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AQUATIC COMMUNITY MONITORING FOLLOWING THE
EXCLUSION OF CATTLE FROM A SMALL WATERCOURSE IN
EASTERN ONTARIO
Michelle Nunas
Department of Natural Resource Sciences
McGill University, Montreal
August 2010
A thesis submitted to McGill University in partial fulfillment of the requirements
for the degree of Master of Science
© Michelle Nunas
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ABSTRACT
Previous studies on the impacts of cattle on the aquatic environment have
mainly focused on cold water systems with high intensity grazing, and may be of
limited relevance for assessing impacts of cattle grazing on low gradient, low
intensity sites such as those in eastern Ontario. The present study looks at
changes to the aquatic habitat following the removal of cattle from a watercourse.
Biomonitoring was completed at an aquatic restoration site over a four year period
encompassing pre- and post-implementation conditions. The initial results
indicated modest improvements in the habitat and in the benthic
macroinvertebrate and fish communities following exclusion of cattle from the
watercourse. Trends over time suggested an increase in the proportion of
sensitive benthic macroinvertebrates, a decrease in tolerant benthic species and an
increase in fish density. Longer-term monitoring is required to observe changes
to the aquatic communities following the growth of woody riparian vegetation.
.
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RÉSUMÉ
Par la passé, plusieurs scientifiques ont étudié l’effet du bétail sur le
milieu aquatique utilisant des sites où se trouve une haute densité de vaches dans
des endroits où les cours d’eaux ont une forte pente. Puisque nous utilisons des
sites dans l’Est Ontarien, les résultats de ces recherches auront peu de pertinence
en ce qui concerne cette présente étude car la majorité des sites de la région sont
ceux où l’on retrouve peu de vaches et des cours d’eaux ayant une faible pente.
Cette thèse examine les changements du milieu aquatique suivant l’enlèvement
des vaches à proximité du cours d’eau, et ce, depuis les quatre dernières années,
incluant les conditions pré et post implémentation. Les résultats indiquent une
amélioration modeste d’habitat et des communautés de macroinvertébrées
benthiques et de poissons. Les tendances au fil du temps ont suggérées qu’il y eu
une amélioration des proportions de macroinvertébrées benthiques sensible, une
diminution de macroinvertébrées benthiques insensible et un accroissement dans
le nombre de poissons. Plusieurs années seraient nécessaires pour étudier les
effets de la croissance des arbres et arbustes sur les communautés aquatiques.
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ACKNOWLEDGMENTS
This work could not have been completed without the help of a number of
organizations and people. All efforts pertaining to the designing of the restoration
works, obtaining of funds, monitoring of construction and follow-up meetings
with the landowners were performed by the Raisin Region Conservation
Authority and the Stewardship Council of South Stormont, Dundas & Glengarry,
in particular Chris Critoph, Jim Hendry and Normand Genier. I would like to
thank the landowners, Mr. and Mrs. Legroulx who agreed to participate in the best
management practices program and allowed access to the site over the four-year
period. Land access was also granted by their downstream neighbours, Mr. and
Mrs. Chatelaine, over the same time period. I would also like to thank my
supervisor Dr. Mark Curtis and committee member Dr. Nicholas Jones for their
assistance throughout the project and willingness to take on a part-time graduate
student. Both Dr. Curtis and Dr. Jones offered invaluable assistance in the project
design and comments during the data collection and analysis of the report. I
would like to thank Dr. Brian Hickey from the St. Lawrence River Institute of
Environmental Science for his assistance with statistical analysis. Also wish to
offer thanks to my various field assistants during the past four years: Tamara
Hartrick, Shaun St. Pierre and Flaubert Santullo who assisted in data recording
and collection of fish. Shaun St. Pierre also assisted in benthic macroinvertebrate
sorting during the final years. Thank you to Rob Capell with Aqua-Tech who
completed the benthic identifications during all four years. Thank you as well to
my family during while I worked on this thesis. Funding was provided by the
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Great Lakes Sustainability Fund, Environment Canada and the Ontario Ministry
of Natural Resources.
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Table of Contents ABSTRACT ........................................................................................................ ii
RÉSUMÉ ........................................................................................................... iii
ACKNOWLEGMENTS ...................................................................................... v
GENERAL INTRODUCTION ............................................................................ 1
Research Objectives ......................................................................................... 3
Chapter 1 – Introduction and Literature Review ................................................... 5
Potential Impacts of Cattle on Watercourses .................................................... 5
Aquatic Ecosystem Biomonitoring using Benthic Macroinvertebrates.............. 8
Aquatic Ecosystems Biomonitoring using Fish Communities ......................... 11
Chapter 2 - Methods .......................................................................................... 14
Data Collection .............................................................................................. 16
Statistical Analysis ........................................................................................ 19
Chapter 3 - Results ............................................................................................ 21
Channel Morphology and Riparian Habitat Description ................................. 21
Benthic Macroinvertebrates ........................................................................... 25
Fish ...................................................................................................... 27
Chapter 4 – Discussion ...................................................................................... 29
Riparian Habitat and Channel Morphology .................................................... 29
Benthic Macroinvertebrates ........................................................................... 32
Fish ...................................................................................................... 33
CONCLUSION AND SUMMARY ................................................................... 35
REFERENCES .................................................................................................. 38
List of Tables
Table 1 List of fish species recorded as occuring in the beaudette river sub-
watershed ...................................................................................... 44
Table 2 Upstream and downstream coordinates for the six sampling sites. 45
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Table 3 Summary of observed stressor and implemented aquatic
rehabilitation per site ..................................................................... 45
Table 4 Description of the vegetation and rocky material ranking criteria.. . 46
Table 5 Summary of average yearly channel morphology per site. . .......... 47
Table 6 Median (lower and upper confidence interval) of in-stream cover .. 49
Table 7 Median (lower and upper confidence interval) of riparian habitat
cover ............................................................................................. 50
Table 8 Summary of the proportion of benthic invertebrates collected per
year. .............................................................................................. 51
Table 9 Mean percent (±SD) benthic macroinvertebrates ........................... 55
Table 10 Median (lower and upper confidence interval) of benthic
macroinvertebrates ........................................................................ 56
Table 11 Summary of the proportion of fish species captured per year ......... 57
Table 12 Median (lower and upper confidence interval) of fish Results ........ 58
List of Figures
Figure 1 Location of the St. Lawrence River (Cornwall) AOC..................... 59
Figure 2 Location of project area ................................................................. 60
Figure 3 Location of the sampling sites ....................................................... 61
Figure 4 Correspondence analysis results on abundance data for benthic
macroinvertebrates grouped by habitat. ......................................... 62
Figure 5 Correspondence analysis results on abundance data for fish grouped
by habitat ...................................................................................... 63
List of Photographs
Photo 1 View upstream from the downstream end of Site 2; 2005. ............. 64
Photo 2 View upstream from the downstream end of Site 2; 2007. ............. 64
Photo 3 View downstream from upstream on Site 5, 2005 .......................... 65
Photo 4 View downstream from upstream on Site 5, 2008 .......................... 65
CONNECTING STATEMENT ......................................................................... 66
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Appendix A – Comparison of 100-fixed Count and Whole Count Results.......... 67
Introduction ................................................................................................... 67
Methods ...................................................................................................... 68
Macroinvertebrate Sampling .................................................................. 68
Data Analysis ......................................................................................... 69
Results ...................................................................................................... 70
Discussion ..................................................................................................... 71
References ..................................................................................................... 72
List of Tables
Table 1 Upstream and downstream coordinates for the six sampling sites.
Coordinates are in UTMs............................................................... 74
Table 2 Summary of the total benthos percent composition for the 100-count
and whole count. ........................................................................... 74
Table 3 Mean (±SD) EPT values for the 100-count and whole-count.......... 79
Table 4 Median (upper and lower confidence interval) values for the 100-
count and whole-count .................................................................. 79
1
GENERAL INTRODUCTION
The St. Lawrence River Area of Concern (Cornwall AOC) is one of 43
Great Lakes AOCs established by the International Joint Commission (IJC)1
(Whitaker 2004). The Cornwall AOC is located between the Moses-Saunders
power dam in Ontario and the Beauharnois Canal and dams in Quebec and
includes the Canadian tributaries to the St. Lawrence River within these
boundaries. An AOC is an area which is environmentally degraded. Each of the
Great Lakes AOCs has a Remedial Action Plan which outlines the steps and
criteria necessary to restore the area. Following restoration each AOC is then to
be delisted (Environment Canada 2010). There are several criteria that must be
met in order to delist the Cornwall AOC, including that the tributary fish habitat
and communities within the AOC must resemble those outside the AOC
(Whitaker 2004). One method of meeting this criterion is through the
implementation of projects that enhance the aquatic environment. Over the past
several years, the Raisin Region Conservation Authority (RRCA)2 has completed
a number of projects designed to rehabilitate aquatic habitats within the
agricultural areas of the Cornwall AOC. These projects have included fencing
cattle from watercourses, providing alternative water sources, upgrading manure
storage facilities, bank stabilization and planting riparian buffers. While the work
1 IJC is an international organization that was formed to manage waterways located along the
border between Canada and United States. 2 The RRCA is a local government agency whose aim is the protection, enhancement and
restoration of the natural environment. They are responsible for flood control, plan review, habitat
management and enhancement, water quality enhancement and pollution prevention.
2
is anticipated to have a positive benefit to the aquatic environment, there has been
no opportunity to complete a pre- and post- implementation study. This is not an
unusual circumstance. Shields et al. (2007) indicated that fewer than 10% of
habitat restoration works in the United States are being monitored for success.
O’Donnell and Galat (2008) also found that few (only 34%) river enhancement
projects were being monitored. This created an opportunity to study the changes
expected to occur to the aquatic system following the implementation of aquatic
enhancement measures. The chosen site was a beef farm located along tributary
of the Beaudette River. The conservation authority restricted the cattle from the
watercourse, planted riparian buffer and provided an alternative water source.
The study occurred over a four year period encompassing pre- and post-
implementation conditions.
Although there is some published literature on the impacts of cattle on
creeks, this has primarily dealt with cold-water, steep gradient sites dominated by
salmonids (Gary et al. 1983, Long and Medina 2006, Ballard and Krueger 2005a,
Ballard and Krueger 2005b, Scrimgeour and Kendall 2003). It is questionable
that results from such studies would be applicable to eastern Ontario where
aquatic systems typically consist of warm-water fish communities in habitats with
low flow and turbid water. Since the initiation of this thesis project, additional
studies have been published dealing with the success of aquatic habitat
enhancement projects in Ontario and southeastern United States (Yates and Bailey
2007, Shields et al. 2007).
3
The present study also offered an opportunity to look for ways to enhance
small-scale environmental impact assessment (EIA) methodologies for follow-up
monitoring. Due to the varying level and type of impacts associated with small-
scale EIAs there are no established methodological guidelines. A field study
completed for small-scale EIAs is typically restricted by very limited study areas
(frequently consisting of the immediate subject lands), available funds and short
timeframes. Regardless, the responsible authorities are usually expected to
determine the success of the mitigation and/or compensation measures through
follow-up monitoring. Depending on the scope of the monitoring program
success may be measured by either maintaining the status quo or by a
demonstrable improvement.
