and reptiles in the riparian zone,
using Visual Encounter Surveys (VES) (Crump and Scott 1994). At the beginning
of the stream reach, two members of the crew performed a timed survey of the
stream banks, stream and adjacent habitats such as oxbows and ephemeral
puddles looking for egg masses, tadpoles, or adult amphibians. We attempted to
capture all amphibians and reptiles encountered and used a standard snout to
vent length (SVL) measurement. We identified all reptiles and amphibians to
species and recorded length and life stage, though that was not used as a metric
in the index (Crump and Scott 1994, p. 91). Amphibians observed or captured
during the fish sampling were recorded as incidental observations and
contributed to the total abundance score for the site.
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Macroinvertebrate Survey
Benthic macroinvertebrates (BMI) were sampled using modified Level 2
protocols from Harrington and Born (2000). We took nine total samples from
within each 50 m fish sampling reach, using a D-net, preferentially sampling
riffles but also collecting from distinctive habitats throughout the reach. The
samples were combined and each was placed in a white enamel pan and the
major debris removed. We sorted and identified live invertebrates to family in the
field and returned them to the stream afterwards. We identified the first ~300
invertebrates in each sample. Invertebrates with questionable identification were
preserved in 70% ethanol for later identification in the laboratory. Three complete
samples were brought back for traditional laboratory processing to validate field
sorting accuracy.
Habitat Survey
We used the EPA (Barbour et al. 1999) Habitat Assessment Sheet for low
gradient streams to assess the habitat structure and geomorphological conditions
of the meadow streams. This assessment is based on ten instream, bank
stability, and vegetation parameters, each scoring between 0 (worst) and 20
(best) for a total possible score of 200 (See Appendix 1). Each of the ten metrics
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-point spread in each. The categories consist of verbal
descriptions of pertinent habitat features that distinguish observable human
impacts to the stream and riparian area. The numeric values within each
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category allow the observer to quantify their observations through scoring each
site on a scale of 0 to 200, combining the scores from each of the ten metrics.
Vegetation Survey
We designed our terrestrial vegetation survey as a modified version of
Weixelman et al. (2003). The Weixelman assessment was based on seral status
(as a proxy for recovery from past disturbance), depth to rooting frequency of
>100 roots/dm2, percentage of bare soil in the meadow, and vegetation functional
guilds (i.e., wetland indicator status, growth or rooting habit). Our methods
differed in that Weixelman used line transects and quadrats throughout the
meadow, whereas we surveyed only the vegetation within 10 meters of the
-meter
from the banks, riparian vegetation permitting. We estimated the percent cover
for all species within the survey area by breaking the 50-meter reach into ten
5x10 meter transects. We walked each plot and noted the species present,
making a visual estimate of their percent cover within that 5x10-meter plot, then
combined the results to get an overall species list and percent cover for the entire
1000 m2 survey area. We identified all plants to the lowest possible taxonomic
level. Unknown plants were either preserved or photographed for later
identification. We assumed multiple canopies within each plot (i.e., a shrub layer
with forbs in the understory); therefore percent coverage did not have to equal
100. We measured the percent of bare ground exposed (as a measure of
disturbance) in each 5x10-meter plot, and also noted the percentage of rocks,
11
and cryptogams in the survey reach. Vegetation functional guilds were
determined following Weixelman et al. (2003), and from information in the USDA
Plants Database on identification, habitat, distribution, growth forms, and function
of plants (http://plants.usda.gov/wetland.html).
Calculated Indices of Biotic Integrity
We used five multi-metric indices to assess the condition of the 74 study
meadows. The three indices focusing on fish, amphibians, and
macroinvertebrates were original to the project. The vegetation and habitat
indices came from previously published assessment methods. The vegetation
index was constructed and calculated according to Weixelman (2003) in order to
be consistent with the vegetation assessments commonly used by natural
resource managers in state and federal agencies, though field data collection
differed slightly (Table 4). The physical habitat index created by the EPA and
described above provided a commonly-used qualitative habitat assessment to
compare with the IBIs and the Weixelman vegetation index (Barbour et al. 1999,
Harrington and Born 2000, Ode et al. 2005). The fish-only IBI measures habitat
suitability and productivity for fishes regardless of whether the fish was of native
or introduced origin. The native fish and amphibian index measures the habitat
suitability and productivity of native fishes and amphibians and reflected long-
term human impacts to native communities. The invertebrate index measures
water quality, habitat productivity and availability, and community structure. The
vegetation index measures terrestrial and stream bank vegetation as a reflection
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disturbance and hydrologic conductivity. The habitat index measures available
habitat, stream bed condition, and disturbance.
