legacy effects of alternative riparian forest harvest
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Legacy Effects of Alternative Riparian Forest Harvest Practices on Legacy Effects of Alternative Riparian Forest Harvest Practices on
Headwater Streams and Forest Ecosystems in Western Maine Headwater Streams and Forest Ecosystems in Western Maine
Mitchell Paisker University of Maine, [email protected]
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LEGACY EFFECTS OF ALTERNATIVE RIPARIAN FOREST HARVEST PRACTICES
ON HEADWATER STREAMS AND FOREST ECOSYSTEMS IN WESTERN MAINE
By
Mitchell Paisker
B.S. University of Maine, 2017
A THESIS
Submitted in Partial Fulfillment of the
Requirements for the Degree of
Master of Science
(in Ecology and Environmental Science)
The Graduate School
The University of Maine
May 2020
Advisory Committee:
Dr. Hamish S. Greig, Associate professor of Stream Ecology, Co-Adviser
Dr. Amanda Klemmer, Assistant Professor of Food-Web Ecology, Co-Adviser
Dr. Stephen M. Coghlan Jr., Associate Professor of Freshwater Fisheries Ecology
LEGACY EFFECTS OF ALTERNATIVE RIPARIAN FOREST HARVEST PRACTICES
ON HEADWATER STREAMS AND FOREST ECOSYSTEMS IN WESTERN MAINE
By Mitchell Paisker
Thesis Advisors: Dr. Hamish S. Greig and Dr. Amanda Klemmer
An Abstract of the Thesis Presented
In Partial Fulfillment of the Requirements for the
Degree of Master of Science
(in Ecology and Environmental Sciences)
May 2020
Abstract
Riparian buffers are used as a management tool for sustaining freshwaters during forest
harvest, but the log-term (decadal) outcomes of riparian management practices are not well
understood. To address the impact that variable riparian timber harvest has on stream
invertebrates, fish, terrestrial invertebrates and forest composition, environmental data were
collected from streams that experienced a range of harvests; clear-cuts with 0m, 11m, or 23m
buffers, selection harvest or no harvest at all. Riparian zones were initially harvested during the
winter of 2000-2001. Results within this thesis are from data collected during the summers of
2018 and 2019. Aquatic invertebrates were collected via Surber samples and leaf bags. Aerial
invertebrates within the terrestrial landscape, including emerged adult aquatic invertebrates, were
collected using sticky traps placed within the stream corridor. Fish were collected using
quantitative electrofishing. Lastly, forest composition was calculated using forest inventory plots
with diameter at breast height (DBH) and species was recorded for every stem over 1.37 meters
tall. Results show a shift in forest community composition towards smaller stems and more
deciduous trees with more disturbance to the riparian zone. Aerial invertebrates show a shift in
community composition both seasonally and with respect to treatment type. This was noted as a
change in proportion of adult aquatic to terrestrial invertebrate as well as composition of adult
aquatic invertebrate. Aquatic invertebrates experienced a shift in community composition as well
with an increase towards shredders, scrapers and more diverse Diptera taxa in streams with more
disturbed riparian habitat. Abundance and condition of brook trout also increased with greater
disturbance. Importantly, we observed stronger effects of riparian forest harvest on community
composition for each our response variables than total abundance and diversity of taxa. This
suggests that the scale to which data was analyzed is important in detecting ecosystem responses
to riparian management, and also that aggregate measures like diversity and total abundance may
be more robust to environmental change than specific taxa. In this case, current management
guidelines could prove ineffective to maintaining stream and riparian zone community
composition.
ii
ACKNOWLEDGEMENTS
I would like to thank my advisers, Dr. Hamish Greig and Dr. Amanda Klemmer. Without
their support and confidence in me as a budding ecologist, I would not be where I am today.
Thank you for taking a chance with me and bringing me into the lab group. Additionally, I would
like to thank my third committee member Dr. Stephen M. Coghlan Jr. for guidance and help
along the way. All three, were at multiple times, professors of mine during undergrad at UMaine
and I looked up to them all immensely since then. Having the opportunity to work alongside my
committee as well as the broader research team of Dr. Shawn Fraver, Dr. Mindy Crandall, and
Dr. Robert Northington, Dr. Brian Roth, The Cooperative Forest Research Unit (CFRU) and
Ethel Wilkerson, has been a truly wonderful experience. I would like to thank my incredible
support system, field technicians and support along the way. Mom and Dad, thank you for
always believing in me and supporting me throughout the years while at UMaine or any other
endeavor I set forth to accomplish. Ethan Cantin, Kathleen Brown (Tang Gang), Kate O’Connor,
Michaela Forbes, Josh Smith, Brad Erdman, Joe Mohan, Isaac Shepard and Zach Wood: this
project would not have been possible without your incredible help and dedication of so many
hours doing arduous field work and tiresome lab processing. You were all incredible workers
and I cannot thank you enough for the time and effort that you dedicated to helping me. Zach and
Isaac, thank you for always being willing to selflessly help me with my project and analyses.
Without all of you this project would not have been possible. To the Gremmer Lab, thank you
for helping me narrow my project down and for all the advice you gave me over my time here. I
was very fortunate to be able to conduct this project and strengthen my skills as an ecologist in
the beautiful state of Maine and with strangers that became friends, and friends that became
family.
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TABLE OF CONTENTS
ACKNOWLEDGEMENTS……………………………………………………………………….ii
LIST OF TABLES………………………………………………………………………………...v
LIST OF FIGURES…………………………………………………………………………...….vi
CHAPTER 1: THE RESPONSE OF FOREST AND AERIAL INVERTEBRATE
COMMUNITIES TO ALTERNATIVE RIPARIAN TIMBER HARVEST PRACTICES………1
Introduction ................................................................................................................................. 1
Methods ....................................................................................................................................... 3
Forest Metrics .......................................................................................................................... 4
Aquatic Subsidies and Terrestrial Invertebrates ...................................................................... 4
Statistical Analyses .................................................................................................................. 5
Results ......................................................................................................................................... 6
Forest Attributes ...................................................................................................................... 6
Flying Invertebrate Communities ............................................................................................ 7
Discussion ................................................................................................................................. 14
Effects of Forest Harvest on Riparian Forest Composition ................................................... 15
Effect of Harvest on Adult Invertebrates ............................................................................... 16
Connections Between Forest Structure and Adult Insect Communities ................................ 17
Conclusion ................................................................................................................................. 19
CHAPTER 2: AQUATIC INVERTEBRATES AND BROOK TROUT RESPONSES TO
ALTERNATIVE RIPARIAN FOREST HARVEST.................................................................... 20
iv
Introduction ............................................................................................................................... 20
Methods ..................................................................................................................................... 23
Stream Habitat and Invertebrate Sampling ............................................................................ 23
Fish Monitoring ..................................................................................................................... 24
Statistical Analysis ................................................................................................................ 25
Results ....................................................................................................................................... 27
Leaf Bags ............................................................................................................................... 27
Surber Samples ...................................................................................................................... 27
Fish Data ................................................................................................................................ 28
Discussion ................................................................................................................................. 33
Responses of Stream Invertebrates ........................................................................................ 34
Responses of Stream Fish to Riparian Forest Harvest .......................................................... 36
Conclusions ............................................................................................................................... 38
CHAPTER 3: CONCLUSION ..................................................................................................... 40
REFERENCES ............................................................................................................................. 43
APPENDIX…..…………………………………………………………………………………..50
BIOGRAPHY OF AUTHOR ....................................................................................................... 57
v
LIST OF TABLES
Table 1.1 Statistical outputs of CCA and ANOVA’s of forest inventory and sticky traps…….14
Table 2.1 Statistical outputs of CCA and ANOVA’s of invertebrate and fish data…..………..33
Table A. 1 Total abundances of all invertebrates collected in fourteen study streams over
two years via Surber sample………..…………………………………………………..50
Table A. 2 Total abundance of all trees measured from the fourteen streams…………...……..53
Table A. 3 Total abundances of all invertebrates sampled via all leaf bags from fourteen
streams………………………………………………………………………….………54
Table A. 4 Total abundances of invertebrates of their respective order per stream from
both sampling years of sticky traps………………………………………………..…...56
vi
List of Figures
Figure 1.1 A conceptual diagram of study design with the location of forest inventory
plots and sticky traps …………………………………………………………………..…5
Figure 1.2 Cannonocal Correspondence Analysis (CCA) ordinations on Hellinger-
transformed tree abundance data collected from fourteen streams spanning
five harvest treatment types (polygons)……………………………………………….......8
Figure 1.3 CCA ordinations on deciduous tree communities across harvest types…………….....9
Figure 1.4 CCA ordination on small trees (<1.37 m tall) across harvest types………………….10
Figure 1.5 Mean stem counts for total stems (A.) and (B.) deciduous tree stems per
treatment type……..…………………….………………………………………….........11
Figure 1.6 Cannonocal Correspondence Analysis (CCA) ordinations of aerial
invertebrate abundance across the five harvest treatments………………………..….…12
Figure 1.7 Mean total abundances of (A.) all invertebrates and (B.) EPT taxa collected
on sticky traps per harvest type………………………………………………………….13
Figure 2.1 A conceptual diagram of the study system with the locations of each
sampling technique; Surber samples, leaf bags and electrofishing.......…………………25
Figure 2.2 Cannonocal Correspondence Analysis (CCA) ran on Hellinger-transformed
invertebrate abundances from leaf bags (A, B) and Surber samples (C, D)……………..29
vii
Figure 2.3 Total log transformed abundances of invertebrates per Surber sample per
streams without riparian harvest (11m, 23m, Control) and with riparian
harvest (0m, Partial) in their respective riparian harvest category…………………….30
Figure 2.4 Total fish biomass collected per stream (A) and total number of fish
collected per stream (B) in the presence and absence of riparian harvest.…………......31
Figure 2.5 Condition of fish is represented using Fulton’s fish condition of fish collected
in Riparian harvest treatments…………………………………...…........……………...32
1
CHAPTER 1: THE RESPONSE OF FOREST AND AERIAL INVERTEBRATE
COMMUNITIES TO ALTERNATIVE RIPARIAN TIMBER HARVEST PRACTICES
Introduction
Disturbance in small streams can often be characterized by natural factors like floods or
droughts and alterations within the catchment, such as riparian timber harvest (Malmqvist,
2002). Riparian harvest often leads to a change in stem density of trees and shrubs which can
alter the level of complexity in the forest understory or canopy (Boothroyd et al., 2004). Many
emergent stream taxa are relatively poor flyers in their adult stage (i.e. Trichoptera, Plecoptera
and Ephemeroptera) and are therefore vulnerable to changes in forest structure that affect
dispersal corridors. For example, Curry and Baird (2015) showed that caddisflies were limited in
their dispersal abilities depending on the complexity of the surrounding vegetation. Dispersal
increased in open areas and grassy riparian zones and decreased in zones with heavy tree cover.