Research Objectives
The present study assesses the effects of cattle restriction from a warm
watercourse as part of a typical aquatic habitat rehabilitation project in eastern
Ontario. This was accomplished by monitoring the benthic macroinvertebrate and
fish communities and their habitat over a four year period in the headwaters of a
stream located in Glen Roy, Ontario. The field study was representative of small-
scale EIAs and this provided an opportunity to improve monitoring methods for
these.
The hypothesis was that the restriction of cattle from creeks in eastern
Ontario would result in a significant improvement to the aquatic environment. It
was anticipated that there would be an increase in the amount of riparian
4
vegetation, in-stream cover, the diversity and proportion of sensitive benthic
macroinvertebrate species, and in the diversity, density and proportion of sensitive
fish species. The null hypothesis is that such changes would not occur during the
years of monitoring.
The following thesis provides a summary of the field study and discusses
its findings in terms of improvements in the aquatic environment and on the
relevance and applicability of the chosen methods for small-scale EIA monitoring
programs.
5
Chapter 1 – Introduction and Literature Review
Potential Impacts of Cattle on Watercourses
The riparian zone is often described as a highly important area for aquatic
and terrestrial fauna and flora (Clary and Kinney 2002, Hoover et al. 2001,
Moring et al. 1985). It provides important resources for fauna in terms of cover
and food and is typically associated with a high species diversity of flora (Clary
and Kinney 2002, Hoover et al. 2001, Moring et al. 1985). Vegetated riparian
zones are also important for stabilizing banks and erosion control (Fitch and
Adams 1998). Many published studies have indicated that cattle with unrestricted
access to waterbodies tend to spend a disproportionate amount of effort grazing
within the riparian zone as compared to surrounding areas (McIver and McInnis
2007, Saunders and Fausch 2007, Braccia and Voshel 2006). This has been
attributed to their attraction to the more succulent vegetation found along the
water’s edge as well as the presence of a water source for both cooling and
drinking purposes (Saunders and Fausch 2007). Intuitively, one would expect
there to be negative impacts associated with unrestricted cattle access to
waterbodies. The most obvious impacts are the loss of riparian vegetation
through trampling and consumption, followed by the discharge of urine and feces
directly and indirectly into affected aquatic habitat (Argent and Zwier 2009,
Braccia and Voshell 2006, Howard et al. 1983).
6
These impacts of cattle can result in changes to water quality and the
physical conditions of available habitats for aquatic biota. Several studies have
observed increased soil compaction, soil erosion, channel widening, summer
water temperatures and bacteria content (Weigel et al. 2000, Saunders and Fausch
2007, Fitch and Adams 1998), while others have noted a decrease in water quality
and percent available in-stream cover (Clary and Kinney 2002, Barton 1996,
Amour et al. 1994). Since different benthic macroinvertebrate and fish species
vary in tolerance to these changes, their diversity and abundance can also be
affected. For example, Schofield et al. (1990) indicated that benthic
macroinvertebrates and fish respond to low dissolved oxygen concentrations by
avoidance behaviour and Wohl and Carline (1996) found that the density of
benthos decreases when the amount of sediment load increased. Karr (1987)
demonstrated that clearing of vegetation, analogous with its trampling and
consumption by cattle, can shift the fish community towards one with a higher
percentage of omnivores and herbivores and fewer invertivores and piscivores.
Not all published studies have shown that grazing negatively impacts the
aquatic environment (Rinn 1988), and explanations for the variations in the
outcomes may be attributed to other causes. Site characteristics such as soil type,
climate, channel morphology and habitat may play a key role in the sensitivity of
the aquatic environment to impacts stemming from unrestricted cattle (Rinn 1988,
Weigel et al. 2000, Scrimgeour and Kendall 2003). The level of impact could be
dependent on the type of livestock operation; for example the number of animals
per hectare, timing and duration of grazing or type of operation (McIver and
7
McInnis 2007, Rinn 1988, Platts 1982). Furthermore, cumulative impacts
stemming from upstream urban areas, fertilizer application and/or road salts may
also play an important role.
The sensitivity of the site’s characteristics and associated aquatic
communities to habitat degradation is an important consideration for this research
as the published literature primarily dealt with steep, cold-water environments
(Gary et al. 1983, Long and Medina 2006, Ballard and Krueger 2005a, Ballard
and Krueger 2005b, Scrimgeour and Kendall 2003). There are several studies
which look at the impacts to salmonids in the western part of North America
(Long and Medina 2006, Ballard and Krueger 2005, Scrimgeour and Kendall
2003) and many on sites with high cattle grazing intensities (McIver and McInnis
2007, Scrimgeour and Kendall 2003, Weigel et al. 2000). It is questionable that
results from such studies would be applicable to eastern Ontario where aquatic
systems typically consist of warm-water fish communities in areas with low flow,
turbid water and low intensity grazing. Since the initiation of this thesis project,
new studies have been published dealing with the success of aquatic habitat
enhancement projects including Yates et al. (2007) who looked at stream quality
in first and second order basins in Ontario and Shields et al. (2007) who
completed an 11 year monitoring program on bank rehabilitation in warm water
systems in the south eastern United States.
In areas with steep, cold-water systems where Plecoptera (stonefly larvae)
and salmonids form an important part of the aquatic community, evidence of
impacts from sedimentation and trampling of the banks would be more obvious
8
and easily documented. Saunders and Fausch (2007) found that trout production
and recruitment diminished as a result of a decrease in streambank stability and
associated channel morphological changes brought on by overgrazing. This
supported earlier work by Wohl and Carline (1996) linking salmonid recruitment
failure with habitat degradation caused by livestock grazing. However, in areas
with low gradient and warm-water communities already dominated by species
tolerant of turbidity or sedimentation, such as chironomids (midge larvae) and the
fish species central mudminnow (Umbra limi) or creek chub (Semotilus
atromaculatus), it may be more difficult to quantify any change.
Aquatic Ecosystem Biomonitoring using Benthic Macroinvertebrates
Many authors have suggested that benthic macroinvetebrates be used for
biomonitoring studies (Piscart et al. 2006, Kilgour et al. 2004, Linke et al. 1999,
Barbour et al. 1999, Resh and Jackson 1993). The characteristics that favour their
use include ease of sampling, presence in all environments, high diversity and
relatively high longevity (Braccia and Voshel 2006, Piscart et al. 2006, Weigel et
al. 2000, Linke et al. 1999). Benthos are particularly suitable for the present study
as many factors that affect their composition are the same as those that can be
impacted by unrestricted cattle. Nonetheless, there has been an ongoing
discussion around the use of benthic macroinvertebrates in biomonitoring and in
relation to their aggregated spatial distributions, lack of good taxonomic keys and
cost of sample processing (Jones 2008, Bowman and Bailey 1997, Karr 1987).
9
The distribution of riverine benthos is variable in both time and space
(Linke et al 1999) which can cause difficulties in the comparison of sites with
controls or overtime. This can lead to significant differences between two
communities even when careful pre-selection of sites based on abiotic
characteristics has been completed (Braccia and Voshel 2006, Piscart et al. 2006,
Weigel et al. 2000, Linke et al. 1999). Furthermore, the comparison between
samples can also be influenced by upstream factors. Weigel et al. (2000)
indicated that upstream watershed influences can explain between 61% and 98%
of the variance in the benthic macroinvertebrate communities and that the stress
could be detected as far as 300 m downstream. Increasing the number of samples,
the size of the samples, types of habitat sampled and conducting studies during
the same period for multi-year studies are all important considerations.
The identification of riverine benthic macroinvetebrates to species or even
genus level is difficult and time consuming (Karr 1987) due to a lack of
taxonomic keys and agreement among taxonomists (Bowman and Bailey 1997).
Jones (2008) considers that in freshwater systems only 25-50% of the species can
be identified with accuracy and suggests that the need to identify organisms to
species would depend on the question being asked and the type of analysis being
proposed. For example, identification to species would be necessary when
comparing the species richness before and after as there would be a bias in the
results if one completes the comparison using only genus due to the differing
number of species found per genus. On the other hand, identification to the
lowest possible level would not provide additional information nor would it
10
warrant the extra cost if the analysis only involved comparisons of the Family
Biotic Index (FBI). The idea that the level of identification that is suitable
depends on the question being asked was supported by Resh and Jackson (1993)
who found that identification to family was sufficient when the analysis was
restricted to FBI3, Marglef’s Index of diversity, ratio of scrapers to total number
of individuals, and three richness measures (number of EPT4 taxa, number of
families). Moreover, Bowman and Bailey (1997) stated that identification to
species can result in too much “noise” in the results. Some authors have included
a mixed level of identification (Bowman et al. 2006, Braccia and Voshel 2006). It
is obvious that the level of identification required remains unresolved. In the case
of many small-scale environmental impact assessment (EIA) studies, family level
may suffice given the questions being asked.
Because cost can often be the overriding factor in deciding on the level of
detail as well as the size of sample to process (Jones 2008, Hilsenhoff 1988, Lenat
1988, Karr 1987), this has led to the creation of Rapid Bioassessment Protocols
(Courtemanch 1996). RBPs limit costs by targeting a specific habitat (usually
riffle habitats) and using fixed counts in order to minimize the number of samples
and processing time (Piscart et al. 2006). Piscart et al. (2006) pointed out that
sampling one particular habitat and then relating the findings to all habitats may
result in biasing the results if the impact only affects a portion of the habitat or a
3 FBI refers to Family Biotic Index which provides an average tolerance value for benthic
macroinvertebrates to organic pollution on a family basis 4 EPT refers to Ephemeroptera plus Plecoptera plus Trichoptera (the total number of what are
considered sensitive species)
11
particular functional group. Furthermore, there is a large portion of the benthic
macroinvertebrate community found in habitats other than riffles (Lenat 1988).
This issue of the number of organisms that should be counted has also been
debated (Doberstein et al. 2000, Barbour and Gerritsen 1996, Coutemanch 1996).
Other methods such as those used by the Ontario Benthic Biomonitoring Network
(Jones et al. 2005) and the Ecological Monitoring Assessment Network (EMAN)
(Rosenberg et al. 1997), sample from bank edge to bank edge over a fixed length
or for a fixed period of time. This helps minimize the effects of sampling only
one habitat type. One suggestion to help determine what a coarser level of
identification or sub-sampling would mean to the results would be to conduct
more detailed processing on a sub-set of samples. For this present study a
comparison of the results from a 100-count versus a total count was completed
during the first year prior to making the decision to proceed with 100-count
(Appendix A).
Aquatic Ecosystems Biomonitoring using Fish Communities
Fish share many of the same benefits for biomonitoring as benthic
macroinvertebrates because they are easy to sample, found in many environments
and relatively long lived (Kwak and Peterson 2007, Karr 1987). Fish are also
affected by many of the same factors as benthos including stream size, flow rate,
distance from source, substrate, water temperatures, dissolved oxygen, and water
chemistry as well as availability of in-stream cover (Kwak and Peterson 2007).
Karr (1987) suggested that fish be preferred for biomonitoring over benthic
12
macroinvertebrates since identification is simple, there is much information
available on most species’ life history, they occupy variety of trophic levels
including top of the food chain, and fish recruitment can be used for time series
analysis (Karr 1988). Fausch et al. (1990) added that because fish are sensitive to
a variety of stresses they can provide information on the combined effects (i.e.
changes to habitat or to the benthic macroinvertebrate community). They also
added that fish can help explain social impacts (i.e. loss of a commercial or
recreational fishery) to the general public better than benthos.
