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Survival of Translocated Clubshell and Northern Riffleshell in Illinois

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Translocation of freshwater mussels is a conservation tool used to reintroduce extirpated populations or augment small populations. Few studies have evaluated the effectiveness of translocations, mainly because estimating survival is challenging and time-consuming. We used a mark-recapture approach to estimate survival of nearly 4,000 individually marked Clubshell (Pleurobema clava) and Northern Riffleshell (Epioblasma rangiana) translocated to eight sites over a five-year period into the Salt Fork and Middle Fork Vermilion rivers in central Illinois. Survival differed among sites and between species; Clubshell were approximately five times more likely to survive than Northern Riffleshell. Survival also increased in the fourth year following a release and decreased following high-flow events. Translocating numerous individuals into multiple sites over a period of years could spread the risk of catastrophic high-flow events and maximize the likelihood for establishing self-sustaining populations.
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Freshwater Mollusk Biology and Conservation 20:89–102, 2017
ÓFreshwater Mollusk Conservation Society 2017
REGULAR ARTICLE
SURVIVAL OF TRANSLOCATED CLUBSHELL AND
NORTHERN RIFFLESHELL IN ILLINOIS
Kirk W. Stodola, Alison P. Stodola, and Jeremy S. Tiemann*
Illinois Natural History Survey, 1816 South Oak Street, Champaign, IL 61820 USA
ABSTRACT
Translocation of freshwater mussels is a conservation tool used to reintroduce extirpated
populations or augment small populations. Few studies have evaluated the effectiveness of
translocations, mainly because estimating survival is challenging and time-consuming. We used a
mark-recapture approach to estimate survival of nearly 4,000 individually marked Clubshell
(Pleurobema clava) and Northern Riffleshell (Epioblasma rangiana) translocated to eight sites over a
five-year period into the Salt Fork and Middle Fork Vermilion rivers in central Illinois. Survival
differed among sites and between species; Clubshell were approximately five times more likely to
survive than Northern Riffleshell. Survival also increased in the fourth year following a release and
decreased following high-flow events. Translocating numerous individuals into multiple sites over a
period of years could spread the risk of catastrophic high-flow events and maximize the likelihood for
establishing self-sustaining populations.
KEY WORDS: reintroduction, freshwater mussel, high flow, PIT tag, unionids
INTRODUCTION
North American freshwater mussels have undergone
drastic population declines during the past century and are
one of the most imperiled groups of animals in the world
(Williams et al. 1993; Lydeard et al. 2004; Strayer et al. 2004).
Translocation has been used for decades to augment
populations or reintroduce mussels into regions where species
have declined or are extirpated (Coker 1916; Ahlstedt 1979;
Sheehan et al. 1989). Much time and effort is placed on
collecting, marking, and transporting mussels for transloca-
tion, but few studies have evaluated the effectiveness of
mussel reintroductions. More than a quarter of all translocation
projects conducted prior to 1995 failed to report on the
efficacy of those efforts (Cope and Waller 1995).
Obtaining precise and unbiased estimates of mussel
survival is challenging, even for translocated individuals.
Mussels often burrow beneath the substrate surface when not
actively feeding or reproducing, making them difficult to
detect (Amyot and Downing 1998; Watters et al. 2001; Strayer
and Smith 2003). Furthermore, an unequal proportion of the
population is often sampled, such as larger individuals, those
found in easy-to-sample areas, or those at or near the surface
(Strayer and Smith 2003; Meador et al. 2011). Reliable
estimates of survival can be obtained using capture-mark-
recapture techniques (Hart et al. 2001; Meador et al. 2011).
Capture-mark-recapture methods are often time-intensive due
to the effort needed to capture and mark a large number of
individuals, but marking individuals already captured for
translocation can be easily incorporated.
The federally endangered Clubshell (Pleurobema clava)
and Northern Riffleshell (Epioblasma rangiana) were former-
ly widespread in the Ohio River and Great Lakes basins but
have experienced significant range reductions during the last
century. The recovery plan for the Clubshell and Northern
Riffleshell set objectives of reestablishing viable populations
in 10 separate river drainages across the species’ historical
range via augmentation and reintroduction (USFWS 1994).
Bridge construction on the Allegheny River, Pennsylvania,
which supports large populations of both species, prompted a
salvage operation to remove thousands of individuals from the
impacted area. In an attempt to meet recovery plan objectives,
these individuals were translocated to multiple streams within
seven states where the species had declined or had been
extirpated.
*Corresponding Author: jtiemann@illinois.edu
89
Beginning in 2006, the Illinois Department of Natural
Resources and the Illinois Natural History Survey partnered
with the U.S. Fish and Wildlife Service and state agencies in
Ohio, Pennsylvania, and West Virginia to translocate Club-
shell and Northern Riffleshell from the Allegheny River to the
Vermilion River system (Wabash River basin) in Illinois,
where both species occurred historically (Cummings and
Mayer 1997; Tiemann et al. 2007). Pilot translocations (n,
75 individuals) first occurred in 2010 at one site each in the
Salt Fork and Middle Fork Vermilion rivers, and more
widespread translocations occurred at eight sites in 2012,
2013, and 2014. We conducted a five-year capture-mark-
recapture study focusing on those individuals released in 2012,
2013, and 2014 to estimate survival of translocated mussels.
Specifically, our goals were to evaluate (1) how survival
differed according to species, sex, and mussel size, (2) how
survival varied spatially (among sites and between rivers), and
(3) how survival varied temporally after release.
METHODS
Mussel Collection and Transportation
Mussels were collected from the Allegheny River at the
U.S. Highway 62 Bridge, Forest County, Pennsylvania. The
Allegheny River at this site is approximately 200 m wide and
drains an area of approximately 10,000 km
2
. Mean daily
discharge is approximately 56 m
3
/s at the end of August and
nearly 425 m
3
/s at the beginning of April (average of 71 yr;
USGS gage 03016000). We collected 197, 758, and 807
Clubshell and 957, 249, and 777 Northern Riffleshell in 2012,
2013, and 2014, respectively. We measured total length of
each individual as the greatest distance from the anterior to
posterior shell margin (nearest 1 mm), and affixed a 12.5 mm,
134.2 kHz PIT tag (BioMark, Inc., Boise, Idaho) to the right
valve and a uniquely numbered HallPrint Shellfish tag
(HallPrint, Hindmarsh Valley, South Australia) to the left
valve. Northern Riffleshell averaged 45.6 mm long (range 15–
70 mm) and Clubshell averaged 52.2 mm long (range 18–84
mm). We also determined the sex of each Northern Riffleshell
based on shell morphology, although a few smaller individuals
were classified as ‘‘unknown’’ (male:female ratio ¼1.34:1);
Clubshell sexes cannot be differentiated by external shell
morphology and were all classified as ‘‘ unknown.’’ Clubshell
and Northern Riffleshell were placed in coolers between damp
towels and transported in climate-controlled vehicles to
Illinois.
