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THESIS

It{I'lllljiflllil‘ilfllflllllfllljifiII

LIBRARY

Michigan State
University

 

 

 

 

This is to certify that the

dissertation entitled

Some effects of riparian habitat alteration
on lotic invertebrate ecology

presented by

Roger Malcolm Strand

has been accepted towards fulfillment
of the requirements for

Ph .D degree in Entomology

 

 

    

Major prof ssor

Date _lJ_Sep_t_emb_er_._l_9_9_6

MS U is an Affirmative Action/Equal Opportunity Institution 0-12771

«r—«r A —

4.. '—-—-

PLACE It RETURN BOX to remove thte checkout from your record.
TO AVOID FINES return on or before date due.

DATE DUE DATE DUE DATE DUE

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

MSU le An Affirmative Action/Equal Opportunity lnetltulon
W

ans-9.1

 

SOME EFFECTS OF RIPARIAN HABITAT ALTERATION ON LOTIC
INVERTEBRATE ECOLOGY

By

Roger Malcolm Strand

A DISSERTATION

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY

Department of Entomology

1996

ABSTRACT

SOME EFFECTS OF RIPARIAN HABITAT ALTERATION ON LO'TIC
INVERTEBRATE ECOLOGY

By
Roger Malcolm Strand

Three studies were conducted to determine the effects of allogenic alteration of riparian
habitat on benthic invertebrate ecology.

Study 1. Community abundance, richness, and diversity were monitored during two
radically different landscape alterations which occurred simultaneously in a northern
Michigan watershed. This three-year study incorporated analyses of introduced substrata
samples taken from an upstream forested reach undergoing the transition from unrestricted
flow to beaver pond and from a perennially overgrazed pasture reach experiencing the
initial effects of cattle exclusion. Beaver colonization and cattle exclusion both resulted in
initial abundance increases. Recovery from overgrazing led to increased diversity while
disturbance by beaver resulted in decreased diversity. Impoundment produced large
changes in community structure and functional composition. In contrast, eattle exclusion
did not meaningfully affect community composition. Disturbance by beaver fostered
increased abundance of broadly distributed habitat generalists while recovering and
undisturbed habitat harbored less common taxa with specialized habitat requirements. The
forest reach supported many taxa not found in the overgrazed pasture reach, but the pasture
reach contained few taxa not found in the forest reach. Therefore, losses of forest reaches
that result from human activities may have relatively large impacts on watershed
biodiversity.

Study 2. The initial response by predacious stonefly nymphs (Paragnetina media) to
intense sedimentation was examined in field trials with chambers designed to give nymphs
the choice to stay or reposition in response to sediment additions. The results indicate that
the often posited immediate-escape response to the onset of intense sedimentation may not
be commonly enacted by stonefly nymphs or perhaps by other relatively immobile, lotic
insects.

Study 3. Net-spinning caddisflies often thrive in the heavily sedimented waters of
midwestem agricultural catchments despite the presumed costs of sedimentation-induced
net fouling. A laboratory experiment was conducted to determine whether larval growth
and survival of Hydropsyche bettem' and Ceratopsyche spama are affected by intense,
episodic sedimentation. Larvae of both species suffered increased mortality in
sedimentation treatments relative to controls and taxa were differentially affected as H.
betteni significantly outperformed C. spama. Sedimentation did not, however, alter the
relative growth rate of either species.

ACKNOWLEDGEMENTS

D]. N guyen's consistency and amazing technical ability made the study presented in
Chapter 2 far more accurate and interesting than it would have been had she not been
dedicated to its completion. I thank Bill Morgan for helping me with all aspects of my
research. Roger W. Strand worked tirelessly in the field during his annual visits to
Carlson Creek, he assisted in the design and development of the apparatus used in the
Chapter 3 study, and he also contributed some fantastic illustrations that I have used
when presenting this research, one of which appears on page 13. I gratefully
acknowledge the guidance provided by my advisory committee: Drs. W.T. Cooper,
D]. Hall, L. Hedin, D. long, R.J. Stout, E.D. Walker and particularly that offered by
my major advisor, Dr. Richard W. Merritt. Dr. Matt Ayres' much-appreciated
consultation on mtters concerning data analysis, interpretation, and presentation vastly
improved this dissertation. The Department of Entomology and the Center for
Integrated Studies in Science provided me with several great opportunities to develop
my teaching skills. The Center for Water Resources provided research funds. I thank
the staffs of Department of Entomology the Center for Integrated Studies in Science for
their professionalism and enthusiastic assistance. I also thank my colleagues from the
Merritt Lab and Dr. Dan Herms for their advice and assistance in the field and
laboratory. Finally, I am proud to have been part ofthe greatest ofall 1M sports
franchises, the DINOSORES.

TABLE OF CONTENTS

List of Tables

 

List of Figures

 

Chapter 1. Benthic invertebrate ecology in Ameriea's pasture- and
rangeland streams

 

 

Literature Cited

Chapter 2. Benthic invertebrate response to cattle exclusion and beaver

 

colonization in a small, Michigan stream
Overview

 

Introduction

 

 

Methods
Results

 

 

Discussion

 

Summary
Literature Cited

 

Appendix 1. Forest-reach taxa organized by feeding guild and year of
collection. letters after taxa names indicate sample type(s) from which

animals were collected: R=rock; W=wood; L=leaf packs; D=drift ........

Appendix 2. Pasture-reach taxa organized by feeding guild and year of
collection. Letters after taxa names indieate sample type(s) from which

animals were collectect R=rock; W=wood; L=leaf packs; D=drift -----

Chapter 3. The behavioral response of the stonefly Parugnetina media
(Plecoptera: Perliche) to the onset of intense sedimentation
Introduction

 

 

Methods

Page
iv

H

14
14
15
20
24
31

36

52

53

55
55
57

 

Results
Discussion

 

Literature Cited

 

Chapter 4. Effects of episodic sedimentation on the net-spinning
caddisflies Hydropsyche betteni and Ceratopsyche spama
(Trichoptera: Hydropsychidae)

 

Introduction

 

Methods

 

Results

 

 

Discussion
Limture Cited

 

Appendix 1. Contingency analyses testing whether Hydropsyche betteni
and Ceratopsyche spama survival differed in sediment treatment and

 

controls with pupae included

 

Appendix 1. Record of deposition of voucher specimens

 

Appendix 1.1. Voucher specimen data

59
59
61

68
68
7 l
73
74
77

87

88
89

LIST OF TABLES

Chapter 2

Table 1. Baseline abiotic data collected from forest- and pasture-reach

 

sites during 1991. Values indicate means i: 1 se
Table 2. Carlson Creek invertebrate tolerance-index values

(from Resh et al 1996). * = values estimated for this study ------------------

Table 3. AN OVA results comparing invertebrate community

characteristics as estimated with rock, wood, and leaf pack samples

Table 4. Contingency analyses testing whether the proportion of
animals representing different functional feeding groups varied
between reaches in rock, wood, and leaf-pack samples

 

Chapter 3

Table 1. ANOVA results comparing the number of stonefly nymphs that

 

selected each behavioral option
Table 2. Particle-size distribution and chemical composition of bank
sediments used in sediment treatment

 

Chapter 4

Table 1. Particle—size distribution and chemieal composition of bank

 

sediments used in sediment treatments

Table 2. ANOVA results comparing relative growth rates of
Hydropsyche betteni and Cenopsyche spama in two sediment
treatments and two controls

 

iv

41

42

43

65

81

82

Table 3. Contingency analyses testing whether Hydropsyche betteni and
Ceratopsyche spama survival differed in sediment treatments and
controls (two tanks each)

 

83

LIST OF FIGURES

Chapter 1

Figure 1. Some effects on lotic ecology of grazing cattle 1n small- stream
riparian zones

 

Chapter2

Figuee 11.99Tgre pasture reach of Carlson Creek in (A) 1991, (B) 1992, and
( )
Figure 2.99Tg1e forest reach of Carlson Creek in (A) 1991, (B) 1992, and
(C) 1 ~
Figure 3. Mean depth and current velocity :t 1 se in one forest—reach and
two pasture-reach stations. Measurements taken during July, August, and
September, 1991-1993
Figure 4. Suspended solid concentration from forest— and pasture-reach
water samples taken monthly, July- September, 1991- 1993.... .... .......
Figure 5. Mean invertebrate abundance, richness, and diversity 1n rock,
wood, and leaf pack sammes
Figure 6. Community functional-fading group composition in
forest-reach samples. Values indicate total number of animals collected.
Functional-feeding groups that comprised less than 1% of total are not
represented on pie-charts. These are: 3 herbivores in 1992 wood
samples and 5 scrapers in 1992 leaf-packs
Figure ’7. Community functional- —feeding group composition in
pasture-reach samples. Values indicate total number of animals collected.
Functional-feeding groups that comprised less than 1% of total are not

 

 

 

 

 

represented on pie-charts These are: 2, 9, and 9 scrapers in 1991, 1992,

and 1993 wood samples respectively; 1 herbivore in 1991 wood samples;
and 6 scrapers in 1991 leaf-pack samples

 

Chapter 3

 

Figure 1. A sedimentation-response chamber
Figure 2. Mean number of stoneflies (2t 1 se) that chose each behavioral
opfion.

 

13

45
46

47
48
49

50

51

67

Chapter 4

Figure 1. Mean NTU :tl se in two sediment-treatment and
t wo control tanks 84

 

Figure 2. Figure 2. Box plots of mean RGR :l: 1 SE of Hydropsyche betteni
and Ceratopsyche spama in two sedimented and two control tanks.
Mean pre-trial dry mass (i 1 se) indieate relative size of fifth-instar
H. bettem' (n=24) and C. spama (n=21) 85

 

Figure 3. Mean survivorship of larval Ceratopsyche spama and
Hydropsyche betteni in two sedimented and two control which is 0.020 ... 86

BENTHIC INVERTEBRATE ECOLOGY TN AMERICA'S
PASTURE- AND RANGELAND STREAMS

Erosion eats into our hills like a contagion, and floods bring
down the loosened soil upon our valleys like a scourge. Water,
soil, animals, and plants - the very fabric of prosperity - react to
destroy each other and us. Science can and must unravel those
reactions, and government must enforce the findings of science
(Leopold 1923).

Seventy-three years have passed since Aldo Leopold issued this advisory to future
ranchers, ecologists, and policy makers and today almost one-half of Earth's terrestrial
space is grazed by domestic animals (Goude 1994). After more than a century of large-
scale grazing in the United States, overgrazed pastures and range have become common
features on the rural landscape (Fleischner 1994, Wissmar et al. 1994, Waters 1995).
Cattle grazing is by far the most common way to use land the American west, where
more than seventy percent of the landscape is grazed Most of this land is held in public
trust and as evidence of ever-increasing habitat damage mounts, policy makers are
being pressured by scientists and concerned citizens to restrict grazing in particularily
sensitive habitats (Armour 1991, 1994; Fleischner 1994).

There is general agreement among range managers and ecologists that many currently
grazed habitats, for example alpine meadows (Kondolf 1994), can not sustain
conflicting societal demands for maximized cattle production and optimized water- and
wildlife-habitat quality ( Fleischner 1994, Li et al. 1994, Brown and McDonald 1995,
Mosely et al. 1993). Despite this reality, grazing on public lands continues with little
regulatory recognition of differential habitat sensitivity, even as restoration costs often
far outweigh fiscal inputs from grazing fees (Minshall 1989, Fleischner 1994, Kondolf
1994).

It is well-established that habitat types vary in nature and magnitude of response to
cattle grazing (Milchunas and Lavenroth 1993, Li et a1. 1994). For example, plant

2

species richness in grasslands within the historic range of bison ofien increases with
cattle grazing intensity, whereas introduction and intensification of cattle grazing in
grasslands outside of the historical range of bison or other large herbivores is known to
cause plant richness to sharply decline (Milchunas and Lavenroth 1993, Hofstede
1995).

Cattle, unlike bison, spend a disproportionate amount of time feeding and wallowing in
streams. This universal intensification, particularily during hot weather, renders
riparian areas in all grazed habitat types especially vulnerable to overgrazing (Kauffman
and Krueger 1984, Resh et al. 1988, Armour et al. 1991, 1994, Platts 1991, Fleischner
1994, Waters 1995). Most of the concern over grazing-induwd riparian and instream
habitat damage has been centered on the consequent reductions of game-fish production
that typically occur in heavily grazed catchments (Armour et al. 1991, 1994, Fleischner
1994). There is still, therefore, much left to learn about grazing effects on the
numerically (Hynes 1970) and often energetically dominant animals (Waters 1984) that
comprise lotic invertebrate communities.

This chapter provides an overview of cattle-grazing effects on instream and riparian-
zone habitat that are known or suspected to affect benthic invertebrate ecology.
Subsequent chapters include an analysis of management-induced recovery of a
perennially overgrazed, northern-Michigan stream and two studies conducted to fill
voids in understanding of grazing-induced sedimentation effects on aquatic insect
behavior, growth, and survival.

General efiects of livestock grazing on stream ecology

Overgrazing is a contemptuous term that remains undefined by range managers
(Fleischner 1994), but has nonetheless been determined to be the cause a variety of
negative effects in aquatic ecosystems (Minshall et al. 1989, Fleischner 1994, Waters
1995) (Figure 1). Overgrazing (sensu Ieopold 1923) in riparian areas is indicated by
total vegetative denudation along banks and channels which results in increased erosion
(Kauffman et a1. 1983 b, Gamougoun et al. 1984), sedimentation, eutrophieation
(Odion et al. 1988, Waters 1995), and thermal regime variation (Li et al. 1994) (Figure
l). Benthic invertebrates are sensitive to fluctuations in all of these environmental
characteristics both directly through modification of physiological processes (Hynes
1966, 1970; Sweeny 1993) and indirectly through alteration of substrata characteristics
(Kauffman and Krueger 1984, Waters 1995).

Efiiects of riparian vegetation transformation

Intense grazing pressure in riparian forests results in drastically diminished tree
regeneration in aging pastures. This process typically ultimates in transformation of
vegetative community from forest to grassland (Fausch and Bramblett 1991, Li et al.
1994). The consequently reduced inputs of abscised leaves and woody debris to
streams causes the elimination food and shelter for a diverse array of benthic
invertebrates (Merritt and Cummins 1996) and fish (Hynes 1970). Stream structural
and biogeochemical characteristics are also altered by reduction of debris-dam habitat as
habitat heterogeneity is lessened (Sedell 1990), and important sites for carbon (Meyer et
al. 1988) and phosphorus dynamics (Munn 1989) are eliminated.

The structural changes associated with riparian-zone denudation may also remove cues
necessary for habitat recognization by aerial adult aquatic insects on mate-location or
dispersion flights. The shift in dominance between two black fly species was
demonstrated by Tim (1994) to be an example of this type of differential habitat
selectivity by ovipositing females. After timber was cleared from a reach of a Rhine
River tributary, the opportunistic Simulium omatum rapidly displaced S. vernum as
the dominant black fly in the stream. There were no changes in substrata, water
chemistry, or resource availability to account for the differences. However these
species do behave differently when seeking oviposition sites. Simulium vemum
prefers shaded riffles for egg deposition, whereas S. omatum only oviposits in riffle
exposed to full sun. Because many other aquatic insects use physical cues when
seeking oviposition sites (Wallace and Anderson 1996), it is probable that discovery
awaits other such direct effects of grazing-mediated riparian transformation on benthic
community composition.