With only 158 fish species listed as occurring in Ontario (Mandrak and
Crossman, 1992) there is obviously a much smaller number of fish species than
benthic macroinvertebrates. The smaller number of species and the availability of
keys (i.e. Becker 1983, Scott and Crossman 1973) simplify the identification to
species for fish as compared to benthos. There is also a wealth of information
readily available on many species found in Ontario both in print (i.e. Becker 1983,
Scott and Crossman 1973) and online (i.e. Ontario Fish Database by Eakin 2009).
The presence of certain fish species can readily provide useful information on an
aquatic system and help determine potential impacts from stressors. For example,
in Ontario the capture of mottled sculpin (Cottus bairdi) would typically indicate
cool water habitat while slimy sculpin (C. Cognatus) would indicate cold water.
This type of information would make determining thresholds for thermal impacts
quick and cost efficient. While Plecoptera could also be used for assessing the
thermal regime (as an indicator of cool or cold water), Plecoptera tend to require
flow (Kosnicki and Sites 2007) in order to capture their prey and as such would
13
not be appropriate in all habitats. Another example would be the presence of
redside dace (Clinostomus elongates) or cutlip minnow (Exoglossum maxillingua)
which would signify that the area contained very clear water. The absence of
these species may simply signify that other features of the habitat are not
appropriate or their presence may be a result of the fish passing through the area.
This highlights the need to have an experienced ichthyologist involved during
interpretation of the results (Fausch et al. 1990). It also stresses the importance of
gathering physical and chemical variables during monitoring (Fausch et al. 1990).
The use of fish species also has its difficulties, as often additional studies
are required in order to identify the exact cause of the impairment, and
sensitivities may vary in different regions and for different stressors (Fausch et al.
1990). While some researchers have favoured the use as one species as the “key
species” to determine the level of impact for all members of a particular group,
this is not recommended as the various members of the groups often respond
differently (Fausch et al. 1990).
14
Chapter 2 - Methods
The present study is located within the St. Lawrence River (Cornwall)
Area of Concern which includes the river and its tributaries between the Moses-
Saunders power dam in Ontario and the Beauharnois Canal and dams downstream
in Quebec (Figure 1). Potential study sites were located by flying over the region
and identifying those which appeared to have unrestricted cattle access. These
sites were then ground truthed to ascertain which ones allowed unrestricted access
to watercourses and to observe the upstream and downstream influences. The
preferred site was chosen due to the limited upstream access, presence of a long
length of watercourse within the site and the willingness of the land owner to
participate. The preferred site was located at 45°14’15.60” N 74°38’53.52”W in
the village of Glen Roy, Ontario, approximately 8 km to the southeast of the
Town of Alexandria (Figure 2). The farm consists of a small beef cattle operation
with 21 cows plus their calves. Running through the middle of the pastureland
was a branched watercourse, the Glen Roy Drain, a tributary of the Beaudette
River which flows east to enter the St. Lawrence River near the Ontario/Quebec
border (Figures 1 & 2). The Beaudette River sub-watershed covers an area of
15,421 ha and contains 8% wetland cover, including a locally significant wetland
located on the Glen Roy Drain upstream of the project area (RRCA 2003). There
are 24 fish species documented as occurring within the Beaudette River sub-
watershed (Table 1). Both the Beaudette River and the Glen Roy Drain are low-
gradient warm-water systems. The total length of the drain is approximately
15
11.7 km. The beef farm is located along 1.1 km of the drain beginning 5.0 km
upstream of the confluence with the Beaudette River.
Previous fish community sampling on the Glen Roy Drain was completed
by the Raisin Region Conservation Authority (RRCA) in 2000 and 2001 when
they captured a total of eleven fish species: central mudminnow (Umbra limi),
white sucker (Catostomus commersoni), golden shiner (Notemigonus
crysoleucas), common shiner (Notropis cornutus), bluntnose minnow
(Pimephales notatus), creek chub (Semotilus atromaculatus), rock bass
(Ambloplites rupestris), Iowa darter (Etheostoma exile), fantail darter (E.
flabellare), johnny darter (E. nigrum) and logperch (Percina caproides) (RRCA
2000, 2001).
Six sampling sites were established for this study, five of which were
located on the farm property and one downstream (Figure 2, Table 2). The sites
are labelled 1 through 6, from downstream to upstream. Site 1, located offsite,
was considered a downstream control. Sites 2, 3 and 4 were located on the main
branch of the Glen Roy Drain within the farm property. Sites 5 and 6 were
located on the side branch within the farm property. Degraded sites with obvious
cattle use such as cattle crossings, trampled banks and little shrub cover were
targeted, thus sites 2 to 6 were located along the most degraded sections of the
drain. No upstream control site was available as the aquatic habitat alters
immediately upstream of the farm property to a wetland with a substrate of
unconsolidated organic matter. Sampling within that would have necessitated
different techniques and in any case would not have provided a comparable
16
species composition. Each site began and ended at a cross-over5 and was
approximately 7x the channel width. The upstream and downstream ends of the
sites were permanently marked, photographed and their positions recorded by
GPS. The same sites were sampled over a four year period. The first year
sampling was completed prior to the implementation of the aquatic restoration
works (2005), the second year during the implementation (2006) and the two final
years post-implementation (2007-08). With the exception of the fencing, which
had to be re-designed and re-installed, all of the enhancements were successfully
implemented during the fall of 2005. The fencing was fully functional by the end
of July 2006. The aquatic restoration projects for the study area included a low-
level crossing, replacement of a culvert, fencing of the cattle from the
watercourse, providing an alternative water source, protection of banks with
boulders and riparian plantings. The types of cattle impacts observed and
rehabilitation implemented were recorded for each site (Table 3).
Data Collection
Information on channel morphology was always collected in July. The
Ontario Stream Assessment Protocol (OSAP) point observation technique
(Stanfield 2003) was used to describe the channel morphology. Ten evenly
spaced transects, each with six evenly spaced observation points were established
at each site. The channel width and active (wetted) width were recorded at each
5 Cross-over is the location where the deepest part of the watercourse is located in the middle of
the channel. In natural watercourses cross-over is easily distinguished by the presence of riffles.
17
transect. Depth (cm), substrate size (cm), and in-stream cover were recorded at
each observation point. Within the project area, in-stream cover consisted of
aquatic vegetation, overhanging vegetation6, undercut banks, algae or rocky
substrate as cover.
The riparian habitat was described based on the extent of the riparian
vegetation and rocky material cover as per American Fisheries Society (AFS)
protocols (Stevenson and Mills. 1999). The riparian habitat descriptions occurred
during July in base flow conditions. A measuring tape was used to create a
transect which was placed parallel to the water flow, between the water’s edge
and the floodplain7. Each transect was divided into thirty segments and the
percent vegetation and rocky cover were recorded along the transect. The riparian
data was ranked based on the AFS protocol with slight modifications made to the
vegetation section. The protocol assigned four rankings for vegetation cover. A
rank of 1 signified that less than 50% of the segment was covered by vegetation, a
rank of 4 signified that the vegetation cover was greater than 90% (Stevenson and
Mills 1999). This ranking system did not allow any differentiation for sites that
contained woody vegetation and since riparian plantings formed part of the
rehabilitation measures an additional four rankings were added (Table 4).
“Protocols for Measuring Biodiversity: Benthic Macroinvertebrates in
Fresh Waters” (Rosenberg et al. 1997) was followed for the benthic community
sampling. Benthic sampling was conducted every October. The site was sampled
6 Overhanging vegetation includes herbaceous and woody vegetation which are overhanging the
watercourse within a vertical distance of 1 m from the water’s surface. Does not include woody
vegetation that is providing canopy cover. 7 The floodplain was considered to be the land located adjacent to the stream and inundated less
than annually
18
by walking perpendicular to the flow from one bank to the other for a period of 2-
minutes. The substrate was disturbed using the travelling kick methodology for
benthic macroinvertebrate sampling and a dip net held downstream collected the
dislodged organisms. All habitats encountered were sampled. The site was
divided into 3 equal lengths and three 2-minute samples were collected from each
site. The sampling began at the downstream end of the downstream site and
continued upstream. The invertebrates were preserved with 70% ethanol alcohol.
A sub-sample of a minimum of 100 individuals was identified to family.
The fish community was sampled during June using a standard backpack
electrofisher in keeping with the OSAP methods for multi-pass sampling
(Stanfield 2003). The beginning and the end of the site were blocked using seine
nets; 5 m long by 1.8 m high seine net with a 0.2 cm mesh (stretched). Three
passes were made at each site. Effort was kept at 5-10 s/m2. The stunned fish
were removed immediately from the water and placed in buckets. Once a pass
was completed and the fish processed, they were transferred into large holding
bins. The holding bins were covered to provide shade and were monitored
regularly with new water added to the bins after every pass. Once all three passes
were completed the fish were released. All fish were identified to species and
fork lengths recorded.
19
Statistical Analysis
The majority of the data was analysed using non-parametric tests as
recommended by several authors (Rosenberg and Resh 1993, Brown and Guy
2007) given the low number of samples and/or the use of rank data or indices.
The only exception was benthic macroinvertebrate percent and count data which
was analysed using parametric methods, as it is generally accepted that parametric
tests can be used in this case after applying the appropriate transformations (i.e.
arcsine) (Norris and George 1993). Results were compared on a yearly basis
using Friedman tests, two-way ANOVA, correspondence analysis and/or
Pearson’s correlation. The hypotheses were that the habitat and benthic
macroinvertebrates and fish communities would be the same during all years of
monitoring. The null hypotheses would be rejected for any items that had an
alpha p value < 0.05. Except where otherwise indicated, all statistical analysis
was completed using Minitab15.
For the nonparametric Friedman tests, the data was grouped by year and
by site, with year as the treatment and site as the block. The channel morphology
data was evaluated based on a yearly comparison of the different in-stream cover
types: aquatic vegetation, overhanging vegetation, undercut banks and rocky
material as cover. The riparian habitat data included both the ranked vegetation
and rocky material results. The benthic macroinvertebrate data tested with
Friedman tests were the Modified Family Biotic Index (MFBI) (Mandaville 2002)
and the Shannon-Wiener Index of Diversity. The fish community data analysis
20
consisted of the number of species, fish density, Shannon-Wiener Index of
diversity and trophic guilds (percent omnivores, insectivorous cyprinids and
herbivores). Note that while a comparison of the percent carnivores was
considered, they were only captured in low numbers throughout the sampling
period and as such were not included in the analysis.
Two-way ANOVA was used for the percent benthic macroinvertebrate
data (percent sensitive families (EPT), percent tolerant families (chironomids) and
the count data (number of families). The percentage data was transformed using
arcsine square root and the count data following log10 (x+1) (Norris and George
1993, Guy and Brown 2007). The ANOVA data was grouped by year and site.
A comparison of the benthic macroinvertebrate and fish species
compositions using abundance (by year and by site) was then made using
correspondence analysis (CA) from the Biplot add-in for Excel (Lipkovich and
Smith 2002).
Trends over time were analysed using Pearson’s correlation (Pearson
product-moment correlation). Nonparametric data was ranked first.