Mussel Translocation and Release
We selected release sites based on the presence of
presumably suitable habitat for Northern Riffleshell and
Clubshell, which consisted of clean, stable sand, gravel, and
cobble riffles (Watters et al. 2009), abundant and diverse
mussel populations (INHS 2017), and presence of suitable host
fishes (i.e., darters and minnows) for both mussel species
(Cummings and Mayer 1992; Tiemann 2008a, 2008b; Watters
et al. 2009). Based on these criteria, we selected four sites each
in the Salt Fork and Middle Fork Vermilion rivers in east-
central Illinois (Fig. 1). These streams are an order of
magnitude smaller than the Allegheny River, each 30–40 m
wide and draining approximately 1,100 km
2
. Mean daily
discharge in the Salt Fork is 0.4 m
3
/s at the end of August and
4.3 m
3
/s at the beginning of April (average of 45 yr; USGS
gage 03336900); mean daily discharge in the Middle Fork is
0.9 m
3
/s at the end of August and 8.5 m
3
/s at the beginning of
April (average of 38 yr; USGS gage 03336645).
We released 3,745 mussels (both species combined)
among all eight sites from 2012 to 2014 (Table 1). Mussels
were released in the late summer, following a quarantine and
acclimatization period (14 d for 2012 mussels and 4–5 d for
2013–2014 mussels, differences between years due to
logistics). We hand-placed mussels into the substrate at each
site within an area demarcated by site-specific landmarks (such
as trees, boulders, water willow beds, or other discernible
feature) to facilitate recapture surveys. The size of marked
release areas varied with site and were between 3–10 m wide
and 20–100 m long. Sites with greater suitable area received
more mussels, but all sites were stocked at less than 50% of
the density observed at the collection site on the Allegheny
River, which is 5.5/m
2
for Northern Riffleshell and 7.5/m
2
for
Figure 1. The Clubshell and Northern Riffleshell release sites in the Vermilion
River basin (Wabash River drainage), Illinois.
STODOLA ET AL.90
Clubshell (Enviroscience, Inc., personal communication);
these densities are similar to those seen for these species at
other locations (Crabtree and Smith 2009). We stocked
Clubshell at greater densities than Northern Riffleshell due
to presumed historical presence based on historical shell
collection records (INHS 2017). Logistical constraints (e.g.
land access, previous stocking, mussel availability) largely
dictated which sites received mussels in multiple years.
Field Surveys
We surveyed for PIT-tagged Clubshell and Northern
Riffleshell during 12 sampling periods from 2012 to 2016
(Appendix 1). We used a robust design sampling protocol that
included primary and secondary samples (Fig. 2; Kendall and
Nichols 1995; Kendall et al. 1997). We attempted to conduct
primary samples every 3–4 mo to represent each season
(spring, summer, autumn, winter), but environmental condi-
tions prevented us from collecting all samples during every
year. We used two to three observers during each primary
sample. Each observer was considered an independent sample
and represented a secondary sample in the robust design
framework. We detected PIT-tagged mussels using BioMark
FS2001F-ISO or BioMark HPR Plus receivers with portable
BP antennas (BioMark). Each observer independently tra-
versed the stream in a systematic manner from a unique
starting point while slowly sweeping the streambed with an
antenna. Surveys continued until the release site was covered
completely and extended 5–10 m downstream after detections
ceased. Each sample typically required 2–3 h/site.
Statistical Analyses
We used the Huggins Robust Design model (Huggins
1989, 1991) to estimate apparent survival while accounting for
imperfect detection and to estimate of the numbers of
individuals remaining after each sampling period. Population
estimates from the Huggins Robust Design model (Huggins
1989, 1991) are derived using the actual number of individuals
observed during a primary sample and detection probability.
We were interested in the influence of individual traits (sex,
length, and species), environmental factors (site within river
and whether or not flood events had occurred between primary
sampling periods), and number of years following release on
survival. We fit a single model that included all covariates
instead of fitting a suite of models and comparing model fit
(Burnham and Anderson 2002). Consequently, we attained
estimates for each species released at each site during each
year by estimating a species effect, site effect, and an effect of
years following release, along with the individual covariates of
sex and length and the environmental covariate of the presence
of a flood. We did not include group (site or species) by
sampling period interactions because we had no reason to
believe that survival would vary along that spatio-temporal
scale (Anderson and Burnham 2002). We constrained our
model so there was no immigration or emigration between
primary samples, which we believed was biologically
reasonable given the limited vagility of freshwater mussels
(Amyot and Downing 1998; Schwalb and Pusch 2007). We fit
detection as a function of sampling period and site to
encompass differences in sampling efficiency due to variation
in flow, temperature, and depth among dates and variation in
habitat conditions among sites. We did not account for
species-specific differences in detection because we used PIT
tags and hand-held readers for both species and did not believe
detection would differ by species when using this method.
Table 1. Number of Clubshell and Northern Riffleshell released into the Salt Fork and Middle Fork Vermilion rivers in 2012, 2013, and 2014.
Site
2012 2013 2014
Clubshell Riffleshell Clubshell Riffleshell Clubshell Riffleshell
Salt Fork
1 - 291 - - - -
2 106 196 258 - - -
3 91 470 250 - - -
4 - - 50 50 277 290
Middle Fork
5--5050--
6 - - 50 50 175 180
7 - - 50 50 181 174
8 - - 50 49 174 133
Totals 197 957 758 249 807 777
Figure 2. Robust design as employed in this study, with primary samples
(seasons) and secondary samples (observers).
SURVIVAL OF TRANSLOCATED MUSSELS 91
Post hoc analyses indicated that inclusion of species-specific
detection had very little influence on survival probabilities
(i.e., estimates were within 0.01%). We determined if a flood
occurred between primary samples using the Indicators of
Hydrologic Alteration software package (IHA; Richter et al.
1996) and discharge data for both streams from the U.S.
Geological Survey National Water Information System
(https://waterdata.usgs.gov/il/nwis/rt; gages 03336900 and
03336645). We did not differentiate between small floods
and large floods as identified by IHA, and anything equivalent
to or greater than a 2-yr flood event was considered a flood.
We used the Huggins’ p and c extension in Program MARK
(White and Burnham 1999) with initial capture probability (p,
probability of detecting an individual at least once during a
primary sample) equal to recapture probability (c, probability
of detecting an individual during a primary sample given it is
detected) because secondary samples occurred via the same
method on the same day. We interpreted the strength and
biological meaning of each model covariate using the beta
coefficients (b) and their 95% confidence intervals and log-
odds ratios, which approximate how much more likely it is for
an event (survival) to occur based on the beta coefficient (log-
odds ratio ¼e
b
, Gerard et al. 1998; Hosmer and Lemeshow
2010).