Additionally, because aquatic invertebrates tend to disperse upstream along the stream corridor,
dispersal can be influenced by tree density in the immediate stream bank of riparian areas
(Romaniszyn et al., 2006). These landscapes also vary along a temporal gradient as a forest
undergoes succession. For example, dispersal can be hindered by dense stands of woody shrub
and pioneer species (i.e. witch hazel, speckled alder) obstructing their flight path along the
stream corridor (Petersen et al., 2013). This is important because invertebrate presence is often a
product of large scale processes and disturbances more than local population dynamics
(Peckarsky et al., 2000) meaning alterations downstream can affect invertebrate communities
upstream. Moreover, these aspects suggest riparian forest structure could alter the movement of
aquatic subsidies into the terrestrial ecosystem.
2
Forest permeability also plays a large factor in the dispersal and movement of flying
invertebrates within the terrestrial landscape, which can have repercussions for aquatic
ecosystems. Galic et al., (2013) showed that low permeability resulted in lower dispersal of
terrestrial invertebrate species from their natal grounds. These mechanisms could decrease the
transfer of terrestrial invertebrates to the stream ecosystems and therefore increase reliance of
aquatic food webs on aquatic primary and secondary production. Conversely, a more permeable
forest could result in a larger transfer of invertebrates between the terrestrial and aquatic
landscape.
Forest harvest within the riparian zone has the ability to alter both the quantity and
quality of terrestrial subsidies that could enter the stream system. This can be seen by a reduction
in trees within the riparian zone and successive plants taking their place. Hagen et al. (2006)
compiled forest inventory data following the original forest harvests of this study and found that
there is a distinct shift among plant communities as distance from stream increases. He found the
first five meters of the riparian zone directly adjacent to the streams, while small, was largely
comprised of herbaceous shrubs. After this distance, trees began to be much more prominent.
This change in vegetation is a potential shift in terrestrial subsides entering the stream systems.
Changes in riparian forests and associated subsidies can lead to a change in the aquatic
invertebrate communities and subsequently a change in emergence from the stream (Leberfinger
et al., 2011). Kominoski (2017) showed a shift in invertebrate taxa and forage preference in the
presence of broadleaves and needles in aquatic invertebrates. Additionally, plant herbivory via
terrestrial invertebrates following forest harvest increases due to a shift in vegetation to early
3
successional species (Luttge et al., 2007). Thus, adult stages of aquatic and terrestrial
invertebrates are likely to be influenced by differences in riparian forest harvest.
The goal of this chapter is to determine whether forest harvest within the riparian zone
affects the composition of adult aquatic and terrestrial invertebrate communities present within
stream corridors, and whether these patterns are consistent with differences in the forest structure
of riparian zones. Additionally, I determined the long-term (<15 year) effects of riparian harvest
on the structure and composition of trees within the buffers. I predict to see a higher terrestrial
invertebrate abundance as well as overall abundance of invertebrates along streams that were
clear-cut without a buffer and streams harvested with partial harvest compared to those with
more intact riparian forests. I also predict to see a larger proportion of adult aquatic taxa that are
more susceptible to environmental change (Ephemeroptera, Plecoptera, Trichoptera) compared
to streams that had no riparian forest harvest.
Methods
I assessed the importance of riparian forest harvest on forest structure and adult aquatic
and terrestrial invertebrate communities in the riparian forests of fourteen streams in the Western
mountains of Maine seventeen years after harvest occurred. Fifteen streams, with known harvest
history, were selected to determine the long-term effects of riparian zone timber harvest
conducted in the winter of 2001-2002 (Wilkerson et al., 2006). Five harvest types were
established over three streams each; control, clear-cut with 0m riparian buffer, clear-cut with
11m buffer, clear-cut with 23m buffer and clear-cut with partial harvest within the buffer to a
residual basal area of 13.7m2/ha. All clear-cuts were 200m wide by 300m long (6 ha) along both
sides of the stream with no harvest at least 20 years prior. Streams fell within the Upper
Kennebec, Lower Kennebec, Dead River, Upper Androscoggin and Lower Androscoggin
4
watersheds. Five sample locations were equally spaced along each stream at 0m, 75m, 150m,
225m, and 300m (Figure 1.1). I was able to sample fourteen of the fifteen streams as one control
stream did not contain surface water through the duration of this study. Sticky traps were placed
along the stream margins to measure amounts of adult aquatic invertebrates and aerial terrestrial
invertebrates within harvested zones and standard forest inventories to assess the structure of the
riparian forest.
Forest Metrics
Sampling protocol for forest condition consisted of five 200 m2 (7.9 m radius) forest
inventory plots equally placed along each side of the harvested stream for a total of ten plots
each on the control, 0m, 23m and partial harvest streams. Streams with 11m buffers were
inventoried with twenty 100m2 (5.64m radius) plots to follow protocol established in the original
sampling event in 2001. Within these plots, every tree over 10 cm diameter at breast height (1.37
meters high; DBH) was identified and measured. At the center of each plot, a 20 m2 (2.5 m
radius) plot was established in which every stem that was over 1.37 meters tall and under 10 cm
DBH was recorded.
Aquatic Subsidies and Terrestrial Invertebrates
We used sticky traps to estimate the abundance and diversity of both adult aquatic and
terrestrial invertebrates within the riparian landscape. Each trap consisted of two 8.5 x 11-inch
transparencies with Tanglefoot® spread across each side. Traps were oriented perpendicular to
the stream to collect invertebrates flying up and down stream. Two traps were placed at sites
75m, 150m and 225m on each stream within the study in June and July of 2018 for
approximately one month each. Sticky traps were again placed at the same locations in May of
2019 and retrieved after one month. Upon retrieval, traps were wrapped in plastic wrap to
5
prevent the addition or loss of invertebrates from the traps and frozen until lab processing. In the
lab, invertebrates were enumerated and identified to order, with the exception of Diptera, which
was identified to the sub-order Brachyacera and Nemotocera).
Figure 1.1
A conceptual diagram of study design with the location of forest inventory plots and sticky traps.
(Note: This design was used for streams with 0m, 23m, partial and control treatments. Streams
with 11m buffers had twice as many forest inventory plots equally spaced along the stream
reach.)
Statistical Analyses
Total abundance of invertebrates on sticky traps were aggregated to obtain a total amount
of invertebrates per order per year per stream. Thus, the unit of replication was a stream in each
sampling year. ANOVA’s were used on the total abundances to determine if there was a
difference between treatment type and riparian harvest. Response variables for these tests were
the total number of invertebrates (both aquatic and terrestrial) per trap as well as the total EPT
taxa per trap, while their predictors were ‘treatment type’, ‘riparian harvest’ and season. An
6
abundance matrix was made and subsequently Hellinger transformed (using ‘Decostand’ in
Vegan) to down weight the effects of highly abundant taxa and meet assumptions of normality in
subsequent analyses. Canonical correspondence analyses (CCA) were then performed on
invertebrate communities based on two measurements of harvest type. The first, treatment type,
focused on treatment (clear-cuts with buffers of, 0m, 11m, 23m, partial harvest and controls)
whereas the second, riparian harvest, combined treatments into those in which the riparian zone
was harvested (0m and partial harvest), and those where the riparian zone was left unharvested
(11m, 23m and control). Permutational multivariate analysis of variance (perMANOVA) tests
were then run to assess statistical significance of differences among related treatments, using
either treatment type or riparian harvest as categorical predictors.
Forest inventory data were similarly processed, aggregated, transformed and analyzed
using the same categorical predictors. Data was analyzed as both total stems as well as small
stems (>10cm DBH) and deciduous vs coniferous stems. CCA were then run on forest inventory
data in the same fashion described above. All analyses were run in R version 3.5.2 using the
“Vegan” package for ordinations.
Results
Forest Attributes
3,725 woody stems were measured for diameter at breast height (DBH) and species along
both sides of fourteen headwater streams. Treatment type had a significant effect on tree
diversity as well as density (p= 0.011; Table 2.1; Figure 1.2). Control streams were classified by
a large abundance of red spruce while riparian zones with more intensive harvest had a larger
portion of trees such as speckled alder. Deciduous composition within riparian zones of different
treatment types differed (p=0.023; Figure 1.3). Treatments with a high level of disturbance all
7
showed a large portion of stems being attributed to speckled alder while those less disturbed
contained species like silver maple. Treatment type showed total numbers of stems under 10cm
DBH and under 1.37m tall differed as well (p=0.004; Figure 1.4). Strong differences in these
communities showed speckled alder and sugar maple being significant drivers of community
composition as harvest intensity increased while treatments like the control were driven by
species like cedar.