The methods for building IBIs were originally established by Karr (1981) to
evaluate Midwestern fish populations using metrics such as species richness,
functional feeding groups, and life stage. The IBI concept evolved under the
premise that combining multiple community metrics that respond to different
stressors provides a far more reliable indicator of overall ecosystem integrity than
a single criterion. A second premise of IBIs was that a scoring system could be
devised that was easily interpreted by the public and natural resource managers.
To that end, Karr (1981) utilized a large data set of many community criteria from
an assortment of rivers with a broad range of conditions. Each metric was scored
using a 1, 3, or 5 to indicate a range of values for poor, moderate, or good
condition. Metric values were determined subjectively by individuals familiar with
stream impairment in the region. Since the initial introduction of the IBI concept,
the USEPA (Barbour et al. 1999), California Department of Fish and Game (Ode
et al. 2005), and Moyle and Marchetti (1999) have developed more quantitative
regional IBIs for some parts of the western United States, but as of yet there is
no published IBI specifically for Sierra Nevada meadow systems.
Because our study aimed to analyze differences in rapid bioassessment
procedures, the IBIs were built on metrics that have been consistently shown by
other studies to be important indicators of stream impairment (e.g., Karr and Chu
1997, Barbour et al. 1999, Harrington and Born 2000, Ode et al. 2005). We set
individual metric values for the IBIs we built
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scores of 1, 3, and 5, respectively to indicate low, moderate, or high condition
values. The breaks between values were determined by visually inspecting
frequency histograms of each metric from two years of field data (Moyle and
Marchetti 1999). We used natural breaks present in the upper and lower ends of
the histograms for all of our sites combined to determine the breaks between
metric values because they likely represented important ecological thresholds
better than an arbitrary percentage. The scores for all of the metrics in a given IBI
were combined and then normalized to get a final IBI score as a proportion of
100 total score. This allowed comparison of IBIs from different meadows using a
consistent scoring rubric.
Fish-only IBI
Due to the low species and functional diversity of fishes encountered in
most of the Sierra Nevada, we used only three simple metrics for the fish IBI:
biomass/m3 of habitat, species richness, and total abundance (Table 1). These
metrics indicate how well the stream supported fish regardless of if they were
native or alien species. The index assumed that the presence of fish in the
stream was an indicator of good condition, i.e., that the habitat was of high
enough quality to support fish fauna. However, the presence of non-native and
hatchery origin fishes is a perturbation to native aquatic communities and
represents a departure from the historical condition (Knapp 2005, Eby et al.
2006, Schilling et al. 2009). Therefore, we created a second IBI that focused on
native fishes and amphibians and regarded non-native fishes and amphibians as
detrimental to the condition of the ecosystem.
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Fish-Only IBI Metric Value 1 3 5
Biomass g/m3 <30 30-100 >100 Abundance <10 10-50 >50 Species Richness 0-1 2-3 4+ IBI Score = [Total points/number of metrics] x 20 Table 1. Fish-only IBI showing the three metrics used to obtain a final IBI score, Biomass, Abundance, and Species Richness. The metric value at the top indicates the score for each of three ranges of values. The scores for the three metrics are then summed, divided by three (the number of metrics in the IBI) and multiplied by 20 to provide a final IBI score out of 100 possible.
Native Fish and Amphibian IBI
The native fish and amphibian IBI used eight metrics that included the
presence of native trout, percent native species in the sample, number of native
species present, number of age classes of native species, fish abundance, fish
taxa richness, number of native amphibians, and amphibian taxa richness. We
combined the native fishes and the amphibians into a single index because,
while we felt it was critical to represent amphibians in the survey of meadow
conditions, their presence on the landscape was so rare that they could not
support their own index. However, since the historical literature indicates that
native amphibians were once common in the ranges that we sampled and
presumably co-occurred with native fishes, we combined the two taxa into a
single index (Grinnell and Storer 1924).