RESULTS
Detection rate averaged 0.78 across both species (range of
averages ¼0.66–0.90; Appendix 1). Detection was generally
greatest in autumn. Average detection in autumn samples was
about 1.25 times greater than for spring and summer samples;
we had only one winter sample because of high flows and
frozen conditions. However, detection probabilities were
highly variable among sites and sampling periods (Appendix
1).
Monthly survival varied among species, sites, and
sampling periods. Average monthly survival was 0.981 for
Clubshell and 0.905 for Northern Riffleshell; these values
translate to an approximate annual survival of 0.79 for
Clubshell and 0.30 for Northern Riffleshell, irrespective of
site, individual traits, and years following release. The b
coefficient and log-odds ratio showed that, overall, Clubshell
was approximately 5 times more likely to survive than
Northern Riffleshell, but the precision of this estimate was
low (95% confidence interval ¼1.57–18.003; Table 2). There
was no difference in survival among males, females, and
mussels of unknown sex; confidence intervals included zero
for all coefficients (Table 2). There was no appreciable effect
of size on survival. The log-odds ratio indicated that
individuals were 1.009 times more likely to survive (95%
confidence interval ¼1.003–1.016) for every mm increase in
length (Table 2).
Survival was greatest at Sites 1 and 4 on the Salt Fork and
lowest at Site 7 on the Middle Fork (Figs. 3–6). Log-odds
ratios showed that mussels were nearly 6 times less likely to
survive at Site 7 than Site 1, and mussels were 2–4 times less
likely to survive at Sites 2, 3, 5, and 6 (Table 2). Survival was
reduced following floods. The log-odds ratio showed that
Table 2. Parameter estimates (bcoefficients), standard errors (SE), log-odds (e
b
), and log-odds lower and upper 95% confidence limits (CL) of monthly survival of
translocated Clubshell and Northern Riffleshell relative to site, years following release, species, sex, mussel length, and presence of flood between primary
samples. Parameter estimates should be interpreted in relation to the baseline, which was Northern Riffleshell of average length and unknown sex at Site 1, four
years postrelease, and during a period with no flooding, as indicated.
Parameter Estimate SE Log-odds Lower CL log-odds Upper CL log-odds
Intercept 4.760 0.891
Individual traits
Clubshell versus Riffleshell 1.670 0.623 5.312 1.567 18.011
Male versus unknown 0.207 0.620 1.230 0.365 4.150
Female versus unknown 0.117 0.621 0.890 0.263 3.004
Length 0.009 0.004 1.009 1.003 1.016
Environmental factors
Site 2 versus Site 1 0.853 0.085 0.426 0.361 0.504
Site 3 versus Site 1 1.402 0.079 0.246 0.211 0.287
Site 4 versus Site 1 0.007 0.165 0.993 0.718 1.374
Site 5 versus Site 1 0.999 0.130 0.368 0.286 0.475
Site 6 versus Site 1 1.063 0.132 0.345 0.267 0.448
Site 7 versus Site 1 1.757 0.128 0.173 0.134 0.222
Site 8 versus Site 1 0.958 0.142 0.384 0.290 0.507
Flood versus No Flood 0.530 0.077 0.589 0.506 0.685
Years following release
Year 1 versus Year 4 1.260 0.658 0.284 0.078 1.030
Year 2 versus Year 4 1.666 0.661 0.189 0.052 0.691
Year 3 versus Year 4 1.228 0.660 0.293 0.080 1.066
STODOLA ET AL.92
Figure 3. Derived estimates of proportion of Clubshell remaining at each release site in the Middle Fork from 2012 to 2016. Gray boxes indicate when a flood
occurred. Numbers of individuals released per year per site can be viewed in Table 1.
Figure 4. Derived estimates of proportion of Clubshell remaining at each release site in the Salt Fork from 2012 to 2016. Gray boxes indicate when a flood
occurred. Numbers of individuals released per year per site can be viewed in Table 1.
SURVIVAL OF TRANSLOCATED MUSSELS 93
Figure 5. Derived estimates of proportion of Northern Riffleshell remaining at each release site in the Middle Fork from 2012 to 2016. Gray boxes indicate when a
flood occurred. Numbers of individuals released per year per site can be viewed in Table 1.
Figure 6. Derived estimates of proportion of Northern Riffleshell remaining at each release site in the Salt Fork from 2012 to 2016. Gray boxes indicate when a
flood occurred. Numbers of individuals released per year per site can be viewed in Table 1.
STODOLA ET AL.94
mussels were 1.70 times less likely to survive after floods
(95% confidence interval: 1.46–1.98) than after periods with
no floods; this is equivalent to a reduction of monthly survival
from 0.950 to 0.917 (average of all species and sites). The
occurrence of a flood on the Middle Fork during June–July
2015 was associated with a sharp decline in population size for
both species (Figs. 3, 5), but the influence of other flood events
was not associated with similar declines. We did not model
river as a separate factor (see Methods), but survival appeared
to be greater in the Salt Fork than in the Middle Fork. An
average of 62% of Clubshell and 19% of Northern Riffleshell
were alive in the Salt Fork in 2016 compared with only 21% of
Clubshell and 4% of Northern Riffleshell in the Middle Fork in
2016 (Figs. 3–6). This difference was apparent despite the fact
that most mussels were translocated to the Salt Fork 1–2 yr
earlier than in the Middle Fork (Table 1).
Number of years following release was an important
determinant of survival. Survival was greatest in the fourth
year following a release; individuals were 3.52 times more
likely to survive in the fourth year following release (95%
confidence interval: 0.97–12.80) compared to the first year
following release (Table 2). Survival was lowest in the second
year following release; individuals were 1.50 times less likely
to survive (95% confidence interval: 1.30–1.70) compared to
the first year (Table 2).
DISCUSSION
The long-term efficacy of a reintroduction program
depends on the establishment of a self-sustaining population,
which requires translocated individuals to survive until they
reproduce and replace themselves. It is too early to tell if the
Clubshell and Northern Riffleshell reintroduction program into
Illinois has been a success because no recruitment has been
documented. Reintroduction of the Clubshell appears to have
been more successful initially than reintroduction of Northern
Riffleshell. Reintroduced Clubshell survived at a much greater
rate and represented the majority of individuals remaining after
five years of monitoring. Annual survival for Clubshell (0.79)
is within the estimated range for other mussel species in the
wild, (0.50–0.99, Hart et al. 2001; Villella et al. 2004) and near
the estimates of the closely related Southern Clubshell
(Pleurobema decisum) (0.91, Haag 2012). However, annual
survival for Northern Riffleshell (0.30) was well below those
values, those reported from French Creek, Pennsylvania,
which averaged 0.60 (Crabtree and Smith 2009), and those of
the closely related Oystermussel (Epioblasma capsaeformis)
(0.73, Jones and Neves 2011; Haag 2012).