21
Chapter 3 - Results
Channel Morphology and Riparian Habitat Description
The channel morphological features were summarized on a yearly basis
for each site (Table 5). Sites 1, 2 and 6 were more characteristic of natural or
naturalized sites with pool and riffle/run sequences and variable depths and
substrate types. Sites 3 to 5 were typical channelized sites with pool habitat,
primarily fines as substrate and even depths and bank heights.
Site 1
Site 1 was located immediately downstream of the farm property within a
rural residential area that was forested on both banks. The channel was confined
and the banks showed signs of erosion. The median channel width was 7.37 m
(range 6.89-7.52 m) and active width was 4.01 m (range 3.79-4.79 m). The
median depth was 11 cm (range 9-12 cm). The median substrate size was 3 cm
(range 0.7-6.9 cm). The average percent in-stream cover was 85% (range 70-
93%) and was composed primarily of rock (median 80.0%; range68-95%).
22
Site 2
Prior to the implementation of the aquatic habitat restoration measures,
this site was accessible by the cattle year-round. The channel was confined and
the banks tall and steep with signs of erosion and trampling. The vegetation was
grazed all the way to the edge of the channel. Large boulders had been placed
along the banks at the downstream end by the landowner, many years prior to the
commencement of this study. The median channel width was 4.57 m (range 4.30-
4.85 m) and active width was 3.30 m (range 2.96-3.50 m). The median depth was
15 cm (range 12-18 cm). The median substrate size was 6.5 cm (range 5.1-
9.0 cm). The average percent in-stream cover was 90% (range 78-97%) and was
composed primarily of rock (median 79.0%; range 70-92%).
Site 3
The banks were low, hummocky, undercut and failing. The median
channel width was 4.75 m (range 4.08-5.10 m) and active width was 4.34 m
(range 4.26-4.79 m). The median depth was 68 cm (range 65-69 cm). The
median substrate size was 1.9 cm (range 0.7 cm-2.9 cm). The average percent in-
stream cover was 55% (range 20-60%) and was composed primarily of
overhanging vegetation (median 33%; range 7-43%) and undercut banks (median
26%; range 12-28%).
23
Site 4
Site 4 was established at the upstream end of the project area, downstream
of a cattle crossing site. Following the implementation of the aquatic restoration
activities, the upstream cattle crossing was altered into a confined low-level
crossing. The banks were low. Some portions of the banks were hummocky
while others were grazed to the bank edge. The median channel width was
4.38 m (range 4.25-4.47 m) and active width was 4.08 m (range 3.98-4.15 m).
The median depth was 54 cm (range 49-57 cm). The median substrate size was
1.0 cm (range 0.3-1.5 cm). The average percent in-stream cover was59% (range
43-78%) and was composed primarily of overhanging vegetation (median 37%;
range 27-43%) and aquatic vegetation (median 27%; range 13-47%).
Site 5
Site 5 was on the side branch, downstream of a cattle crossing site. The
banks were low. Some portions of the banks were hummocky others were grazed
to the bank edge. Boulders were placed along the right bank during the
implementation of the aquatic restoration works. The culvert located immediately
downstream of this site was replaced by a larger culvert and fencing was
continued along both sides to allow cattle to pass over the culvert but preventing
access to the creek. The median channel width was 2.80 m (range 2.62-3.16 m)
and active width was 1.80 m (range 1.63-2.13 m). The median depth was 26 cm
(range 21-30 cm). The median substrate size was 0.1 cm (range 0.1-2.0 cm). The
average percent in-stream cover was 50% (range 48-62%) and was composed
24
primarily of aquatic vegetation (median 28%; range 22-43%) and overhanging
vegetation (median 24%; range 8-45%).
Site 6
Site 6 was also located on the side branch. A cattle crossing site was
located within this site. The banks were steep. Some portions of the banks were
hummocky, while others were grazed to the bank edge. Boulders were placed
along the right bank during the implementation of the rehabilitation works. The
median channel width was 2.27 m (range 2.03-2.79 m) and active width was
1.43 m (range 1.28-1.55 m). The median depth was 8 cm (range 7-11 cm). The
median substrate size was 4.3 cm (range 1.9-6.1 cm). The average percent in-
stream cover was 72% (range 62-80%) and was composed primarily of rock
(median 48%; range 33-52%) and aquatic vegetation (median 15%; range 8-30%).
There was a significant difference found in the channel morphology data
for the amount of overhanging vegetation and rocky material as in-stream cover
on a yearly basis (Friedman test: p=0.029 and p= 0.034, respectively. n=6), due to
an increase in the amount of overhanging vegetation between 2005 and 2007 and
a slight decrease in 2008 (Table 6). Despite the decrease between 2007 and 2008
results, overhanging vegetation cover remained higher following the
implementation of the rehabilitation works. The amount of rocky material as a
component of in-stream cover increased slowly from 2005 to 2008 (Table 6).
25
Changes in all other in-stream cover types (undercut banks, aquatic vegetation
and algae) were found to be insignificant (Table 6).
The riparian habitat data supported the channel morphology results in that
there was also a significant difference in the amount of riparian vegetation over
the years of the study (Friedman test: p=0.031, n=6). Riparian vegetation
increased in 2005 to 2007 and remained the same in 2008 (Table 7). Note that the
riparian ranking reached a rank of 4 in 2007 and remained at the same level in
2008. The rank would not be able to improve above a 4 until sufficient time
elapsed to allow the planted woody vegetation to grow large enough to be
included in the percent cover. There was no significant change in rocky material
coverage on the banks.
None of the Pearson’s Correlations were significantly correlated with time
(Tables 6 & 7).
Benthic Macroinvertebrates
A total of 70 families recorded over the four year study period (Table 8).
All samples combined, the dominant families were Chironimidae (30%), Elmidae
(20%) and Coroxidae (8%) in 2005; Caenidae, Chironimidae and Elmidae (20%,
each) in 2006; Elmidae (25%), Chironomidae (24%) and Hydrosychidae (9%) in
2007; and Elmidae (25%), Caenidae (13%) and Hyalillidae and Chironomidae
(12% each) in 2008.
The number of families decreased over the study period from a median of
16 in 2005 to 13 in 2008. The percent EPT increased from a mean of 18% in
26
2005 to 36% in 2008 while the percent chironomids decreased from 31% in 2005
to 13% in 2008. The median values for the Shannon-Wiener diversity index were
highest in 2006 and lowest in 2007 (1.92 and 1.82, respectively). The median
MFBI values decreased from 6.05 in 2005 to 5.48 in 2008.
Over this time there was a significant difference found for the proportion
of sensitive macroinvertebrate species (EPT) (ANOVA: F=20.95, p<0.0001,
DF=3, n=18) and of tolerant species (chironomids) (ANOVA; F=12.14,
p<0.0001, DF 3, n=18). The proportion of EPT varied throughout the four years
with 2005 having the least and 2006 the most (Table 9). The proportion of
chironomids decreased slowly between 2005 and 2008 (Table 9). No other metric
was significantly different (Table 10).
The correspondence analysis demonstrated that there were no clear trends
between the downstream control (Site 1) and the impacted sites regardless of the
year (Figure 4). The control site (Site 1) and impacted Sites 2 and 6 were all
located in the same section of the plot as the impacted Sites 3, 4 and 5 (Figure 4).
Pearson’s correlation indicated that changes in both the percent sensitive
species (EPT) and the tolerant species (chironomids) were significantly correlated
with time. The percent sensitive species had a positive correlation (p=0.012,
r=0.295) and the percent tolerant species a negative correlation (p<0.0001, r=-
0.416) (Tables 9 & 10).
27
Fish
Twenty-one species of fish were captured over the four year study (Table
11). All samples combined, the same three species were dominant during 2005,
2006 and 2008. These species were fantail darter (78% in 2005, 54% in 2006,
46% in 2008), creek chub (4% in 2005, 12% in 2006 and 13% in 2008) and
johnny darter (3% in 2005, 7% in 2006, 9% in 2008). In 2007 the dominant
species were fantail darter (45%), creek chub (14%) and bluntnose minnow and
central mudminnow (7% each).
There was a significant difference over time in the number of fish species
(Friedman test: p=0.044, n=6), the density of fish (Friedman test: p=0.004, n=6)
and in the Shannon-Wiener index of diversity (Friedman test: p=0.014, n=6). The
number of species increased yearly, with a slight decrease in 2008 (Table 8).
There was also a steady increase in the density of fish captured (Table 12). The
Shannon-Wiener values increased yearly, with a slight decrease in 2008, all
though its values were still above those of 2005 (Table 12).
The correspondence analysis results for the fish followed the same trends
as those for the benthos. The control site (Site 1) and impacted Sites 2 and 6 were
always similar regardless of the year, as were impacted Sites 3, 4 and 5 (Figure 5).
Three of the metrics were found to be significant using Pearson’s
correlation. Changes in the density of fish (p=0.020, r=0.472) and percent
herbivores (p=0.019, r=0.474) were both found to be positively significant with
time (Table 12).
28
The comparison of presence/absence data (Table 11) over the years did
not demonstrate any important shifts in fish species composition. One exception
was the presence of the Iowa darter in 2008. This species was absent during the
first three years and was captured in very low numbers in 2008. Iowa darter is
considered to be one of the least adaptable and most intolerant of the darters (Karr
1987, Fausch et al. 1990). This is not the first recording of the species for the
Glen Roy Drain, as Iowa darters had been previously captured by the RRCA in
2001 near Site 2 (RRCA 2001).
29
Chapter 4 – Discussion
Riparian Habitat and Channel Morphology
The hypotheses of this study were that the implementation of the aquatic
restoration would result in a significant increase in the availability of in-stream
cover, an increase in riparian vegetation and a decrease in exposed rocky material
along the banks. The results identified a significant difference between years in
the amount of overhanging and riparian vegetation. No other in-stream cover or
riparian bank characteristics were found to significantly differ over the study
period. There was also no significant correlation between overhanging or riparian
vegetation and time. There are several explanations for the lack of a trend over
time including the possibility that the Friedman tests identified yearly variations
as opposed to a significant improvement following the implementation of the
restoration works. As such there would be no improvement over time. I believe
that there was a change in the amount of vegetation, both in terms of overhanging
and riparian. During the pre-implementation period there was a large amount of
herbaceous cover throughout the area, by it was simply cropped close to the
ground by cattle grazing. Viewing the photographic records one can clearly see
an increase in herbaceous vegetation height between 2005 and 2007 (Photos 1 &
2). The height of the vegetation was not measured. Thus, while the data recorded
may have only recorded the yearly variation in the percent of cover, there may
have been a real increase in the herbaceous vegetation height following the
30
implementation of the restoration work. This would agree with the results from
Hoover et al (2001) who indicated that they found a significant increase in
vegetation height and a decrease in bare ground within 2 years of cattle exclusion.
Furthermore, the rank of 4 was obtained during 2007 and maintained in 2008. In
order to increase above a rank of 4 the woody vegetation would need to grow
(Table 4). The plantings were all small seedlings and would require several years
to grow to chest height. As such the short-term nature of this study was not
expected to evaluate the success of the riparian plantings. Provided that the cattle
are restricted from the riparian area, it is reasonable to assume that some of the
seedlings will grow to tree height, at which time, the riparian vegetation ranking
would increase.