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Fish and Amphibian IBI
Metric Value 1 3 5
Presence of Native Trout None
Mixed native/ non-native
Native trout only
Percentage of native species (# native individuals/total) <25% 26 75% >75% Number of Native Species present 0-1 2 3+ Number of Age Classes of Native Species 0-1 2 3+ Fish Abundance (#/50m reach) <10 10-50 >50 Fish Species Richness 0-1 2-3 4+ Number of Native Amphibians 0 1-3 4+ Native Amphibian Taxa Richness 1 2 3+
IBI Score = [Total points/number of metrics] x 20
Table 2. Fish and Amphibian IBI with each of the eight metrics and the ranges of values used to obtain a final IBI score.
Invertebrate IBI
The invertebrate IBI consisted of 7 metrics: 1) The Hilsenhoff family-level
index, 2) the EPT index, 3) percent Plecoptera (stoneflies), 4) percent predators,
6) percent Diptera (true flies) , and 7) percent Elmidae (riffle beetles). The
Hilsenhoff family-level index (Hilsenhoff 1988) provided a measure of organic
pollution based on the tolerance values established for individual taxa and their
proportionate representation in the sample (see Table 3 for scoring and
interpretation). The formula for calculating the Hilsenhoff index is HI =
i*ti)/(n)(100), where xi = number of individuals within a species, ti = tolerance
value of a species, and n = total number of organisms in the sample. The second
metric was the EPT index, the percent of Ephemeroptera, Plecoptera, and
Trichoptera individuals in a sample; this metric should increase with improved
site condition. These three taxa are considered to be the most sensitive to
disturbance and the least tolerant of poor water quality; they also have broad
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array of functional morphologies and habitat use. Therefore, the higher the
percentage EPT individuals, the better the water quality and habitat complexity
(Barbour et al. 1992). Plecoptera were also used separately because stoneflies
are consistently the taxon most intolerant of sedimentation and organic pollution
and are not necessarily present even though a stream might have a high EPT
index (Surdick and Gauphen 1978). Use of Plecoptera abundance twice in the
IBI was justified as a way to increase IBI sensitivity to stream degradation. Taxa
Richness provided a measure of diversity, another metric expected to increase
with improved water quality. Percent predators provided a metric of ecosystem
condition by describing how well the community supported top predators. While
taxa of the predatory guild have varying responses to water quality, their
presence was an indication of an environment capable of supporting a multilevel
food web (Gross et al. 2009). Percent Diptera, a highly tolerant taxonomic
grouping, generally increases with stream degradation or water quality
impairment (Barbour et al. 1999, Harrington and Born 2000). Percent Elmidae, a
taxon shown to be particularly responsive to mining effluent, was expected to
decrease with decreased water quality (Garcia-Criado and Fernandez-Alaez
2001).
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Biotic Index Water Quality Degree of Organic Pollution
0.00-3.50 Excellent No apparent organic pollution
3.51-4.50 Very good Possible slight organic pollution
4.51-5.50 Good Some organic pollution
5.51-6.50 Fair Fairly significant organic pollution
6.51-7.50 Fairly poor Significant organic pollution
7.51-8.50 Poor Very significant organic pollution
8.51-10.00 Very poor Severe organic pollution
Table 3. Scoring rubric for the Hilsenhoff family-level index from Hilsenhoff (1988).
Invertebrate IBI
Metric Value 1 3 5
Hilsenhoff Index 6-10 4.25-6 <4.25 Percent Ephemeroptera/Plecoptera/Trichoptera Individuals (EPT Index) <20% 20-40% >40% Percent Plecoptera <5% 5-12% >12% Taxa Richness (Families) <17 17-22 >22 Percent Predators <10% 10-21% >21% Percent Diptera >26% 10-30% <10% Percent Elmidae <10% 10-30% >30%
IBI Score = [Total points/number of metrics] x 20
Table 4. Invertebrate IBI with each of its seven component metrics and the ranges of values used to obtain the final IBI Score.
Vegetation Index
The vegetation index developed by Weixelman (2003) measured condition
by a combination of seral status (e.g., later seral status indicating better condition
than early), functional groupings (e.g., obligate wetland plants indicate better
condition than facultative or upland plants), amount of bare ground (bare ground
being both vulnerable to erosion, and an indicator of disturbance), and plant
18
rooting depth (an indicator of both soil compaction and seral status). See
Weixelman (2003) for metric values developed for different meadow types and a
more in depth description of methods.