Some species may be inherently more difficult to
translocate. There is high variability in the success of
translocation projects, ranging from nearly all individuals
remaining after a few years to very few if any (e.g., Ahlstedt
1979; Sheehan et al. 1989; Cope et al. 2003). Some of this
variation may be explained by inherent life history differences
among species, and Clubshell probably lives longer than
Northern Riffleshell. For instance, the Southern Clubshell, a
congener of Clubshell, can reach 45 yr of age (Haag and Rypel
2011), while Northern Riffleshell is a relatively short-lived
species with a maximum age reported in French Creek,
Pennsylvania, of 11 yr (Crabtree and Smith 2009). Based on
these differences, Northern Riffleshell is expected to have
lower survival than Clubshell even in wild populations, and
our data show that translocated populations may have even
lower survival. Consequently, translocation of short-lived
species such as Northern Riffleshell may require larger
numbers of individuals and repeated translocations to
overcome high mortality and ensure that translocated individ-
uals experience conditions favorable for recruitment.
Differences in hydrology, either between rivers or even
within the same river, may play an important role in
determining the suitability of sites for freshwater mussel
reintroduction (Cope et al. 2003; Carey et al. 2015). The
hydrology, land use, and watershed size of the Vermilion
River basin differ from the source location of the Allegheny
River (Larimore and Smith 1963; Smith 1968; Larimore and
Bayley 1996; White et al. 2005), thus some discrepancy in
survival between the source and recipient basins may be
expected. However, the Salt Fork Vermilion and Middle Fork
Vermilion rivers are comparable in size and have similar land
use and hydrology, yet we found that survival varied even
among sites within a river. Local-scale differences among
sites, such as substrate or gradient, can lead to biologically
significant differences that influence survival (McRae et al.
2004). We selected release sites based on the best available
habitat and species assemblage data, yet unmeasured habitat
differences and stochastic events appeared to have a large
effect on survival. Similar results have been observed in other
translocations, such as siltation due to bank failure following
flow diversion (Bolden and Brown 2002), possible washout
due to earthen causeway removal (Tiemann et al. 2016), or
diminished recovery of relocated individuals in sites with high
current velocity in the two years following relocation (Dunn et
al. 2000).
High-discharge events present an ongoing threat to the
reintroduction of Clubshell, Northern Riffleshell, and similar
translocation projects. High-flow events have been problem-
atic in other translocation projects (e.g., Sheehan et al. 1989;
Carey et al. 2015) and were clearly detrimental for
translocated Clubshell and Northern Riffleshell. Following
the flood in June–July 2015, we examined the nearest
downstream gravel bar at a few sites and found numerous
stranded and dead individuals. Existing native mussel
communities in the Salt and Middle Fork Vermilion rivers
have persisted throughout similar high-flow events, but
translocated mussels may be at a disadvantage. PIT tags
can decrease the burrowing rate of individuals (Wilson et al.
2011), and translocated mussels may have lower energetic
status (Patterson et al. 1997), which could reduce their ability
to anchor themselves in the substrate or rebury after a flood
event (Killeen and Moorkens 2016). Additionally, the native
mussel community represents individuals that have found
optimal locations to withstand scouring and dislodging. The
SURVIVAL OF TRANSLOCATED MUSSELS 95
Clubshell and Northern Riffleshell we translocated may not
have had enough time to find optimal locations, which may
have made them more vulnerable to dislodgement and may
partly explain why individuals survived at a greater rate 4 yr
following release.
We provide the following recommendations for conducting
and monitoring reintroduction efforts. The best time to
monitor Clubshell and Northern Riffleshell was during
autumn, when stream flows were low and we observed the
greatest probability of detection. Sampling was difficult or
impossible during the spring because of high stream flows,
which resulted in reduced detectability using handheld readers;
sampling also was difficult in winter because of high flows and
occasional ice cover. Spreading reintroduction efforts over
several geographically separate river systems could lessen risk
of failure due to stochastic events such as floods, chemical
spills, and biological invasion (e.g., Griffith et al. 1989; Trdan
and Hoeh 1993). Translocating individuals over a period of
several years might also reduce the overall risk of failure due
to isolated events occurring in a particular year. For instance,
many Clubshell and Northern Riffleshell, especially in the
Middle Fork, were lost during a late spring/early summer high-
flow event in 2015. Finally, stocking greater numbers of
individuals in multiple translocations for species with naturally
low annual survival, such as Northern Riffleshell, may be
necessary to maximize chances for natural recruitment.
ACKNOWLEDGMENTS
This project is a collaborative effort among the U.S. Fish
and Wildlife Service (USFWS); Pennsylvania Fish and Boat
Commission (PFBC); Pennsylvania Department of Trans-
portation; Illinois Department of Natural Resources (IDNR),
including the Illinois Nature Preserves Commission and the
Illinois Endangered Species Protection Board; Illinois
Natural History Survey; University of Illinois, Urbana-
Champaign; Champaign County Forest Preserve District;
the Ohio State University; Columbus Zoo and Aquarium;
West Virginia Department of Natural Resources; Indiana
Department of Natural Resources; Kentucky Department of
Fish and Wildlife; and EnviroScience, Inc. Permits were
provided by the USFWS (no. TE73584A-1); PFBC (e.g., no.
2014-02-0837, no. 2013-756); IDNR (e.g., no. SS16-047, no.
S-10-30); the Illinois Nature Preserves Commission; and the
University of Illinois. Funding was provided in part by the
USFWS (through the IDNR’s Office of Resource Conserva-
tion to the Illinois Natural History Survey, Grant no.
R70470002 and no. RC09-13FWUIUC); the USWFS’s Ohio
River Basin Fish Habitat Partnership (Award no.
F14AC00538); the IDNR (through the Natural Resource
Damage Assessment settlement: Heeler Zinc–Lyondell
Basell Companies, Reference Document no. OREP1402
and no. OREP1504); the Illinois Wildlife Preservation Fund
(Grant no. RC07L25W); and the Illinois Department of
Transportation.
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SURVIVAL OF TRANSLOCATED MUSSELS 97
Appendix 1. Estimates of detection for each site and during each period; 95% confidence intervals are provided in parentheses.