Flying Invertebrate Communities
Traps were responsible for capturing 63,297 invertebrates identified to 9 orders (Diptera
taxa were classified to the suborder level of Nematocera and Brachycera). These data were
analyzed using a CCA to determine differences in flying invertebrate abundances along the
gradient of forest harvest. Comparing invertebrate community composition, there was a
statistical difference seasonally (p=0.001; Figure 1.6A) as well as compared across harvest types
(p=0.013; Figure 1.6B). Additionally, when testing the interaction of the two variables, season
and harvest type, there was significant interaction (p=0.017). Ordinations showed a change in
taxa per season being driven largely by Ephemeroptera in the early summer, Trichoptera in mid-
summer and then assorted Diptera taxa in late summer.
Stream averages of invertebrates caught on sticky traps ranged from 968 invertebrates per
trap (partial) to 3,412 invertebrates per trap (23m). Averages per harvest type were highest in the
partial harvest treatment type (Figure 1.7A). EPT taxa stream averages ranged from 58
individuals per trap (23m) to 7 individuals per trap (control) and treatment averages were 18.4
individuals per trap (11m) to 38 individuals per trap (23m) (Figure 1.7B).
8
Figure 1.2:
Cannonocal Correspondence Analysis (CCA) ordinations on Hellinger-transformed tree
abundance data collected from fourteen streams spanning five harvest treatment types
(polygons). A negative score for CCA1 corresponds with big tooth aspen while a positive score
corresponds with red spruce and silver maple. Positive CCA2 scores relate to white spruce and
speckled alder while negative scores relate to elm and pin cherry.
9
Figure 1.3:
CCA ordinations on deciduous tree communities across harvest types. Positive values
correspond to speckled alder for both axes while a negative CCA1 score relates to silver maple
and negative CCA2 relates to elm.
10
Figure 1.4:
CCA ordination on small trees (<1.37 m tall) across harvest types. Those subject to harvest type
showed a significant difference of 0.001. Positive scores for CCA1 and CCA2 are red spruce and
cedar, and sugar maple and speckled alder respectively. Negative scores for CCA1 and CCA2
both correspond to black spruce and elm.
11
Figure 1.5:
Mean stem counts for total stems (A.) and (B.) deciduous tree stems per treatment type.
12
Figure 1.6:
Cannonocal Correspondence Analysis (CCA) ordinations of aerial invertebrate abundance across
the five harvest treatments. Panel A through C represent sites plotted in early summer, mid-
summer and late summer respectively. Polygons enclose the points for each harvest type in a
given season showing important taxa within each season being A. Ephemeroptera, B.
Trichoptera, and C. a mix of Diptera and Trichoptera.
A B C
13
Figure 1.7:
Mean total abundances of (A.) all invertebrates and (B.) EPT taxa collected on sticky traps per
harvest type.
A
B
14
Table 1.1
Statistical outputs of CCA and ANOVA’s of forest inventory and sticky traps. Response
variables were tested with predictors of treatment type (0m, 11m, 23m, partial, control) and
riparian harvest (with- 0m, partial, without- 11m, 23m, control).
Data source Response variable Predictor Statistical
Test dF F P
Forest
Inventory
Riparian Forest
Community
Composition
Treatment
type CCA 4, 23 1.54 0.011
Total Stems Treatment
type ANOVA 4, 23 0.49 0.744
Deciduous Tree
Community
Composition
Treatment
type CCA 4, 23 1.53 0.023
Deciduous Stems Treatment
type ANOVA 4, 23 1.44 0.254
Small Stem
Community
Composition
Treatment
type CCA 4, 9 1.99 0.004
Small Stems Riparian
harvest ANOVA 1, 12 1.46 0.107
Sticky Traps
Invertebrate
Community
Composition
Season CCA 2, 23 11.63 0.001
Invertebrate
Community
Composition
Treatment
type CCA 4, 23 1.91 0.008
Invertebrate
Community
Composition
season x
treatment
type
CCA 8, 23 1.65 0.017
Total
Invertebrates/trap
Treatment
type ANOVA 4, 33 0.58 0.676
Total EPT/ trap Treatment
type ANOVA 4, 33 1.40 0.255
Discussion
The goal of this chapter was to study the long-term-effects of forest harvest on the
riparian forest and the flying adult invertebrate communities of headwater streams throughout
western Maine. Unlike response observed within streams (Chapter 2), significant differences in
riparian tree and invertebrate communities were observed at the treatment types level rather than
15
riparian harvest. Results show a shift in composition of total stems and deciduous stems in
riparian zones along streams that had 0m buffers or partial harvest. Furthermore, this shift in
stem density associated with an increase in small stems characteristic of early regeneration (i.e.
speckled alder, red maple). Invertebrate composition also showed significant variation among
treatment type and these differences varied seasonally. Together these results indicate riparian
environments exhibit signatures of riparian harvest and management many years after the
conclusion of harvest.
Effects of Forest Harvest on Riparian Forest Composition
Forest harvest, by definition reduces the number of standing stems within a forest.
However, after a forest begins to regenerate, stem density can increase to levels exceeding pre-
harvest conditions (Carswell et al., 2007). Streams with less intensive harvesting in riparian
zones had fewer stems when compared to those that had higher harvest intensities (i.e. 0m and
partial harvests). Those more intensively harvested treatments, however, had a higher average
amount of trees, particularly smaller stems. This can be explained by the extreme amount of
forest regeneration taking place (Finegan, 1984; Harvey and Bergeron, 1989; Pierce et al., 1993).
Finegan (1984) describes this process as a shift in vegetation and a change in the plant species
that vary based on the time it takes to germinate and grow as well as those that are tolerant to
biotic driven (e.g. high competition and shading) and abiotic driven (e.g. high temperatures and
scarified soils) stressors. Post-harvest, forest floors can reach extreme temperatures, soils can be
scarified and unfertile and water regimes can be altered; all characteristics of an early succession
forest (Finegan, 1984; Pierce et al., 1993). These factors could explain why we found a higher
abundance in streams that were subject to intensive riparian forest harvest in this experiment.
16
Shifts in tree composition often occur following forest harvest. Harvey and Bergeron
(1989), recorded a shift of up to 92% reduction of softwood regeneration in a softwood dominant
stand to a mixed or hardwood-shrub dominated ecosystem following whole-tree removal. This
shift towards hardwood regeneration was also noticed within our data (Figure 1.2). A change in
tree composition from coniferous to mixed or hardwood shrub dominated can result in a change
in leaf quality, especially a change in phenolics (Harvest and Bergeron, 1989). Tannins, a type of
phenolic, is associated with a decreased level of herbivory by both terrestrial invertebrate
herbivores and mammals. Reduced hervbivory can then lead to an increase in shading over the
streams (Barber and Marquis, 2010) and theoretically provide more habitat for terrestrial
invertebrates to live. This increase in terrestrial invertebrates can have an adverse effect on
stream invertebrates (Paisker Chapter 2, 2019).
Effect of Harvest on Adult Invertebrates
Results from sticky trap analysis show a strong effect of both season and harvest type on
invertebrate community composition as well as the interaction of season and harvest type.
Emergence of aquatic invertebrates, especially taxa such as Ephemeroptera, Plecoptera, and
Trichoptera, is often heaviest in early months of the summer (Petersen et al., 1999; Shualia,
Salmah, Al-Shami 2011) because larval stages will overwinter and emerge once water
temperatures begin to rise. Following this, a secondary peak can be observed later in the summer
months (Nakano and Murikami, 2001). While there are peak times in emergence, individual taxa
emergence can vary greatly leading to an overall low background emergence throughout the
summer (Petersen et al., 1999). The invertebrate communities captured by our sticky traps were
consistent with this trend (Figure 2.5). Although there are peaks in aquatic emergence and
background level of emergence, the abundance of terrestrial adult aquatic invertebrates’
17
abundance is typically dwarfed by the overall abundance of terrestrial invertebrates. This was
seen in our study when numbers of Diptera per trap were in excess of 1000-2000 individuals and
peaked over 4000 while the maximum combined EPT abundance reached just over 170 per trap.
The only species coming remotely close to not being over shadowed by terrestrials during a
particular season were Diptera taxa (Banks et al., 2007). Romaniszyn et al., (2006) did report that
there was also a large flux of terrestrial invertebrates in early summer compared to spring and
late summer. These seasonal trends have the potential to shift if harvest intensifies, which could
lead to other changes among the riparian ecosystem.
Connections Between Forest Structure and Adult Insect Communities
There are multiple ways that forest harvest can affect aquatic invert communities (Paisker
Chapter 2, 2019). For example, a shift in riparian forest composition will undoubtedly result in a
change in leaf input into nearby streams. A shift from a softwood, needle bearing forest to a
hardwood, broadleaf bearing forest could mean a change in invertebrate assemblages within the
stream (Paisker Chapter 2, 2019). Softwood tree species typically are of lesser quality, and when
entering stream systems, needles are typically not chosen as food sources when the option of
broadleaves are available (Kominoski, 2017). Whiles and Wallace (1997), reported that
invertebrate assemblages change greatly in the presence of varied leaf litter with communities
having higher diversity in the presence of broadleaves over needles. Additionally, riparian
derived CPOM is often more rich in nutrients such as nitrogen and thus are often favored by
invertebrate taxa (Yoshimura, 2012). An opening of canopy cover in headwater streams can also
alter primary productivity occurring within the aquatic habitat.
A reduction in canopy cover allows increased light penetration and warming of stream
water (Conrad and Hatten, 1995, Culler et al., 2018) which can lead to an increase in periphyton
18
and algal growth (Pierce et al., 1993; Banks et., al 2007). Subsequent increases in high algal
cover can support higher abundances of scrapers (Yoshimura, 2012) which are present within the
riparian forest after emergence.
