1 Introduction
The muskrat (Ondatra zibethicus ) is seen as a fixture of wetlands
across North America, being one of the most common and widely
distributed furbearer species on the continent. They are found primarily
in marshes, but also in ponds, sloughs, lakes, ditches, rivers and
streams from the east coast to the west, and from the Mackenzie Delta in
Canada’s north to northern Mexico in the south (Boutin and Birkenholz,
1987). The muskrat also ranks as the most heavily harvested wild
furbearer in North America of the 20th century (Obbard et al., 1987) and
has contributed more than any other animal to the combined income of
North America’s trappers over the past 150 years (OFMF, 2019). In the
early 1900s, millions of muskrats were trapped and sold across North
America and, though harvest numbers are lower today, muskrats remain a
major source of income for fur trappers and are still among the most
prevalent species trapped for fur (Fur Institute of Canada, 2019).
While the muskrat played a major role in the early fur trade and
colonization of North America by Europeans (White et al., 2015), the
species has been of cultural significance as a traditional clothing and
food item and as a spiritual symbol among Indigenous people long before
European explorers arrived on the continent. For example, in an
Anishinaabe story of creation, the muskrat (Wa-zhushk) comes to the
rescue to help re-build the Earth after a great flood and decimation of
life, and is said to embody humility, courage, and determination
(MacGregor, 2013).
As a wetland obligate and dominant herbivore, the muskrat plays several
important roles in wetland ecosystems. Their foraging, travel, and
house-building activities create numerous small openings in marsh
vegetation, thereby increasing the interspersion of open water and
emergent vegetation which often results in increased structural
diversity and plant species richness in wetlands (Nyman et al., 1993;
Connors et al., 2000; Keddy, 2010). Different biotic communities are
known to respond positively to such enhancements in habitat diversity
(Wilcox and Meeker, 1992); in particular, a greater density and
diversity of marsh birds and waterfowl has been found in wetlands with
an equal ratio of open water to emergent vegetation (Weller and
Fredrickson, 1973; Kaminski and Prince, 1981; McDonnell, 1983). Muskrat
houses (both new and old) can also create important loafing and nesting
sites for marsh birds. For example, the black tern (Chlidonias
niger ) will use muskrat houses and feeding platforms as nesting
substrate (Hickey and Malecki, 1997) and it is quite common to find
Canada geese (Branta canadensis ) and sometimes trumpeter swans
(Cygnus buccinator ) loafing or nesting on old muskrat structures.
Less recognized are the benefits that muskrats provide to snakes and
turtles, many of which are species at risk. In an Illinois study, a
large number of spotted turtle (Clemmys guttata ) captures
occurred in deep open-water pools associated with muskrat lodges where
muskrat grazing decreased the vegetative cover, and it was believed that
these pools served as refugia for the turtles during periods of high
temperature and/or drought (Litzgus and Brooks, 2000). Numerous
vertebrate species, including more than ten herptile species, have been
observed using muskrat houses, burrows, cleared pathways and other
features for thermoregulation, nesting, cover and ease of travel (S.
Gillingwater, unpublished data; Kiviat, 1978). Furthermore, muskrat
activity can influence mussel abundance (Diggins and Stewart, 2000),
invertebrate communities (De Szalay and Cassidy, 2001; Nummi et al.,
2006), microbial activity (Wainscott et al., 1990) and nutrient cycling
(Connors et al., 2000). Muskrats are also an important food item for
many predators, such as red foxes (Vulpes vulpes ), coyotes
(Canis latrans ), raccoons (Procyon lotor ), raptors and
especially mink (Neovison vison ) (McDonnell, 1983).
Early research on muskrats by Paul Errington and others illustrated the
density-dependent nature of muskrat populations and a variety of
abundance cycles (Errington, 1951; Clark and Kroeker, 1993; Erb et al.,
2000). In their analysis of almost one hundred individual time series of
muskrat harvest data from the Hudson’s Bay Company in Canada, Erb et al.
(2000) found that the mean period length of muskrat population cycles
differed between ecozones, ranging from 3.7 to 8.6 years, with the
shorter periods tending to occur at higher latitudes and in eastern
regions. In some cases, however, the time series did not exhibit any
periodicity. Evidence exists for both population-intrinsic factors
(e.g., social factors) and extrinsic factors (e.g., disease,
environmental variability and trophic interactions) explaining the
observed patterns of these cycles (Errington, 1963 and Errington et al.,
1963; Bulmer, 1974; Weller and Fredrickson, 1973). Despite some
commonalities these patterns lack consistency across geographic areas,
and in many cases muskrat population dynamics and their mechanisms of
regulation remain unclear.
The muskrat is a prolific species; females typically have 2-3 litters
per year and an average litter size of 6.5 kits (Boutin and Birkenholz,
1987). As well, most muskrats are able to breed the same year they are
born and have high dispersal capabilities. These demographic
characteristics make muskrats relatively resilient to harvest and other
population pressures (e.g., disease, predation) as a small number of
individuals can quickly multiply and enable population recovery. They
are also reasonably flexible in their habitat requirements and, as noted
by Errington (1951), muskrats often demonstrate a remarkable ability to
stay alive up to the very edges of what may be considered habitable
range.
Today, however, this marsh denizen may not be thriving like it once was.
There is a growing body of literature suggesting declines in muskrat
populations from myriad locations across North America over the past
10-20 years. Most recently Gregory et al. (2019) reported that muskrat
harvest on Prince Edward Island declined by more than half when
comparing the average harvest from 1977-1988 to 1988-2016. More broadly,
Ahlers and Heske (2017) analyzed harvest data from 1970 to 2012 across
the United States and, after controlling for pelt prices, found strong
evidence that muskrat populations declined during this time period.