Sample Period
Middle Fork Salt Fork
Site 1 Site 2 Site 3 Site 4 Site 5 Site 6 Site 7 Site 8
Summer 2012 --------
Autumn 2012 0.71 (0.68–0.74) 0.67 (0.64–0.71) 0.68 (0.64–0.72) - ----
Summer 2013 0.72 (0.68–0.75) 0.68 (0.63–0.73) 0.69 (0.63–0.74) - ----
Autumn 2013 0.79 (0.77–0.81) 0.76 (0.74–0.70) 0.76 (0.72–0.80) 0.87 (0.85–0.89) 0.83 (0.80–0.85) 0.77 (0.73–0.80) 0.81 (0.77–0.85) 0.85 (0.82–0.88)
Winter 2014 - - - 0.80 (0.76–0.84) 0.84 (0.80–0.88) - 0.83 (0.78–0.87) -
Spring 2014 ----0.76 (0.72–0.80) 0.69 (0.63–0.74) 0.71 (0.66–0.76) 0.79 (0.75–0.84)
Summer 2014 0.70 (0.67–0.72) 0.66 (0.63–0.69) 0.67 (0.64–0.71) 0.81 (0.77–0.84) 0.75 (0.71–0.78) 0.67 (0.63–0.72) 0.73 (0.68–0.78) 0.78 (0.74–0.82)
Autumn 2014 - 0.75 (0.72–0.78) - 0.85 (0.81–0.87) 0.80 (0.76–0.83) 0.73 (0.68–0.77) 0.78 (0.73–0.82) 0.82 (0.78–0.86)
Spring 2015 - - - 0.72 (0.67–0.77) 0.77 (0.73–0.82) 0.70 (0.64–0.75) 0.75 (0.69–0.81) -
Summer 2015 0.80 (0.78–0.82) 0.78 (0.75–0.80) 0.78 (0.74–0.82) 0.88 (0.86–0.90) 0.84 (0.81–0.87) 0.78 (0.74–0.82) 0.83 (0.78–0.86) -
Autumn 2015 0.86 (0.84–0.87) 0.83 (0.81–0.85) 0.84 (0.80–0.87) 0.92 (0.90–0.93) 0.88 (0.86–0.91) 0.84 (0.80–0.87) 0.87 (0.84–0.90) 0.90 (0.88–0.92)
Spring 2016 0.78 (0.74–0.81) 0.75 (0.71–0.79) - 0.87 (0.83–0.89) 0.82 (0.78–0.86) - 0.81 (0.75–0.85) 0.85 (0.81–0.88)
STODOLA ET AL.98
Appendix 2. Monthly apparent survival estimates for Clubshell. Years (2012–2014) represent the year animals were released. Numbers in parentheses beside
primary sample indicate the number of months since the preceding sample; 95% confidence intervals are provided in parentheses beside survival estimates. Bold
rows indicate a flood occurred during that period (e.g., between Su 2013 and Au 2013). Sp ¼spring, Su ¼summer, Au ¼autumn, Wi ¼winter.
Primary
Samples (mo)
Salt Fork Vermilion River
Site 1 Site 2 Site 3 Site 4
2012 2013 2012 2013 2012 2013 2014
Su 2012–Au 2012 (2) 0.994
(0.993–0.995)
- 0.977
(0.974–0.981)
- 0.987
(0.984–0.989)
--
Au 2012–Su 2013 (9) 0.990
(0.989–0.992)
- 0.962
(0.956–0.967)
- 0.978
(0.973–0.982)
--
Su 2013–Au 2013 (2) 0.992
(0.990–0.993)
0.994
(0.993–0.995)
0.966
(0.962–0.971)
0.977
(0.974–0.981)
0.980
(0.976–0.984)
0.994
(0.992–0.996)
-
Au 2013–Wi 2014 (4) 0.992
(0.990–0.993)
0.994
(0.993–0.995)
0.966
(0.962–0.971)
0.977
(0.974–0.981)
0.980
(0.976–0.984)
0.994
(0.992–0.996)
-
Wi 2014–Sp 2014 (2) 0.992
(0.990–0.993)
0.994
(0.993–0.995)
0.966
(0.962–0.971)
0.977
(0.974–0.981)
0.980
(0.976–0.984)
0.994
(0.992–0.996)
-
Sp 2014–Su 2014 (2) 0.992
(0.990–0.993)
0.994
(0.993–0.995)
0.966
(0.962–0.971)
0.977
(0.974–0.981)
0.980
(0.976–0.984)
0.994
(0.992–0.996)
-
Su 2014–Au 2014 (4) 0.995
(0.993–0.996)
0.992
(0.990–0.993)
0.978
(0.973–0.982)
0.966
(0.962–0.971)
0.987
(0.983–0.990)
0.991
(0.988–0.994)
-
Au 2014–Sp 2015 (5) 0.995
(0.993–0.996)
0.992
(0.990–0.993)
0.978
(0.973–0.982)
0.966
(0.962–0.971)
0.987
(0.983–0.990)
0.991
(0.988–0.994)
0.994
(0.992–0.996)
Sp 2015–Su 2015 (3) 0.991
(0.988–0.993)
0.986
(0.983–0.988)
0.963
(0.955–0.97)
0.944
(0.934–0.953)
0.979
(0.972–0.983)
0.986
(0.980–0.990)
0.990
(0.986–0.993)
Su 2015–Au 2015 (3) 0.995
(0.993–0.996)
0.992
(0.990–0.993)
0.978
(0.973–0.982)
0.966
(0.962–0.971)
0.987
(0.983–0.990)
0.991
(0.988–0.994)
0.994
(0.992–0.996)
Au 2015–Sp 2016 (6) 0.997
(0.990–0.999)
0.991
(0.988–0.993)
0.989
(0.961–0.997)
0.963
(0.955–0.970)
0.994
(0.977–0.998)
0.991
(0.986–0.994)
0.986
(0.98–0.990)
SURVIVAL OF TRANSLOCATED MUSSELS 99
Appendix 2, extended.
Middle Fork Vermilion River
Site 5 Site 6 Site 7 Site 8
2013 2014 2013 2014 2013 2014 2013
-------
-------
0.985
(0.980–0.988)
- 0.984
(0.979–0.988)
- 0.968
(0.959–0.975)
- 0.985
(0.981–0.989)
0.985
(0.980–0.988)
- 0.984
(0.979–0.988)
- 0.968
(0.959–0.975)
- 0.985
(0.981–0.989)
0.985
(0.980–0.988)
- 0.984
(0.979–0.988)
- 0.968
(0.959–0.975)
- 0.985
(0.981–0.989)
0.985
(0.980–0.988)
- 0.984
(0.979–0.988)
- 0.968
(0.959–0.975)
- 0.985
(0.981–0.989)
0.977
(0.971–0.982)
- 0.976
(0.969–0.981)
- 0.953
(0.940–0.963)
- 0.978
(0.972–0.983)
0.977
(0.971–0.982)
0.985
(0.980–0.988)
0.976
(0.969–0.981)
0.984
(0.979–0.988)
0.953
(0.940–0.963)
0.968
(0.959–0.975)
0.978
(0.972–0.983)
0.962
(0.950–0.971)
0.974
(0.966–0.981)
0.960
(0.946–0.97)
0.973
(0.964–0.980)
0.922
(0.898–0.941)
0.947
(0.931–0.959)
0.964
(0.951–0.973)
0.977
(0.971–0.982)
0.985
(0.980–0.988)
0.976
(0.969–0.981)
0.984
(0.979–0.988)
0.953
(0.940–0.963)
0.968
(0.959–0.975)
0.978
(0.972–0.983)
0.975
(0.966–0.982)
0.962
(0.950–0.971)
0.974
(0.963–0.981)
0.960
(0.946–0.97)
0.953
(0.940–0.963)
0.922
(0.898–0.941)
0.976
(0.967–0.983)
STODOLA ET AL.100
Appendix 3. Monthly apparent survival estimates for Northern Riffleshell. Years (2012–2014) represent the year animals were released. Numbers in parentheses
beside primary sample indicate the number of months since the preceding sample; 95% confidence intervals are provided in parentheses beside survival estimates.