Changes in temperature in more open canopy streams can also have potentially
detrimental effects in aquatic invertebrates and alter their phenology. For example, Zwieniecki
and Newton (1999) recorded temperature changes of up to four degrees Celsius in surface
temperatures over the course of one day when comparing harvested to unharvested streams. This
increase in temperature can also make living conditions more hostile for in-stream invertebrates
leading to decreased in-stream and emerging insect abundance (Hoover et al., 2007). However, if
conditions do not exceed thermal thresholds, a warming stream can also shift the phenology of
species emergence (Greig et al., 2012). An increase in temperature can accelerate the
development of aquatic invertebrates, shortening larval life history duration and advancing
emergence (i.e. Chironomidae; Maier et al., 1990, Plante and Downing, 1989). These factors
could explain differences in the seasonal patterns in aerial invertebrates among different riparian
harvest treatments. Conversely, other aquatic taxa can be sensitive to heat in different life stages:
heated water can lead to compromised eggs, decreased hatch rate of invertebrates, decreased
metabolic efficiency of immature stages and decreased emergence of aquatic invertebrates
(Yoshimura, 2012) as well as increased susceptibility to toxic metals (Patnode and Schrank,
1997). If stream temperatures remain conducive to invertebrate survival and emergence, the
surrounding riparian zone can play a big factor in dispersal.
Several aspects of this study could use further exploration. For example, additional
harvest that has occurred within stream watersheds since the initial study design and
experimental harvest could limit inferences linked to the initial riparian harvest practices.
19
Additionally, patterns may have been stronger if higher resolution of taxonomic classification of
invertebrates found on sticky traps. In our processing, invertebrates were identified to the order
level, which could lead to us missing finer scale compositional change. Also some orders (e.g.,
Diptera) contained both aquatic and terrestrial invertebrate taxa so we do not know whether
forest or stream processes are driving factors to their differences.
Conclusion
Our results show a difference in forest communities along the gradient of different
riparian harvest types. Our results support also previous literature showing a change in forest
communities caused by natural events or anthropogenic disturbances, such as forest harvest, can
result in a change of both aquatic and terrestrial invertebrate communities (Wikars and
Schimmel, 2001 Stone and Wallace, 2002, York, 1999). Changes in forest composition can be
linked to changes in overall invertebrate presence because it serves as forage and habitat for both
aquatic and terrestrial invertebrates. Our results also showed a change in community composition
of aerial invertebrates throughout seasons and treatments. Additional sampling and experiments
should be conducted to further disentangle the effects of forest harvest on emergence of aquatic
invertebrates and presence of terrestrial invertebrates within a harvested riparian forest. When
doing this, it will be important to remember the scale to which data is analyzed. As shown above,
stem count and species alone will not suffice in describing the status of a riparian forest.
Consideration of size and relative abundance is also needed.
20
CHAPTER 2: AQUATIC INVERTEBRATES AND BROOK TROUT RESPONSES TO
ALTERNATIVE RIPARIAN FOREST HARVEST
Introduction
Riparian zones are important because they influence stream temperature, light
availability, dissolved oxygen concentrations, nutrient loads, and overall stream complexity
(Martin and Smith, 1997; Culler et al., 2018; Macdonald et al., 2003; Baldigo et al., 2004).
Forests and streams are both known to receive inputs of energy from adjacent ecosystems, with
the riparian zones between the two, as a hotspot for the transfer of energy (Baxter et al., 2005).
Retaining riparian buffers during forest harvest or land-use changes can help to stabilize soil and
filter toxins from surrounding catchments before they can enter the aquatic system (Keim and
Schoenholtz, 1998). Therefore, riparian zones can be essential to the maintenance of stream
structure and function. However, lack of knowledge on the long-term (decadal) impact of
riparian buffers, especially in the Northeastern United States, prevents harvesters and regulating
agencies from making ecologically-sound management decisions.
Timber harvest can potentially alter the physical habitat of stream ecosystems that are
inhabited by a wide range of invertebrates. Following harvest events, streams often have less
complex stream beds and have larger bank-full widths while also becoming more channelized in
bank structure (Davies et al., 2005). Additionally, nutrient fluxes in riparian zones and streams
can be altered (Davies et al., 2005) resulting in changes within stream and riparian the food webs
(Wootton, 2012). Noticeable changes in food webs occur following forest harvest, for example,
when primary producers are released from light or nutrients limitation. This increase in primary
production by algae and periphyton can have positive effects on herbivores (Rosemond et al.,
1993). Forest harvest has the ability to alter invertebrate communities and density within streams
21
as well as those leaving and entering aquatic systems. Changes in stream taxa occur because of
shifts in food sources (Stout et al., 1993) or altered habitats (Ballie et al., 2005; Medhurst et al.,
2010). Often, this will result in the loss of taxa that are sensitive to variables such as low oxygen,
or high temperatures (Stout et al., 2010) which ultimately lowers taxonomic diversity (Hoover et
al., 2007).
The input of leaf subsidies into streams can be a major energy source for inhabitants like
shredders and grazers. Immediately following forest harvest, sensitive taxa such as mayflies,
caddisflies and stoneflies (EPT taxa), decrease due to changes in the environment, specifically
altered temperature, flow rate and increased sedimentation. However, after the early succession
phase of the forest takes place following harvest, invertebrate abundances, particularly those that
are detritivores and shredders, can increase (Ballie et al., 2005). Functional feeding groups, like
shredders, depend heavily upon the input of leaves and other vegetation to grow and complete
their life history (Haggerty et al., 2004). Post-management, forests can undergo succession and
with this, streams can experience a difference in leaf quality. Early succession plants will
typically produce leaves with less lignin and lower carbon to nitrogen ratios that are more
palatable, and therefore, more easily processed by shredders (such as some caddisflies and
stoneflies). This can lead to an increased abundance of shredder taxa that are more resilient to
disturbances (Stout et al., 1990). The composition of feeding groups and species within a stream
can be used to estimate the relative energetic resources available. An increase in shredder taxa
can indicate that there is a large amount of free energy within stream systems (Stout et al., 1993).
High abundance of shredder taxa, post-forest management, can increase taxa by contributing to
the particulate organic matter within the streams (Medhurst et al., 2010).
22
Changes in riparian forest structure are likely to impact food resources for fish, such as
brook trout. Terrestrial invertebrates are an important energy subsidy for fish in small headwater
streams, where they can equal the amount of in-stream invertebrate production (Mason and
MacDonald, 1982). These additions of terrestrial invertebrates can be vital to small, headwater
streams when compared to higher order streams, where they have the potential to constitute
nearly 50% of fish diets in summer months while comprising 10-15% of all drift (Hubert and
Rhodes, 1989, Young et al., 1997).
Aside from altering the origin (terrestrial vs aquatic) and quality of prey species, forest
harvest can affect stream dwelling fish by altering physical habitat. Forest harvest can channelize
stream beds (Davies et al., 2005), add various woody debris (Vaz et al., 2011) and shift in-stream
structure and habitat patches that fish and their prey can use (Parkstrom, 2002). Altered
streambeds can also change the percentage of suitable spawning grounds that are available for
fish in the system (Buffington et al., 2011). Timber harvest reduces overhead vegetation and
cover, increasing light penetration and water temperature, which could decrease the oxygen
availability, especially in lower gradient streams, that fish and invertebrates need to survive
(Martin-Smith, 1998). Depending on stream temperature before harvest and canopy opening, the
subsequent increase in temperature has the potential to exceed thermal maxima thresholds for
brook trout living within (Trumbo et al., 2011).
Due to a myriad of factors determining buffer width, including slope, substrate,
vegetation, and discharge, final riparian buffer size is most often up to the discretion of the
harvester. Understandably, it is in the harvesters’ best interest to harvest the maximum amount of
timber without disrupting the stream. However, making these decisions requires data that
explains and displays the impact of various approaches on stream and riparian ecosystems. While
23
short-term effects of buffers on water quality, sedimentation, erosion, and nutrient loading is
increasingly understood, what remains unknown is the legacy effects of timber harvest on
subsidy exchange between streams and riparian forests. In this study, I investigated long-term
(18-year) impacts of varying riparian buffers on stream invertebrates and fish communities. I
hypothesized: 1a) Invertebrate abundance will decrease and invertebrate composition will change
with an increase in riparian forest harvest, and 2a) Riparian forest harvest will lead to a decrease
in fish. Alternatives to these hypotheses are 1b) there will be an increase in invertebrate
abundances due to a change in forage quantity and quality and 2b) harvest within the riparian
zone will increase fish numbers and condition. Additionally, there is the chance that no change
between riparian buffer widths will produce no change in invertebrates or fish. To test these
hypotheses, I sampled stream invertebrate community composition, and fish presence/absence
and fish condition at sites in Western Maine 18 years after a previous experiment that removed
large blocks of riparian forest from headwater streams.
Methods
This study took place in the Western Maine, a landscape characterized by logging and
rugged mountainous terrain. Within this mountainous terrain are numerous head water streams
draining the landscape. Fifteen of these streams, with known harvest history, were selected to
determine the long-term effects of riparian zone timber harvest conducted in the winter of 2001-
2002 (See Chapter 1 for a description of the study system and harvest experimental design).
Stream Habitat and Invertebrate Sampling
Surber samples were used to collected benthic invertebrates in June 2018 and May 2019.
Samples were taken in cobble-gravel substrate within 10 meters of each site location if the
immediate substrate was not conducive to Surber samples (i.e. large boulders, too low of flow,
24
etc.). Once collected, specimens and debris/sediment were preserved in 75% ethanol. The entire
sample was spread out in an observation tray where invertebrates were separated from the
mixture of organic and inorganic matter and placed into a separate vial in the lab. This separation
process continued for twenty-five minutes after which the entire sample was searched for thirty
seconds to find any additional organisms. If an organism was found within thirty seconds,
another 30 second search was started until no further invertebrates were found. After the
invertebrates were collected from the sampling tray, they were identified to lowest practical
taxonomic level (usually genus) using Peckarsky et al., (1990) for all taxa except for the family
Chironomidae which was not identified further than family.
Leaf bags were used to sample invertebrate communities inhabiting detritus
accumulations. Three leaf bags were placed at sites 75m, 150m, and 225m (5g +- 0.1g Acer
rubrum) and were collected upon 3 successive visits in 2018. Leaf bags were allowed to be
colonized and broken down for varying lengths of time (20-118 days). When collected, bags
were placed into a Ziploc bag which was then put on ice in the field and then frozen in the lab
until processing. Invertebrates and organic matter were separated in the lab and invertebrates
were placed into 75% ethanol. Invertebrate identification followed the same procedure as the
Surber samples.