Vegetation Index Metric Value 1 3 5
Hydric Type Meadow Seral Status/Functional Guild
>50% Low function
>50% Moderate Function
>50% High Function
Percent Bare Ground 0-4% 5-9% >9% Root Depth (>100 roots/dm2) <10 cm 10-19 cm >19 cm Mesic Type Meadow Seral Status/Functional Guild
>45% Low Function
>55% Moderate Function
>45% High Function
Percent Bare Ground >13% 7-13% 0-6% Root Depth (>100 roots/dm2) <10 cm 10-17 cm >18 cm Xeric Type Meadow Seral Status/Functional Guild
>45% Low Function
>55% Moderate Function
>45% High Function
Percent Bare Ground >13% 8-13% <8% Root Depth (>100 roots/dm2) 0-3 cm 4-6 cm >6 cm IBI Score = [Total points/number of metrics] x 20 Table 5. Vegetation Index providing metric values for each of the three types of montane meadows encountered from Weixelman (2003).
Habitat Index
The EPA habitat assessment index developed by Barbour et al (1999)
was designed to provide an assessment of general habitat conditions. While the
original index was based on total possible score of 200, we adjusted the scoring
to match our other IBIs on a 100-unit scale. (See Appendix 1 for metrics and
scoring rubric).
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Index Interpretation
We calculated scores for each of the 5 indices for all sites with complete
records for all of the parameters measured (n=70). For the purposes of
comparing the IBIs, meadow sites were excluded if they were either fishless or
not all components of the survey were conducted (9 sites excluded). This helped
to ensure that sites were not penalized for lacking fish, because most fishless
sites were ephemeral and could not support fish, rather than reflecting a
disturbance that had extirpated a historical fish population. We felt that
comparing ephemeral streams to perennial streams would introduce more bias to
the analysis than omitting a small number of streams that did not have year
round water but would confound the results. Following the habitat index of
Barbour et al. (1999), we broke index scores into four equal categories to provide
a general verbal interpretation of the habitat conditions found in that range of
scores. This was merely to provide a broader context for IBI scores and to
facilitate a gross comparison of overall site condition of the montane meadows
we surveyed, a particularly important feature in explaining results to stakeholders
and managers. Under this scheme, an index score of 20-40 indicated poor
ecological condition with heavy degradation that was either active or
unrecovered; 41-60 indicated marginal ecological condition in which considerable
degradation had occurred, but the site had either stabilized with loss of function,
or had not yet reached a severe level of degradation; 61-80 indicated fair
ecological condition in which observable degradation was present and the habitat
was capable of supporting all but the most sensitive taxa, and 81-100 indicated
20
good ecological condition in which there was little observable degradation to
habitat and the most sensitive taxonomic groups were represented.
Results
We compared the five IBIs based on their mean scores, ranges, and the
percent of sites in each condition class. The five indices provided different
indications of meadow condition (Table 5). The native fish and amphibian index
indicated the lowest proportion of sites in good condition, while the vegetation
index indicated the highest proportion of sites in good condition. The vegetation
index and habitat index provided very similar results and were significantly
correlated (r= 0.60, p<0.01, Table 6). The invertebrate index was weakly
correlated with the habitat index (r=0.42, p<0.01), but indicated fewer total
meadows in good condition than the habitat index.
Score 20-40 41-60 61-80 81-100
Index Poor Marginal Fair Good Fish-only 8 67 16 9 Native Fish and Amphibian 40 41 16 3 Invertebrate 13 31 36 20 Habitat 3 13 34 50 Vegetation 1 17 31 51 Table 6. Percentage of meadows (n=70) for each of the five meadow ecological condition indices, in
Fish-only IBI
The fish-only IBI had a mean score of 58.7, (SE =1.9), a maximum of 100,
and a minimum of 20 (Table 6). The majority of the sites (67%) scored in the
marginal category, while 9% scored in the good category, 16% scored in the fair
category, and 8% scored in the poor category. We captured 6030 fishes of 23
21
species over the two field seasons. The most abundant taxon captured was
speckled dace (37% of fish captured), followed by trout species (31%), and
Paiute sculpin (17% ). Of the trout species, brook trout were the most numerous
(18% of fish captured), followed by brown trout (9%), and rainbow trout (5%).