Bold rows indicate a flood occurred during that period (e.g., between Su 2013 and Au 2013). Sp ¼spring, Su ¼summer, Au ¼autumn, Wi ¼winter.
Primary Samples (months)
Salt Fork
Site 1 Site 2 Site 3 Site 4
2012 2013 2012 2013 2012 2013 2014
Su 2012–Au 2012 (2) 0.971
(0.907–0.991)
- 0.891
(0.706–0.965)
- 0.934
(0.806–0.98)
--
Au 2012–Su 2013 (9) 0.951
(0.852–0.985)
- 0.828
(0.586–0.942)
- 0.893
(0.711–0.966)
--
Su 2013–Au 2013 (2) 0.957
(0.867–0.987)
0.971
(0.907–0.991)
0.844
(0.614–0.949)
0.891
(0.706–0.965)
0.904
(0.735–0.97)
0.970
(0.904–0.991)
-
Au 2013–Wi 2014 (4) 0.957
(0.867–0.987)
0.971
(0.907–0.991)
0.844
(0.614–0.949)
0.891
(0.706–0.965)
0.904
(0.735–0.97)
0.970
(0.904–0.991)
-
Wi 2014–Sp 2014 (2) 0.957
(0.867–0.987)
0.971
(0.907–0.991)
0.844
(0.614–0.949)
0.891
(0.706–0.965)
0.904
(0.735–0.97)
0.970
(0.904–0.991)
-
Sp 2014–Su 2014 (2) 0.957
(0.867–0.987)
0.971
(0.907–0.991)
0.844
(0.614–0.949)
0.891
(0.706–0.965)
0.904
(0.735–0.97)
0.970
(0.904–0.991)
-
Su 2014–Au 2014 (4) 0.972
(0.909–0.991)
0.957
(0.867–0.987)
0.894
(0.71–0.967)
0.844
(0.614–0.949)
0.936
(0.809–0.98)
0.956
(0.862–0.987)
-
Au 2014–Sp 2015 (5) 0.972
(0.909–0.991)
0.957
(0.867–0.987)
0.894
(0.71–0.967)
0.844
(0.614–0.949)
0.936
(0.809–0.98)
0.956
(0.862–0.987)
0.970
(0.904–0.991)
Sp 2015–Su 2015 (3) 0.953
(0.855–0.986)
0.928
(0.793–0.978)
0.832
(0.59–0.944)
0.762
(0.483–0.916)
0.896
(0.715–0.967)
0.928
(0.785–0.979)
0.951
(0.846–0.986)
Su 2015–Au 2015 (3) 0.972
(0.909–0.991)
0.957
(0.867–0.987)
0.894
(0.71–0.967)
0.844
(0.614–0.949)
0.936
(0.809–0.98)
0.956
(0.862–0.987)
0.97
(0.904–0.991)
Au 2015–Sp 2016 (6) 0.986
(0.923–0.997)
0.953
(0.855–0.986)
0.944
(0.746–0.99)
0.832
(0.59–0.944)
0.967
(0.836–0.994)
0.952
(0.849–0.986)
0.928
(0.785–0.979)
SURVIVAL OF TRANSLOCATED MUSSELS 101
Appendix 3, extended.
Middle Fork
Site 5 Site 6 Site 7 Site 8
2013 2014 2013 2014 2013 2014 2013
-------
-------
0.924
(0.78–0.977)
- 0.920
(0.768–0.975)
- 0.851
(0.624–0.952)
- 0.927
(0.785–0.978)
0.924
(0.78–0.977)
- 0.920
(0.768–0.975)
- 0.851
(0.624–0.952)
- 0.927
(0.785–0.978)
0.924
(0.78–0.977)
- 0.920
(0.768–0.975)
- 0.851
(0.624–0.952)
- 0.927
(0.785–0.978)
0.924
(0.78–0.977)
- 0.920
(0.768–0.975)
- 0.851
(0.624–0.952)
- 0.927
(0.785–0.978)
0.890
(0.702–0.966)
- 0.884
(0.688–0.963)
- 0.792
(0.525–0.929)
- 0.894
(0.709–0.967)
0.890
(0.702–0.966)
0.924
(0.78–0.977)
0.884
(0.688–0.963)
0.920
(0.768–0.975)
0.792
(0.525–0.929)
0.851
(0.624–0.952)
0.894
(0.709–0.967)
0.827
(0.578–0.943)
0.878
(0.675–0.961)
0.818
(0.563–0.94)
0.871
(0.66–0.959)
0.691
(0.391–0.887)
0.771
(0.493–0.921)
0.833
(0.587–0.946)
0.890
(0.702–0.966)
0.924
(0.78–0.977)
0.884
(0.688–0.963)
0.920
(0.768–0.975)
0.792
(0.525–0.929)
0.851
(0.624–0.952)
0.894
(0.709–0.967)
0.881
(0.679–0.963)
0.827
(0.578–0.943)
0.874
(0.665–0.961)
0.818
(0.563–0.940)
0.776
(0.498–0.924)
0.691
(0.391–0.887)
0.885
(0.687–0.964)
STODOLA ET AL.102
... The importance and vulnerability of mussels has led to many attempts to augment and reintroduce populations using translocated individuals (Haag and Williams 2014); however, the success of these translocations are highly variable, long term monitoring data are lacking, and factors determining the success or failure of said translocations are poorly understood (Stodola et al. 2017, Tiemann et al. 2016, Bolden and Brown 2002, Cope and Waller 1995. ...
... These components may lend themselves to differential translocation success between sites. In addition to site differences, biology of different species may also impact translocation success (Stodola et al. 2017). ...
... A 1995 literature review reported an average 51% survival in translocated mussels (Cope and Waller 1995). Stodola et al. (2017) found that estimated survival of translocated mussels differed between their eight study sites (in Illinois) and two species (Clubshell (Pleurobema clava) and Northern Riffleshell (Epioblasma rangiana)). Similarly, in a short-distance relocation, species predicted survival and detection probabilities (Tiemann et al. 2016). ...
... For example, it is well known that relocating mussels to suitable habitat using appropriate handling techniques can improve survival of relocated populations (Hamilton, Brim Box & Dorazio, 1997;Chen, Heath & Neves, 2001;Cope et al., 2003;Tsakiris et al., 2017). However, there are cases in which suitable habitat is selected but low survivorship still occurs (Dunn, Sietman & Kelner, 1999;Tsakiris, 2016;Stodola, Stodola & Tiemann, 2017). This indicates that other factors can adversely affect performance of a relocated population. ...