The results from the overhanging vegetation analysis over time may have
been influenced by the presence of overhanging vegetation at Sites 3 and 4 in
2005. These sites had highly trampled, hummocky and undercut banks. This
created a lack of stability which likely restricted cattle from accessing the
watercourse in these locations even prior to fencing. From the photographic
record, there was an increase in the overhanging vegetation at the three remaining
impacted sites (Sites 2, 5 and 6, Photos 3 & 4). With a p-value of 0.06 for
Pearson’s correlation, the results for overhanging vegetation were nearly
statistically correlated with time.
The percent rocky material along the banks was expected to decrease over
time. This was difficult to assess as there was little exposed rocky material at
Sites 3 or 4 even at the beginning of the project. Furthermore, rocky material was
31
added to the banks at Sites 5 and 6 as a bank stabilizing measure, during the
implementation of the enhancement measures.
Overall changes in the channel morphology and riparian habitat would
appear to need more time to or more aggressive and costly restoration efforts
become statistically significant. This is supported by others such as Shields et al.
(2007) who followed a site over a ten year period. They had protected the banks
and during the follow up monitoring were able to document that there was an
increase in water depth. The protection of the banks likely forced the water’s
energy down and allowed the channel morphology to change. Once the riparian
plantings on the beef farm mature, they will provide more stability to the banks
and may help re-direct the water’s energy thereby allowing the channel
morphology to change here as well. It should also be noted that the Glen Roy
Drain within the study area is located within a low gradient area, downstream of a
wetland. These factors reduce the water velocity and flow and its ability to carve
the channel.
With the exception of some stabilization of the banks at Sites 5 and 6, the
aquatic habitat restoration measures implemented did not intend to modify the
exposed rocky material along the banks or the severely damaged, hummocky
banks over the short term. It seems that the aquatic habitat enhancement
measures implemented have begun to demonstrate a net improvement over the
pre-existing conditions in terms of the height of vegetation and overhanging
vegetation. Over the next 10 years it would be expected that the riparian
plantings will continue to mature and provide stability to the banks. This may
32
allow for the channel morphology to change, although it would be likely to take a
long time, in the order of 20 years or more, for any significant differences to the
channel shape to occur.
Benthic Macroinvertebrates
It was hypothesized that the benthos would show a significant
improvement in the diversity and/or presence of the sensitive benthic
macroinvertebrates following the implementation of the aquatic habitat
enhancement works. The noticeable increase in the percent of sensitive species
(EPT) and a decrease in the percent tolerant species (chironomids) over the four
year period, confirmed by both the Friedman tests and Pearson’s correlation,
indicated that that there was a significant change following the implementation of
the restoration activities. This agrees with the findings of Weigel et al (2000)
who found that there were negative impacts to benthic macroinvertebrates in a
grazed area as compared to nearby wooded areas. Weigel et al. (2000) also found
that there was a lower percent of EPT in the grazed sections. McIver and McInnis
(2007) similarly reported that percent EPT was lower in grazed than ungrazed
sites.
It was intended that Site 1 be used as a downstream control site with a
different benthic invertebrate community from the impacted sites during the first
year. Following the implementation of the restoration activities it was expected
that the communities within the impacted sites would begin to resemble Site 1.
However, this was not the case. The CA results suggest that the only differences
33
between Site 1 and the impacted sites were related to the type of available habitats
and these differences did not change over the course of the study period. The
three sites that contained pool and riffle sequences (Sites 1, 2 and 6) contained
similar community structure as did the three sites which were pool habitat (Sites
3, 4 and 5). This indicated that Site 1 could not be used as an appropriate control
and, although the statistical analysis of the time series at the sites which received
habitat restoration measures suggested that the hypotheses were true, data from
the control site could not confirm the improvements were a result of the
restorations.
Fish
The protection of the aquatic habitat caused by the exclusion of cattle was
expected to improve the fish community in terms of the quantity, diversity and/or
presence of sensitive fish species. The fish data supported this when compared on
a yearly basis using Friedman tests, and indicated that we could reject the null
hypotheses. The CA results for the fish data were similar to those of the benthos
data (control site being similar to impacted Sites 2 and 6 and impacted Sites 3, 4
and 5 being similar to each other, during all four years). Results from the
Pearson’s tests suggested that the density of fish increased yearly, implying that
there may have been a change in the fish productivity following the
implementation of the aquatic habitat enhancements. This agreed with findings
from Saunders and Fausch (2007), who also demonstrated an increase in the fish
density within 5 years of cattle removal. Saunders and Fausch (2007) work was
34
conducted on a cold-water system and they attributed the increase in salmonids to
an increase in the numbers of terrestrial invertebrates available from the
overhanging vegetation. They did not see a similar increase in the other fish
species however, those consisted primarily of benthic feeders (longnose dace
(Rhinichthys cataractae), white sucker (Catostomus commersoni) and longnose
sucker (C. Catostomus). This leads to an interesting hypothesis which may apply
to the present study. With a predominance of omnivorous species, it is possible
that the increased riparian growth may have resulted in an increase in terrestrial
input and that this additional food source created an increase in fish productivity.
Yet, there was no increase in the abundance of sensitive fish species except for
the appearance of the Iowa darter.
Based on the discussion in Karr (1987) it was expected to see an increase
in insectivorous cyprinids and piscivores and a decrease in omnivores and
herbivores. The only trophic guild with significant change over time was
herbivores, which increased. It may be that the fish community requires
additional time to adjust to the new habitat during which sensitive species would
find their way to the site, reproduce and grow large enough to be captured with
the backpack electrofisher. As in the benthic macroinvertebrate data, the fish data
results suggest that the hypotheses were true in terms of the changes in fish
density but it cannot discount the possibility that the increased production is the
result of yearly variation.
35
CONCLUSION AND SUMMARY
The first purpose of this study was to determine if the aquatic habitat
enhancement measures currently being implemented within eastern Ontario
provide a benefit to the aquatic environment. While initial changes to the
herbaceous vegetation within the riparian habitat were observed it was clear that
additional monitoring years would be required in order to fully document
improvements. Since it is well established that a healthy riparian area supports a
diverse community of flora and fauna, it would be anticipated that the increase in
herbaceous vegetation would result in a positive impact to the aquatic community.
The increase in the EPT, decrease in chironomids and increase in fish density
supports that the habitat is improving.
Unfortunately, longer-term monitoring was not feasible for this study or
other similar environmental impact assessment studies, but it is possible to
anticipate the improvements that one would expect to see over a longer timeline.
These improvements would begin with a more diverse riparian area followed by
changes to the channel morphology. In the next 5 to 8 years, it would be
anticipated that the planted trees and shrubs would continue to grow and provide
stabilization to the banks as well as some shade cover. Canopy cover would be
expected to improve in 10-20 years. This shading and eventual canopy cover
would be anticipated to lower the water temperature within the watercourse. The
increased stability of the banks from the trees should allow changes to the channel
36
morphology to occur during periods of high flows. The denser herbaceous cover
combined with the presence of woody vegetation would also decrease soil
erosion. These improvements would allow sensitive species to colonize the area;
changing the benthos and fish communities. Following even more time one
would also expect some of the trees and shrubs to die or break and provide small
woody debris and large woody debris to the watercourse which in turn would not
only provide additional structure but could help change the channel morphology
by altering the direction of flow. Thus while some changes to the riparian
vegetation was observed during this study, it is anticipated that these represented
only the beginning and that in another 20+ years, the riparian area would be treed
and the aquatic communities more diverse.
The second goal of this study was to determine if the commonly applied
methods could provide sufficient information for a small-scale EIA monitoring
program. In such monitoring, success is usually regarded as the ability to
maintain status quo or provide an improvement in habitat conditions. In this
study the results indicated an improvement for the habitat (overhanging and
riparian vegetation). It could be argued that the data point to an improvement in
the benthic macroinvertebrate quality (increase in EPT and decrease in
chironomids) and in fish production. While these improvements could be random
yearly variations, they do suggest that there was no deterioration in conditions.
As such I would suggest that the methods, over the period of time applied, can
observe initial changes but are insufficient to conclusively assess the effects
unless they are large.
37
The original design of the study did not include monitoring of plant
height. This would have been a valuable addition to the study design in that it
would have documented the changes that were observed in the photographic
records. Also, the Ontario Benthic Monitoring Network methodology (Jones et
al. 2005) published following the commencement of this study, includes not only
a timed sampling period for the travelling kicks but also a set stream length.
Including a fixed distance would have enhanced the ability to compare sites over
time.
38
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44
Table 1 List of fish species recorded as occuring in the beaudette river sub-
watershed
Common Name Latin Name
Northern pike Esox lucius
Central mudminnow Umbra limi
White sucker Catostomus commersoni
Northern redbelly dace Chrosomus eos
Brassy minnow Hybognathus hankinsoni
Golden shiner Notemigonus crysoleucas
Common shiner Luxilus cornutus
Rosyface shiner Notropis rubellus
Bluntnose minnow Pimephales notatus
Fathead minnow Pimephales promelas
Creek chub Semotilus atromaculatus
Northern Pearl dace Margariscus nachtriebi
Brown bullhead Ameirus nebulosus
Stonecat Noturus flavus
American eel Anguilla rostrata
Brook stickleback Culaea inconstans
Rock bass Ampbloplites rupestris
Pumpkinseed Lepomis gibbosus
Iowa darter Etheostoma exile
Fantail darter Etheostoma flabellare
Johnny darter Etheostoma nigrum
Logperch Percina caproides
Mottled sculpin Cottus bairdii
45
Table 2 Upstream and downstream coordinates for the six sampling sites.
Coordinates are in UTMs based on NAD83.
Site Downstream End Upstream End
Easting Northing Easting Northing
1 527924 5009254 527870 5009252
2 527769 5009257 527738 5009260
3 527636 5009381 527607 5009413
4 527444 5009524 527414 5009515
5 527468 5009556 527453 5009576
6 527430 5009609 527431 5009619
Table 3 Summary of observed stressor and implemented aquatic rehabilitation
per site
Stressor/Aquatic Rehabilitation Site
1 2 3 4 5 6
Stressors
Decrease in water quality from upstream
land-uses ● ● ● ● ● ●
Trampling ○ ● ● ● ● ●
Cattle crossing ○ ● ○ ○ ●
Runoff from fields ○ ● ● ● ● ●
Failed banks ○ ● ● ● ● ●
Aquatic Rehabilitation
Fencing ○ ● ● ● ● ●
Riparian planting ○ ● ● ● ● ●
Low-level crossing placed upstream of site ○ ○ ○ ● ○ ○
Culvert replacement ○ ○ ○ ○ ● ○
Boulder protection of banks ○ ○ ○ ○ ● ●
46
Table 4 Description of the vegetation and rocky material ranking criteria.
Vegetation ranks 1- 4 and rocky material cover ranks were taken from
Stevenson and Mills 1999. Vegetation rankings 5-8 were added as part
of this study.
Rank Vegetation cover Rocky Material
1
<50% of segment is
covered by herbaceous
vegetation.
<20% is stony material
(>2,5cm)
2
50-70% of segment is
covered by herbaceous
vegetation
20-40% is stony material
(>2.5cm)
3
71-90% of segment is
covered by herbaceous
vegetation.