Efii'cts of increased erosion

Bank sediments exposed by overgrazing erode into overland flow following spates and
rapid snowmelt. These processes immediately elevate concentrations of suspended
solids (Waters 1995), nitrogen, and phosphorus (Mosely et al. 1993) and eventually
blanket stream beds with deposited sediments (Waters 1995). Substrata burial can be
lethal to benthic invertebrates in extremely high levels (Thomas 1985) or in sensitive life
stages such as immobile pupae (Rutherford and Mackay 1986).

Benthic invertebrate communities often experience functional and taxonomic changes in
response to persistent sedimentation. Typical transitions feature the replacement of
animals that require solid substrata with those that thrive in thick deposits of soft
substrata (Nuttall and Bielby 1973, Quinn et al. 1992, Waters 1995). For example,
Strand and Merritt (Chapter 2) found that the relative proportional abundance of filter-
feeders to omnivorous gatlrerers, a measure of particulate transport versus deposition
and of the availability of solid substrata (Merritt and Cummins 1996), was far greater in
an undisturbed reach of a northern Michigan stream than it was in an overgrazed reach.

Habitat simplifieation through substrata burial also is likely to reduce hyporheic-zone
quality. This sub—flow region is saturated with a mixture stream and ground water and
is inhabited a variety of invertebrates, some of which are hyporheic enderrrics and
others that forage and seek refuge from disturbance in hyporheic habitat (Williams
1984, Meyer et a1. 1988, Dahm and Valett 1996). As the habitat of occasional and
permanent hyporheic residents is vanquished by interstitial filling, biogeochenrical
activity at nutrient upwelling sites is also diminished (Grimm et al. 1991, Findlay 1995,
Dahm and Valett 1996, Grimm 1996). This process imposes limits on stream
metabolism through alteration of nutrient cycling rates (Grimm and Fisher 1984).
Grazing-induced hyporheic burial is expected to be particularily destructive to the
benthic ecology of temporary streams where population persistence for many
invertebrates is dependant upon the survival of individuals that wait out no-flow periods
in damp, subchannel interstitial spaces (Fisher and Gray 1983).

Effects of excrement input

Due to the periodic threat of bacterial pollution of drinking water, cattle excrement input
to streams, perhaps more than any other grazing-related problem, concerns to humans
who live downstream from grazed riparian areas. High loads of excrement input are
also suspected to impose direct threats to aquatic life. Foremost among these is the
potential of nutrient enrichment to affect instream biogeochemical processes (Mosely et
al. 1993, Harris et a1 1994).

In seasonally grazed streams, eutrophication that results from waste-mediated
fertilization, principally through increased phosphorous loading (Allen et al. 1982,
Mosely et al. 1993), can drastically reduce dissolved oxygen concentration (Harris et a1.
1994) and increase algal and macrophyte production (Fleischner 1994). Persistent

5

fertilization and selective herbivory by cows often result in the formation of dense
instream accumulations of unpalatable plants such as Cladophora glommerata (Armour
et al. 1991, 1994, Matthews et al. 1994), a widespread filamentous green alga, mats of
which typically contain few, if any, benthic macroinvertebrates (personal observation).
Substrata domination and benthic shading that result from the presence of thick algal
growths may also reduce benthic secondary productivity by limiting the establishment
of the more readily foraged-on five-kingdom amalgam known as biofilm or periphyton
which coats most exposed substrata (Cumnrins 1973, Lamberti and Feminella 1996).

Extremely high levels of excrement input in seasonally grazed streams may act to
further slow instream secondary productivity through ammonia toxicity to invertebrates
and fish. Although ammonia levels rarely exceed the tolerance threshold of most
pasture-stream inhabitants (Overcash et al. 1983), particularily sensitive organisms may
be displaced due to chronic, ammonia-induced damage to gill membranes (Hazel et al.
1979, DeGraeve et al. 1980). Acute ammonia toxicity (1.2-8 mg/l) (Hazel et al. 1979,
DeGraeve et al. 1980) is also a potential problem associated with heavy grazing,
especially during dry, hot weather when stream flow is low and cattle wallow to
alleviate heat stress.

Efikcts of thermal regime alteration

Although cattle grazing does not universally cause pronounced elevations in stream
temperature (Li et al. 1994), thermal regimes are known to vary more in grazed than in
ungrazed streams (Kauffman and Krueger 1984). Li et al. (1994) reported that
insolation increased with increasing grazing activity as did algal and invertebrate
biomass. These conditions combined to render grazed habitat less amenable to trout as
increased temperature caused a decline in the ratio of trout to invertebrate biomass. This
relationship has been detected elsewhere and has been in part attributed to direct heat
stress to trout (Platts and Rinne 1985, Platts and Nelson 1989, Minshall et. al. 1989)
and partly to elevated temperature—mediated increases in the abundance of less-preferred
invertebrate prey (Tait et al. 1994).

Sweeny (1993) documented 2 to 4°C increases attributable to clearing of northeastern
U.S. forest streams. This increase may seem trivial when compared to daily and
seasonal temperature fluctuations in terrestrial realms; however, in the relatively
constant lotic thermal environments, 2-4'C can be a large-enough difference to affect
the physiology and alter distributions of sensitive species (Sweeny and Vannote 1986).

In streams where grazing elevates insolation, consequent restrictions on intolerant
species and concurrent increases in secondary production by thermally insensitive
species (Murphy and Hall 1981, Murphy et al. 1981, Hawkins et al. 1983, Bilby and
Bisson 1987) are therefore expected to result in changes in benthic community
composition and bioenergetics.

Riparian-forest clearing may also affect the metabolism and reproduction of aerial adult
aquatic insects through thermal regime alteration. Tree removal eliminates shaded
resting structure which is utilized extensively, and perhaps required by, adult aquatic
insects. Most holometabolous aquatic insects rely on resources stored during the larval
period to complete their life cycles (Sweeny 1993). Therefore, resource conservation
prior to reproduction may have positive effects on fecundity. Reduction or elimination
of cool, humid resting habitat could thus limit the reproductive potential of sensitive

species.

Aldo Ieopold warned us long ago that overgrazing of livestock can negatively affect
ecological processes in pastured watersheds. Although there is still relatively little
known about the specific effects of grazing on lotic animals, it is certain that the
physical impacts of grazing and trampling can combine to alter channel hydrology,
vegetative composition, and nutrient-retention capacity. The ultimate synergy of these
effects produces simplified lotic ecosystems, less buffered from climatic disturbance,
undergoing heightened stream-water losses of nutrients. This distressing scenario is far
too common in American watersheds where short-term economic gains are commonly
sought despite the glaring portent of ecological calamity.

LITERATURE CITED

Allen, L.H., Jr., J.M. Ruddell, G.J. Ritter, F.E. Davis, and P. Yates. 1982. Land
use effects on Taylor Creek water quality. Proceedings of the Special
Conference on Environ mentally Sound Water and Soil Management, Orlando,
FL. American Society of Civil Engineering, New York.

Armour, C.L., D.A. Duff, and W. Elmore. 1991. The effects of livestock grazing on
riparian and stream ecosystems. Fisheries 16: 7-11.

Armour, C.L., D.A. Duff, and W. Elmore. 1994. The effects of livestock grazing on
western riparian and stream ecosystems 19: 9-12.

Bilby, RE, and RA. Bisson. 1987. Emigration and production of hatchery coho
salmon (Oncorhynchus kisutch) stocked in streams draining an old-grth and
clearcut watershed. Canadian Journal of Fisheries and Aquatic Sciences 44:
1397-1407.

Brown, J.H. and W. McDonald. 1995. Livestock grazing and conservation on
southwestern rangelands. Conservation Biology 9: 1644-1647.

Cummins, K.W. 1973. Trophic relations of aquatic insects. Annual Review of
Entomology 18: 183-206.

Dahm, C.N., and M. Valett. Hyporheic zones. Pages 107-119 in PR Hauer and
GA. Lamberti. MW. New york: Academic Press.

DeGraeve, G.M., R.L. Overcash, and H.L. Bergman. 1980. Toxicity of
underground coal gasification condenser water and selected constituents to
aquatic biota. Archives of Environmental Contamination Toxicology 9: 543-
555.

Fausch, K.D., and R.G. Bramblett. 1991. Disturbance and fish communities in
intermittent tributaries of a western Great Plains river. Copeia 1991: 659-674.

Findlay, S. 1995. Importance of surface-subsurface exchange in stream ecosystems:
the hyporheic zone. Limnology and Oceanography 40: 159-164.

Fisher, 8.6., and L]. Gray. 1983. Secondary production and organic matter
processing by collector macroinvertebrates in a desert stream. Ecology 64:
1217-1224.

Fleischner, TL. 1994. Ecological costs of livestock grazing in western North
America. Conservation Biology 8: 629-644.

Gamougoun, N.D., R.P. Smith, M.K. Wood, and RD. Pieper. 1984. Soil,
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12

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BENTHIC INVERTEBRATE RESPONSE TO CATTLE EXCLUSION
AND BEAVER COLONIZATION IN A SMALL, MICHIGAN STREAM

Abstract. Benthic invertebrate community abundance, richness, and diversity were
monitored during two radically different landscape alterations which occurred
simultaneously in a northern Michigan watershed. This three—year study incorporated
analyses of introduced substrata samples taken fiom an upstream forested reach
undergoing the transition from unrestricted flow to beaver pond and from a perennially
overgrazed pasture reach experiencing the initial effects of cattle exclusion. Prior to
habitat alteration, abundance was higher and diversity was lower in grand sites.
Taxonomic richness was similar between reaches, but the forest reach had a higher
proportion of endemic taxa which produced substantial inter-reach differences in the
proportions of taxa from each functional-feeding guild. Beaver colonization and cattle
exclusion both resulted in initial abundance increases. Recovery from overgrazing led
to increased diversity while disturbance by beaver resulted in decreased diversity.
Impoundment produced large changes in community structure and functional
composition. In contrast, cattle exclusion did not meaningfully affect community
composition. Disturbance by beaver fostered increased abundance of broadly
distributed habitat generalists while recovering and undisturbed habitat harbored less
common taxa with specialized habitat requirements. The forest reach supported many
taxa not found in the overgrazed pasture reach, but the pasture reach contained few taxa
not found in the forest reach. Therefore, losses of forest reaches that result from
human activities may have relatively large impacts on watershed biodiversity.

OVERVIEW

In 1989, legislation was adopted to combat the problem of overgrazing of Michigan's
headwater riparian areas. The program provided farmers with an 80% savings on
wells, improvements to stream crossings, and fence to exclude cattle. Carlson Creek
was one of four upper-peninsula streams selected for these restoration measures before
the program was discontinued in 1992. During the winter of 1990—1991 the pasture-

14

15

land owner, the Michigan Department of Natural Resources, and the U.S. Soil
Conservation Service reached an agreement to restore the channel by fencing cattle out
before the 1992 pasturing season. We were allowed to analyze the response of the
stream invertebrate community to these restoration efforts.

During 1991, cattle had unrestricted access to the pasture reach of Carlson Creek and
instream and near-bank conditions were degraded by years of trampling and wallowing
(Figure 1). In stark contrast, the upstream forest reach was relatively undisturbed
(Figure 2). Two 20—m long sites in the forest reach and two in the pasture reach, were
selected for analysis through the final year of lower-riparian grazing and the first and
second years of recovery. Forest-reach sites were to serve as comparative analogues to
pro-grazing conditions during the first year of the study and as a presumably static
habitat relative to the pasture reach. However, in June of 1992, beaver colonized the
upstream reach forming a pond that eventually exten®d to include both ”reference"
sites at the same time at which cattle were excluded from the pasture reach. This
serendipitous action presented the unique opportunity to simultaneously examine
radically different riparian-habitat alterations occurring along contiguous reaches.

The primary objective of the study was to characterize the responses of forest- and
pasture-reach benthic invertebrate communities to beaver irnpoundment and cattle
exclusion in an attempt to assess the effects on lotic ecology of the onset of, and
recovery from, landscape disturbances.

INTRODUCTION
Livestock grazing efi’ects on benthic invertebrate ecology

Chapter 1 provides an overview of the effects of livestock grazing on stream-dwelling
benthic invertebrates. In general, the effects of cattle grazing in riparian areas have been
determined to be ecologically negative (e.g. Leopold 1923, 1946, Tarzwell 1938, Resh
et al. 1988, Armour et al. 1991, 1994, Fleischner 1994, Waters 1995). However,
there is disagreement among range managers concerning the severity of the likely
consequences of current practices (e.g. Fleischner 1994, Brown and McDonald 1995).
Most of the concern over the repercussions of grazing-induced riparian and instream

16

habitat alteration has been centered on consequent reduction of game-fish production
(Armour et al. 1991, 1994). There is, therefore, much left to learn about grazing
effects on the numerically (Hynes 1970) and often energetically dominant (Waters
1984) animals that comprise lotic benthic invertebrate communities.

When cattle are allowed unrestricted access to streams, they tend to spend
disproportionate amount of time feeding and wallowing in or near water (Van Vuren
1982, Kauffman and Krueger 1984, Fleischner 1994). Grazing in riparian areas is
known to modify lotic biotic processes through alterations of insolation level, substrata
characteristics, and flow regime (Minshall et al. 1989, Armour et al. 1991, 1994,
Fleischner 1994, Waters 1995). By eliminating vegetative cover (Kauffman et al. 1983
a, Platts and Nelson 1985 a, Odion et al. 1988), grazing livestock at high densities
results in increased rates of erosion (Kauffman et al. 1989, Gamougoun et a1. 1984),
sedimentation, eutrophication (Odion et al. 1988, Waters 1995), and thermal regime
variation (Li et al. 1994). Alterations of all of these physical parameters are known to
affect benthic invertebrates through substrata transformation (Kauffinan and Krueger
1984, Waters 1995) and modification of physiological processes (Sweeny 1993).

Habitat restoration through cattle exclusion is a well-established way to improve
standing stocks of benthic invertebrates and fish (Peters 1962, Bjorn 1973, Armour et
al. 1991, 1994, Waters 1995). In perhaps the first study conducted to quantify the
efficacy of stream improvements which include cattle exclusion, Tarzwell (193 8)
demonstrated that invertebrate production and trout growth rate in southwestern U.S.
streams were greatly enhanced during the eighteen months that followed exclusion.
Several other exclusion studies followed (e.g. VanVelson 1979, Stuber 1985) and like
that of Tarzwell (1938), the primary emphasis was on characterizing trout response. In
all of these cases exclusion proved to be a successful way to improve fish production.