... For example, Dunn, Sietman & Kelner (1999) observed higher mortality rates when mussels were relocated during cooler months, which they hypothesized was due to the inability of individuals to reburrow into compact substrate combined with rapidly declining water temperatures. Stodola, Stodola & Tiemann (2017) observed that translocated mussels in the Salt and Middle Forks of the Vermilion River, Illinois, were less likely to survive after floods than after periods with no floods. Sheehan, Neves & Kitchel (1989) made a similar observation in the North Fork of the Holston River, Virginia, noting that flood events were likely to have been responsible for the loss of translocated individuals. ...
... year of translocation activities are important considerations that may affect translocation success. The results from this study confirm that these factors are important, and that species-specific differences, time of year, large stochastic events such as extreme floods, and occurrence of predators should be considered when planning translocations.Stodola, Stodola & Tiemann (2017), evaluating translocation success of Pleurobema clava (clubshell) and Epioblasma rangiana (northern riffleshell) in the Salt Fork and Middle Fork Vermilion rivers came to a similar conclusion and further argued that some species are inherently difficult to translocate owing to life history differences. The authors suggested for these sp ...
Article
1. Translocation is used to conserve mussels, yet there remains a debate about its merit owing to poor understating of its effects on transported mussels. 2. This study evaluated survivorship, body condition, and total glycogen and lipids for one common and widely distributed species (Cyclonaias pustulosa), two rare species (Cyclonaias petrina; Lampsilis bracteata), and one species complex (Fusconaia sp.-Fusconaia chunii and Fusconaia flava) from the East Fork of the Trinity River and the Llano River of Texas. 3. Survivorship estimates for C. pustulosa and Fusconaia sp. using the Kaplan-Meier estimator were high in the East Fork. Body condition, glycogen, and total lipids varied for C. pustulosa and Fusconaia sp., which may have indicated a short-term impact. For the Llano, survivorship of C. petrina and L. bracteata was high for the resident treatments but significantly reduced for the translocation treatments. 4. The decline in survivorship for C. petrina was mirrored by decreases in the body condition, which may indicate inability to acclimate to novel environments. For L. bracteata, declines in survivorship were due to predation by Procyon lotor, racoon. A large flood of 3,766 m 3 s −1 at the end of the study eliminated both study sites. 5. The findings of this study show that translocating mussels can be successful; however , sublethal effects and mortality may still occur. These effects are rooted in species-specific differences, which is not unexpected because mussel species vary in how they cope with environmental change based on their life-history traits. However, these traits are rarely considered when translocating mussels. 6. To complicate matters, most mussel species have yet to be evaluated on how they respond to translocation, and for species where such information is available, adults are the primary focus. Addressing these knowledge gaps is critical for determining the appropriateness of translocation and improving its efficacy.
... Success may be defined as achieving the project targets, or meeting the conservation objectives for the species in question. Another definition of success is the establishment of a self-sustaining, longterm viable population (Griffith et al., 1989;Gusset, 2009;Tarszisz et al., 2014;McMurray and Roe, 2019) or a self-sustaining population requiring individuals to survive until they reproduce and replace themselves (Stodola et al., 2017). The number of mussel releases which have resulted in recruitment remains low, even after several decades of carrying out these activities (Paul Johnson, pers. ...
... Environmental stochasticity between different release years can have large effects on growth and survival (Jones et al., 2012;IUCN/SSC, 2013;Stodola et al., 2017;Černá et al., 2018) and so smaller releases over a number of years rather than a single release would help protect against high losses due to sub-optimal conditions in one particular year e.g. severe summer droughts, winter floods, pollution incidences etc. ...
... Spreading individuals over several sites and over several years helps to spread the risk of high losses due to stochastic events such as sub-optimal environmental conditions, high flows or pollution incidents (Haag, 2012;IUCN/SSC, 2013;Stodola et al., 2017;Černá et al., 2018). This practice also helps to protect against any genetic bias occurring if a single cohort is very successful in one particular year. ...
Technical Report
This review summarises the literature on freshwater mussel reintroduction/reinforcement studies carried out to date, highlights common themes and lessons learned and puts these in context for potential release strategies and activities in the River Irt. The number of releases of propagated M. margaritifera juveniles to date is low across Europe, although some of the significant examples of large-scale releases have been highlighted as specific case studies in this report. To date the Lutter project in Germany remains the only example of successful restoration of a naturally and sustainably recruiting Margaritifera population. This project involved the purchase of the catchment, removal of unsustainable land use practices and implementation of catchment-wide measures to improve habitat quality, thus restoring natural recruitment. Freshwater mussel conservation projects in the USA were generally started earlier than Margaritifera projects in Europe and thus provide an excellent source of experience, knowledge and information. Despite this, standardised reporting and long-term monitoring of relocated individuals is still not commonplace thus making assessment of suitable methods difficult. This review aims to provide information on the entire conservation translocation process and make it relevant to the current project and the target species, M. margaritifera. Topics covered in this literature review include assessment of appropriate conservation measures (are reintroductions/reinforcements an appropriate option?), planning reintroductions/reinforcements, goal setting, factors affecting the success of releases, release methods, implementation of conservation translocations, project monitoring, adaptive management and reporting. It is hoped that this literature review will provide a suitable starting point from which to begin discussions about appropriate action for the River Irt.
... Whilst not considered in this paper, genetic variation in juvenile cohorts reared for population reinforcements in the River Irt have been found to be genetically representative of the wild population [35]. The release of multiple cohorts across all sites and release dates is advised in order to protect against stochastic effects disproportionately impacting a particular cohort and any potential genetic bias which may arise due to unequal contributions from females in any one cohort [9,25,[36][37][38]. ...
... The use of PIT tags to monitor mussel populations is a powerful tool and one which has enabled us to estimate retention per site two years after releases with relatively few monitoring occasions. The regular monitoring of release sites and downstream areas should be continued so that site-specific retention (and migration during high-flow events) can be modelled in the future [21,33,36,41] to help inform the adaptive management of this population reinforcement project. ...
Article
Full-text available
Freshwater mussel populations are in sharp decline and are considered to be one of the most imperilled groups globally. Consequently, the number of captive breeding programmes has increased rapidly in recent years, coupled with subsequent reintroductions/population reinforcements to reverse these declines. The outcomes of mussel conservation translocations are seldom reported in the primary literature, hindering opportunities for learning and for population recovery at pace. Here, we describe the methods employed to carry out a successful conservation translocation of the freshwater pearl mussel (Margaritifera margaritifera) in a declining population in northwest England. Following a small-scale pilot release in 2017, four release sites were identified for a population reinforcement of over 1300 tagged mussels in 2021. Monitoring during 2022 showed high levels of retention of juveniles at three out of the four release sites, despite the occurrence of a significant flood event during October 2021. Subsequent releases of 1100 juveniles were carried out across the three successful sites in 2023. Ongoing and regular monitoring is essential in order to provide data on the longer-term fate of propagated juveniles in the wild. This will allow for adaptive management of release activities in this river. These data will be useful to design conservation translocation strategies for other imperilled pearl mussel populations in the UK and throughout Europe.