Prior to that study, analyses by Roberts and Crimmins (2010) revealed a
75% decline in muskrat harvest across the northeastern United States
and eastern Canada from 1986 to 2006. This decline was thought to be
indicative of regional declines in muskrat abundance as the study’s
authors found that a previously strong correlation between harvest and
pelt prices had weakened in the latter years of their analysis. In other
words, they felt it was likely that recent changes in muskrat harvests
were reflecting underlying population change (and not simply changes in
harvest effort) because they did not find the strong relationship
between harvest levels and pelt prices which had previously defined the
harvest dynamics of muskrats and other furbearer populations
(Scognamillo and Chamberlain, 2006; Bailey, 1981). As well, in the same
study the authors reported a lack of periodic fluctuations in the modern
muskrat harvest data (as compared to the historic data which exhibited
mild periodicity), providing further support for widespread population
decline (Roberts and Crimmins, 2010).
In Ontario, similar to trends across Canada, provincial fur harvest
records show a clear decline in muskrat harvest over the past 100 years,
most notably over the past 30 years (Fig. 1). In fact, the average
annual number of muskrats harvested in the past 30 years has declined by
more than 90% from the mean in the previous 20-year period (late 1960s
to late 1980s). While some of this decline can be explained by changes
in fur harvest reporting structure that occurred in Ontario in the late
1980s, as well as economic factors and cultural shifts in trapping, the
magnitude and time span of the decline seems too great to be explained
by these factors alone. For example, the spatial pattern of change in
muskrat harvests from 1972 to 2004 suggests an actual decline in muskrat
population may have occurred because the pattern did not conform to the
expectation arising from spatially homogeneous decline in fur price and
therefore trapper effort (Gorman 2007). Furthermore, the time span of
the recent period of low muskrat harvest numbers (30 years) far exceeds
that of the maximum population cycle length reported for muskrats (Erb
et al., 2000). As well, the low annual numbers harvested in recent years
are well below the lower limit of published historic muskrat population
fluctuations (Statistics Canada, 2011).
Since many furbearers (including muskrats) are difficult and costly to
census (Erb and Perry, 2003), harvest data are often the only type of
information related to species’ population trends that are available for
a long time series or across a large geographic area (White et al.,
2015). However, there are problems with relying on harvest data to infer
wildlife population trends, and these have been noted by many of the
same authors that have analyzed harvest trends of furbearers (e.g.,
Gregory, 2019; Roberts and Crimmins, 2010). For example, harvest data
can encompass trapping seasons of different lengths, can be missing data
for some years, or can be deficient due to lack of full reporting. More
importantly, harvest data can be biased by trapping effort which can be
affected by the number of trappers or by changing economic factors
(e.g., pelt price) that may encourage or discourage trappers (Landholt
and Genoways, 2000).
Although long-term census data on muskrat populations are generally
rare, a few researchers have published evidence of declines in local
muskrat populations from direct count surveys. For example, Benoit and
Askins (1999) reported that counts of muskrat houses from marshes of the
Quinnipiac and Connecticut Rivers decreased dramatically (78% and 100%
respectively) between 1965 and 1990. As well, Ward and Gorelick (2018)
analyzed muskrat house count records from 1970 to 2016 for the
Peace-Athabasca Delta in Canada and found a significant decline in
population density over this time frame.
Other recent studies, while not long-term in nature, have reported
finding lower than expected muskrat densities in areas of seemingly
suitable habitat. For example, Toner et al. (2010) reported a mean
muskrat density of 0.71 houses/ha in six coastal wetlands along the
Upper St. Lawrence River surveyed multiple times from 2001 to 2006, and
Greenhorn et al. (2017) found a mean muskrat density of 0.27 houses/ha
in 43 cattail-dominated marshes surveyed along the north shore of Lake
Ontario in 2014. In 2009, Gregory et al. (2019) intensively searched
four marshes on Prince Edward Island for muskrats and reported a mean
density of just 0.07 houses/ha. All of these results are much lower than
the typical densities of 2.1 – 3.6 muskrat houses/ha reported
previously in the literature for cattail-dominated marshes in Canada
(Proulx and Gilbert, 1984; Messier and Virgil, 1992).
Anecdotal and interview-based reports of muskrat decline over the past
few decades have also been reported by trappers (Gregory et al., 2019),
fur managers (Roberts and Crimmins, 2010) and Indigenous people across
Canada (Brietzke, 2015; Straka et al., 2018; S. Mallany, personal
communication, December 2016), and have also been received from a
variety of sources by us. The underlying theme among these reports is
that trappers and other long-term land users are simply not finding
muskrats in the numbers they used to, despite efforts to do so.
Collectively these recent studies and reports point to potential
widespread declines in muskrat populations. However, robust empirical
data on long-term trends in muskrat populations based on field
observations are generally scarce in the literature. Thus, more
information is needed to confirm that declines in muskrat harvest
correspond to real declines in muskrat abundance (Ahlers and Heske,
2017). Fortunately, we learned of annual muskrat field surveys having
been conducted at multiple locations in Ontario, Canada between 1950 and
1990 and sought to exploit this untapped source of historic muskrat
population data.
Our specific objectives in this study were to locate and revisit sites
where historic muskrat survey data exists for Ontario, to replicate the
historic survey methods as closely as possible, and to then compare
contemporary survey results with the historic data to determine if there
have been empirical changes in the muskrat populations in these areas.
We hypothesized that declines observed in muskrat harvest are due at
least in part to real declines in muskrat populations. Therefore, we
expected to observe evidence of fewer muskrats in contemporary versus
historic muskrat surveys.