The effects of beaver activity on benthic invertebrate ecology

Virtually all habitats on earth are controlled to some extent by organisms whose
activities are influential in determining community composition and biogeochemical
properties (Holling 1992, Jones et al. 1994). The ecological consequences of cattle
grazing well demonstrate the human capacity to modify lotic environments. As
allogenic engineers, organisms that alter their environment and control the flow of
resources to other organisms (Jones et al. 1994), humans are rivaled in lotic ecosystems

17

only by beaver (Castor canadensis) (Naiman et a1. 1986, 1988).

The cyclical colonization pattern that typifies beaver inhabitation of a watershed initiates
with woody vegetation removal during pond formation and ultimates in sedge and
littoral vegetation domination early in the post-abandonment phase (Hammerson 1994).
Like other successional processes in terrestrial plant communities, such as those
initiated by forest-to-pasture transformation, beaver colonization results in altered rates
of secondary productivity and decomposition through transportation and
transmogrification of the products of primary producers (Huston 1994).

Damming—induced flow-rate reductions and depth increases activate a typical lotic-to-
lentic habitat transformation. However, due to beaver-induced modification of
substrata characteristics, the resultant pond is quite unlike other lentic habitats (Naiman
et al. 1988). Beaver feeding and construction activity typically result in the formation
of thick deposits of augmented woody material (Naiman et al. 1986). Wallace and
Benke (1984) detected high levels of invertebrate production in these accumulations,
primarily as a result of increases in larval midge abundance (Diptera: Chironorrridae).
These omnivores are also suspected to benefit from often large inputs of organic
material that result from defecation by waterfowl, fish, and other animals attracted to
beaver ponds (Harnmerson 1994). Although invertebrate secondary productivity can be
elevated by organic inputs associated with pond formation, community functional
diversity typically declines as a result of elirrrination of filter-feeders, biofilm scrapers,
and herbivores (Naiman et a1 1986).

Beaver activity can also strongly affect habitat conditions immediately downstream from
impoundments by entraining silt and modulating flow (Naiman et al. 1986, Harnmerson
1994). For example, the clean gravel substrata critical to the success of an endangered
beetle species (Brychius hungerfordi, Coleoptera: Halplidae) is set up and maintained

in its type locality by a series of beaver impoundments. Much of the current rarity of
this beetle is believed to be attributable to the replacement of beaver by humans as the
principal engineer of northern Michigan streams (Strand and Spangler 1994).

18

Disturbance, perturbation, and recovery

The level of disruption of biotic processes exacted by localized fluctuations in abiotic
conditions is indicative of community resistance to landscape alteration-mediated stress
(sensu Underwood 1989). Organismal response to abiotic stress is believed to be
dependent to a large degree upon historical distrubance regime. Events with extreme
consequences to biota are commonly differentiated from those of lesser impact with the
terms "disturbance" and "perturbation" respectively (Huston 1994).

There is strong evidence of a fundamental trade-off between relative resistance and
resilience to temporally isolated episodes of stress (Hutchinson 1953, Connell 1978,
Huston 1979, 1994). Hutchinson ( 1951, 1953) described those organisms that are
relatively easily displaced by perturbation, but able to quickly colonize newly disturbed
habitat, as ”fugitives" due also to their apparent inability to persist once conditions are
stabilized. The relative proportions of fugitive species and high resistance, long-term
occupants, or "occupiers,” are reflective of a habitat's relative position on a stress
gradient from mild perturbation to catastrophic disturbance. Maximum species richness
is expected in habitats that sustain moderately frequent stress episodes of intermediate
intensity because conditions in these habitats allow for the persistence of populations of
fugitive- and occupier species (Hutchinson 1953, Connell 1978, Huston 1979, 1994).

The relative novelty of a stress event is believed to be more important than its magnitude
in determining resultant effects on biota. If stress frequency, regardless of intensity, is
maintained at a high level over a long period of time, then it is expected that the
organisms that comprise the affected community have in common a relatively high
resistance and resilience to that type of stress, or perturbation. Perturbations are
therefore, not expected to exact lasting changes to community composition. However,
if an event is uncommon and intense enough to cause the elimination of organisms to
proceed at a rate greater than that at which they are able to accumulate, then it is
assumed that the effect of the stressor, or disturbance, will directly alter community
structure and can, if severe enough, have lasting consequences (Hutchinson 1953,
Huston 1994).

Within this theoretical construct, benthic invertebrate dynamics in Carlson Creek from
1991 to 1993 are considered the result of the onset of disturbance by beaver in the

19

forest reach and recovery in the pasture reach from initial land-clearing disturbance and
subsequent decades of overgrazing-generated perturbation.

S ire description

Carlson Creek slowly flows from Kaks Lake, a small undeveloped lake in Michigan's
upper peninsula to an impoundment of the Taquamenon River two miles downstream.
Riparian characteristics change dramatically along its course from climax forest, to cattle
pasture, bog, and ultimately to marsh.

I 991

Table 1 contrasts pre-alteration forest- and pasture-reach substrata particle size and
chemical compositional characteristics as well as selected stream water properties.

From Kaks lake, Carlson Creek flowed at a rate of 0.3 m/second over 3 km of rnixed-
hardwood riparian forest. In this forest reach, Carlson Creek was shallow (less than 40
cm) and unevenly shaded. Open stretches were characterized by Typha and other
typically lentic littoral-zone vegetation; closed zones support little or no instream
vegetation (Figure 2A). The predominant substrata in the forest reach was wood and
mixed gravel and cobble with underlying sand.

Riparian vegetation in the pasture reach consisted of patches of tag alder (A [nus sp.)
along a short grass and exposed soil course. Habitat condition varied from a 8 m-wide
channel cutting through cattle congregation sites at about 0.7 ml second to slower flow
spread out over a 15 m-wide channel coursing around dense clumps of sedge (Figure 1
A). The predominant substrata was up to 25 cm of silt overlying gravel and cobbles.
Cattle were allowed to graze through the forest that demarcates the downstream pasture
boundary. The stream spread out in this reach to a width of about 30 m. It then flowed
slowly through a Typha marsh for approximately 1.5 km to a large impoundment of
the Taquamenon River created to improve waterfowl habitat.

1 992-1993

Beaver moved into the forest reach during June, 1992 and built a darn downstream

20

from both sites. Consequently, flow slowed and stream depth and width increased
dramatically (Figures 2, 3). During the same time period, a fence was constructed to
exclude cattle below the high—water mark of the pasture reach. Inter-reach responses of
riparian and instream landscape alteration are depicted in Figures 1 and 2.

METHODS
In vertebrate sampling

In order to collect the most taxonomically representative set of samples possible without
destroying benthic habitat, a combination of artificial substrata types (rocks, wood, leaf
packs) and drift nets were used to sample benthic invertebrates.

Rocks

The invertebrate fauna of rocks was sampled by introducing similarily sized and
textured field stones (mean area is 0.283 m2) allowing approximately 30 days for the
establishment of biofrlm and subsequent invertebrate colonization. In preliminary
studies, a one month colonization period was sufficient for invertebrate colonization yet
brief enough to avoid total burial. Rock size was determined by wrapping cobbles in
foil, weighing the foil, and extrapolating foil area from a predetermined relationship
between foil area and mass. Rock samples were taken during July, August, and
September, 1991-1993 from two pasture and two forest sites. A total of 155 rock
samples were collected.

Wood

Pieces of cottonwood bark (~ 0.175 m2), selected for its relatively uniform and rough
texture, were anchored to the streambed and collected after approximately 30 days.
Area was determined with the same procedure described for rocks. Wood samples
were taken during September and November, 1991-1993. A total of 108 wood
samples were collected.

21

Leaves

Four-gram, white oak (Quercus alba) leaf packs, selected for their intermediate
breakdown time (Petersen and Cummins 1974), were secured to bricks and placed in
erosional zones facing upstream to simulate natural debris dams (as recommended by
Merritt et al 1979). Packs were removed from streams after 31 days in 1991 and 44
days in 1992. The dramatic reduction in beaver pond size that occurred following a
dam break in 1993 destroyed year three samples. A total of 51 leaf packs were
collected.

In vertebrate drift

Drift was monitored with l-m long, 0.5-m wide, and 0.5-mm diameter mesh drift nets
positioned vertically. Four nets were placed in riffle areas in one forest site and two
pasture sites. Nets were set before dusk and pulled the following morning as
recommended by Wates ( 1972). Animals collected with drift nets were used only for
taxonomic information due to difficulty in interpreting forest-reach drift catches during
beave-pond formation. Drift samples were taken monthly from July to September,
1991-1992. A total 59 drift samples were collected.

Analyses of richness and abundance

Invertebrate samples from the entire study included a total of 35,667 individuals
representing 99 taxa (Appendices 1, 2). Whenever possible, insects wee identified to
genus using keys in Merritt and Cummins ( 1996 a); non-insect invertebrates wee
identified to class or lowe. Abundance and richness as used hee refer to total number
of individuals or taxa corrected for substrata area and colonization period. Therefore
values express relative rate of colonization as well as community variety and
abundance.

Community diversity
The hierarchical richness index procedure (HRI) (French 1995) combines

measurements of taxonomic richness and abundance into a single term that describes, in
the case of the studies presented here, the variety and abundance of benthic invertebrate

22

inhabitants of stream reaches undegoing diffeential disturbance pressure. The HRI
procedure is defined by the equation:

HRI=Z (Sixi)

where i are group ranks (the most abundant group rank = 1, secondmost = 2, and so
on) and si are within group scores (mean abundances), such that all si 2 sin (French
1995).

Functional- feeding guilds

Although not strictly defining for most species, benthic invertebrates typically belong to
one of six primary functional feeding-guilds: filteres, gathees, scrapers, predators,
herbivores, and shredders (Cummins 1973, 1980). Estimates of proportional
abundance, biomass, and richness of these groups in benthic communities have been
shown to be demonstrative of lotic habitat condition (Merritt and Cummins 1996 a, b).
Abundance, as used in functional feeding group analyses hee refers to total number of
individuals collected with each substrate type.

Tolerance indices

Benthic invertebrates have widely varying tolerances to ecological disturbance (Resh et
al 1988) and anthropogenic pollutants (Rosent and Resh 1996). Once these
tolerances are established, it becomes possible to rank organisms by their relative
tolerance index values. Analyses of benthic invertebrate community toleance index
values can then be used to assess past and present habitat conditions (Hilsenhoff 1988,
Resh and Jackson 1993, Resh et al. 1996). The fanrily-level toleance index of Resh et
al (1996), a compilation of past values presented in Hisenhoff (1988), and Ienat
(1993), was used to assess relative tolerance of Carlson Creek benthos undergoing
beaver impoundment and cattle exclusion (Table 2).

Depth, flow, and dissolved oxygen

Depth and flow-rate measurements wee taken at each sampling date. Depth was

23

measured in the middle of the channel at each site the position at which substrates wee
placed. Flow rate was measured by determining the rate at which a fishing float travel
over 2 m of channel. Baseline dissolved oxygen concentration was determined with a
portable DO mete at each site on September 17-20, 1991.

S ubstrata sediment composition

Sediment core samples were taken on July 21, 1991 with an 30 cm-long, 5 cm-
diamete, PVC cylinder driven 20 cm into the substratum, coveed from below and
removed. Particle size- and chemical composition of these was deterrrined by the
Michigan State University Plant Nutrient Laboratory (see Table 1).

Suspended sediments

Monthly water samples were taken from July to September, 1991-1993 by submerging
a 0.5-1, glass jars into the stream from one forest-reach and one pasture-reach site.
Samples wee vacuum filtered through a preweighed, 0.4 micron, polycarbonate filter
which was then dried and weighed on a microbalance to determine suspended solid
mass.

Statistical analyses

Community characteristics (abundance, richness, and diversity) were analyzed with an
ANOVA model that included reach (forest vs pasture), year (three years for rock and
wood, two years for leaves), and site nested within reach (two sites per reach). Reach,
year, and site wee treated as fixed effects. Abundance and richness data were log-
transformed to correct for heterogeneity of variance. least squares means comparisons
made at each level of reach and year are used to estimate the relative strength of
treatment effects.

The proportional representation of different functional-feeding guilds (filterers,
gathees, scrapers, predators, and herbivores) was analyzed with a contingency
analysis (CATMOD procedure, SAS 1990) that tested for effects of reach (forest vs
pasture), year (three years for rock and wood, two years for leaves), and the interaction
between reach and year. In all cases, the null hypothesis was that the proportion of

24

individuals representing each feeding guild did not diffe across treatments. Shreddes
were excluded from these analyses due to their low frequency of occurrence.

Two-by-two chi-square tests were used to analyze responses of populations and
feeding guilds to landscape alteration (calculations follow Dowdy and Wearden 1991).
Three null hypotheses were tested: (1) the probability that taxa would persist,
disappear, or appear from 1991 to 1993 was independent of reach and tolerance index,
(2) changes fiom 1991 to 1993 in the numbe of taxa were the same in forest and
pasture reaches, and (3) the probability that taxa from the forest reach also occurred in
the pasture reach equalled the probability that taxa from the pastrue reach also occurred
in the forest reach.

RESULTS
Baseline comparisons: undisturbed vs overgrazed reaches

Measurements of abiotic characteistics in the forest and pasture reaches indicate that
substrata chemical composition, particle-size array, wate temperature, dissolved
oxygen concentration, and suspended solid concentration were similar between reaches
during 1991 (Table 1, Figure 4). However, these contiguous reaches had obviously
diffeent riparian characteristics (Figures 1, 2) which apparently had strong effects on
benthic invertebrate community composition (Appendices l, 2).

During the summer pasturing season, invetebrate abundance was far greate in forest-
reach rock samples than it was on pasture-reach rocks (back-transformed least squares
means = 12.2 vs 7.5 inds./m2, t = 2.67, p = 0.0086) (Table 3, Figure 5). Howeve,
abundance on wood samples (collected afte cattle had been removed) was similar
between reaches, suggesting perhaps that autumn conditions were more similar between
reaches than was the case when cattle wee present. leaf-pack samples, which wee
taken late in the fall, reveal a more abundant forest-reach invertebrate community (back-
transformed least squares means 3.4 vs 1.6 indsJ day, least squares means t = 2.40, p
= 0.021) (Table 3, Figure 5) primarily due to the presence of larval black-fly
(Prosimulium, Diptea: Simuliidae) densities which were 13 times greater in the forest
reach than in the pasture reach.

25

Taxonomic richness and community diversity (HRI) were similar in undisturbed and
overgrazed reaches during 1991(Figure 5). Howeve, community composition and
reach-specific endemism, varied considerably between reaches. Forty-five percent of
the taxa collected in 1991 from the forest reach wee endenrics, whereas all but 14% of
taxa collected in the pasture reach occurred in both reaches (1:2 = 8.79, p S 0.01,
Appendices 1, 2).