40-65% is stony material
(>2.5cm)
4
>90% of segment is
covered by herbaceous
vegetation.
>65% is stony material
(>2.5cm)
5
5-50% of segment is
covered by woody
vegetation
6
51-70% of segment is
covered by woody
vegetation.
7
71-90% of segment is
covered by woody
vegetation.
8
>90% of segment is
covered by woody
vegetation.
47
Table 5 Summary of average yearly channel morphology per site. Channel width refers to the active channel width.
In-stream cover types consist of rocky material, aquatic vegetation (AV), overhanging vegetation (OV), algae
and undercut banks (UC).
Site Date
Site
Length
(m)
Bank
Height
(cm)
Wetted
Width
(m)
Channel
Width
(m)
Depth
(cm)
Substrate
Size (mm)
Obser.
Points
with
Cover
(%)
In-Stream Cover Types
Rock
(%)
AV
(%)
OV
(%)
Algae
(%)
UC
(%)
1 2005 46.0 42.8 379.0 751.5 8.6 9.2 80.0 75.0 3.3 1.7 0.0 0.0
1 2006 46.0 26.5 401.0 688.5 10.5 7.1 70.0 68.3 1.7 0.0 3.3 0.0
1 2007 46.0 32.0 478.5 722.0 12.4 69.3 90.0 85.0 3.3 3.3 0.0 0.0
1 2008 46.0 37.1 400.0 751.5 10.8 56.1 93.3 95.0 5.0 6.7 6.7 0.0
2 2005 28.0 22.1 349.5 451.0 16.5 51.1 78.3 70.0 15.0 1.7 6.7 0.0
2 2006 28.0 26.3 2.96 4.30 12.6 63.6 93.3 70.0 41.7 15.0 38.3 0.0
2 2007 28.0 26.6 3.35 4.64 17.5 90.0 86.7 88.3 25.0 6.7 5.0 0.0
2 2008 28.0 49.3 3.25 4.85 11.9 67.1 96.7 91.7 40.0 11.7 11.7 1.7
3 2005 33.0 20.0 4.79 5.10 68.3 0.7 20.0 0.0 5.0 6.7 0.0 11.7
3 2006 33.0 35.7 4.26 4.08 67.2 25.1 60.0 1.7 31.7 36.7 0.0 25.0
3 2007 33.0 25.2 4.32 4.69 68.6 12.9 55.0 3.3 15.0 43.3 0.0 26.7
3 2008 33.0 36.3 4.36 4.81 64.9 29.1 55.0 6.7 16.7 30.0 3.3 28.3
4 2005 32.0 20.8 3.98 4.47 57.7 2.6 43.3 8.3 13.3 26.7 0.0 1.7
48
Site Date
Site
Length
(m)
Bank
Height
(cm)
Wetted
Width
(m)
Channel
Width
(m)
Depth
(cm)
Substrate
Size (mm)
Obser.
Points
with
Cover
(%)
In-Stream Cover Types
Rock
(%)
AV
(%)
OV
(%)
Algae
(%)
UC
(%)
4 2006 32.0 42.1 4.14 4.25 49.5 14.9 56.7 1.7 25.0 38.3 3.3 8.3
4 2007 32.0 31.3 4.15 4.38 57.0 9.1 61.7 5.0 28.3 35.0 0.0 13.3
4 2008 32.0 33.7 4.02 4.38 50.9 10.3 78.3 1.7 46.7 43.3 3.3 13.3
5 2005 21.0 22.9 1.77 2.89 29.9 0.1 50.0 0.0 43.3 8.3 0.0 1.7
5 2006 21.0 26.2 1.63 2.62 25.1 0.1 48.3 0.0 26.7 26.7 0.0 6.7
5 2007 21.0 26.4 1.84 2.71 27.5 2.3 61.7 1.7 21.7 45.0 0.0 11.7
5 2008 21.0 36.5 2.13 3.16 20.8 0.1 50.0 3.3 28.3 21.7 0.0 3.3
6 2005 17.0 21.5 1.36 2.16 7.6 19.1 61.7 33.3 8.3 3.3 18.3 3.3
6 2006 17.0 25.7 1.28 2.03 9.0 60.8 80.0 50.0 30.0 5.0 5.0 0.0
6 2007 17.0 28.5 1.51 2.37 10.5 44.3 75.0 45.0 11.7 31.7 0.0 1.7
6 2008 17.0 33.4 1.55 2.79 7.4 42.5 68.3 51.7 18.3 10.0 0.0 1.7
49
Table 6 Median (lower and upper confidence interval) of in-stream cover; N=6; * indicates significant result
Metric Year
Friedman
Test
DF=3
Pearson’s Correlation
2005 2006 2007 2008 p-value p-value R
Rocky
material
12.5
(0.00;43.93)
14.5
(0.36;41.64)
15.0
(1.36;52.29)
17.5
(1.36;56.29) 0.034* 0.317 0.213
Aquatic
vegetation
6.5
(2.36;19.93)
17.0
(6.00;22.86)
11.5
(3.79;16.29)
14.0
(5.50;26.57) 0.171 0.299 0.221
Overhanging
vegetation
3.0
(1.00;12.07)
12.5
(1.07;22.64)
20.0
(2.71;23.29)
10.0
(2.71;23.29) 0.029* 0.061 0.388
Algae 0.0
(0.00;8.50)
2.0
(0.00;17.14)
0.0
(0.00;2.64)
2.0
(0.00;6.57) 0.130 0.768 0.064
Undercut
Banks
2.5
(0.36;11.86)
2.0
(0.00;13.50)
4.0
(0.00;12.50)
1.5
(0.36;13.79) 0.995 0.990 -0.003
50
Table 7 Median (lower and upper confidence interval) of riparian habitat cover; N=6; * indicates significant result
Metric Year
Friedman
Test
DF=3
Pearson’s Correlation
2005 2006 2007 2008 p-value p-value r
Vegetation 3.0
(2.00;5.57)
3.5
(1.36;5.29)
4.0
(3.00;5.29)
4.0
(3.00;5.93) 0.031* 0.091 0.353
Rocky
Material on
Banks
1.0
(1.00;2.00)
1.5
(1.00;2.64)
1.5
(1.00;3.00)
1.0
(1.00;2.00) 0.245 0.944 0.015
51
Table 8 Summary of the proportion of benthic invertebrates collected per year.
TAXON 2005 2006 2007 2008
Nematoda 0.0013 0.0000 0.0012 0.0000
Coelenterata
F. Hydridae 0.0000 0.0005 0.0000 0.0000
Turbellaria
F. Dugesiidae 0.0008 0.0000 0.0000 0.0000
F. Planariidae 0.0004 0.0000 0.0000 0.0000
Undetermined Tricladida 0.0000 0.0000 0.0004 0.0000
Annelida
F. Erpobdellidae 0.0004 0.0000 0.0000 0.0000
Oligochaeta
F. Enchytraeidae 0.0000 0.0005 0.0000 0.0000
F. Lumbriculidae 0.0004 0.0000 0.0000 0.0000
F. Naididae 0.0085 0.0232 0.0094 0.0014
F. Sparganophilidae 0.0000 0.0000 0.0012 0.0000
F. Tubificidae 0.0242 0.0319 0.0331 0.0062
Hirundinea
F. Glossiphoniidae 0.0030 0.0015 0.0004 0.0000
Branchiobellida
Undetermined
Branchiobdellida 0.0000 0.0000 0.0012 0.0010
Mollusca
F. Ancylidae 0.0008 0.0029 0.0000 0.0000
F. Hydrobiidae 0.0038 0.0126 0.0082 0.0171
F. Lymnaeidae 0.0076 0.0005 0.0000 0.0000
F. Planorbidae 0.0068 0.0010 0.0000 0.0000
F. Sphaeriidae 0.0331 0.0058 0.0078 0.0038
Isopoda
52
TAXON 2005 2006 2007 2008
F. Asellidae 0.0000 0.0000 0.0000 0.0005
Amphipoda
F. Gammaridae 0.0374 0.0082 0.0140 0.0157
F. Hyalellidae 0.0318 0.0160 0.0491 0.1217
Podocopida
Undetermined Podocopida 0.0004 0.0000 0.0000 0.0010
Decapoda
F. Cambaridae 0.0000 0.0000 0.0019 0.0057
Cladocera
F. Chydoridae 0.0004 0.0015 0.0012 0.0005
F. Daphnidae 0.0004 0.0000 0.0000 0.0000
F. Macrothricidae 0.0000 0.0000 0.0004 0.0000
Undetermined Cladocera 0.0000 0.0000 0.0004 0.0000
Copepoda
F.Cycloipidae 0.0004 0.0019 0.0008 0.0005
Acariformes
Undertermined Acariformes 0.0004 0.0116 0.0031 0.0000
Ephemeroptera
F. Baetidae 0.0047 0.0102 0.0078 0.0171
F. Caenidae 0.0718 0.2023 0.0834 0.1260
F. Ephemeridae 0.0000 0.0010 0.0004 0.0005
F. Heptageniidae 0.0047 0.0053 0.0187 0.0280
F. Leptophlebiidae 0.0004 0.0015 0.0035 0.0252
Undetermined
Ephemeroptera 0.0000 0.0005 0.0004 0.0000
Ondonata
F. Aeshnidae 0.0013 0.0005 0.0008 0.0005
F. Calopterygidae 0.0000 0.0015 0.0016 0.0081
53
TAXON 2005 2006 2007 2008
F. Coenagrionidae 0.0064 0.0024 0.0058 0.0071
F. Libellulidae 0.0008 0.0000 0.0000 0.0000
Hemiptera
F. Belostomatidae 0.0008 0.0024 0.0000 0.0019
F. Corixidae 0.0832 0.0131 0.0175 0.0019
F. Nepidae 0.0000 0.0010 0.0012 0.0010
F. Notonectidae 0.0000 0.0000 0.0004 0.0000
F. Veliidae 0.0017 0.0015 0.0004 0.0014
Plecoptera
F. Capaniidae 0.0000 0.0000 0.0012 0.0000
F. Taeniopterygidae 0.0259 0.0363 0.0347 0.0257
Undetermined Plecoptera 0.0000 0.0005 0.0000 0.0000
Coleoptera
F. Dytiscidae 0.0000 0.0000 0.0008 0.0005
F. Elmidae 0.1987 0.1936 0.2537 0.2510
F. Haliplidae 0.0000 0.0015 0.0004 0.0014
F. Hydrophilidae 0.0008 0.0000 0.0000 0.0024
F. Psephenidae 0.0085 0.0169 0.0355 0.0390
F. Scirtidae 0.0004 0.0000 0.0000 0.0010
Meglaloptera
F. Sialidae 0.0000 0.0005 0.0004 0.0029
Trichoptera
F. Brachycentridae 0.0004 0.0252 0.0027 0.0000
F. Dipseudopsidae 0.0008 0.0000 0.0008 0.0005
F. Helicopsychidae 0.0000 0.0029 0.0004 0.0000
F. Hydropsychidae 0.0229 0.0682 0.0869 0.1022
F. Hydroptilidae 0.0127 0.0213 0.0016 0.0024
F. Leptoceridae 0.0042 0.0039 0.0004 0.0000
54
TAXON 2005 2006 2007 2008
F. Limnephilidae 0.0127 0.0136 0.0097 0.0233
F. Philopotamidae 0.0072 0.0131 0.0230 0.0105
F. Phryganeidae 0.0098 0.0039 0.0000 0.0005
F. Polycentropodidae 0.0034 0.0034 0.0043 0.0067
Undetermined Trichoptera 0.0004 0.0005 0.0004 0.0000
Lepidoptera
F. Pyralidae 0.0000 0.0000 0.0004 0.0005
Diptera
F. Ceratopogonidae 0.0306 0.0169 0.0043 0.0010
F. Chironomidae 0.3028 0.2014 0.2408 0.1212
F. Empididae 0.0004 0.0010 0.0027 0.0000
F. Ephydridae 0.0013 0.0005 0.0000 0.0000
F. Psychodidae 0.0008 0.0000 0.0066 0.0000
F. Simuliidae 0.0004 0.0068 0.0004 0.0000
F. Stratiomyidae 0.0004 0.0000 0.0000 0.0000
F. Tabanidae 0.0055 0.0010 0.0008 0.0005
F. Tipulidae 0.0102 0.0048 0.0117 0.0138
55
Table 9 Mean percent (±SD) benthic macroinvertebrates; N=18. * indicates significant result
Metric Year Friedman Test (DF=3_ Pearson’s Correlation
2005 2006 2007 2008 p-value F p-value r
EPT 24.52
(1.68)
39.51
(2.06)
30.47
(2.46)
36.65
(2.07) 0.000* 20.95 0.012* 0.295
Chironomids 32.52
(2.60)
25.86
(1.88)
25.15
(2.70)
19.61
(1.95) 0.000* 12.14 0.000* 0.416
56
Table 10 Median (lower and upper confidence interval) of benthic macroinvertebrates; N=6; * indicates significant
result
Metric Year
Friedman Test
DF=3 Pearson’s Correlation
2005 2006 2007 2008 p-value p-value r
MFBI 6.07
(4.75; 6.23)
5.76
(4.43; 6.35)
5.72
(4.39; 6.60)
5.27
(4.41; 6.56) 0.896 0.745 -0.070
Shannon
Wiener
Diversity
Index
2.05
(1.90; 2.24)
2.20
(1.78; 2.59)
2.06
(1.66; 2.36)