Fish Monitoring
During the 2019 sampling season, a 40m reach in each stream was sampled for brook
trout (Salvelinus fontinalis) via backpack electrofishing using a Smith-root LR-24 Electro fisher
set to 900 V AC at 35% duty cycle. Reaches were selected to include a series of pools, riffles,
and runs. Stop nets were placed 20 meters up and downstream from the center of the reach to
prevent fish from entering or exiting the stretch. Each stretch was shocked from downstream to
25
upstream three times to effectively deplete all fish from the reach. Only nine of fourteen streams
contained brook trout, of which all but one, contained only brook trout. Fish were anaesthetized
using MS-222 (50mg/L), weighed (g), measured (length in mm) and returned to the stream
below the sample reach.
Figure 2.1
A conceptual diagram of the study system with the locations of each sampling technique; Surber
samples, leaf bags and electrofishing.
Statistical Analysis
Total abundance of invertebrates per Surber and per leaf bag were aggregated to obtain a
total amount of invertebrates per taxa per stream per year as well as invertebrate abundances per
square meter. Thus, our unit of replication was a stream in each sampling year. The abundance
matrix was then Hellinger transformed (using ‘Decostand’ in Vegan) to down weight the effects
of extremely abundant species and meet assumptions of normality in subsequent analyses.
Canonical correspondence analyses (CCA) ordinations were performed on invertebrate
26
communities based on two categories of harvest type. The first, treatment type, focused on
treatment (clear-cuts with buffers of 0m, 11m, 23m, partial harvest and controls) whereas the
second, riparian harvest, combined treatments into those in which the riparian zone was
harvested (0m and partial harvest), and those where the riparian zone was left unharvested (11m,
23m and control). These ordinations were performed separately on Surber sample and leaf bag
data to assess whether patterns differed between communities in inorganic vs detritus substrate.
Permutational multivariate analysis of variance (perMANOVA) tests were then run to assess
statistical significance, using either treatment type or riparian harvest as categorical predictors.
This analysis uses centroids to determine whether treatment polygons drawn around replicate
date differed significantly from each other in a 2D space. Overall stream density
(invertebrates/0.9m2) was calculated by averaging the invertebrates per Surber at the five sites
along each stream and then averaged per harvest type.
Data on brook trout collected via electrofishing was put into the Fulton’s fish condition
model (K= (W/L3) x 100) to determine the fish condition in each stream containing fish. Total
fish biomass and abundance per riparian harvest was calculated by averaging the biomass per
streams from streams with and without riparian harvest.
Linear models were used to determine differences in invertebrate abundances, density
and community composition, as well as, fish biomass, abundance, and condition between
treatment types and riparian harvest. All analyses were run in R version 3.5.2.
27
Results
Leaf Bags
Leaf bags accounted for 2,182 individual invertebrates belonging to 58 genera. CCA on
invertebrate abundance in leaf bags indicate communities differed between treatment type
(p=0.001; Table 2.1, Figure 2.2A). This separation of polygons was non-directional but showed
that there was a large amount of core invertebrates found in most treatment types and polygons
were separated at their maxima by more novel invertebrates (i.e. Brachycentridae, Simuliidae,
Ephydridae, Heptgeniidae, Molannidae and Sericostomatidae). There was also significant a
difference in invertebrate community and abundance between sites with and without riparian
harvest (p=0.001; Figure 2.2B). Streams within each category again shared core taxa however,
streams with riparian harvest were separated by taxa such as Staphylinidae and Dystiscidae
(Coleoptera) while those without riparian harvest were influenced by taxa such as
Ephemerellidae, Brachycentridae and Peltoperlidae (Ephemeroptera, Trichoptera and
Plecoptera).
Surber Samples
Over both summers, our 140 Surber samples collected 14,556 individuals belonging to
102 genera and 23 orders. CCA on invertebrate abundance in Surber samples indicated
separation of communities (Figure 2.2C). Control sites separated from all other streams along the
first axis, and the second axis reflected a transition from 0m and partial harvested streams to
those with 11m and 23m riparian buffers. This effect of treatment type on invertebrate
community was statistically significant (P= 0.032). There was no significant effect of riparian
harvest in permutation tests, however, communities appeared to separate in ordination space
(P=0.35; Figure 2.2D). Although there were a large number of taxa that were found in all
28
treatments, treatments with riparian harvest were influenced heavily by Trichoptera and
Ephemeroptera, such as, Calamoceratidae, Psychomiidae, and Caenidae, Heptageniidae.
Plecoptera family of Perlidae also influenced communities within harvested streams. Those
without riparian harvest were also influenced by different Trichoptera heavily (Brachycentridae,
Hydropsychidae, and Sericostomatidae) but, less so by Ephemeroptera. Instead, we notice a
growing influence of Diptera and Coleoptera taxa; Culicidae, Dixidae, Elmidae and Carabidae.
Averages per stream ranged from 423.5 inverts/0.09m2 in a 23m buffer stream to 3190.8
inverts/0.09m2 in a partial harvest stream. There was no statistically significant difference in
density (p= 0.108) between treatment type however, streams with harvest in the riparian zone
had higher densities of invertebrates than those in unharvested treatments (p= 0.0282). Total
number of invertebrates per Surber per treatment type and riparian harvest were significant
(p=0.0355 and 0.0117 respectively; Figure 2.3).
Fish Data
Mean total biomass of fish collected from streams with riparian zone harvest was 161.1g
compared to fish from streams without riparian harvest; 132.1g. Mean number of fish sampled
from streams with riparian forest harvest was 23.7g, while mean number of fish collected from
streams without riparian forest harvest was 13.8 (Figure 2.4). However, variability among
riparian harvest meant that these differences in fish biomass (p=0.139) and abundance (p=0.222)
were not statistically significant. Using Fulton’s fish condition equation, fish condition (K) was
found for each individual. Results of an ANOVA run on fish condition and riparian harvest (0m
buffers and partial harvests) supported fish with significantly higher condition (p= 0.0155) than
streams without riparian harvest (11m buffer, 23m buffer and control streams; Figure 2.5).
29
Figure 2.2
Cannonocal Correspondence Analysis (CCA) ran on Hellinger-transformed invertebrate
abundances from leaf bags (A, B) and Surber samples (C, D). Panels A and C represent the
comparisons of streams invertebrates related to the five treatment types (0m, 11m, 23m, partial,
and control). On the left (B, D), stream invertebrates were compared based on presence of
riparian harvest; Yes (0m, partial) and No (11m, 23m, control). All CCA’s showed significant
difference among treatment type and riparian harvest (p = < 0.05) except panel D which had a p
value of 0.35.
B
C D
A
30
Figure 2.3
Total log transformed abundances of invertebrates per Surber sample per streams without
riparian harvest (11m, 23m, Control) and with riparian harvest (0m, Partial) in their respective
riparian harvest category. ANOVA analysis resulted in a p=0.0155.
31
Figure 2.4
Total fish biomass collected per stream (A) and total number of fish collected per stream (B) in
the presence and absence of riparian harvest.
32
Figure 2.5
Condition of fish is represented using Fulton’s fish condition of fish collected in Riparian
harvest treatments. Fish condition analyses resulted in streams with riparian harvest showing a
higher statistically significant condition (p= 0.0155).
33
Table 2.1
Statistical outputs of CCA and ANOVA’s of invertebrate and fish data. Response variables were
tested with predictors of treatment type (0m, 11m, 23m, partial, control) and riparian harvest
(with- 0m, partial, without- 11m, 23m, control).
Data source Response
variable Predictor
Statistical
Test dF F P
Leaf Bags
Invertebrate
Community
Composition
Treatment
type CCA 4, 92 1.432 0.001
Invertebrate
Community
Composition
Riparian
Harvest CCA 1, 95 1.452 0.006
Surber
sample
Invertebrate
Community
Composition
Treatment
type CCA 4, 23 1.216 0.032
Invertebrate
Community
Composition
Riparian
Harvest CCA 1, 26 1.044 0.35
Invertebrate
Density
Treatment
type ANOVA 4, 23 2.145 0.108
Invertebrate
Density
Riparian
Harvest ANOVA 1, 26 5.402 0.0282
Total Invertebrates Treatment
type ANOVA 4, 133 2.703 0.0355
Total Invertebrates Riparian
Harvest ANOVA 1, 136 6.294 0.0133
Electrofishing Fish Biomass Riparian
Harvest ANOVA 1, 59 2.210 0.139
Fish Abundance Riparian
Harvest ANOVA 1, 7 1.797 0.222
Fulton's Condition Treatment
type ANOVA 4, 156 4.438 0.002
Fulton's Condition Riparian
Harvest ANOVA 1, 159 5.983 0.0155
Discussion
In an effort to investigate the long-term effects of riparian zone timber harvest on in-
stream invertebrates and fish (brook trout). I studied fourteen headwater streams in the western
mountains of Maine. Results allude to an importance of forest harvest on stream communities.
34
Significant results point to a difference in community composition of in-stream invertebrates,
and an increase in invertebrate abundance and high fish condition with an increase of forest
harvest within the riparian zone. There was a non-significant trend towards higher fish biomass
and abundance in streams with riparian harvest.