Although richness and divesity wee similar between reaches, reach-specific endernism
differences directed substantial inter-reach diffeences in the proportion of taxa
representing different functional-feeding guilds (Table 4; Figures 6, 7). Gatherers were
much more abundant in pasture reach rock and leaf-pack samples relative to forest-reach
samples, but wee similar in wood samples. There was little difference in filtee
abundance on rocks, but they wee more abundant in forest-reach wood and leaf-pack
samples as winter-developing populations of Pmsimulium were far more abundant in
forest-reach, particularin in leaf-pack samples (Figures 6, 7). The relative proportion
of filtees to gatherers, a measure of particulate organic matte in transport versus
deposition (Merritt and Cummins 1996 b), was greater in the forest reach on rocks
(0.23 vs 0.11), wood (0.16 vs 0.04), and leaf packs (2.47 vs 0.06). This upstream-
downstream relationship, indicates that riparian-zone grazing affected the ratio of
eosional to depositional benthic habitat in Carlson Creek. Biofilrn scrapes were more
abundant in the forest-reach than in the pasture reach on rock (2.36 vs 1.99 inds./m2),
wood (6.36 vs 1.83 inds.lm2) and leaf-pack samples (0.07 vs 0.03 inds./day) due
principally to differential abundances of heptageniid mayfly nymphs (Ephemeoptera:
Heptageneiidae). The relative proportion of predators to other groups, an estimate of
predator-prey ratio (Merritt and Cummins 1996 b), was more than two-fold higher in
forest-reach rock samples than it was in pasture-reach rock samples (20 vs 10% of total
abundance). This was largely due to the presence of five predator taxa found only in
the forest reach (Appendices 1, 2). The proportion of predators to other groups was
similar between reaches on wood samples and lowe in leaf-pack samples (Figures 6,
7).

General efi’ects of beaver impoundment and cattle exclusion

Figures 1 and 2 depict the transition of the forest-reach from free flow in 1991 to
impoundment by 1993. Increased depth and reduced flow were accompanied by a

26

change in substrata from mixed fme-rrrineral composition (Table 1, Figure 3)
supporting a variety of macrophytes to a benthic habitat consisting principally of
beave—masticated wood material with little rooted vegetation. Cattle exclusion actuated
rapid recolonization by plants on previously barren banks and channel. Depth and flow
remained constant and suspended solids increased from 1991-1993 (Figures 1-4).

The eflects of landscape alteration on abundance

Beaver activity resulted in a large increase in abundance from 1991 to 1992 in rock and
wood samples (Table 4, Figure 5). These responses wee largely the result of
combined abundance increases in non-Tanypodinae Chironomidae (Chironomidae
hereafter) which increased 38% on rocks and 41% on wood; Tanypodinae
Chironomidae which increased 7% on rocks and 20% on wood; and Hirudinea which
experienced increases of 14% on rocks and 6% on wood. In leaf-pack samples,
abundance losses (largely Prosimulium), overwhelmed the effect of Chironomidae
increases leaving no net effect of the first stage of impoundment. Abundance decreased
markedly during the second year of heave activity on both rock and wood samples
(Figure 5). In general, those taxa that experienced proportional increases from 1991 to
1992, moreased from 1992 to 1993. Chironomidae abundance decreased by 59% on
rocks and 66% wood, Tanypodinae decreased by 17% on rocks and 33% on wood,
and Hirudinea populations declined by 23 % on rocks.

As was the case in the free-flow to impoundment transition, relaxation of instream and
stream-bank trampling in the pasture reach resulted in large abundance increases during
the first season of restoration (Table 3, Figure 5). Abundance increased more than 17-
fold on rocks (back-transformed least squares means ~ 8 vs 138 inds.lm2 in 1991 and
1992 respectively, t = 8.28, p = 0.0001), tenfold on wood (is 40 vs 400 inds.lm2, t =
7.56, p = 0.0001), and two-fold in leaf packs (~2 vs 4 animals pe day, t = 3.66, p =
0.0007). These effects were largely attributable to Chironomidae abundance increases
(75, 67, 56% of total increases on rocks, wood, and leaf packs respectively). From
1992 to 1993, abundance decreased somewhat on rock samples and considerably on
wood samples (back-transformed least squares means *3 400 vs 137 indslm2, t = 4.19,
p = 0.0001). Howeve, despite these declines, thee remained a significant overall
increase on both rocks (~8 vs 66 inds.lm2, t = 7.39, p = 0.0001) and wood (from =1 40
to 140 inds.lm2, t = 4.18, p = 0.0001) as a result of large increases from 1991 to 1992

27

(Table 3, Figure 5).

Abundance increased at least temporarily with both environmental changes largely as a
result of concurrent Chironomidae increases in forest and pasture reaches. Although
abundance patterns wee similar between reaches in rock and wood samples, the effect
of cattle exclusion was far greater and less transient than that of beaver impoundment
(Figure 5). Abundance changed from being higher in 1991 in the undisturbed forest
reach, to being far lower than that in pasture reach by 1993 on rocks (back-transformed
least squares means e 16.5 vs 66 inds.lm2, t = 2.99, p = 0.003) and somewhat lower
on wood (18 52 vs 60 inds.lm2, t = 3.56, p = 0.0006) (Figure 5).

Efiects of landscape alteration on richness

Taxonomic richness changed little on rocks and leaf packs during the frrst year of
beave activity, but increased considerably on rocks by 1993. This effect was largely
the result of increases in occurrence of all of the relatively few taxa present in 1993, for
the total numbe of taxa actually declined markedly as a result of beave activity
(Appendix 1). Beaver impoundment also strongly affected taxonomic richness on
wood samples, but the pattern was diffeent from that detected with rock samples
(Figure 5). Richness increased from 1991 to 1992 primarily as a result of the
diffeence between taxon additions (23 taxa comprising 18% of 1992 abundance) and
eliminations (12 taxa comprising 46% of 1991 abundance). This increase was followed
by a sharp decrease in 1993, which was largely accounted for by 1993 taxa eliminations
(Appendix 1). The total numbe of taxa in all samples combined decreased sharply
from 54 taxa prior to impoundment to 29 and 21 taxa in the subsequent two years. This
decline occurred in all trophic guilds (Appendix 1), but was most pronounced in taxa
that require relatively silt-free, solid substrata (reductions of 78, 60, and 62% of
filteers, scrapers, and predators respectively) (Figure 6).

In contrast to the effects of impoundment on richness, cattle exclusion resulted in
taxonomic richness increases, particularily in rock samples (Table 3, Figure 5).
Although richness changed little on wood and leaf-pack samples, richness in all
substrate types combined increased from 41 to 59 taxa after one year of cattle exclusion,
then returned to 42 by 1993 (Appendix 2). Beaver impoundment had a much greate

28

overall effect on richness than did cattle exclusion. While richness declined steadily
during pond formation, the net effect of cattle exclusion on richness was the addition of
just one taxon (x2 = 8.40, p < 0.001; 1991-1993 totals = 54, 29, and 21 in forest-
reach samples and 41, 59, and 42 in pasture-reach samples). Response to cattle
exclusion and beaver impoundment included taxa that persisted, disappeared, and
appeared in differing proportions between reaches (x2 = 15.62, p = 0.004).

Compared to cattle exclusion, beave impoundment resulted in more disappearance (48
vs 9%), less persistence (16 vs 32%), and less appearance (35 vs 59%) from 1991 to
1993 (total taxa = 69 and 66 for forest- and pasture-reach respectively).

Toleance index values varied little between reaches or years (all means between 4.2
and 5.6), organismal response to changing landscapes was detected in an analysis of
the effects of habitat alteration in geneal (impoundment and exclusion combined).
Taxa assigned low tolerance index values (T1 = 1-5), compared those with high values
(T1 = 6—10), were less likely to persist (19 vs 37%), more likely to disappear (53 vs 34
%), and equally likely to appear following perturbation (x2 = 4.93, p = 0.085).

The efiects of landscape alteration on community diversity

Diversity, as estimated with the HRI procedure, increased on wood samples over the
frrst year of beaver activity (Figure 2) (least means squares, t = 5.74, p = 0.0001) then
returned to near free-flow levels by 1993. The same trend was evident in rock samples
although high variance resulted in insignifieant differences from 1991 to 1992. As was
the case overall, leaf pack samples indicate that beave activity had no detectable effect
on divesity (Table 3, Figure 5).

Community diversity increased with cattle exclusion but showed only transient changes
with beave impoundment. Prior to beaver activity and cattle exclusion, forest-reach
divesity was similar to that in the pasture-reach in rock and wood samples (Figure 1).
By 1993, however, HRI values indicate that diversity was much higher in the pasture
reach relative to the impounded forest-reach on these substrates. This was both the
result of declines experienced in the forest reach and strong, positive effects of pasture-
reach restoration.

29

Efieas of landscape alteration on fimctional-fleeding groups

Diffeential declines in richness and overall increases in abundance attributable to the
lotic-to-lentic transition produced strong treatment effects on proportional abundance of
functional-feeding guilds (Table 4, Figures 6, 7). The general pattern of this transition
was that scraping and filtering taxa wee replaced by collector-gatheers able to exploit
vast deposits of organic material that form as a result of beaver feeding activity and
impoundment of fine particulates (Naiman et al. 1986, Harnmerson 1994).

In free-flowing conditions, filter feeders comprised 12, 8, and 64% of invertebrates
collected from rock, wood, and leaf-pack samples respectively. Filtere abundance then
declined sharply from 1992-1993 (to 3, 2, and 31 %) as beave activity slowed flow
(Figure 4) and buried substrata. The 1993 beave pond community as sampled with
rocks was devoid of filterers; wood samples were composed of just 2% filterers which
wee zooplankton that do not require current to facilitate filte-feeding (Figure 6,
Appendix 1).

The proportion of gatlrees to other taxa increased on all substrata. On rocks and
wood, increases were primarily as a result of Chironomidae dynamics. The proportion
of Chironomidae on forest rocks increased from 10% of the community in 1991 to 31%
in 1992 and up again in 1993 to 52% of the total community. On wood, Chironomidae
accounted for 19% of the total abundance in 1991, then increased to 43% in 1992 and
dropped to 27% in 1993. The proportional abundance of gatherers on leaf packs
increased as a result of an 11% increases in chironomid abundance from 1991 to 1992
(Figure 6).

Biofrlm scrapers declined as a proportion of the community fi'om 1991 to 1992 (Figure
3). On rocks, Stenacron, Stenonema, Macdunnoa, Arthroplea (all Ephemeoptera:
Heptageniidae) and Helicopsyche (Trichoptera: Helicopsychidae) declined dramatically
while Nixe (Heptageniidae) and Gastropoda populations experienced proportional
increases. Gastropoda and Stenacron population increases influenced an overall
scraper abundance increase on wood samples as a result of impoundment. By 1993,
Stenonema and Stenacron were the only scrapers collected from the then fully formed
pond (Figure 6, Appendix 1).

Stream-to—pond transition caused the replacement invertebrate predators adapted for
lotic existence with those that typically inhabit standing wate. Proportional increases
of predators from 1991 to 1992 on rocks (Figure 6) were primarily the result of

30

changes in Tanypodinae abundance which comprised 3% of the community in 1991 and
increased to 6% by 1992, and Hirudinea which increased from 4% to 11% from 1991
to 1992. Tanypodinae abundance dynamics also strongly influenced the relative
abundance of predators on wood samples which increased from 13 to 21%. The

proportion of predators changed little on leaf packs from 1991 to 1992 as Wes in
0ecetis (Trichoptera: Ieptoceridae) and Hemerodromia (Diptea: Empididae) were
directly offset by additions of Chauliodes (Megaloptera: Corydalidae) and
Nyctiophylax (Trichoptera: Polycentropodidae) (Appendix 1). Although predator
abundance generally increased with beave activity, predator richness declined sharply
(Appendix 1) largely as a result of local elimination of lotic specialists. On rocks,
Nyctiophylax, Polycentropus (Trichoptera: Polycentropodidae), Nigronia
(Megaloptera: Corydalidae), and Probew'a (Diptera: Ceratopogonidae) wee
eliminated. Hernerodromia (Diptea: Empididae) disappeared from all substrata. Leaf-
pack samples indicated that Isoperla (Plecoptea: Pelodide) also abandoned the forest
reach as beaver activity proceeded (Appendix 1).

The strong effects of cattle exclusion on abundance wee disproportionally distributed
between functional-feeding guilds, but not to the same degree as they were as a result of
beaver impoundment (Figures 5-7). Filteres decreased gradually on rocks due to
reductions of fingenail clams (Sphaeriidae) and net-spinning caddis larvae
(Cheumatopsyche) and increased on wood as a result of increases in Shaeriidae and
Hydropsyche, thus producing little overall frmctional change. Filteres increased in
numbers on leaf-packs largely as a result of proportional increases in Prosimulim

which increased from 14 to 50 % of total abundance. At the catchment level (forest and
pasture reaches combined), increases in the pasture compensated for upstream declines
that resulted from impoundment.

Collector-gatherer response to cattle exclusion wee largely attributable to
Chironomidae abundance dynamics. Gathees increased in relative proportion on
wood and leaf packs from 1991 to 1992 and decreased on wood from 1992 to 1993.
On rocks, gathee abundance and taxonomic composition changed little from 1991 to
1993 (Figure 7, Appendix 2). The ratio of filtere and gathee proportional
abundances, a measure of relative levels of transport and deposition of particulate
organic matte (Merritt and Cummins 1996 b), was greate in the forest reach on rocks
(0.23 vs 0.12), wood (0.16 vs 0.044), and leaf packs (2.47 vs 0.06). This upstream-
downstream relationship, typical of comparisons made between low- and mid-orde

31

streams (Merritt and Cummins 1996 b), indicates that pasturing may have affected the
relative amounts of erosional and depositional benthic habitat in Carlson Creek.

The proportion of predators to other taxa changed little from 1991 to 1993 on rocks and
decreased on wood despite increases in all predacious taxa present (with the exception
of Hemerodromia which decreased slightly on wood). Theefore, it was the
concurrent large increases in othe taxa (principally Tanypodinae) that drove observed
differences in predator proportions between years. In contrast to this patten, predator
decreases on leaf packs wee the result of large decreases experienced in populations of
Hirudinea and Hemerodromia.

The lotic to lentic transition that resulted from beave activity had a far greate effect on
invertebrate community functional composition than did pasture-reach restoration
efforts (Figures 6, 7). This is perhaps the result of differential levels of landscape
transformation, for relaxation of trampling did not affect channel characteristics nearly
as much as did beave impoundment (Figure 3, 4).

DISCUSSION

The onset of beave-mediated disturbance and recovery from ovegrazing in Carlson
Creek geneated pronounced abundance increases in response to both landscape
alterations. These increases were largely the result of concurrent large increases in
chironomid rnidge populations and therefore exemplified the commonly observed early
dorrrinance of post-alteration communities by small, rapidly developing generalists (e. g.
Nuttall and Bielby 1973, Wallace and Gurtz 1986). In a manne suggestive of a
delayed density-dependant response, abundance declined from 1992 highs by 1993 in
both forest- and pasture reaches (Figure 5). The principal inter-reach diffeence
between compensatory responses was that, from 1991 to 1993, abundance increased in
the pasture-reach and decreased in the forest-reach.

