i

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).

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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.

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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).

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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)

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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).

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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