1.96
(1.84; 2.23) 0.590 0.350 -0.199
57
Table 11 Summary of the percent of fish species captured per year
Common Name Latin Name 2005 2006 2007 2008
Central mudminnow Umbra limi 2.71 3.39 6.53 3.83
White sucker Catostomus commersoni 1.35 2.76 1.48 0.89
Northern redbelly dace Chrosomus eos 1.18 1.25 2.49 7.93
Finescale dace Chrosomus neogaeus 0.00 0.50 0.39 1.78
Brassy minnow Hybognathus hankinsoni 0.51 0.39 1.40 1.42
Golden shiner Notemigonus crysoleucas 0.00 0.75 1.48 1.51
Common shiner Luxilus cornutus 0.85 3.26 5.59 5.79
Blacknose shiner Notropis heterolepis 0.17 0.00 0.16 0.00
Bluntnose minnow Pimephales notatus 2.20 4.14 6.53 4.99
Fathead minnow Pimephales promelas 0.34 4.14 1.17 0.27
Creek chub Semotilus atromaculatus 4.23 12.05 14.37 13.36
Brown bullhead Ameirus nebulosus 0.17 0.75 0.31 0.18
Stonecat Noturus flavus 0.00 0.00 0.00 0.09
Tadpole madtom Noturus gyrinus 0.00 0.00 0.08 0.00
Brook stickleback Culaea inconstans 0.85 0.50 1.55 0.80
Rock bass Ampbloplites rupestris 2.20 2.51 3.81 1.78
Pumpkinseed Lepomis gibbosus 1.86 2.13 2.33 0.45
Iowa darter Etheostoma exile 0.00 0.00 0.00 0.27
Fantail darter Etheostoma flabellare 77.50 53.95 44.68 45.68
Johnny darter Etheostoma nigrum 3.75 7.15 5.05 8.55
Logperch Percina caproides 0.17 0.38 0.62 0.45
58
Table 12 Median (lower and upper confidence interval) of fish Results; N=6; * indicates significant result
Metric Year
Friedman
Test
DF=3
Pearson’s
Correlation
2005 2006 2007 2008 p-value p-value r
Fish Density 0.91
(0.12;1.79)
1.81
(0.43;2.25)
2.08
(0.90;3.92)
2.52
(0.74;3.51) 0.044* 0.020* 0.472
Number of
Species 9.0 (3.7;11.6)
10.5
(6.4;15.0)
13.0
(9.4;17.3)
12.5
(7.0;16.0) 0.004* 0.067 0.380
Shannon
Wiener
Diversity
Index
1.23
(0.43;2.19)
1.51
(0.82;2.39)
1.88
(1.21;2.52)
1.53
(1.32;2.35) 0.014* 0.181 0.283
Percent
Insectivorous
Cyprinid
0.01
(0.000;0.058)
0.02
(0.005;0.108)
0.04
(0.008;0.121)
0.08
(0.004;0.128) 0.077 0.060 0.389
Percent
Omnivores
0.04
(0.005;0.388)
0.15
(0.029;0.461)
0.28
(0.047;0.470)
0.016
(0.037;0.355) 0.494 0.413 0.175
Percent
herbivore
0.00
(0.000;0.080)
0.02
(0.000;0.072)
0.08
(0.002;0.158)
0.10
(0.020;0.153) 0.120 0.019* 0.474
59
Figure 1 Location of the St. Lawrence River (Cornwall) AOC. (Figure produced by RRCA)
60
Figure 2 Location of project area (Figure produced by Raisin Region Conservation Authority)
61
Figure 3 Location of the sampling sites
62
Figure 4 Correspondence analysis results on abundance data for benthic macroinvertebrates grouped by habitat (Site 1
= downstream control, Sites 2-6 = impacted). In order to reduce the clutter, the taxa have been grouped and
coded as follows: A =Nematoda, B = Coelenterate, C = Turbellaria, D = Annelida, E = Oligochaeta, F =
Hirundinea, G = Branchiobellida, H = Mollusca, I = Isopoda, J – amphipoda, K = Podocopida, L =
Decapoda, M = Cladocera, N = Copepoda, O = Acariformes, P = Ephemeroptera, Q = Ondonata, R =
Hemiptera, S = Plecoptera, T = Coleoptera, U = Megaloptera, V = Trichoptera, W = Lepidoptera, X =
Diptera. The families that are represented by each of the above groups are available in Table 8.
63
Figure 5 Correspondence analysis results on abundance data for fish grouped by habitat (Site 1 = downstream control,
Sites 2-6 = impacted). 141 = central mudminnow, 163 = white sucker, 182 = northern redbelly dace, 183 =
finescale dace, 189 = brassy minnow, 194 = golden shiner, 198 = common shiner, 200 = blacknose shiner,
208 = bluntnose minnow, 209 = fathead minnow, 212 = creek chub, 233 = brown bullhead, 235 = stonecat,
236 = tadpole madtom, 281 = brook stickleback, 311 = rock bass, 313 = pumpkinseed, 338 = Iowa darter, 339 = fantail darter, 341 = johnny darter, 342 = logperch
64
Photo 1 View upstream from the downstream end of Site 2; 2005.
Photo 2 View upstream from the downstream end of Site 2; 2007.
65
Photo 3 View downstream from upstream on Site 5, 2005
Photo 4 View downstream from upstream on Site 5, 2008
66
CONNECTING STATEMENT
As discussed in Chapter 1 the processing of benthic macroinvertebrate samples is
costly and has lead to the creation of Rapid Bioassessment Protocols
(Courtemanch 1996) which minimize costs by using fixed counts (Piscart et al.
2006). One suggestion to help determine what a coarser level of identification or
sub-sampling would mean to the results would be to conduct more detailed
processing on a sub-set of samples. The results of a comparison of the
interpretation of results from a 100-count and a total count was completed on the
2006 benthic samples are presented in the following section.
67
Appendix A – Comparison of 100-fixed Count and Whole Count
Results
Introduction
While the cost of collecting benthic macroinvertebrate samples is
relatively small the processing time for each sample can be quite large, and
prohibitively so for small scale environmental impact assessments (EIAs). Fixed-
count methods are widely used in benthic macroinvetebrate sampling both the
United States (USEPA) and Canada (Ontario Stream Assessment Protocol
(OSAP), Ontario Benthic Biomonitoring Network (OBBN) in order to reduce
processing time. Examples of both total counts (Kosnicki and Sites 2007) and
fixed counts (Somers et al.1998, Pond et al. 2008) are documented in recent
published literature. Doberstein et al. (2000) tested a variety of sizes of fixed-
counts versus a whole-count using computer generated results for the fixed
counts. They concluded that whole samples provided more reliable information
with less variability (Doberstein et al. 2000). However, others argue that fixed
count methods are reliable when describing species richness (Barbour and
Gerritsen 1996) and for metrics which use percentage data (i.e. percent
Ephemeroptera plus Plecoptera plus Trichoptera) (Barbour and Gerritsen 1996,
Courtemanche 1996). Courtemanche (1996) recommended that fixed counts not
be used for comparing the number of taxa as they were inaccurate.
68
Understanding how the use of fixed-counts versus whole-counts affects
the interpretation of results is important. One way of determining what different
conclusions could be drawn depending on the amount of sample that is processed
is to compare the results of the fixed-counts and the whole-count on a subset of
samples. As part of a four-year study of the aquatic environment before and after
the implementation of aquatic habitat restoration a comparison of a 100-count and
a whole-count was conducted. The purpose was to determine if different
conclusions would be made following a fixed-count or a whole-count processing.
The hypothesis was that the conclusions drawn from the data analysis would be
statistically the same for the 100-count and the whole-count data.
Methods
Macroinvertebrate Sampling
Six sampling sites were established for this study, five of which were
located on the farm property and one downstream (Figure 2, Table 1). The sites
are labelled 1 through 6, from downstream to upstream. “Protocols for Measuring
Biodiversity: Benthic Macroinvertebrates in Fresh Waters” (Rosenberg et al.
1997) utilized by the Ecological Monitoring and Assessment Network (EMAN)
(http://www.ccmn.ca) was followed for benthic community sampling. Sampling
was conducted in October 2006 whe benthic macroinvertebrates were collected
using a 0.3 m wide D-frame kicknet with 500 µm mesh. A total of 27 samples
were collected between October 10 and 12, 2005. A sampled was collected by
walking perpendicular to the flow from one bank to the other until a period of 2-
69
minutes elapsed. The substrate was disturbed using the travelling kick
methodology for benthic macroinvertebrate sampling and a dip net held
downstream collected the dislodged organisms. All habitats encountered were
sampled. Each site was divided into 5 equal lengths and five 2-minute samples
were collected. The sampling began at the downstream end of the downstream
site and continued upstream. The invertebrates were preserved with 70% ethanol
alcohol. The samples were processed first for the 100 fixed-count. The
remainder of the sample was then processed and the data from the 100 fixed-
count was added for the total-count. The benthic macroinvertebrates were sorted
and identified to family.