Lahontan redsides made up 6% of the catch. Biomass ranged from 0.1 to
447g/m3. Abundance ranged from 2 to 840 fish per reach.
Figure 1. Score distribution of Fish-only IBI. n=70
Native Fish and Amphibian IBI
The native fish and amphibian index had a mean of 48.5 (SE =1.9), a
maximum of 85, and a minimum of 20 (Table 6?). The scores for this index were
more heavily weighted towards the lower end of the index with 41% of sites rated
in marginal condition, 40% in poor condition, 16% in the fair condition, and only
3% in good condition. This IBI was structured to give high scores to sites that
mainly contained native fishes and amphibians. Overall, the index indicated that
native fish and amphibian populations are generally not abundant or even
present in the meadows we studied. Amphibians in particular were very rare and
22
thus drove the index values down. While many of the survey sites occurred in
historic mountain yellow-legged frog (Rana sierrae) and Cascade frog (Rana
cascadae) habitats that were shown by the Grinnell surveys to contain many
amphibians, none were encountered in either the 2006 or 2007 surveys (Grinnell
and Storer 1924). In the 2006 field season, only 25 of the study sites had reptiles
or amphibians present. We observed Pacific treefrogs (Pseudacris regilla) at 16
of the sites. Non-native bullfrogs (Lithobates catesbeianus) occurred at 3 of the
sites. Three percent of the sites contained California toads (Bufo boreas
halophilus). Reptile observations included western terrestrial garter snakes
(Thamnophis elegans) (7 of the sites), western aquatic garter snakes
(Thamnophis couchii) (21 of the sites), gopher snakes (Pituophis catenifer) (1
site), alligator lizards (Elgaria coerulea) (1 site), and western fence lizards
(Sceloporus occidentalis) (1 site). In the 2007 field season, we found P. regilla at
1 site, and T. elegans at 4 of the 11 sites.
Figure 2. Score distribution of Fish and Amphibian IBI. n=70
23
Invertebrate IBI
The invertebrate IBI scores indicate that 20% of sites were rated as being
in good condition, 36% in fair condition, 31% in marginal condition, and 13% in
poor condition. The invertebrate index had a mean of 65.4 (SE, 2.1), a maximum
of 100, and a minimum of 27. The mean taxa richness (families) was 19.4,
ranging from a minimum of 9 to a maximum of 29 families. The mean EPT index
was 0.53, ranging from 0.02 to 0.89. The Hilsenhoff index (Hilsenhoff 1988), a
frequently used index that indicates organic pollution through taxa tolerance
values and relative frequency, had a mean of 3.9 and ranged from 2.2 to 7.4,
indicating significant variability in water quality throughout the survey meadows.
The ephemeropteran family, Baetidae, was the most commonly encountered
abundant taxon, dominating (most abundant) the community in 26 of the sites.
The dipteran family, Chironomidae, was the next most abundant family,
dominating at 18 of the sites. Other abundant taxa included the dipteran family
Simuliidae, and the ephemeropteran families, Heptageniidae and Tricorythidae,
dominant in 12, 10, and 3 of the sites respectively.
Figure 3. Score distribution of Invertebrate IBI. n=70
24
Habitat Index
The results of the habitat index (using the RHA sheet in Appendix 1)
indicate that overall meadow condition is better than indicated by either the fish
or invertebrate indices. According to the habitat index, 51% the sites were in
good condition, 31% of the sites were in fair condition, 17% of the sites were in
marginal condition, and 1% of the sites were in poor condition. The habitat index
had a mean of 76.0 (SE, 1.7), a maximum of 97, and a minimum of 25 (Table 7,
Figure 6). The results were strongly skewed to the right with 82% of the sites
rated as in either good or fair condition.
Figure 4. Score distribution for Habitat Index. n=70
Vegetation Index
The results of the vegetation index indicated that 51% of the meadow sites
were in good condition, 31% of the sites were in fair condition, 17% of the sites
were in marginal condition, and 1% of the sites were in poor condition. The
vegetation index had a mean of 79.0 (SE 1.8), a maximum of 100, and a
minimum of 33 (Table 7, Figure 6). The survey sites were predominantly (57)
mesic (moist) type meadows, with 13 hydric (wet) meadows, and 3 xeric (dry
25
meadows). In many cases, moisture class was not consistent throughout the
entire meadow; 12 of the meadows were mixed hydric/mesic type, 9 were mixed
mesic/xeric type, and 1 site had all three types, hydric/mesic/xeric, represented.
