A Comparison of Terrestrial and Aquatic … MS Thesis_Montane...based assessments of meadow condition. We surveyed 1) fish, 2) reptiles, 3) amphibians, 4) aquatic macroinvertebrates,

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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

ii

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.

1

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

2

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

3

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

4

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

5

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

6

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

7

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

8

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.

9

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

-

-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

10

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

12

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

13

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.

14

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).

15

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

16

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).

17

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).

19

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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