Similarities in inte-reach patterns of response to beave impoundment and cattle
exclusion could be evidence of inter-annual abundance patterns within the catchment
(forest and pasture reaches combined) and therefore not the result of concurrent
landscape alterations. Howeve, community richness- and compositional changes
varied consideably between reaches and thus evidence very different landscape-

32

alteration processes.

During 1991, differing riparian characteristics wee reflected in the feeding ecology of
reach-specific endemic taxa. Forest-reach endemics included four taxa that consume
coarse particulate organic material (Mystacides, Taeniopteryx, Acentn'a, Isopoda) and
no strictly fine-particle fwdes. The opposite relationship was detected in the pasture
reach whee endemics included two fine-particle feeders (Ameletus and Hexagenia)
and no large-particle specialists (Appendices 1, 2). This invertebrate taxonomic
distribution reflects the expected patten in particle size composition when moving from
tributary to mouth, or as is apparent in Figure 1, from areas of dense to sparse
allochthonous vegetation (Vanotte et al. 1980, Merritt and Cummins 1996 b).

Prior to disturbance by beaver, the forest-reach benthic community was richer, more
diverse, and more evenly distributed across functional-feeding guilds than was the
pasture-reach community. This resulted in large inte-reach differences in feeding-guild
richness. In the pre-impoundment forest reach, filterer-, predator-, and shredder
richness estimates wee greater; gatlrere richness was lower; and scrapers and
herbivores were proportionally similar between reaches. These relationships shifted
entirely as beave activity progressed and cattle exclusion was initiated By 1993, all
forest-reach functional-feeding guilds were less rich than their pasture counteparts due
to the combined net effect of richness decreases in response to impoundmelt and
richness increases in response to habitat restoration.

Although the effects of beaver impoundment on abundance, richness, and divesity
wee more transient than those of cattle exclusion, lotic-to—lentic transition directed
consideably more substantial functional changes in the invetebrate community than did
restoration efforts. Unlike the strong effects of impoundment, relaxation of trampling
perturbation did not cause lasting changes in the frmctional structure of the benthic
communities (Figures 6, 7). Restoration could, howeve, have potentially lasting
effects on reach-level bioerergetics as a result of the general pattern of increases in
abundance, richness, and diversity from 1991 to 1993 (Figure 5).

The benthic invertebrate community response to impoundment of the forest reach
typified well-established effects of beave disturbance. Beaver pond formation typically

33

results in replacement taxa that depend on flow to deliver food, scrape biofrlm from
solid substrata, or cling to vascular hydrophytes with taxa adapted to exploit vast
accumulations of entrained allochthonous organic material.(Naiman et al. 1986, 1988).
Theefore, because the sampling program was clearly sensitive to changes caused by
beave, the effects detected in response to excluding cattle from Carlson Creek may
allow a broadening of scope of inference to include restoration efforts undertaken
elsewhere.

Fencing and bank stabilization efforts were endeavored principally to enhance substrata
and water quality in effort to slow eosion and reestablish the pasture reach as a brook
trout (Salvelinusfontinalus) nursery. Benthic invertebrate standing stocks increased
on all substrata in response to exclusion. It is therefore reasonable to assume that
insectivorous fish may have also benefitted from enhanced food resources. However,
because increases were predominantly in taxa that burrow into soft sediment, high
levels of deposition and transport (Figure 4) were still indicated by 1993 which would
undoubtedly limit brook trout redd success (Waters 1995).

Inter-reach and inter-year habitat diffeences wee evidenced by visual inspection and
detection of significant invertebrate responses to disturbance and recovey (Tables 5, 6;
Figures 1, 2, 5). However, despite obvious physical change in both reaches, tolerance
index values varied little between reaches or throughout landscape alteations.
Published benthic invertebrate tolerance values wee therefore arguably insensitive to
stress induced by livestock grazing and beave activity in Carlson Creek. Howeve, at
the catchment level (forest and pasture reaches combined), tolerance index values
indicated relative resistance and resilience to the effects of changing environments.
Low-tolerance to perturbation was associated with relative inability to pesist, higher
likelihood of disappearance, and equal ability to appear following peturbation. This
pattern would, in recently (or continually) disturbed reaches, result in a higher
proportion of high-tolerance taxa than would be expected in relatively stable habitat.
The data therefore indicate that although mean TI values were not responsive to the
effects of beave activity or cattle exclusion, it seems that the toleance index may be
useful in assessing an organism's relative resilience to changing physical conditions.

As was the case in a study conducted by Reed et al. (1994) in three Australian streams

34

undergoing different levels of livestock grazing, functional-feeding group composition
in Carlson Creek was the most illustrative of initial habitat condition (undisturbed vs
overgrazed) and thus potentially the best on-site method of determining habitat
condition in grazed streams. Community response to habitat restoration was, however,
best characterized by changes in diversity as estimated with the hierarchical richness
procedure (Figure 5).

The relative contributions of local and regional processes on richness reveals that local
(reach-level) spatial heterogeneity contributed immensely to regional (catchment-level)
richness in Carlson Creek. Prior to beave colonization and cattle exclusion, reach-
specific endemism was far more common in the free-flowing forest reach than it was in
the ovegrazed pasture reach. Howeve, by the second year of impoundmert and
restoration, the opposite relationship was detected as most beave-pond inhabitants
could be found in both reaches. Disturbances by beave and cattle therefore facilitated
the establishment of communities rich in habitat genealists and conversely, undisturbed
and recovering conditions encouraged the establishment of lubitat specialists.
Catchment-level biodiversity was thus enhanced through maintenance and augmentation
of landscape heterogeneity, a phenomenon expected to ensure maximum stability in
agricultural ecosystems (Tilman 1996).

SUMMARY

(1). Beaver colonization and cattle exclusion both resulted in initial abundance
increases in chironomid rrridge populations, thus exemplifying the commonly observed
dominance of post-alteation communities by small, rapidly developing ecological
generalists.

(2). Recovery from ovegrazing led to increased community diversity while
disturbance by beave impoundment resulted in decreased divesity.

(3). Impoundment by beave produced very large changes in community structure and
functional composition. In contrast, recovey from cattle-grazing disturbance did not
meaningfully affect community composition.

(4). Disturbance by beave fostered increased abundance of broadly distributed habitat
generalists while recovering and undisturbed habitat harbored less common taxa with
specialized habitat requirements.

(5). The forest reach supported many taxa not found in the ovegrazed pasture reach,

35

but the pasture reach contained few taxa not found in the forest reach. Theefore,
losses of forest reaches that result from human activities have relatively large impacts on
watershed biodiversity.

36

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39

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41

Table 1. Baseline abiotic data collected from
forest- and pasture-reach sites during 1991.
Values indicate means i: 1 se.

 

FOREST PA STURE

Substrata

% Sand 48.4 :1: 4 49.4 :1: 27
% Silt 11.7i1 15.721: 11
% Clay 39.9 :1: 5 34.9 :t 16
pH 8.4 i 0.4 8.55 i: 0.05

CBC (me/ 100g) 20.4 i 1.45 18.4 :t 4.2

p (ppm) 10.5 i 0.5 23 i: 18

K (ppm) 108 :1: 27 77.5 :1: 32.5
Ca (ppm) 3452 i 262 3141 :1: 668
Mg (ppm) 340 j: 10 297 i 93

Na (ppm) 260 i 59 212.5 :1: 72.5
14114 (ppm) 2.25 :1: 0.05 2.45 :1: 0.75
N03 (ppm) 2 :t 1 3.5 3; 2.3
H20

Temp. (° C) 19.4 i 1.8 15.7 :1: 0.8

DO (mg/l) 7.85 :t 0.17 8.3 :1: 0.24

 

Table 2. Carlson Creek invertebrate tolerance-index values (from Resh et a1 1996).

42

* = values estimated for this study.

COLLEMBOLA* 4 TRICHOPTERA COLEOPTERA
PLECOPTERA Brachycentridae 1 Dytiscidae* 7
Taeniopterygidae 2 Glossosomatidae 0 Elmidae 4
Perlodidae 2 Helicopsyche 3 Gyrinidae 7
Hydropsychidae 4 Haliplidae* 4
EPHEMEROPTERA Hydroptilidae 4
Baetidae 4 Lepidostomatidae 1
Caenidae 7 Leptoceridae 4 DIPTERA
Ephemerellidae 1 Limnephilidae 4 Ceratopogonidae 6
Ephemeridae 4 Polycentropodidae 6 Chaoboridae 3
Heptageniidae 4 Psychomyiidae 2 Chironomidae 6
Leptophlebiidae 2 Dixidae 4
Siphloneuridae 7 Empididae 6
Simuliidae 6
ODONATA MEGALOPTERA Tipulidae 3
Calopterygidae 5 Corydalidae 0
Coenagrionidae 9 Siahdae 4 AMPHIPODA 4
Cordulegastridae 3 ISOPODA 8
Gomphidae 1 LEPIDOPTERA ACARI 4
Iestidae 9 Noctuidae* 5 GASTROPODA 7
Libellulidae 9 Pyralidae 5 CNIDARIA“ 3
NEMATOMORPHA 10
NHVIA'IDDA 10
HIRUDINEA 10

OLIGOCHAET A 10

 

43

Table 3. AN OVA results comparing invertebrate community characteristics as
estimated with rock, wood, and leaf-pack samples (see Figure 1).

 

 

 

ROCKS Log (Abundance)
Source df MS F
Reach 1 1.122 5.242 *
Year 2 8.129 37.996 ****
Reach x Year 2 2.375 1 1.092 ****
Site (Reach) 2 0.298 1.391
Site x Year (Reach) 4 0.367 1.716
Error 137 0.214

WOOD
Reach 1 1.172 10.198 ***
Y” 2 6.411 55.773 ****
Reach x Year 2 0.353 3.072 1.
Site (Reach) 2 0.268 2333
Site It Year (Reach) 4 0.146 1.267
END! 92 0.115

LEAVES

Reach 1 0.040 0.323
Year 1 0.291 2.340
Reach x Year 1 1.656 13.305 ****
Site (Ranch) 2 1.624 13.051 “*1:
Site x Year (Reach) 2 0613 4.925 *
Error 43 0.124

 

 

 

 

 

 

Log (Richness)
MS F
1.492 12.277 ***
8.075 66.429 ****
0.458 3.771 *
0.228 1.876
0.173 1.426
0.122
0.194 4.257 *
1.309 28.787 ****
0.234 5.136 **
0.179 3.930 *
0.017 0.369
0.045
0.000 0.000
0.174 6,555 ***
0.069 2.606
0.209 3.937 *
0.132 2,489
1.396

 

 

 

HRI
MS F
1823539 19.837 ***
1754082 19.081 ****
1066447 11.601 ****
1033.4 0.112
6203.4 0.675
9192.8
2472.4 9.348 ***
16075.0 60.777 ****
980.8 3.709 *
565.4 2.138
138.2 0.522
264.5
05 0.058
6.3 0.807
74.1 9.563 **
72,3 9.359 ***
372 4.788 *
7.8

 

'1' P < 0.10, * P < 0. 05, ** P < 0.01, *** P < 0.001, **** P< 0.0001

44

Table 4. Contingency analyses testing whether the proportion of animals
representing different functional feeding groups varied between reaches
in rock, wood, and leaf-pack samples (see Figures 2, 3).

 

ROCKS

d_f 82
Reach 4 109.12***
Yw 8 262.44***

Reacthear 8 77.38***

WOOD LEAVES

 

4 125.09*** 4 3.42
8 272.42*** 4 117.30***
8 129.36*** 4 22.09***

 

*** p < 0.001

45

 

Figure 1.’T‘he;smre reach of Carlson Creek in (A) 1991, (B) 1992, and (C) 1993.

46

 

Figure 2. The forest reach of Carlson Geek in (A) 1991,—(B) 1992’and (C) 1993.

47

 

 

 

100 -
90 : DEPTH p0
80: . K
A . 0 Forest Statlon
E 70 - A Pasture Station 1
v 60 ‘ V Pasture Station 2
.C q .x’
‘6. ‘
0 5° 3 ix ...........................
a 40 - £f:::. ,,,,,,,,, n/flx ................ z
30: ,9W”flfl
20 - a
10 r T 1
a? .35 -
8 1
Q _3- VELOCITY
g Q, :1.'_
>. .25 — -‘- 1 """"""""
:1: ‘ ......
8 eeeee 1 """"""""
'6 2 ‘ m """"" """"" i
-- 15 - " """"
5 1
L- ,_
3 .
0

 

 

1 991 1 992 1 993

Figure 3. Mean depth and flow rate i 1 se in one forest-reach and two pasture-reach
stations. Measurements taken during July, August, and September, 1991-1993.

Suspended Solids (mg/L)

 

1 991 1 992 1 993

Figure 4. Suspended solid concentration from forest— and pasture-
reach water sampless taken monthly, J uly—September, 1991-1993.

's91dums need-5291 pure ‘poom 61001 H; Amie/11p pue ‘ssauqop ‘aouepunqe armqanaAur meow 'g mar

49

Abundance
log (inds. lday) log (indsJ m2)

0 p __. to o .1
our-tour-tcnromwocn-tmm

3661 166|-

866 l

 

Richness

log (taxa/day) log (taxa! 0.1 m2)

 

 

266 l

 

 

 

866 l

 

3661 1661-

866 l-

 

SO

 

FilterersO Gatherers . Scrapers edators . HerbivoresO Shredders.

UNDISTURBED BEAVER YR. 1 BEAVER YR. 2
ROCKS 5°2 109° 495

      

3335 705

 

g 1727

   

Figure 6. Community functional-fading group composition in forest-reach samples.

Values indicate total number of animals collected. Functional-feeding groups that comprised
less than 1 % of total are not represented on pie-charts. These are: 3 herbivores in 1992
wood samples and 5 scrapers in 1992 leaf-packs.

51

Filterers. Gatherers . Scrapers. Predators . Herbivores O Shredders.

OVERGRAZED YR. 1 RECOVERY

YR. 2 RECOVERY

 

LEAVES ,3, we.

 

Figure 7. Community functional-feeding group composition in pasture-reach samples.
Values indicate total number of animals collected. Groups that comprised less than
1% of total count are not represented on pie-charts. These are: 2, 9, and 9 scrapers in

1991-1993 wood samples respectively; 1 herbivore in 1991 wood samples; and 6 scrapers
in 1991 leaf-pack samples.

APPENDIX 1

52

Appendix 1. Forestjreach taxa organized by feeding guild and year of collection. letters
after taxa names rndreate sample type(s) from which animals were collected: R=rock;

W=wood; L=leaf packs; D=drift.