Figure 5. Score distribution for Vegetation Index. n=70
Statistic Fish-only IBI
Native Fish-Amphibian IBI
Invertebrate IBI
Habitat Index
Vegetation Index
Mean 58.7 48.5 65.4 76.0 78.9 St.Err. 1.9 1.9 2.1 1.7 1.8 Kurtosis 0.76 -0.65 -0.76 2.3 0.17 Skewness 0.51 0.25 -0.16 -1.31 -0.61 Range 80 65 73 72 67 Minimum 20 20 27 25 33 Maximum 100 85 100 97 100 Table 8. Summary statistics for each of the five meadow condition indices including mean, standard error, kurtosis, skewness, range, minimum, and maximum.
26
Score (20-100)
20 40 60 80 100
Fish-only IBI
Native Fish and Amphibian IBI
Invertebrate IBI
Habitat index
Vegetation Index
Box and Whisker Plot of Index Means and Ranges
Figure 6. Box and whisker plots of means, minima, and maxima of each of the indices.
Discussion
The five indices did not consistently give meadows the same ecological
condition ratings (Figure 6). The vegetation and habitat indices tended to rate
meadows as being in better condition than the aquatic indices, especially the
native fish and amphibian index. The habitat index was correlated with the
invertebrate IBI and the vegetation index, but neither of the fish IBIs correlated to
any other index (Table 6). The lack of correlation between some indices
suggested that each index is responding to different drivers or that they are
responding at different temporal or spatial scales, or that the meadows
themselves are sufficiently heterogeneous that they have different capacities to
support ecosystem functions and services (Stoffels et al. 2005). For example,
invertebrate communities might respond negatively to pollution of the water by
27
livestock (manifested as increased nitrogen and phosphorus as well as increased
turbidity), while the surrounding vegetation might respond positively to the input
of additional nutrients. The mechanisms that drive structure and organization in
invertebrate communities range from regional climatic drivers to microhabitat
changes at the reach level and below (Stoffels et al. 2005). Fish, with their
greater mobility, respond predominantly to influences outside of the reach level
such as flow, temperature, and land-use at the watershed level (Lammert and
Allen 1999). However, their presence and distribution within a given reach
indicate localized habitat preferences (Lammert and Allen 1999). Overall, the
different responses of fish and amphibian indices, invertebrate indices, stream
habitat, and vegetation indices represent the differential results of legacy effects,
on-going changes (e.g., recovery from anthropogenic effects), watershed effect,
variable natural conditions, and management actions.
Native fish and amphibians
The two fish-based indices indicated the poorest condition of the
meadows sampled (Figure 6). There are likely several factors contributing to this.
Widespread stocking of both native and non-native hatchery trout over the last
century has resulted in either fish being present in historically fishless streams or
streams that no longer support the native fish fauna. The streams surveyed in the
study were predominantly small first and second order streams with small
catchment areas. The native trout in our study area consisted of several
subspecies of rainbow and redband trout (Oncorhynchus mykiss, O.m.
aquilarum, and O.m. stonei) on the west slope and northern Sierra/southern
28
Cascades, and Lahontan and Paiute cutthroat trout (O. clarki henshawi and O. c.
seleniris) on the east slope, and were only present at a small proportion of the
survey sites.
The streams surveyed in the more northerly meadows (i.e. north of Lake
Tahoe) tended to have intact native fish faunas and none were historically
fishless, whereas the streams surveyed in the more southern areas tended to be
in areas that were likely either historically fishless (due to steep gradients
downstream) or no longer support the historic native fish fauna due to extensive
stocking. The dominant trout taxon in many of the sites was non-native brook
trout (Salvelinus fontinalis), a species that maintains large populations in
headwater streams ,excludes other species, and has high densities of individuals
with small body sizes (Moyle 2002, Letcher 2007).