Responses of Stream Invertebrates
A hypothesis for increases in invertebrate densities in the presence of increased riparian
forest harvest is increased primary production in the form of algae and periphyton within the
streams (Kiffney et al., 2003, Perrin & Richardson, 1997). An increase in forest harvest typically
results in an increase in annual water runoff containing terrestrial derived nutrients (Ca2+, Mg2+,
Na2+, K+, Cl-, TN and TP) (Lamontagne et al., 2011). (Although the release of these elements
from soils is highest in the year immediately following harvest, higher than normal releases have
been recorded to persist several years after forest harvest occurs.) Sheridan et al., (1998)
recorded an increased level of runoff into clear-cut and selectively harvested streams when
compared to streams with less harvest intensity, following harvest. Additionally, differences in
forest composition will lead to altered runoff and nutrient releases into adjacent streams
(Lowrance and Sheridan, 2005). Along with potential limiting elemental nutrients, light can also
be a major limiting factor to periphyton growth (Allan and Castillo, 2007). This combination of
increased nutrients and light can be a perfect mixture for accelerated periphyton accumulation
within the stream (Kiffney et al., 2003). A limitation of this study is not measuring stream algal
productivity.
Another food source that can influence the density and community composition of
aquatic invertebrates is terrestrial leaf detritus. In the presence of leaf input in streams,
invertebrate taxa (i.e. Diamesinae, Chironomidae, Limnephilus, Trichoptera and Podomosta,
35
Plecoptera) can increase in densities (Melody and Richardson, 2004). We saw similar responses
in our data with invertebrate communities in streams with riparian harvest being distinguished
by specific Trichoptera and Ephemeroptera taxa (Limnephilius, Heptageniidae, respectively)
while streams without riparian harvest showed a wider range of invertebrates driving their
communities, such as, Diptera, Coleoptera, and Trichoptera. However, not all detritus is similarly
palatable; stream invertebrates have been documented to preferentially feed on leaves from
deciduous trees rather than needles from conifers (Hisabae et al., 2009). Regeneration of riparian
zones following forest harvest can seem to favor this preference by showing an increase in
deciduous tree percentage (Piccolo and Wipfli, 2002).
Results support my original alternative hypothesis that there will be an increase in
invertebrates with varying levels of forest harvest. For example, the polygons in Figure 2C, show
clear differences in invertebrate communities across forest harvest treatments and invertebrate
density was higher in harvested streams. Additionally, leaf bag data also showed differences in
community composition across treatment type and between the presence and absence of riparian
zone harvest. One explanation could be the subsidy-stress hypothesis, an increase in invertebrate
or fish abundance can be seen due to an increase in primary production caused by an increase in
a subsidy entering the system (Niyogi et al., 2007). Alternatively, the increase of invertebrates
could possibly also be explained by the intermediate disturbance hypothesis. Streams in this
study were subjected to varying degrees of disturbance when their surrounding forests were
harvested in the early 2000’s. It could be that the response of different invertebrate communities
was a result of how disturbed they were by the forest harvest; those more disturbed being streams
with a 0m buffer and partial harvest. Streams that were harvested more heavily within the
riparian zone could also been more subject to natural disturbances, such as flooding, debris
36
inundation, and wind damage (Tang et al., 1997) due to their lack of buffer when compared to
streams that were less altered, i.e. those with 23m buffers of the controls.
Responses of Stream Fish to Riparian Forest Harvest
Harvest within the riparian zone resulted in higher means of both total biomass and total
abundance of fish. This, along with the increased condition of fish, supports my alternative
hypothesis that increased riparian forest harvest will experience an increase in suitable conditions
for fish. This increase in suitable conditions can be take place through a few different
mechanisms.
All fish collected were brook trout (Salvelinus fontinalis) which are native to the state of
Maine. Being a salmonid, they are a cold-water species with a critical thermal maximum of
around 26-27° C in juveniles (Grande and Andersen, 1991) and optimal temperature for growth
at roughly 16 °C (Hokanson et al., 1973). Model predictions, as well as empirical observations,
made by Brown (1969) showed that small forested streams saw regular daily high temperatures
of around 19.5 °C. However, these temperatures were only experienced for very short periods of
time during the day. Moore et al., (2006) reported temperatures of stream beds exposed to clear-
cuts reaching in excess of 21 °C, with even higher streambed temperatures depending on
substrate. Although forests and streams with high percentages of canopy cover are recorded to
have cooler daytime temperatures (2-4 °C) (Hagan and Whitman, 2000a) and higher nighttime
temperatures (1-2 °C). Temperature changes post riparian harvest were recorded in Wilkerson et
al. (2006) with increases over 4°C in streams experiencing the most intensive treatment type.
Streams with less cover also have the potential to stay within that heightened metabolically-
optimal range described by Nislow and Lowe (2006), especially if those streams were already
within a higher thermal range.
37
Although streams with riparian forest harvest see an increase in average temperature, this
could be beneficial to brook trout in the stream. Nislow and Lowe (2006) proposed the idea that
an increase in stream temperature could drive metabolic rates for brook trout and energetic
forage efficacy and thus increase overall condition of the fish within a stream. Although streams
with riparian harvest may experience an increase in temperature in open areas, stream
temperatures can decrease to pre-harvest temperatures very quickly once they experience
shading again (Pierce et al., 1993). With a series of opened areas from forest harvest and shaded
areas from regeneration or uncut timber, a brook trout could utilize the high structural
complexity of a stream as either feeding grounds or thermal refuge (Schlosser, 1995), however,
increased shading can lead to temperatures that are too cold for organisms to persist (Dyer et al.,
2010)
Hypothetically, with an increase in light penetration to the stream, increased foraging
efficiency and prey, detection should increase (Nislow and Lowe, 2006). This line of thought can
be backed by results concluded by Sweka and Hartman (2011). They found that the distance of
detection, or the distance a prey item was from a trout before it made any movement to obtain
said prey item, decreased as stream turbidity increased. However, once a prey item was spotted,
even in turbid waters, there was no effect on ingestion rates of prey. It can be understood that in
clearer, more illuminated water, foraging could be easier for trout due to increased detection
distance.
Stream side riparian forest harvest can indirectly affect the presence of brook trout aside
from the direct effects listed above by increasing food stocks. Described earlier, riparian forest
harvest can lead to an increase in aquatic invertebrate presence due to increased periphyton
(Hornick et al., 1981). Additionally, Nakano et al., (1999c) found that areas with larger numbers
38
of terrestrial inputs resulted in larger abundances of aquatic invertebrates. This bottom-up effect
can be strengthened in the spring and summer seasons into fall when detrital input from forests is
typically lower (Hornick et al., 1981). A more complex ecosystem can be supported because
highly productive areas can support high levels of primary consumers, like stream invertebrates
(Polis, 1999).
Limitations to our results and conclusions can come from a loss of samples due to storm
event. Multiple leaf bags were lost during large storm events throughout the sampling period;
namely stream 11 which experienced a flood large enough to displace almost all leaf bags from
the study. Additionally, we were unable to find the third control stream used by Manomet
leaving us with fourteen streams total instead of fifteen. Lastly, brook trout were only present in
nine of fourteen streams with uneven replication within treatment type.
Conclusions
My study showed that streams can vary in response to riparian forest harvest, when
considering variables like benthic invertebrate community composition and fish condition.
Surber samples and leaf bags showed these changes in community composition in streams with a
0m or selectively harvested buffer from communities dominated by scrapers and filter-gatherers
(i.e. Calamoceratidae, Heptageniidae) to communities with less intense treatment types (11m,
23m, and control) containing more shredders (i.e. Lepidostomatidae). Streams with 0m and
selectively harvested buffers also contained more brook trout per unit area as well as overall
higher condition fish when compared to streams without riparian harvest. Continued research
and tests isolating specific responses are needed to make stronger conclusions on the effect of
riparian forest harvest on streams in western Maine. My findings show that it will be important
for managers and conservationists to be mindful of which response variables are measured and
39
how they are analyzed when looking to draw conclusions, as a change in response could be
overlooked if comparing abundances versus community between buffer widths and intensity of
harvest.
40
CHAPTER 3: CONCLUSION
In an effort to assess the long-term impacts of riparian timber harvest on stream
invertebrates and fish as well as the terrestrial forest and invertebrates, fourteen streams were
sampled eighteen years after varied levels of harvest intensity. Immature aquatic invertebrates
were sampled using Surber samples and leaf bags while emerged adult aquatic invertebrates and
aerial terrestrial invertebrates were both collected using sticky traps placed within the stream
corridor. Brook trout were sampled using electrofishing equipment and forest inventory plots
were conducted to assess the riparian habitat. Data collected were then compared between
treatment type and riparian harvest. By using these two different comparisons (a gradient of
harvest intensity and presence of riparian harvest), we were able to see different responses to
riparian forest harvests on the variables previously discussed.
The common theme found throughout the data was that little to no difference was found
when looking at total abundances of trees, invertebrates (aquatic or terrestrial), or fish across
either treatment types or riparian harvest. However, differences were found when comparing
community composition of all variables against treatment type and riparian harvest. When
assessing the forest inventory data, a shift from trees that take more time to establish within a
system to trees that regenerate quickly and readily was seen (i.e. red spruce to speckled alder).
With this, a shift from needle to broadleaf trees was recorded (red spruce to red maple) as well as
a shift in size class of trees with those streams with riparian harvest sporting a higher percentage
of smaller stems.
Stream corridors experienced a shift in aerial invertebrates (emerged adult aquatic and
aerial terrestrial invertebrates). This was seen not only across a temporal scale but also across
41
different treatment types. The shift was characterized by a large importance of Ephemeroptera
present in the early season which shifted to Trichoptera in the mid-season and then shifted
towards a larger importance of Diptera taxa in the late season. Total abundances of invertebrates,
however, did not change between treatment types. Streams showed a shift in aquatic invertebrate
community composition along the series of treatment types, but not in sheer abundance of
invertebrates.
Abundance and weight of brook trout showed no correlation to treatment type or riparian
harvest. Brook trout condition also showed no difference among treatment type but, when
compared between riparian harvest, condition of brook trout increased with riparian harvest.
Limitations to this project came in multiple forms. First, unfortunately we were unable to
locate the third control stream from the original Manomet project and thus had an unequal
number of streams per treatment type. Additionally, all streams, even though they were selected
under the same guidelines by Manomet, differed in overall morphology (i.e. depth, width, pitch,
forest community composition). This made comparisons between the streams more difficult due
to slightly different individual baselines. Lastly, extreme weather events were the driving cause
for loss of some samples. Comparisons between streams may have benefitted with more samples
and/or sampling events of each type.