 

 

 

 

 

 

 

 

FEEDING 1991 1992 1993
GUILD UNDISTURBED BEAVER 1N BEAVER YR. 2

Anhroplca R Ceratopryche D Cladocera W
FILTERERS Bmchycenm L Cheumatopsyehe R W L D Copepoda W

Cerawpuyche R W Hydropsyche R W D

Cheumatopsyche R L D Pmsimuliwn R W L D

Hydropsyche R W L D Neureclipsis W D

Neurcclips'is W D Bivalvia R

Prosimulim W L D Hydra R LD

Bivalvia RWD Cladocera W L D

Cladocera D Copepoda D

Padura D Padura R W D Padum R
GATHERERS Baetis RWLD Baetis RWLD Cami: W

Cami: R W D Cami: R W Ephemerella W

Pamleptophlebia L D Ephemerella R W Paralepcophlebla W

Mystacides R Pmleptophlcbia W L D grammar W

thomyi'a R W Mystacr'des W ironornidae R W

Dubiraphia W Dubiraphia W D Dixella W

Macromrchus W L D Macmnychw W D Oligochaeu R

Opa'oservus L Optioservus W Acari R

Stenelmis R Stenclnu's W L

Chironomidae RWLD Chironomidae RWLD

Nematoda R L Nematode R

Acari R W D Acari R W D

Macdwmoa R Nix: R Stenacmn W
SCRAPERS Stenacron RD Stenacron RW LD Sma W

Stenonema R W L D Stenonema R W L D

Helicopsy‘he R D Helicopsyche R

Gastropoda R D Gastropoda R W

Amphiagn’an D Argia D Argia W
PREDATORS Argia D Calopteou W Stalls W

Bayen'a R D Isoperla L Oeccds W

Chromagflon D Bayer-fa W Probezzia W

Cordulegasm' R Chauh'odes L Tan inae R W

Calopteoa R L D Nigronia R H ' R W

Pmmogomphu: D Stalls R Nematomorpha W

Isoperla L 0ecetis R W D

Mgram'a R Nym'ophylax R W L

Nyca'ophylax R Polycenu'opw R W

0ecetis R L D waua's D

[’01me R Tanypodinae R W L D

alaoborus D Hemerodromia R W L D

Tanypodinae RWLD Hirudinea RWD

Dicranota W Nematomorpha R W

Probezzia R

Heater-04mm W L

Hirudinea R W

Hydmpa'la R W
HERBIVORES OWN w
Peltodytes D

Taenio W Taeniopteryx L Amphlpoda° W
SHREDDERS ”mm; D ”40...... no

Lepidostoma R LD Pycnopsyche R

humming R W Trianodes W D

Acemria D Amphipoda R W D

Tipula. D

Amphxpoda D

IsopodaL

 

53

Appendix 2. Forest—reach taxa organized by feeding guild and year of collection. letters
alter taxa names indicate sample type (s) from which animals were collected: R=rock;
W=wood; L=leaf packs; D=drift.

APPENDIX 2

54

 

 

 

 

 

 

 

FEEDING 1991 1992 1993
GUILD OVERGRAZED CATTLE EXCLUDED RECOVERY YR. 2
Chemnawps'yche R WD Brachycentrus R Brach centrus W
FILTERERS Hydro psyc che RL Ceratopsyche R LD Chewzatopsyche RW
Neureclipsi: D h p he RWLD HydropsycheW
Prosimulium R L D Hydmpsyche R W L D Simuliidac R
Bivalvia R W L D Neureclipsis Bivalvia R
Prosimuliwn W L D Hyira R W
Bivalvia R W D
Hydra D
Cladocera R L D
Cepepoda D
Daphuia RD
Collembola W Collembola LD Podura R
GATHERERS Ameletus D Podura D Baeu's RW
Baetis-RWLD BaedsRWLD Caenis
Caem‘sRLD CaenisRWLRDD ParaleptogvhlebiaRW
Ephemerella W Ephemerella Myslacides W
Paraleplophlebia L Paraleptophlebia R W L D sychomyia RW
Hexagenia R Pswhomyia RW D Dubiraphia RW
Siphloneuridae R Dubira aR L D Maeronychus W
Psychomyia R Macronychu: W D Opaoservus RW
Dubiraphia R L Stenelmis R WD S Inn's
Macronychus W D Dixella RD Chironomidae RW
Opfioservus L D Chironomidae R W L D Nematode R W
Stenelmi: W Anlocha Acari RW
DuellaLD Nematode RWLD CopepodaW
Chironomidae R W L D Oligochaeta R
NematodaRWLD Acari RWLD
Oligochaeta L Copepoda RW
Acari W
Pseudocloeon R Heptagerua R Stenacron RW
SCRAPERS Stenacron RL R Srenonema R
Slenonema R L D Stenacron R D Helicopsyche R W
Glossosoma D Steuonema R W LD
D Psphenidae D
Helicopsyvhe R
Neophylax W
Neophylax W
a D
Calopteryx D Aeschna D Boyer-in RW
PREDATORS Polycentrapus W L Calopteryx D Calopte
Dytiscida eLD Corixidae DD 0ecetis R
Tanypodinae R W L D Nigrmia Polycentmpus W
Probezzza D Oecen's I? W D Dineuu‘s R
Hemerodromia R W L D Nyctiaphylax L Tanypodinae R W
Hirudinea R LD Palycemropus W D limnophora R
Agabus D Hememdromra RW
Dineun‘ 115D Nematomorpha W
Tanypodinae R W L D Hirudinea R W
chranota R
Hexawma R
Hemerodromia R W L D
Hirudinea R W LD
Nematomorpha RW
' Hm‘laRWLD HydroptilawRW
HERBIVORES ”mm W L 081.2 R w L D 0...)...”
Pamponyx W Pampanyx RW
Haliplus R Petroplu'la R
Pelrodytes D
Le ‘ stoma D [epidostoma D [epidoswma RW
SI—IREDDERS Pygwdfpryche L D Pycnopsyche R W L D Pyvnopsyche
Grensia D Necwpsyche RL HydawPhylax W
11thny L Trianodes RD Amphipoda RW
Tipula D Hydatophylax D
Amphipoda LD Tipulidae R
Amphipoda R W L D
R

Decapoda
Isopoda

 

3

THE BEHAVIORAL RESPONSE OF THE STONEFLY PARA GNE TINA MEDIA
(PLECOPTERA: PERLIDAE) TO THE ONSET OF INTENSE SEDIMENTATION

Abstract. The initial response by Paragnetz'na media (Plecoptera: Perlidae) nymphs to
increased suspended sediment concentration was examined in field trials with chambers
designed to give nymphs the choice to stay or reposition in response to experimental
sediment additions. The results indicate that the often posited immediate-escape
response to the onset of intense sedimentation may not be commonly enacted by
stonefly nymphs or perhaps other relatively immobile, lotic insects.

INTRODUCTION

What do stream—dwelling animals do when they experience rapid increases in
suspended sediment concentration? The answer to this question is fundamental to
understanding the effects of the massive amount of sediment entering flowing water
globally as a result of human activities (Waters 1995). However, except for the notable
exception of several widely studied Salmonidae species (e.g. Tagert 1984, Reiser and
White 1988), little is known about how, or if, high concentrations of suspended
sediments affect aquatic organisms (Cordone and Kelly 1961, Reiser and White 1988,
Servizi and Martens 1991, 1992, Waters 1995).

The enhancement of suspended sediment levels associated with human agricultural,
silvicultural, and industrial activities is widely suspected to force benthic invertebrates
to flee impacted habitats (e.g. Hynes 1973, Newbury 1984, Williams and Feltmate
1992), however, virtually no empirical evidence has been presented to support this
contention (Waters 1995). Extreme sedimentation events such as those that follow
heavy rains in overgrazed river—bottom land (Tarzwell 1938), reservoir flushing (Gray
and Ward 1982), road construction (Ogbeibu and Victor 1989), and riparian zone
clear-cuts (Webster et al. 1992) definitely have the effect of removing benthic
invertebrates. However, in most cases it is not known whether these responses were
instantaneous reactions to stress created by elevated suspended solid concentrations or

55

56

an eventuality associated with habitat degradation (Waters 1995).

Several field experiments have been conducted to determine the direct effects of heavy
sedimentation on benthic invertebrate behavior (Brunskill et al. 197 3, Rosenberg and
Snow 1975 a, b, Rosenberg and Wiens 1978, Culp et al. 1986). For example,
Rosenberg and Wiens (1978) observed that introducing unsorted bank sediments to a
Northwest Territories river caused a marked increase in drifting invertebrates five hours
after the episode. They concluded that sediment addition at 30 mg/ I initially ”strips"
benthos from substrata and that the most sensitive and most exposed organisms leave
first. However, because samples were not taken immediately after sediment addition,
initial stripping was not actually measured. After pouring sand into riffles of a British
Columbia stream, Culp et. al ( 1986) also cited unmeasured scouring effects of saltating
sediment particles as the causative factor of, in this case, instantaneous drift of
invertebrates that inhabit stone surfaces.

One technical problem common to traditional field studies of invertebrate response to
sedimentation is that specific drift—initiating factors remain undetectable with standard
drift—monitoring techniques (Waters 1995). After producing and amlyzing data such as
that reviewed above, White and Gammon (1977) concluded that sedimentation-
mediated drift may not be a consequence of stress caused by sediment (i.e. catastrophic
drift) but rather to light-attenuation—induced behavioral drift (sensu Waters 1972,
Mfiller 1974). Therefore, drift in response to rapid increases in suspended solid
concentration may be interpreted as a response to local overpopulation or resource
scarcity (density-dependent drift) that would not necessarily result in negative effects on
benthic communities (Waters 1995).

The primary objective of this study was to determine if nymphs of the relatively
immobile, predacious stonefly Paragnetina media (Plecoptera: Perlidae) would
immediately reposition in response to the onset of intense sedimentation. Exposure to
sediments was limited to a brief period of time in order to focus on assessing the
immediate, direct effects of sedimentation on stonefly behavior such as integument
scouring and gill fouling.

57

METHODS
Stoneflies

Paragnetina media (Walker) nymphs are common inhabitants in eastern U. S. streams
(Frison 1935, Stewart and Stark 1988). Characteristically associated with fast-flowing
streams, their flattened body shape and gripping tarsal claws allow nymphs to crawl
over relatively silt-free substrata and through accumulations of allochthonous debris in
search of a wide variety of invertebrate prey (Stewart and Harper 1996).

In mid—Michigan, P. media nymphs hatch from early summer-laid eggs and reach
maturity by the spring of their second year (Heiman and Knight 1970). Tightly
synchronous emergence (late May in mid-Michigan) enhances the probability of mating
success of these weak—flying, short-lived adults (Feltmate and Pointing 1986) and
consequently allows determination of nymphal age due to the great disparity in size
when age classes overlap (Heiman and Knight 1970). Species determination was made
with keys in Stewart and Stark (1988). Stoneflies in the autumn of their second year
were used in experiments; they were collected and trials were conducted in a riffle of a
second-order, mid-Michigan pasture stream that flowed over cobble and gravel
substrata (Prairie Creek, Ionia Co., MI; 43°N, 85 °W).

Experimental pmcedure

Nymphs were collected with soft forceps and placed into a pan of aerated stream water
that contained a small stone for them to cling to. When six nymphs were collected (five
for one set of trials), the stone (with clinging nymphs) was removed and placed into a
sedimentation-response chamber (Figure 1). Paragnaina media nymphs are relatively
sessile once a position is established (Felmate and Pointing 1986). This allowed for
trials to get underway before a move unrelated to the treatment was likely to be
attempted (approximately 10 seconds after introduction).

Sedimentation-response chambers are compartmentalized, 2-mm thick plastic boxes
with l mm-diameter mesh walls (Figure 1). Chambers were designed to provide
nymphs with four options in response to exposure to a sediment and water slurry
delivered directly into the chamber from an upstream reservoir. Response options

58

include (I) maintain position in the (220 x 85 x 35 mm) overlying compartment (=
stayed); (2) move down through 20 x 5 mm slots to a like-sized underlying chamber
(=down); (3) move upstream into a 200-mm long, 0.5-mm diameter, mesh net (=
upstream); and (4) move downstream into a into 200-mm long, 0.5 mm mesh net (=
downstream) (see Table 1, Figure 2).

The sediment-delivery system is composed of a 23 1, plastic pale fitted with 3 m of 40—
mm diameter, plastic tube through which the sediment and water slurry (or just water
for control trials) is delivered to a chamber from an upstream position. When the
sediment and wate“ slurry (or water in control trials) had completely passed through the
chamber (approximately 2 minutes after the onset of a trial), it was disconnected from
the sediment-delivery system and removed from the stream. Nymphs were then
collected from the nets and compartments and preserved in 70% ethanol. Trials were
conducted on 19 September, 1993 and 23 September, 1994. A total of eight sediment
and nine control trials were conducted.

Sediments

A slurry consisting of a 1.5 1 container of bank sediments and 23 l of stream water
produced a level of sedimentation downstream from chambers that was similar to that
measured during a heavy spate in Prairie Creek (24 mgll on 15 September, 1993); a
level that is known to produce negative, yet sublethal effects on benthic invertebrates
(Newcombe and MacDonald 1991). The actual suspended sediment level in chambers
was approximately at 56 mgll as determined with a a Hach 2100 A NTU meter and a
known mgll : NTU relationship. Trials were conducted on days when the stream was
running relatively clear of suspended sediments. Therefore, control trials were
conducted with low-sedirnent (> 1 mgll) water samples. Sediment particle-size and
chemical composition was determined by the Michigan State University Plant and Soil
Nutrient Laboratory (Table 2).

Statistical analysis

A single-factor, repeated-measures ANOVA model was used to test the effects of the
two treatments on behavioral response. Response alternatives include maintaining

59

position, moving below the main compartment, moving upstream, or downstream (see
Table 1).

RESULTS

Results presented in Figure 2 indicate that P. media nymphs do not immediately flee
habitat in response to intense sedimentation. Nymphs were expected to reposition at a
higher frequency in sedimentation trials, however they actually moved less in sediment
trials than in controls (mean % 21 70.8 vs 50.7) (ANOVA: P > 0.053, Table 1, Figure
2). There were no significant differences between repositioning options within trials or
between treatments (Table 1, Figure 2).

DISCUSSION

The experimental population of mid-Michigan P. media nymphs were apparently
resistant to, or experimental conditions did not reproduce, gill-fouling (Hynes 1970)
and integument shearing effects (Culp et al. 1986) believed to be associated with high
levels of bank sediment in flow. Perhaps frequent exposure in the recent past has led to
the establishment of the tendency to maintain position during high-sedimentation
episodes due to the commonality of such incidents in the pastured experimental stream
reach. The selective pressure against enacting movements unrelated to resource
acquisition is likely reinforced in Prairie Creek by selection against relocation during
daylight hours due to heightened risk of being eaten by visual-feeding insectivorous
fish (Waters 1972). This pattern was also detected in unquantified preliminary trials
conducted by Cheumatopsyche and Hydropsyche larvae (Trichoptera:
Hydropsychidae) in a Minnesota stream with thriving insectivorous fish populations.