The fact that the dominant salmonids throughout the study were not native
indicates considerable alteration of the historic fish fauna in meadow systems.
the result of stocking at some point. In any case, the nearly ubiquitous stocking of
native and non-native trout throughout the Sierra Nevada has been associated
with declines of native amphibian populations, especially those of frogs (Knapp
and Matthews 1996, 2000; Knapp 2005), although aerial drift of pesticides and
non-native diseases may also have played a role (Davidson et al. 2002). Thus
the marked absence of amphibians in the meadows sampled provides a clear
case of legacy effects on meadow-associated taxa, rather than being a result of
specific contemporary meadow habitat conditions. However, it may be that the
29
extensive grazing that characterized the late 19th and early 20th century and the
associated erosion and incision that occurred on many of the meadows also had
negative impacts on amphibian populations prior to the introduction of non-native
fishes. The legacy effects of stream channel changes may continue to affect
amphibian populations and confound the effects of fish stocking, thus preventing
recolonization.
Invertebrates
The invertebrate IBI provided a slightly more positive assessment of
meadow condition than the fish/amphibian-based IBIs. Using invertebrate
communities to assess habitat condition is a commonly-used tool; however, most
invertebrate indices are designed for high gradient, cold temperature streams
(Harrington and Born 2000). In using invertebrates to assess meadow condition,
we took into account the distinctive habitat conditions encountered in many
meadow systems. Montane meadows are by definition mostly low gradient
systems where the substrate is often sandy or silty, an inherent condition that will
naturally limit production of coarser substrate-associated individuals which are
often the species associated with better water and habitat quality in high-gradient
systems. Water velocity in meadows is commonly low and there is frequently little
woody or shrubby riparian cover, which creates conditions of high solar radiation
and warm temperatures, particularly in low-volume streams with small catchment
areas. This will cause the invertebrate community to contain more tolerant taxa,
which can give the impression of impairment, but may actually represent the
unimpaired community for that type of habitat. The ranges for scoring the
30
invertebrate metrics were designed to capture the range of variability we
observed in the field.
While the invertebrate community appeared to be a robust indicator for
meadow condition, and differentiated between sites well, there are several
limitations to relying solely on using aquatic invertebrates as an indicator of the
condition of the entire meadow system. The first limitation is that once the stream
system has stabilized, even if it has entered an alternative state, invertebrate
communities may not reflect historical impacts. For example, our data indicated
that a meadow stream that is actively eroding with either substrate scouring or
siltation occurring will generally be reflected very accurately by low scores for the
invertebrate IBI. However, meadow streams that have significant gullying that
has stabilized may have been recolonized by the historic invertebrate community,
providing a high index score. Yet the meadows themselves in such situations
often have a lowered water table and a shift of the vegetation community towards
more xeric plants. . Therefore, invertebrate sampling does not necessarily reveal
legacy effects that may be reflected in the other indices. If the substrate has not
been greatly altered in the incision process, the invertebrate community will
generally recover within months to years after an impact. This will indicate the
current status of the stream bed and new channel, but will not provide a signal for
the historic impacts represented by the loss of the majority of the non-stream
meadow habitat. However, the index does provide an accurate assessment of
existing instream habitat status.
31
Another limitation is that upstream conditions impact invertebrates (as well
as the other aquatic indicators), but upstream conditions cannot easily be
causally separated from local habitat conditions. Erosion from an upstream
timber harvest area might have a profound impact on the invertebrate community
downstream through sedimentation, but the cause of those impacts may not be
present in the meadow itself.
Invertebrate communities are sensitive to changes in condition on smaller
temporal and spatial scales and respond to a variety of factors, which can make
it potentially challenging to tease apart the key factors that shape the community
(Lammert and Allan 1999). For example, increased nutrient loading from
throughout the watershed can increase primary production in meadow streams
and result in invertebrate communities that are less driven by local habitat
conditions in the meadow (Jackson et al. 2007). Temperature increases from
agricultural return water can also alter the community structure toward a more
tolerant community (Jackson et al. 2007). Sedimentation favors some taxa over
others (Anagradi 1999). Determining the predominant influences to an
invertebrate community requires collecting additional information on water
quality, stream geomorphology, potential upstream factors, temperature, and
substrate. However, these data are also important in understanding the overall
condition of the system, are used in several of the other indices, and are not
overly difficult to obtain.
The invertebrate IBI had a greater range and more varied results than
either the vegetation index or the habitat index. The invertebrate IBI best
32
described short-term conditions within a meadow stream system and showed
particular sensitivity to the effects of scouring, siltation and sedimentation,
organic pollution, and thermal changes, all important components of overall
ecological condition and function in meadow systems.