A major takeaway from this project is the importance of the scope to which data is
analyzed. It was shown that abundances alone did not show the full picture of changes between
different treatment types and instead community composition became important. It is also
important to not only consider the presence or absence of riparian harvest along streams.
Treatment type did prove to be important in determining the changes seen within a stream or
42
forest community. Additionally, the response variables that are looked at are important as well.
This is demonstrated while considering the brook trout data where no response was seen in
treatment type but the presence of riparian harvest proved to be important.
Further research in these systems are important moving forward to continue to achieve
full clarity on the importance of riparian forest harvest on streams. Although this study did show
a continued shift in community composition over time based on treatment type, research on
recovery time of ecosystems back to the original state could also be useful to managers and
conservation efforts.
43
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50
APPENDIX
Table A. 1
Total abundances of all invertebrates collected in fourteen study streams over two years via Surber sample.
Order family
Stream 7 8 3 4 5 1 2 6 9 10 12 11 13 14
Treatment Control Control 0m 0m 0m 11m 11m 11m 23m 23m 23m Partial Partial Partial
Arachnid Araneae 1 0 2 1 2 1 1 1 0 0 2 2 1 2
Diptera 4 11 4 4 2 10 8 11 19 15 2 6 6 19
Diptera Brachycera 0 0 2 2 0 1 4 2 0 0 0 1 5 4
Diptera Cyclorrhapha 3 0 0 0 0 0 0 0 0 0 0 0 0 0
Diptera Phoridae 1 0 0 0 0 0 0 0 0 0 0 0 0 0
Diptera Muscidae 0 0 0 0 1 0 0 0 0 0 0 0 0 0
Diptera Syrphidae 0 0 0 0 0 0 0 0 0 0 0 0 0 0
Diptera Scathophagidae 3 1 0 0 0 0 0 0 0 0 0 0 0 0
Diptera Sciomyzidae 1 0 0 0 0 0 0 0 0 0 0 0 0 0
Diptera Ephydridae 3 0 0 3 2 1 0 0 0 1 1 1 0 0
Diptera Orthorrhapha 0 0 0 0 0 0 0 4 2 0 3 0 0 0
Diptera Athericidae 0 1 2 0 0 1 0 3 2 5 7 25 0 1
Diptera Dolichopodidae 0 0 0 0 0 0 0 0 1 0 1 0 0 0
Diptera Epididae 0 0 0 0 0 0 0 10 1 2 2 1 0 1
Diptera Pelecorhynchidae 0 0 0 0 0 0 0 0 0 0 0 0 0 1
Diptera Stratiomyidae 0 0 0 0 0 0 0 0 0 0 1 0 0 0
Diptera Tabanidae 2 3 0 0 37 1 0 3 2 2 0 0 0 0
Diptera Ceratopogonidae 1 4 3 1 0 16 9 0 1 10 3 2 3 10
Diptera Chironomidae 170 93 295 160 1045 297 529 146 417 67 341 134 1489 123
Diptera Culicidae 1 2 0 2 1 3 0 0 1 0 1 0 2 0
Diptera Dixidae 0 1 0 2 0 0 0 0 0 0 0 0 0 0
Diptera Simuliidae 12 3 107 8 100 9 13 17 36 8 68 80 28 116
Diptera Tipulidae 13 17 13 2 0 19 4 7 14 10 15 8 47 9
Diptera Thaumaleidae 3 0 0 16 0 0 0 0 0 0 0 0 0 0
51
Table A.1 Continued
Diptera Tanyderidae 0 0 0 5 1 0 0 0 0 1 0 0 0 0
Trichoptera Beraeidae 0 0 5 0 0 0 1 0 0 0 2 1 3 2
Diptera Brachycentridae 0 0 0 0 0 0 0 1 0 1 0 0 0 0
Trichoptera Calamoceratidae 0 0 0 1 0 0 2 0 0 0 2 0 0 0
Trichoptera Glossosomatidae 2 0 0 0 0 0 1 0 0 0 0 4 1 1
Trichoptera Helicopsychidae 0 0 0 0 0 0 0 0 0 0 0 0 0 0
Trichoptera Hydrophilidae 0 0 0 0 0 0 0 1 0 0 1 0 0 2
Trichoptera Hydropsychidae 5 1 1 18 53 0 2 3 5 0 0 1 0 4
Trichoptera Hydroptilidae 11 1 0 0 4 0 1 1 0 2 0 1 1 0
Trichoptera Lepidostomatidae 4 3 1 20 4 3 13 0 0 1 9 3 69 12
Trichoptera Leptoceridae 0 0 0 0 1 0 0 0 0 0 0 0 0 0
Trichoptera Limnephilidae 5 2 2 4 21 5 2 0 1 3 11 3 7 2
Trichoptera Molannidae 0 0 0 1 0 0 2 0 0 0 0 4 0 0
Trichoptera Odontoceridae 13 0 0 1 2 2 2 0 0 0 0 0 0 0
Trichoptera Philopotamidae 12 1 1 2 1 0 1 16 0 9 3 0 6 2
Trichoptera Phryganeidae 3 0 1 1 0 0 2 0 1 1 0 1 19 0
Trichoptera Polycentropodidae 3 0 9 2 0 0 4 14 0 0 21 3 1 9
Trichoptera Psychomyiidae 0 2 1 1 6 3 2 5 4 5 9 8 0 17
Trichoptera Rhyacophilidae 4 0 0 0 0 0 0 2 1 1 0 1 0 0
Trichoptera Trichoptera 7 3 3 16 23 8 25 1 15 2 13 10 61 7
Trichoptera Uenoidae 1 0 1 0 2 3 0 0 11 3 1 0 0 0
Ephemeroptera 1 8 1 3 1 1 18 0 0 0 4 12 0 3
Ephemeroptera Baetidae 0 2 0 5 69 0 42 0 13 2 2 43 1 7
Ephemeroptera Baetiscidae 9 4 0 0 0 0 0 0 0 0 0 0 0 0
Ephemeroptera Ephemerellidae 30 27 7 15 175 18 94 0 0 4 44 63 57 121
Ephemeroptera Ephemeridae 2 3 0 0 0 0 0 0 0 0 0 0 0 0
Ephemeroptera Heptageniidae 35 283 4 18 46 3 38 37 0 9 68 130 0 88
Ephemeroptera Leptophlebiidae 1 9 0 1 0 0 54 0 0 0 57 34 0 32
Ephemeroptera Polymitarcydiae 0 2 0 0 0 0 0 0 0 0 0 0 0 0
Ephemeroptera Siphlonuridae 17 8 12 1 12 11 45 9 63 6 20 102 5 1
Coleoptera Cantharidae 0 1 0 0 0 0 0 0 0 0 0 0 0 0
52
Table A.1 Continued
Coleoptera Chrysomelidae 0 0 0 0 0 0 0 0 0 0 0 1 0 0
Coleoptera Coleoptera 0 0 0 0 0 0 1 0 1 1 5 0 0 0
Coleoptera Dytiscidae 1 3 2 0 1 2 0 1 0 0 3 1 1 3
Coleoptera Elmidae 5 29 7 18 1 1 207 2 0 48 167 26 10 101
Coleoptera Gyrinidae 0 0 0 0 0 0 0 0 0 0 1 0 0 0
Coleoptera Haliplidae 1 1 0 0 0 0 0 0 0 0 0 0 0 0
Coleoptera lampyridae 0 0 0 1 0 0 0 0 0 0 0 0 1 0
Coleoptera Limnichidae 0 0 0 0 0 0 0 0 0 0 1 0 0 0
Coleoptera Psephenidae 0 1 0 0 0 0 0 0 0 0 0 0 0 0
Coleoptera Ptilodactylidae 0 0 0 0 0 0 0 0 0 1 0 0 0 0
Coleoptera Staphylinidae 0 1 0 0 0 0 0 0 0 0 0 0 0 0
Plecoptera 3 7 3 2 2 0 8 2 0 4 9 12 1 1
Plecoptera Capniidae 3 5 1 6 7 1 0 6 0 7 2 1 2 2
Plecoptera Chloroperlidae 33 27 26 41 21 29 29 41 5 6 190 97 20 109
Plecoptera Leuctridae 79 26 15 86 14 16 51 51 20 32 138 252 27 154
Plecoptera Nemouridae 48 89 38 13 12 15 0 47 18 54 41 16 35 34
Plecoptera Peltoperlidae 9 16 0 0 0 0 0 0 0 0 0 0 0 0
Plecoptera Perlidae 3 10 0 3 6 0 0 15 1 12 2 0 0 0
Plecoptera Perlodidae 4 4 0 8 35 2 0 6 2 0 5 16 1 13
Plecoptera Pteronarcyidae 0 2 0 0 0 0 0 0 0 0 5 0 0 2
Collembola 0 0 1 0 0 1 1 0 1 0 0 0 0 1
Collembola Poduridae 0 6 0 0 2 0 3 0 0 7 0 2 0 1
Collembola Isotomidae 0 0 0 0 1 0 0 1 1 1 6 2 0 2
Hemiptera Gerridae 0 0 0 0 0 0 6 0 0 0 0 0 0 0
Hemiptera 0 0 0 0 0 0 2 0 0 0 0 0 0 0
Copepod 0 0 0 0 0 1 0 0 0 0 0 0 36 0
Decopod 0 0 0 0 0 1 0 0 0 0 0 0 0 0
Odonata Gomphidae 0 0 0 0 0 0 6 0 0 0 1 0 0 0
Trombidiforme Prostigmata 0 0 0 0 0 0 0 0 1 0 1 0 0 0
Lepidoptera 0 0 0 1 0 0 0 0 0 0 0 0 0 0
Oligocheata 1 117 359 97 24 12 12 102 492 64 187 285 60 85
Unknown 1 0 0 0 1 0 0 0 0 0 0 0 1 0
53
TABLE A.2
Total abundance of all trees measured from the fourteen streams.