Data Analysis
These same analyses that were to be used for the four year before and after
study on the aquatic environment were selected for the comparison between 100-
count and whole-count, these were: the number of families, (percent sensitive
families (EPT8), percent tolerant families (chironomids), and Family Biotic Index
(Mandaville 2002)9. Normality was tested using the Anderson-Darling test and a
comparison of the variances of the two data sets was tested using the Levene test.
The Anderson-Darling results found that the arcsine-square root transformation of
the EPT ratio had a normal distribution and the Levene Test found that all
variables had a similar variation. The Paired T-Test was used to compare the EPT
8 EPT refers to Ephmeroptera plus Plecoptera plus Trichoptera 9 FBI refers to Family Biotic Index which provides an average tolerance to organic pollution for
benthic macroinvertebrate families
70
ratios. The number of families in both the 100-count and the whole-count were
normally distributed however they did not have equal variance (p=0.000) and as
such the non-parametric Mann-Whitney Test was used to compare the number of
families. Despite transforming the chironomid ratios using log10(x+1), log(x+1),
arcsine square root and square root, the data could not be normalized. Norris and
Georges (1993) state that indices such as MFBI should be analyzed using non-
parametric tests as the distributions are unknown. As such the Mann-Whitney
Test was also utilized for the MFBI and the chironomid ratio. The analyses were
completed using Minitab version 15 software. A p-value of p<0.05 was set as the
threshold for significance.
Results
A total of 3522 individuals from the 100-count samples and 20699
individuals from the whole count samples were tallied. There were 61 families
recorded during the 100-count and 77 in the whole count (Table 2), a 20%
difference. All 100-count samples combined, the dominant families were
Chironomidae (33%), Elmidae (18%), Coroxidae (7%) and Caenidae (7%). All
whole count samples combined, the dominant families were Chironomidae (36%),
Elmidae (16%), Caenidae (8%) and Corixidae (5%).
The paired t-test results (t=-1.66, DF =26, P=0.109) for the EPT ratio
indicated that there was no significant difference in the ratios when comparing the
100-count or the whole-count methods (Table 3). The Mann-Whitney results for
the chironomid ratio (DF=26, P=0.597) and the MFBI (DF=26, P=0.574) values
71
were also found to be not significant. Only the results for the number of families
was highly significant (DF=26, p<0.0001) indicating the 100-count highly
underestimated the number of families within a sample (Table 4).
Discussion
Our purpose was to determine if conducting a 100-count would lead us to
the same conclusions as a whole count for our data sets. The four analyses which
were tested (number of families, EPT ratio, chironomid ratio and modified family
biotic index ratio) yielded the same interpretation of the data with the exception of
the number of families. This supports the work carried out by Barbour and
Gerritsen (1996) and Courtemanche (1996) who indicated that data analysis
involving ratios would not be impacted by fixed-count versus whole-counts. The
highly significant difference in the higher number of families counted in the
whole-count was not unexpected as it is commonly accepted that the larger the
sample size, the larger number of species will be enumerated (Kwak and Peterson
2007). This suggests that the use of 100-counts for the purposes of the three
analyses would provide the same conclusions at a much lower cost. While
completing a total count would provide a higher number of families and would
allow for a comparison of abundance, it was determined that this did not warrant
the additional time and expenses associated with whole-counts.
72
References
Barbour, M. T. & Gerritsen J. (1996) Subsampling of benthic samples: a defence
of the fixed-count method. J. N. Am. Benthol. Soc. 15(3), 386-391.
Barbour, M.T., J. Gerritsen, B.D. Snyder, & Stribling J. B. (1999). Rapid
Bioassessment Protocols for Use in Streams and Wadeable Rivers: Periphyton,
Benthic Macroinvertebrates and Fish, Second edition. EPA 841-B-99-002.
U.S. Environmental Protection Agency; Office of Water; Washington, D.C.
Courtemanche D. L. (1996) Commentary on the subsampling procedures used for
rapid bioassessments. J. N. Am. Benthol. Soc., 15(3), 381-385.
Doberstein, C. P., Karr J. R., & Conquest L. L. (2000). The effect of fixed-count
subsampling on macroinvertebrate biomonitoring in small streams. Freshwater
Biology, 44, 355-371.
Kosnicki, E. & Sites R. W. (2007) Least-desired index for assessing the
effectiveness of grass riparian filter strips in improving water quality in an
agricultural region. Environ. Entomol. 36(4), 713-724.
Kwak, T. J. & Peterson J. T. (2007) Community indices, Parameters, and
Comparisons. Pages 677-764 in C.S. Guy and M.L. Brown, editors. Analysis
and interpretation of freshwater fisheries data. American Fisheries Society,
Bethesda, Maryland.
Mandaville, S. M. 2002. Benthic macroinvertebrates in freshwaters – taxa
tolerance values, metrics, and protocols. Soil & Water Conservation Society of
Metro Halifax. 48 + appendices.
Norris and Georges (1993) Chapter 7 Analysis and Interpretation of Benthic
Macroinvertebrate Surveys in Freshwater Biomonitoring and Benthic
Macroinvertebrates Ed. Rosenberg and Resh. Chapman and Hall. New York.
Piscart, C., Moreteau J. C. & Beisel J. N. (2006). Salinization consequences in
running waters: use of a sentinel substrate as a bioassessment method. J. N.
Am. Benthol. Soc., 25(2), 477-486.
73
Pond, G. J., Passmore, M. E., Borsuk, F. A., Reynolds, L. and Rose, C. J. (2008).
Downstream effect of mountaintop coal mining: comparing biological
conditions using family- and genus- level macroinvertebrates bioassessment
tools. J. N. Am. Benthol. Soc., 27(3), 717-737.
Rosenberg, D.M., Davies I. J., Cobb D. G. & Wiens A. P. (1997). Protocols for
measuring biodiversity: benthic macroinvertebrates in fresh waters.
Department of fisheries and Oceans, Freshwater Institute, Winnipeg,
Manitoba.
Somers, K. M., Reid R. A. & David S. M. (1998) Rapid biological assessments:
how many animals are enough? J. N. Am. Benthol. Soc., 17(3), 348-358.
74
Table 1 Upstream and downstream coordinates for the six sampling sites.
Coordinates are in UTMs (NAD83).
Site Downstream End Upstream End
Easting Northing Easting Northing
1 527924 5009254 527870 5009252
2 527769 5009257 527738 5009260
3 527636 5009381 527607 5009413
4 527444 5009524 527414 5009515
5 527468 5009556 527453 5009576
6 527430 5009609 527431 5009619
Table 2 Summary of the total benthos percent composition for the 100-count
and whole count.
TAXON 100-
count
Whole
count
Nematoda
Undetermined Nematoda 0.11 0.05
Turbellaria
F. Dugesiidae 0.06 0.43
F. Planariidae 0.03
Undetermined Turbellaria 0.03 0.03
Annelida
F. Erpobdellidae 0.03 0.01
Oligochaeta
F. Enchytraeidae 0.04
F. Lumbriculidae 0.03 0.01
F. Naididae 1.05 1.81
F. Tubificidae 2.44 3.80
75
TAXON 100-
count
Whole
count
Hirundinea
F. Glossiphoniidae 0.23 0.17
Branchiobellida
Undetermined
Branchiobdellida <0.001
Mollusca
F. Ancylidae 0.09 0.14
O. Basommatophora <0.001
F. Hydrobiidae 0.34 0.17
F. Lymnaeidae 0.60 0.39
F. Planorbidae 0.62 0.61
F. Physidae 0.06 0.03
F. Sphaeriidae 3.63 4.63
Undetermined Mollusca 0.01
Isopoda
F. Asellidae <0.001
Amphipoda
F. Gammaridae 4.46 2.29
F. Hyalellidae 2.87 2.20
Podocopida
Undetermined Podocopida 0.03 0.02
Decapoda
F. Cambaridae 0.04
Cladocera
F. Chydoridae 0.03 0.03
F. Daphnidae 0.03 0.01
F. Sididae <0.001
Copepoda
76
TAXON 100-
count
Whole
count
F. Cycloipidae 0.03 <0.001
Acariformes
Undertermined Acariformes 0.06 0.12
Ephemeroptera
F. Baetidae 0.31 0.15
F. Caenidae 6.70 7.99
F. Ephemeridae <0.001
F. Heptageniidae 0.88 0.57
F. Leptophlebiidae 0.03 <0.001
Undetermined
Ephemeroptera 0.06 0.05
Ondonata
F. Aeshnidae 0.11 0.02
F. Calopterygidae 0.03 0.03
F. Coenagrionidae 0.68 0.58
F. Gomphidae <0.001
F. Libellulidae 0.09 0.08
Undetermined Zygoptera 0.01
Hemiptera
F. Belostomatidae 0.06 0.06
F. Corixidae 6.93 5.25
F. Nepidae 0.03 0.01
F. Notonectidae <0.001
F. Pleidae <0.001
F. Veliidae 0.11 0.02
Plecoptera
F. Taeniopterygidae 2.73 3.15
Undetermined Plecoptera 0.01
77
TAXON 100-
count
Whole
count
Coleoptera
F. Chrysomelidae 0.01
F. Dryopidae 0.03
F. Dytiscidae 0.03 0.06
F. Elmidae 18.00 15.88
F. Gyrinidae <0.001
F. Haliplidae 0.03 0.05
F. Hydrophilidae 0.09 0.06
F. Psephenidae 0.74 0.25
F. Scirtidae 0.03 0.05
Meglaloptera
F. Sialidae 0.03 0.07
F. Sisyridae <0.001
Trichoptera
F. Brachycentridae 0.06 0.05
F. Dipseudopsidae 0.11 0.04
F. Helicopsychidae 0.11
F. Hydropsychidae 2.02 1.87
F. Hydroptilidae 1.31 1.56
F. Leptoceridae 0.43 0.23
F. Limnephilidae 1.16 0.99
F. Philopotamidae 0.60 0.26
F. Phryganeidae 0.85 0.63
F. Polycentropodidae 0.45 0.27
Undetermined Trichoptera 0.03 0.01
Lepidoptera
F. Pyralidae 0.01
Diptera
78
TAXON 100-
count
Whole
count
F. Ceratopogonidae 3.69 4.00
F. Chaoboridae 0.09
F. Chironomidae 32.96 36.39
F. Dixidae <0.001
F. Empididae 0.06 0.06
F. Ephydridae 0.14 0.05
F. Muscidae 0.03 0.03
F. Psychodidae 0.06 0.08
F. Simuliidae 0.06 0.33
F. Stratiomyidae 0.03 0.04
F. Tabanidae 0.45 0.42
F. Tipulidae 0.94 0.91
Undetermined Diptera 0.01
79
Table 3 Mean (±SD) EPT values for the 100-count and whole-count. Statistical
comparisons were completed using paired t-test with arcsine square root
transformed data, n=27
Metric
Mean Paired
t-test
p-value 100-count Whole-count
EPT ratio 24.25 (7.25) 24.98 (7.41) 0.109
Table 4 Median (upper and lower confidence interval) values for the 100-count
and whole-count, n=27
Metric
Mean Mann-
Whitney
p-value 100-count Whole-count
Chironomid ratio 0.37
(0.32; 0.42)
0.34
(0.29; 0.41) 0.598
Number of families 16
(14.97; 17.00)
26
(23.97; 29.00) <0.0001
Modified Family
Biotic Index
5.97
(5.72; 614)
6.04
(5.86; 6.18) 0.574
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