Habitat Index
The habitat index showed that most meadows in the study were either in
good condition or had significantly recovered from past degradation. The heavily
skewed results of the habitat index result from it being designed to be very broad
and take into account the full spectrum of stream conditions from pristine to
catastrophically impacted. For meadows, the habitat index measures physical
changes to geomorphology--particularly incision and erosion--which is a fairly
narrow range of the conditions measured by this index. Even the most altered
meadow system will not score as low as a heavily degraded urban stream with
considerable channel alterat
sensitivity for assessing streams in a comparatively natural state, there were
measurable differences between entrenched, eroding meadow streams versus
the meandering streams connected to their floodplains typical of meadows
regarded as being in good condition. This index was significantly correlated with
the vegetation index (r=0.60, p<0.05), and provided an almost identical
assessment of overall conditions by category. It was also significantly correlated
to the invertebrate IBI (r=0.42, p<0.05), thus tying the aquatic and terrestrial
systems together in a tangible way (Table 6).
33
Vegetation Index
As the standard quantitative method of meadow condition assessment
used by the USDA Forest Service (Weixelman et al. 2003), we were very
interested to see how this index related to the other indices, although we
deviated somewhat from standard procedure by selecting survey sites within 10
m of our stream sampling areas. The index uses only three general metrics to
determine vegetation condition and does not address how plant species richness
affects condition. Depending on the meadow type (hydric, mesic, or xeric)
diversity may factor differently. The literature indicates that mesic meadows (the
main type in this study) are frequently the most speciose, while hydric meadows
are often dominated by just a few species of sedges and other water-loving
plants (Winward 2000, Castelli et al. 2000, Chambers et al. 2004).
The limitations of the vegetation index center on the fact that detectable
shifts in seral stages generally only occur with significant hydrologic alteration
(i.e., drops in the water table). It is not clear how many of our sites have shifted
meadow type due to slight changes in the water table, given a general lack of
historical data. Moderate drops in the water table due to a minor amount of
stream down-cutting can result in subtle shifts in vegetation communities,
particularly at the meadow upland/meadow ecotone, farthest from the water table
and stream, which would not be reflected in the index, especially as we used it
(Chambers et al. 2004, Darrouzet-Nardi et al. 2006).
Overall, the vegetation index provided the most positive assessment of
Sierra-wide meadow condition. However, it lacked the responsiveness of the fish
34
and invertebrate indices (as evidenced by their wider range of score values),
although once down-cutting has passed a certain threshold, there can be a
detectable change in the vegetation community (Micheli and Kirchner 2002). The
results of the vegetation surveys were nearly identical to those of the habitat
index in terms of means and ranges (Table 7). The narrow range of responses
indicated that more metrics are needed to make the index more sensitive to
meadow condition.
Conclusions
Our study indicates that the five indices do not provide the same
evaluation of meadow condition. Instead, index values appear to be controlled by
different spatial and temporal factors and contexts. Many of these interactions
are complex and cannot be explained in the context of this study, but
nonetheless, the differences in response among the indices are an indication that
multi-scalar factors are influencing community structure and organization in
meadows. This heterogeneity leads to the conclusion that we cannot use a single
method of assessment in such complex systems.
A basic premise of this study, validated by our results, is that the meadow
component most sensitive to human-driven change is the stream. Stream banks
are most likely the first location to degrade, especially from cattle grazing and
vehicle use, but they are also the area where meadow recovery is often most
evident (Winward 2000, Castelli et al. 2000, Micheli and Kirchner 2002).
Therefore, vegetation surveys should include this sensitive zone, but as well as
monitor upland vegetation as an indicator of meadow condition in relation to
35
hydrologic change. While the each of the five indices we used appeared to be
responding to different impacts at different temporal, spatial, and organizational
scales, each provided essential information about the condition of the meadow
and its resources. While the assessment of meadow condition changed based on
which index was used, each provided an assessment of different components
important to the overall condition of a meadow system. Indeed, each is integral to
understanding cumulative effects of past events, trajectories of recovery, and
opportunities to change management before additional impacts occur. Utilizing a
multimetric approach that accounts for both terrestrial and aquatic habitats is the
best opportunity to assess meadow condition, particularly given disproportionate
importance of these systems in the Sierra Nevada landscape. To accept the
results of just a single index in the absence of the others is potentially misleading
and costly.
36
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