The relatively immobile P. media, may have also acquired a position-maintenance
strategy early in its evolutionary history. In order for an organism to successfully
colonize a novel habitat, it must be able to respond to its most severe conditions.
Almost universally, the behavior to deal with novel stress such as that which seemingly
accompanies habitat shifts and rapidly changing environments, must be expressed prior
to morphological adaptation (Mayr 1982 ). For example, protoplecoptems, terrestrially
adapted animals believed to have secondarily invaded aquatic habitats some 300 mya

60

(Imms 1957, Illes 1965, Ross 1965), had to maintain position while exposed to high
levels of flow and sedimentation events prior to acquiring lotic adaptations such as
streamlined morphology and gripping tarsal claws (Wooten 1972). Therefore,
throughout their history, P. media nymphs and their direct progenitors may have
evolved a strategy of position maintenance in response to episodes of intense
sedimentation.

Evidence presented here indicates that one species with limited mobility, relatively long
development time, and patchily distributed resources is more likely to endure
sedimentation rather than attempt to avoid the stress of exposure by attempting to
reposition. It is clear that benthic invertebrates do not respond to anthropogenic
sedimentation in a consistent manner, thus reinforcing the exigency for clarification the
roles of phylogeny, morphology, and regional ecology in influencing invertebrate
behavioral response to the widespread problem of anthropogenic sedimentation in
streams.

61

LITERATURE CITED

Brunskill, G.J., D.M. Rosenberg, N.B. Snow, G.L. Vascotto, and R. Wagemann.
1973. Ecological studies of aquatic systems in the Mackenzie-Porcupine
drainages in relation to proposed pipeline and highway developments, volume
1. Northern Pipelines Task Force, Environmental-Social Committee. Northern
Oil Development Report 73-40, Winnipeg, Manitoba.

Cordone, A.J. and D.W. Kelley. 1961. The influences of inorganic sediments on the
aquatic life of streams. California Fish and Game 47: 189-228.

Culp, J .M., P.J. Wrona, and R.W. Davies. 1986. Response of stream benthos and
drift to fine sediment deposition versus transport. Canadian Journal of Zoology
64: 1345-1351.

Felmate, B.W., R.L. Baker, and P.J. Pointing. 1986. Distribution of the stonefly
nymph Paragnetina media (Plecoptera: Perlidae): influence of prey, predators,
current speed, and substrate composition. Canadian Journal of Fisheries and
Aquatic Sciences 43: 1582-1587.

Frison, TH. 1935. The stoneflies, or Plecoptera, of Illinois. Bulletin of the Illinois
Natural Survey 20: 281-471.

Gammon, LR. 1970. The effect of inorganic sediment on stream biota. USEPA
Water pollution Research Series 18050 DWC 12170. ‘

Gray, LI. and J .V. Ward. 1982. Effects of sediment releases from a reservoir on
stream macroinvertebrates. Hydrobiologia 96: 177-184.

Heiman, DR and A.W. Knight. 1970. Studies on growth and development of the
stonefly Paragnetina media Walker (Plecoptera: Perlidae) American Midland
Naturalist 84: 274-278.

Hynes, H.B.N. 1973. The effects of sediment on the biota of running water. Pages
653-663 in W. Canadian Department of the
Environment, Ottawa.

Illes, J. 1965. Phylogeny and zoogeography of the Plecoptera. Annual Review of
Entomology 10: 117-140.

Imms, AD. 1957. W (revised by O.W.
Richards and R.G. Davies). New York: E.P. Dutton.

Mayr, E. 1982. W Cambridge: Harvard University
Press.

62

Miiller, K. 1982. The colonization cycle of freshwater insects. Oecologia 52: 202-
207.

Newbury, R.W. 1984. Hydrologic determinants of aquatic insect habitats. Pages
323-357 in V.H. Resh and D.M. Rosenberg (eds.). Ecology of Aquatic
Iii—sects. New York: Praeger

Newcombe, GP. and DD. MacDonald. 1991. Effects of suspended sediments on
aquatic ecosystems. North American Journal of Fisheries Management 11: 72-
82.

N utall, RM. and G.H. Bielby. 1973. The effects of china-clay wastes on stream
invertebrates. Environmental Pollution 5: 77-86.

Ogbeibu, A.E. and R. Victor. 1989. The effects of road and bridge construction on
the bank-root macrobenthic invertebrates of a southern Nigerian streams.
Environmental Pollution 56: 85-100.

Reiser, D.W., and R.G. White. 1988. Effects of two sediment size-classes on
survival of steelhead and Chinook salmon eggs. North American Journal of
Fisheries Management 8: 432-437.

Rosenberg, D.M., and A.P. Wiens. 1978. Effects of sediment addition on
macrobenthic invertebrates in a northern Canadian river. Water Research 12:

753-763.
Ross, H.H. 1965. A Textbook of Entomolggy, 3rd Edition. New York: John Wiley
and Sons

Servizi, J.A. and D.W. Martens. 1991. Effect of temperature, season, and fish size on
acute lethality of suspen®d sediments to Coho Salmon (Oncorhynchus
kisutch). Canadian Journal of Fisheries and Aquatic Sciences 48: 493-497.

Servizi, J .A. and D.W. Martens. 1992. Sublethal responses of Coho Salmon
(Oncorynchus kisutch) to suspended sediments. Canadian Journal of Fisheries
and Aquatic Sciences 49: 1389-1395.

Slaney, P.A., T.G. Halsey, and H.A. Smith. 1977. Some effects of forest harvesting
on salmonid rearing habitat in two streams in the central interior of British
Columbia. British Columbia Fish and Wildlife Branch, Fisheries Management
Report 71, Victoria.

Stewart, K.W. and RP. Harper. 1996. Plecoptera. Pages 217-261 in Merritt, R.W.

and K.W. Cummins (eds.). An Introductign m the Aggtic Ingts of North
America. Dubuque, IA: Kendall Hunt.

63

Stewart, K.W. and B.P. Stark. 1988. Nymphs of North American stonefly genera
(Plecoptera). Thomas Say Foundation Series, Entomological Society of
America 12: 1-460.

Tagart, J .V. 1984. Coho salmon survival from egg deposition to fry emergence.
Pages 173-182 in J .M. Walton and DB. Houston (eds.). Proceedings of the
Olympic Wild Fish Conference. Peninsula College, Port Angeles, Washington.

Tarzwell, CM. 1938. Factors influencing fish food and fish production in
southwestern streams. Transactions of the American Fisheries Society 67 :
246:255.

Tebo, L.B., Jr. 1955. Effects of siltation on trout streams. Proceedings of the Society
of American Foresters 1956: 198-202.

Wagener, S.M., and J .D. LaPierrere. 1985. Effects of placer mining on the
invertebrate communities of interior Alaska streams. Freshwater Invertebrate
Biology 4: 208-214.

Waters, T.F. 1972. The drift of stream insects. Annual Review of Entomology 17:
253-272.

Waters, T.F. 1995. Sediment in Streams. American Fisheries Society Monograph 7,
Bethesda, Maryland.

Wcislo, W.T. 1989. Behavioral environments and evolutionary change. Annual
Review of Ecology and Systematics 20: 137-169.

Webster, J .R., S.W. Golladay, E.F. Benfield, J.L. Meyer, W.T. Swank, and J.B.
Wallace. Catchment disturbance and stream response: an overview of stream
research at Coweeta Hydrological Laboratory. In Boon, P.J., P. Calow, and
GE. Petts (eds.). WWW New York: John Wiley
and Sons.

White, D.S. and J .R. Gammon. 1977. The effect of suspended solids on
macroinvertebrate drift in an Indiana creek. Proceedings of the Indiana
Academy of Sciences 86: 182-188.

Williams, DD. and B.W. Felmate. 1992. MW. Wallingford Oxen U.K.:
C.A.B. International.

Wootten, R]. 1972. The evolution of insects in freshwater ecosystems. Pages 69-82
in R.B. Clark and RJ. Wootten (eds.). W. Exeter:
University of Exeter Press.

64

Table 1. ANOVA results comparing the number of stonefly nymphs that selected each
behavioral option (see Figure 2).

 

 

STAYED DOM U STRE W
Source df MS F MS F MS F MS F
Sed.Ievel 1 1709.68 4.427 1 0.079 1.931 0.013 3.529 0.000 0.007
Error 15 386.16 0.041 0.004 0.003

 

1‘ P: 0.053

65

Table 2. Particle-size distribution and chemical composition of
bank sediments used in sedimentation treatments.

 

Soil Type: Loam, pH 7.1

Mineral Component: 70.7% Organic Component: 29.3%

45.8 % Sand Na 258 ppm
32.7 % Silt Cl 620 ppm
21.4 % Clay N03 5.4 ppm

NH4 74.7 ppm

 

66

 

Figure 1. A sedimentation-response chamber.

Number Of Individuals

67

 

 

 

 

 

 

5..
- [3 Controls
4- Sediment Treatments
3—
,
2-
1 _
0 .

 

 

 

 

 

I 13" 54v:
Stayed Down Upstream D °Wnstream

Figure 2. Mean number of stoneflies (i 1 se) that chose each
behavioral option in 9 control and 8 sedimentation trials.

4

EFFECTS OF EPISODIC SEDIMENTATION ON THE NET-SPINNING
CADDISFLIES H YDROPS YCHE BETTENI AND CERA TOPS YCHE SPARNA
(TRICHOP'TERA: HYDROPSYCI-IIDAE)

Abstract. Net-spinning caddisflies often thrive in the heavily sedimented waters of
midwestern agricultural catchments despite the presumed costs of sedimentation-
induced gill- and net-fouling. We conducted a laboratory experiment to determine
whether larval growth and survival of two mid-Michigan Hydropsychidae species
(Hydropsyche betteni and Ceratopsyche spama) are affected by daily exposure to high
levels of sedimentation. Larvae of both species had a decreased likelihood of survival
in sedimentation treatments relative to controls and taxa were differentially affected as
H. bettem' significantly outperformed C. spama. Sedimentation did not alter the
relative growth rate of either species although slight losses by H. betteni and gains by
C. spama produced significant differences in relative growth rates between species.

INTRODUCTION

Sedimentation to North American streams, most of which is human generated (Waters
1995), is widely believed to harm stream-dwelling organisms through the effects of
increased deposition and transport of sediments (Newcombe and MacDonald 1991,
Waters 1995). High levels of sediment deposition and suspension have been
demonstrated to negatively affect several game-fish species (principally salmonids) by
burying eggs and fouling gills (Newcombe and MacDonald 1991; Sevizi and Martens
1991, 1992; MacDonald and Newcombe 1993; Waters 1995). The situation is not as
clear, however, for the often numerically (Hynes 1970) and energetically dominant
(Waters 1984) animals that comprise benthic invertebrate communities. In fact, the
abundance of many benthic invertebrates typically increases in response to heavy
sedimentation (Hamilton 1961, Hynes 1966, Gammon 1970, Learner et al. 1971,
Nuttall and Bielby 1973, Waters 1995).

68

69

Invertebrate abundance increases that result from high levels of sediment deposition are
typically experienced by relatively small, burrowing species adapted to exploit deposits
of fine-particulate organic material. The conditions that favor these burrowing
organisms often result in the exclusion of species that require solid substrata. Loss of
the relatively large animals that inhabit exposed benthic habitat often results in overall
invertebrate biomass declines (Waters 1995). Further biomass reductions also typically
follow exclusion of exposed-substrata inhabitants as a result of secondary productivity
declines in insectivorous fish populations (Tait et a1. 1994).

Extremely high levels of sedimentation that result from acute disturbances, such as
streambed-suction mining (Thomas 1985), riparian clearcutting (Webster et al. 1992),
and road construction (King and Ball 1967) commonly have large, but impermanent
effects on benthic communities. However, there is of yet only limited understanding of
the effects of chronic sedimentation at levels moderate enough to increase suspended-
solid concentration but preclude total habitat transformation through substrata burial
(Waters 1995 ).

In this study, a presumably agricultural-pollution tolerant species (Hydropsyche
bettem) and a relatively intolerant species (Ceratopsyche spama) (as determined by
Schmude and Hilsenhoff 1986) were exposed to daily episodes of heavy sedimentation
to determine: ( 1) whether larval growth and survivorship are affected by episodic
sedimentation (Fleischner 1994, Waters 1995), and (2) whether taxa would perform
differentially as predicted by present distribution patterns (Ross 1944, Schmude and
Hilsenhoff 1986).

Efi’ects of sedimentation on filter-feeders

Invertebrate species are known to be differentially affected by fluctuations in suspended
sediment concentration. For example, Culp et al. (1986) found that sedimentation-
mediated drift response was most pronounced in invertebrates that inhabit exposed
substrata surfaces. Invertebrates are also expected to differ in response to suspended
sediments as a result of relative exposure as dictated by feeding strategy. Due to their
near uniform reliance on the availability of stable substrata, filter-feeding invertebrates

70

are suspected to be particularily sensitive to suspended-sediment increases (Hynes
1973). However, little evidence has been presented in support of this widely held view
(Waters 1995).

Filter feeders that utilize morphological filtration structures are apparently particularily
sensitive to suspended sediment increases. For example, unionid clams (Aldridge et al.
1987), and cladocerans (McCabe and O'Brien 1983) are known to experience feeding
limitation when exposed to high levels of suspended sediment. However, filter feeders
that utilize external structures such as hydropsychid nets may not be as hindered by
high levels of suspended sediment as are those that feed with anatomical filtration
structures.

Some Hydropsychidae species occupy both high- and low-sediment environments and
are therefore potentially unaffected by moderate to heavy sedimentation (Schmude and
Hilsenhoff 1986). learner et al. (1971) reported that Hydropsyche pellucidula larvae
in clean reaches of a UK. stream were much larger than conspecifics in a reach
receiving suspended solids from coal-mining operations. They concluded that
development was delayed by sedimentation. However, because this hypothesis was
not, and has not been tested (Waters 1995), it remains unknown if larvae in high-
sedimentation environments grow, survive, and reproduce at rates similar to those in
optimal habitat such as the low-sediment environment below many impoundments
(Fremling 1960).

Hydropsychidae nets trap sediment as well as a wide array of potential food items
(Wallace and Sherberger 1974, 1975, Wallace and Merritt 1980). It thus seems
probable that as sedimentation increases, nets require more maintenance and frequent
replacement. The result of higher net-maintenance costs may be the expenditure of
energy that would otherwise be used for respiration and growth. In addition, because
silk production requires expenditure of lipid reserves critical to adult reproductive
success (Wallace and Malas 1976, Petersson and Hasselrot 1994), more-frequent net
replacement may affect caddisfly fitness.

71

METHODS
Microcosms

Four 45 -l aquaria were fitted with stream-flow simulators (BioQuip Products,
Pasadena, CA) and surface-sanitized with 70% ethanol. A 10 x 10 x 40 cm, split-face
concrete block, 5 g of ponderosa pine needles (used for retreat construction), 2 Kg
aquarium sand, 1 Kg gravel, and 8.5 l stream water were added to each tank. All
substrata were autoclave sterilized. An air compressor was used to generate flow.
Water was allowed to circulate for 10 days prior to larval hydropsychid introduction in
order to allow time for the current to create nearly uniform bed morphology between
tanks.