Tree species
Stream 7 8 3 4 5 1 2 6 9 10 12 11 13 14
Treatment Control Control 0m 0m 0m 11m 11m 11m 23m 23m 23m Partial Partial Partial
A. Beech 0 3 55 0 0 83 34 0 3 58 15 0 88 0
B. Ash 0 0 0 0 0 0 4 0 0 13 0 0 6 36
B. Spruce 0 0 0 0 0 0 3 0 0 0 1 8 2 0
Bt. Aspen 0 0 0 0 0 0 0 2 0 0 0 0 2 0
Cedar 0 1 0 0 0 0 4 0 1 0 0 0 1 0
E. Hemlock 0 0 0 0 0 0 2 0 0 0 0 0 0 1
Elm 0 0 0 0 0 0 0 0 0 0 0 0 18 0
G. Birch 1 0 0 0 2 0 0 0 0 1 0 0 20 0
Pin Cherry 0 0 1 0 0 0 0 0 0 1 1 0 0 63
Q. Aspen 0 0 2 0 6 0 0 1 0 9 0 0 62 10
R. Maple 3 31 52 44 2 5 34 0 31 24 12 3 60 15
R. Spruce 122 89 2 18 4 5 30 0 89 3 49 38 0 11
Sp. Alder 0 0 0 31 0 0 23 6 0 0 0 0 3 0
St. Maple 11 2 43 31 257 32 42 41 2 8 6 2 41 37
Su. Maple 0 0 38 34 35 25 38 83 0 20 12 7 303 174
Sv. Maple 0 6 0 0 0 0 0 0 6 0 0 0 0 0
W. Ash 0 0 13 2 0 5 15 0 0 0 0 0 3 3
W. Birch 43 22 0 11 0 0 1 2 22 3 0 35 0 0
W. Spruce 0 0 0 0 2 0 0 0 0 0 0 0 0 0
Y. Birch 26 55 14 12 22 5 40 27 49 26 36 45 41 41
54
TABLE A.3
Total abundances of all invertebrates sampled via all leaf bags from fourteen streams.
Order Family
Stream 7 8 3 4 5 1 2 6 9 10 12 11 13 14
Treatment Control Control 0m 0m 0m 11m 11m 11m 23m 23m 23m Partial Partial Partial
Arachnidae Aranea 1 1 2 1 4 0 0 2 1 0 0 0 0 0
Diptera ceratopogonidae 0 0 0 0 0 0 0 0 0 0 0 0 1 0
Diptera Chironomidae 77 38 25 8 67 29 84 127 119 38 22 0 12 23
Diptera Culicidae 0 1 1 0 0 2 0 0 0 0 0 0 0 0
Diptera Cyclorrhapha 0 0 0 1 0 0 0 0 0 0 0 0 0 0
Diptera Dixidae 0 1 0 0 0 0 0 0 0 0 0 0 0 0
Diptera Ephydridae 0 0 0 0 0 0 0 0 0 1 0 0 0 0
Diptera Hymenoptera 0 0 0 0 0 0 0 0 1 0 0 0 0 0
Diptera Simuliidae 0 0 0 0 0 0 0 0 0 0 5 0 0 0
Diptera Tipulidae 0 3 0 0 0 0 0 6 0 0 0 0 0 1
Ephemeroptera Baetidae 0 10 0 1 10 0 0 0 6 0 0 0 0 0
Ephemeroptera Beraeidae 0 0 0 2 1 0 0 0 0 0 0 0 2 0
Ephemeroptera Caenidae 0 0 1 0 0 0 0 0 0 0 0 0 0 0
Ephemeroptera Ephemerellidae 21 47 18 13 57 8 10 11 0 15 19 0 45 52
Ephemeroptera Ephemeridae 0 0 0 0 0 0 0 6 0 0 0 0 0 0
Ephemeroptera Heptageniidae 0 0 0 0 1 0 0 0 0 0 0 0 0 0
Ephemeroptera Leptophlebiidae 1 14 0 0 2 0 1 6 0 2 4 0 0 1
Trichoptera Brachycentridae 0 0 0 0 0 0 0 0 0 0 3 0 0 0
Trichoptera Calamoceratidae 0 0 1 0 0 0 0 0 0 0 0 0 0 0
Trichoptera Hydrophilidae 0 0 0 0 0 0 0 2 0 0 0 0 0 0
Trichoptera Hydropsychidae 2 0 0 0 3 0 1 2 0 1 0 0 0 0
Trichoptera Hydroptilidae 0 2 19 0 0 3 0 0 0 1 0 0 1 0
Trichoptera Hyrdopsychidae 0 0 0 0 0 0 0 1 0 0 0 0 0 0
Trichoptera Limnephilidae 0 0 0 4 11 0 1 1 8 2 1 0 3 0
Trichoptera Molannidae 0 0 0 0 8 0 0 0 0 0 0 0 0 0
55
Table A.3 Continued
Trichoptera Lepidostomatidae 49 2 61 7 47 35 8 22 9 13 4 0 12 6
Trichoptera Polycentropodidae 1 2 4 0 2 0 4 0 0 7 1 0 0 3
Trichoptera Psychomyiidae 0 0 0 3 0 0 0 0 0 1 0 0 0 11
Trichoptera Rhyacophilidae 0 0 0 0 5 0 0 0 3 0 0 0 0 0
Trichoptera Sericostomatidae 1 0 0 0 0 0 0 0 0 0 0 0 0 0
Trichoptera Trichoptera 0 0 0 0 1 0 0 0 0 0 0 0 0 0
Trichoptera Uenoidae 0 0 0 0 2 0 0 0 0 0 0 0 0 0
Plecoptera Capniidae 0 1 0 0 9 0 0 0 3 0 1 0 0 0
Plecoptera Chloroperlidae 21 18 0 2 5 0 0 39 1 3 5 0 0 19
Plecoptera Leuctridae 27 15 8 5 37 2 2 24 14 11 14 0 0 6
Plecoptera Nemouridae 0 0 6 1 9 0 1 1 8 2 3 0 0 1
Plecoptera Peltoperlidae 0 1 0 0 0 0 0 0 0 0 0 0 0 0
Plecoptera Perlidae 0 0 0 0 5 0 0 0 0 0 0 0 0 0
Plecoptera Perlodidae 0 14 25 2 36 2 0 1 4 0 1 0 0 2
Plecoptera Pteronarcydiae 0 0 0 0 5 0 0 0 0 0 1 0 0 0
Plecoptera Siphlonuridae 2 6 0 0 0 0 1 0 2 0 0 0 0 0
Coleoptera Carabidae 1 0 0 0 0 0 0 0 0 0 0 0 0 0
Coleoptera Chrysomelidae 0 0 0 0 0 0 4 0 0 0 0 0 0 0
Coleoptera Dytiscidae 1 0 2 1 0 3 1 1 0 0 0 0 0 0
Coleoptera Elmidae 0 1 0 0 0 0 0 1 1 1 9 0 0 0
Coleoptera Staphylinidae 0 1 0 1 0 0 0 0 0 0 0 0 0 0
Collembola Poduridae 0 0 0 7 0 2 0 0 0 0 1 0 0 0
Collembola Isotomidae 0 9 0 10 0 2 0 1 1 3 1 0 0 0
Odonota Gomphidae 0 0 0 0 0 0 0 0 0 1 0 0 0 0
Trombidiforme Prostimata 0 0 0 0 3 0 0 1 2 0 0 0 0 1
Oligacheata Oligachaeta 0 0 0 0 0 0 0 7 0 0 0 0 0 0
Hydrachnidia Oxidae 0 0 0 0 0 0 0 0 0 0 0 0 1 0
Hirundinea Piscicolidae 1 0 0 0 0 0 0 0 0 0 0 0 0 0
56
TABLE A.4
Total abundances of invertebrates of their respective order per stream from both sampling years of sticky traps.
Order
Stream 7 8 3 4 5 1 2 6 9 10 12 11 13 14
Treatment Control Control 0m 0m 0m 11m 11m 11m 23m 23m 23m Partial Partial Partial
Ephemeroptera 5 0 1 5 14 2 25 5 0 2 33 12 0 4
Hemiptera 7 0 12 3 4 3 4 4 2 3 5 2 4 3
Trichoptera 15 1 33 20 7 18 16 6 24 14 34 9 6 3
Arachnid 6 3 4 9 8 7 13 4 12 27 20 12 7 3
Plecoptera 67 11 34 32 34 30 43 21 37 98 107 47 36 82
Brachyocera 159 10 49 206 66 34 65 44 181 267 139 50 32 110
Neotocera 3888 401 2670 6229 1406 2940 4819 5016 3556 9501 5122 2622 2244 4206
Lepidoptera 11 1 5 5 1 2 1 1 3 3 2 0 1 2
Hymenoptera 97 14 51 192 84 102 135 78 53 69 113 72 30 40
Coleoptera 171 13 183 164 126 134 318 93 165 254 149 74 34 125
57
BIOGRAPHY OF AUTHOR
Mitchell Paisker was born in Trenton, New Jersey on June 15, 1995. He was raised in
Pittstown, in Central New Jersey and graduated from North Hunterdon Regional High School in
2009. He attended the University of Maine and graduated in 2017 with a Bachelor of Science
degree in Ecology and Environmental Science with concentrations in Aquatic Ecosystem
Ecology and Forest Ecosystem Ecology accompanied with a Minor in Forest Resource Ecology.
He continued his education at the University of Maine entering the Masters of Science program
later that year. After receiving his degree, Mitch will begin his career as an eager and
enthusiastic ecologist in the state of Alaska. Mitch is a candidate for the Master of Science
Degree in Ecology and Environmental Sciences from the University of Maine in May 2020.