Caddisflies

Final-instar (=fifth instar) Hydropsyche bettem' and Ceratopsyche spama (as
determined by keys in Schmude and Hilsenhoff 1986) were collected from rocks
removed from two second-order streams that flow through a combination of mid-
Michigan agicultm'al and residential land (H. betteni: 21 November, 1994, Prairie
Creek, Ionia Co., 43°N, 85 .W; C. spama 19 November,l994, Flint River, lapierre,
Co., 43°N, 83° 15'W). Larvae were allowed to acclimate to tank temperature (mean =
20°C) from field-collection temperatures of (5°C) before they were introduced into
tanks. Net-spinning activity of both species has been determined to be maximal in 20°C
water (Fuller and Mackay 1980). Twenty-five individuals of each species were added
to each tank.

larvae were fed a daily diet comprised of 0.5 g powdered Tetra MinTM staple food and
0.1 g Red Jungle BrandTM Micro-food. Larvae of both species were observed feeding
on this diet. Feeding was continued for two days following termination of
sedimentation episodes to allow the passage of inorganic material so as to avoid
nonfood items from affecting final weights.

Surviving larvae were placed in 1.5 ml vials, stored at —20°C, thawed, identified,

72

cleansed of debris for 1 minute in a sonic cleaner, dried for 24 hours at 65°C, and
weighed on a microbalance. Relative growth rate of each larva was determined with the
equation:

RGR = (1n final mass - 1n mean initial mass)/ time (days in captivity)
Mean initial mass values were derived from measurements of a representative sample of
each taxa (H. betteni n=24, C. spama n=21) which were frozen upon collection to be
dried and weighed with the experimental animals.

Sediments

Sediment additions were initiated on 26 November, 1996 after all larvae had positioned
themselves and constructed retreats. Sediment was collected from a bank deposit near
the H. betteni collection site. Sediments were autoclave sterilized for one hour,
thoroughly dried at 65°C, and sieved through a 0.6 mm—mesh sieve. Particle size array
and chemical composition were determined by the Michigan State University Plant
Nutrient Laboratory (see Table 1).

Two tanks were randomly chosen to receive daily sedimentation. Each day (16 total),
11 g of sediment were added to experimental microcosms. The initial turbidity (mean =
23 nephelometric turbidity units (NTU)) is considered to be sufficient, if sustained, to
cause adverse, but sub-lethal effects on benthic invertebrates such as reduced growth or
forced abandonment (Newcombe and MacDonald 1991, MacDonald and Newcombe
1993).

Suspended-solid level was measured with a Hach 2100 A NTU meter. NTU
measurements were taken before each trial and several times throughout the first three
hours after the onset of trials, after which levels in sediment-treatment tanks
approximated those in control tanks (Figure 1).

Abiotic conditions

Measurements of water temperature and pH were taken daily. These parameters did not

73

vary between tanks more than the level of aceuracy of the meters (i 1°C, 0.5 pH)
(mean temperature = 208°C, mean pH = 8.5). Current velocity was calibrated
between tanks with a digital flow meter and was set to generate moderately turbid flow
that did not sweep away larvae crawling on exposed substrata. Velocity 5 cm above
retreats was ~5 cm/second.

Statistical analyses

An AN OVA model was used to test the full interaction of the effects of species (two
species), tank (4 tanks), and sediment level (two levels) on relative growth rate (see
Table 2). Survival-rate data were analyzed with a contingency analysis model
(CATMOD procedure, SAS 1990) that tested the same interaction of effects (see Table
3). Survivorship data were analyzed with and without pupae. The addition of pupae to
the analysis of survivorship did not meaningfully affect treatment effects (Appendix 1)
and is therefore excluded from reported values.

RESULTS
Relative growth rate

Sediment treatments had no effect on the relative growth rate of either species (Table 2,
Figure 2). On average, C. spama grew and H. betteni lost mass in sediment
treatments and controls (C. spama: 0.061 vs 0.059 mg dry mass gained per day and
H. betteni: 0.061 vs 0.024 mg dry mass lost per day in control and sediment
treatments respectively) (Figure 2). This interspecific difference is highly significant
(Table 2).

Survival

Both species suffered higher mortality in sediment treatments than in controls, and as
was the ease in relative growth rate measurements, taxa were differentially affected
(Table 3, Figure 3). In contrast to the interspecific trend detected in relative growth
rates, H. betteni vastly outperformed C. spama (Table 3, Figure 3). There was no

74

intraspecific, inter-tank variation with the exception of H. betteni in control tanks
which differed by only one survivor.

Individuals of each species pupated in both treatments (n = 2 H. betteni, 3 C. spama)
prior to completion of the trial. One pupa of undetermined species was consumed by
accidentally introduwd, larval Chironomidae (Eu/defiefiella sp.), a phenomenon also
observed in microcosms by Rutherford (1986) and in nature by Rutherford and Mackay
(1986).

DISCUSSION

Intraspecific effects

Sediments were added at a level predicted to cause sub-lethal effects such as reduced
growth or behavioral avoidance (sensu Newcombe and MacDonald 1991). However,
although the mode of action is unclear, sediment treatments were lethal to a proportion
of experimental populations of both species (Table 3, Figure 3). The two most—often
posited deadly effects of high levels of sedimentation are critically lowered resilience to
suspended solids and enhanced lethality eaused by substrata bmial (Waters 1995).
Entrainment in thick deposits of sediment is known to be lethal to benthic invertebrates,
but only when heavy sedimentation is prolonged enough to completely bury substrata
(Thomas 1985) or when immobile forms like hydropsychid pupae are buried and
ultimately suffocate (Rutherford and Mackay 1986). Either scenario occurring in a
microcosm can be interpreted as representative of potentially negative effects on natural
populations, but certainly not as absolutely lethal ones in habitats where successful
behavioral avoidance is possible.

It is also possible that each microcosm could, perhaps because of spatial limitation, only
support 22 - 23 H. betteni and 10 C. spama (as in control tanks) and, through habitat
denudation, sediment treatments reduced capacity to 15 H. betteni and 8 C. spama (as
in both sediment tanks). Habitat simplification through substrata burial is known to
cause eventual declines in benthic invertebrate populations (e. g. Tarzwell 1938,
Hamilton 1961, Nuttall and Bielby 1973); but as insects were stocked at a small fraction

75

of field densities and many retreats were exposed and unoccupied at the termination of

the 16-day experiment, burial seems unlikely to have been the only cause of larval
death.

Relative growth rates were similar intraspecifically and across treatments suggesting
that surviving larvae were apparently unaffected by sedimentation (Table 2, Figure 2).
Nets were observed to be clogged with sediment after exposure and were cleaned or
replaced prior to the onset of the next trial. It seems reasonable, therefore, to presume
that conditions in the sediment-treatment tanks required higher net maintenance costs
relative to those in controls . However, because sediment treatments had no effect on
relative growth rates (Figure 2), net maintenance costs over 16 days were probably
negligible for survivors of sediment treatments. The possibility does exists that
younger larvae would have responded differently due their typically higher growth rates
(Cudney and Wallace 1980, Mackay 1979, 1984) and net-spinning activity (Fuller and
Mackay 1980). However, other indirect measurements of growth rate have revealed
that final-instar hydropsychids do gain mass (Cuffney and Minshall 1981).

Intetspecific reflects

The expectation that H. betteni are more tolerant to sediment treatments than C. spama
(Schmude and Hilsenhoff 1986) was not corroborated by interspecific comparisons of
relative growth rate and survivorship. Survivorship data suggest that larval mid-
Michigan H. betteni are more manipulation-resistant than C. spama, which may
indieate higher general resistance which is also evidenced by their distribution in
Wisconsin (Schmude and Hilsenhoff 1986). However, although captivity was more
lethal for C. spama larvae overall, survivors actually fared better than the average H.
betteni larva which experienced mass loss.

The overall higher mortality incurred by experimental C. spama populations could also
have been in some way influenced by interspecific interactions, perhaps competition for
high-quality retreat sites. Hydropsychid battles over retreat sites are common, and not
surprisingly, the odds of intruder victory typically favor larger combatants (Jannsson
and Vuoristo 1979). Therefore, because final-instar H. betteni are more than three

76

times larger than final-instar C. spama (Figure 2), the higher probability of winning
battles over retreats may have influenced mortality in the spatially limited microcosm
environments.

In his excellent review of the effects of sedimentation in streams, Waters (1995: 60)
proclaimed that ”on the basis of current knowledge, the direct effect of suspended
sediment upon benthic invertebrates does not appear to be a significant influence upon
stream invertebrate communities.” Results from this study, as well as those presented
in Chapter 3, indicate that some benthic invertebrates are sensitive to, and potentially
harmed by, high levels of sediment in suspension. Although the effects of
sedimentation on invertebrate community composition are typically more the result of
sediment deposition than transport, additional stress inflicted on benthos by elevated
suspended sediment levels may impose chronic, low-level stress and therefore have
indirect effects upon invertebrate communities. Given the enormity of the problem, it is
conceivable that almost every lotic invertebrate population has been exposed to
sedimentation that resulted from human activities. It seems certain, therefore, that
widespread and chronic sedimentation-enhancing activities such as overgrazing of cattle
in small stream riparian areas, have in aggregate, immense ecological consequences.

77

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Aldridge, D.W., B.S. Payne, and A.C. Miller. 1987. The effects of intermittent
exposure to suspended solids and turbulence on three species of freshwater
mussels. Environmental Pollution 45: 17-28.

Cudney, MD. and J .B. Wallace. 1980. Life cycles, microdistribution, and production
dynamics of six species of net-spinning eaddisflies in a large southeastern
(USA) river. Holarctic Ecology 3: 169-182.

Cuffney, T.F. and G.W. Minshall. 1981. Life history and bionomics of Arctopsyche
grandis (Trichoptera) in a central Idaho stream. Holarctic Ecology 4: 252-262.

Culp, J .M., F.J. Wrona, and R.W. Davies. 1986. Response of stream benthos and
drift to fine sediment deposition versus transport. Canadian Journal of Zoology
64: 1345-1351.

Fleischner, TL. 1994. Ecological costs of livestock grazing in western North
America. Conservation Biology 8: 629-644.

Fremling, CR. 1960. Biology of a large mayfly, Hexagenia bilineata (Say), of the
Upper Mississippi River. Iowa Agricultural Experiment Station Research
Bulletin 482: 842-852.

Fuller, R.L. and RJ. Mackay. 1980. Field and laboratory studies of net—spinning
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81

Table 1. Particle-size distribution and chemical composition of
bank sediments used in sedimentation treatments.

 

 

Soil Type: loam, pH 7.1

Mineral Component: 70.7 % Organic Component: 29.3%

45.8 % Sand Na 258 ppm
32.7 % Silt Cl 620 ppm
21.4 % Clay N03 5.4 ppm

N H4 74.7 ppm

 

82

Table 2. ANOVA results comparing relative growth rates of Hydropsyche betteni and
Certaopsyche spama in two sediment treatments and two controls (see Figure 2).

 

 

Source df MS F P
Species l 0.01 1 80.900 0.0001
Sed. Level 1 0.000 0.051 0.8221
Species x Sed. Level 1 0.000 0.291 0.5907
Tank (Sed. Level) 2 0.000 0.295 0.7452
Species x Tank (Sed. level) 2 0.000 1.968 0.1449
Error 103 0.000

 

83

Table 3. Contingency analyses testing whether Hydropsyche betteni and
Ceratopsyche spama survival differed in sediment treatments and controls

(two tanks each) (see Figure 3).

 

 

Source df x2 P

Species l 29.24 0.000
Sed. Level 1 9.54 0.002
Species x Sed. Level 1 4.38 0.036
Tank (Sed. Level) 2 0.16 0.923
Species x Tank (Sed. level) 2 0.16 0.923

 

84

 

25 ’ ....

 

+ Sediment Tanks

 

+ Control Tanks

N
O
l 1 1 L 1

 

 

Nephelometric Turbidity Units
<15 i“

A #1 A 1 I l

 

 

 

o l I l I l

l l l
0 1 45 60 90 120 180 240 360
Minutes After Onset of Sedimentation

Figure 1. Mean NTU :1: 1 se in two sediment- treatment and two
control tanks.

85

'8

 

- C, spama Initial Mace:
° 2.66 mg. i 0.25 'F‘

0.061
: T
(IL ill

0.02': '7'- J

: El

4

.o
.9

 

 

 

 

 

 

 

 

Relative Growth Rate (mg.dry mass/day)

 

 

 

 

 

O
0.02; H. betteni lpl M
f Inltlallhu:
. 3.13 . 0.43
-o.04— "" i ~—
3 ._J
-o.oe 1

 

Sedlment Treatments I Controls

Figure 2. Box plots of mean RGR :i: 1 SE of Hydropsyche
betteni and Ceratopsyche spama in two sedimented and two
control tanks. Mean pre-trial dry mass (:1: 1 se) indicate relative
size of fifth-instar H. betteni (n=24) and C. spama (n=2l).

 

% Survival

86

Sediment
Tanks

 

Control
Tanks

 

 

 

 

 

 

 

 

 

 

 

 

 

C. spama H. betteni

Figure. 3. Mean survivorship of larval Ceratopsyche
spama and Hydropsyche betteni in two sedimented and
two control tanks. Standard error = 0.000 for all but H.
betteni control which is 0.020.

87

Appendix 1. Contingency analyses testing whether Hydropsyche betteni and
Ceratopsyche spama survival differed in sediment treatments and controls with

pupae included (see Table 3).

 

 

Source df x2 P

Species 1 34.84 0.001
Sed. level 1 6.76 0.000
Species x Sed. level 1 3.86 0.049
Tank (Sed. level) 2 0.32 0.850
Species x Tank (Sed. Level) 2 0.32 0.850

 

APPENDIX 1

88
APPENDIX 1

Record of Deposition of Voucher Specimens*

The specimens listed on the following sheet(s) have been deposited in
the named museum(s) as samples of those species or other taxa which were
used in this research. Voucher recognition labels bearing the Voucher
No. have been attached or included in fluid-preserved specimens.

Voucher No.: 1996-8

Title of thesis or dissertation (or other research projects):
Some effects of riparian habitat alteration on lotic invertebrate

ecology

Museum(s) where deposited and abbreviations for table on following sheets:

Entomology Museum, Michigan State University (MSU)

Other Museums:

Investigator's Name (s) (typed)

 

Roger Malcolm Strand

 

 

Date 17 Septembe;l 1996

*Reference: Yoshimoto, C. M. 1978. Voucher Specimens for Entomology in
Nerth America. Bull. Entomol. Soc. Amer. 24:141-42.

Deposit as follows:

Original: Include as Appendix 1 in ribbon copy of thesis or
dissertation.

Copies: Included as Appendix 1 in copies of thesis or dissertation.
Museum(s) files.
Research project files.

This form is available from and the Voucher No. is assigned by the Curator,
Michigan State University Entomology Museum.

 

 

APPENDIX 1.1

89

APPENDIX 1.1
voucher Specimen Data

Pages

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