... Some captive breeding programs maintain species threatened with extinction in zoos and aquaria for multiple generations to ensure their survival, with the goal of future reintroduction once threats to their existence have been removed 15 . Freshwater mussels are also often translocated to different habitats or brought into captive holding facilities for temporary refuge and propagation to mitigate damage from in-stream construction activities, toxic river spills, and zebra mussel infestations [16][17][18][19] . These ex situ management efforts are powerful tools used to maintain or increase biodiversity, but may also cause stress for the animals, making them more vulnerable to factors that directly contribute to translocation failure, such as starvation, disease, and reduced reproductive capacity 20 . ...
Article
Full-text available
Approximately two thirds of freshwater mussel species in the United States and Canada are imperiled, and populations are declining rapidly. Translocation and captive management are commonly used to mitigate losses of freshwater mussel biodiversity, but these conservation tools may result in decreased growth and increased mortality. This study uses RNA-Seq to determine how translocation into captivity affects gene expression in Amblema plicata. Mussels were collected from the Muskingum River in Ohio, USA and brought into a captive holding facility. RNA was extracted from gill tissue 11 months post translocation from mussels in captivity and the Muskingum River on the same day. RNA was sequenced on an Illumina HiSeq 2500, and differential expression analysis was performed on de novo assembled transcripts. More than 1200 transcripts were up-regulated in captive mussels, and 246 were assigned functional annotations. Many up-regulated transcripts were involved in energy metabolism and the stress response, such as heat shock proteins and antioxidants. More than 500 transcripts were down-regulated in captive mussels, and 41 were assigned functional annotations. We observed an over-representation of down-regulated transcripts associated with immune response. Our work suggests that A. plicata experienced moderate levels of stress and altered energy metabolism and immune response for at least 11 months post translocation into captivity.
... Nÿboer et al. (2006) stressed the importance of protecting the physicochemical conditions of the streams where A. pellucida is found. Several restoration activities have recently occurred or are occurring in the Vermilion River basin, including dam removal (Tiemann et al. 2016;Smith et al. 2017) and species reintroductions (Stodola et al. 2017). However, anthropogenic threats are on-going and further investigations are needed to examine their effects on A. pellucida, which should aid in the conservation of species. ...
Article
Full-text available
The Eastern Sand Darter, Ammocrypta pellucida (Agassiz, 1863), has undergone range-wide population declines as a result of anthropogenic disturbances. Within Illinois, the fish historically occurred throughout the Wabash River drainage and Ohio River, but its range was reduced to only the Embarras and Vermilion river basins, including the Middle Fork Vermilion River and North Fork Vermilion River sub-basins. We report the first occurrences of A. pellucida in the Salt Fork Vermilion River sub-basin, thus expanding the known range of this imperiled fish by nearly 50 river-kilometers. The distribution expansion might indicate improved physicochemical conditions in the Vermilion River basin.
Article
Centuries of beaver extirpation, deforestation, and other anthropogenic impacts have disconnected North American rivers from their floodplains and concentrated more hydraulic energy within their channels, degrading aquatic habitat and making the streambed more prone to erosion. Rivers naturally adjust via systematic downcutting, bank erosion, channel widening, bar building, and the gradual recovery of geomorphic equilibrium with well‐connected benches that dissipate hydraulic energy and restore a more natural streambed disturbance regime. The life histories of two endangered freshwater mussels, the fanshell ( Cyprogenia stegaria ) and snuffbox ( Epioblasma triquetra ), suggest they have evolved to and depend on the natural disturbance regime. Young juveniles excyst from their host fish in early summer, burrow into the top few millimeters of the streambed, and then need ca. 3 months of streambed stability prior to growing large enough to be less vulnerable to streambed mobilization. We propose a conceptual model that suggests a potential prerequisite to supporting fanshell and snuffbox populations is a geomorphically recovering (i.e. at least one stable bank and a wide enough channel corridor for at least partially vegetated bars/benches) or recovered (i.e. stable banks and vegetated benches) channel and floodplain corridor that sufficiently dissipates its hydraulic energy to maintain seasonal streambed stability during typical (non‐hurricane) summer/autumns. To explore this conceptual model, we conducted mussel and geomorphic surveys in a reach of the Rolling Fork River that spanned a range of channel conditions from chronically failing streambanks to a geomorphically recovered channel with wide, vegetated benches. Our analyses documented increasing mussel species richness, including the presence of the endangered fanshell or snuffbox, with increasing width of seasonally stable streambed habitat.
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Malacologists took notice of tree snails in the genus Liguus during the last decades of the nineteenth century. Since then, Liguus have undergone repeated shifts in identity as members of species, states, shell collections, backyard gardens, and engineered wildernesses. To understand what Liguus are, this paper examines snail enthusiasts, collectors, researchers, and conservationists—collectively self-identified as Liggers—in their varied landscapes. I argue that Liguus, both in the scientific imaginary and in the material landscape, mediated knowledge-making processes that circulated among amateur and professional malacologists across the United States and Cuba during the twentieth century. Beginning with an examination of early Liggers’ work in Florida and Cuba, this paper demonstrates how notions of taxonomy and biogeography informed later efforts to understand Liguus hybridization and conservation. A heterogeneous community of Liggers has had varied and at times contradictory commitments informed by shifting physical, social, and scientific landscapes. Genealogizing those commitments illuminates the factors underpinning a decision to undertake the until now little-chronicled large-scale and sustained transplantation of every living Floridian form of Liguus fasciatus into Everglades National Park. The social history of Liggers and Liguus fundamentally blurs distinctions between professional scientists and amateur naturalists. The experiences of a diverse cast of Liggers and their Liguus snails historicize the complex character of human-animal relations and speak to the increasing endangerment of many similarly range-restricted invertebrates.
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Pseudunio auricularius (Spengler, 1793) is one of the most threatened unionid species worldwide. Translocation is considered one of the ultimate actions that can save this species from extinction in the Iberian Peninsula. Since 2013, massive mortalities have been recorded in the Canal Imperial de Aragón (CIA), an anthropogenic habitat where the highest density of P. auricularius had been recorded in Spain. An adequacy habitat index was calculated assigning scores to different environmental variables to select the most suitable river stretches receiving the translocated specimens. A total of 638 specimens have been translocated: 291 in 2017, 291 in 2018, and 56 in 2019. The first-year survival in the group of individuals translocated in 2017 was 41.6%. The next year, 95% of these specimens were found alive, suggesting a successful initial establishment. Specimens translocated in 2018 and 2019 showed a survival of c. 69% and 49%, respectively. In contrast, the control group left in CIA in 2017 showed a much lower survival rate of 19.7% after one year, which remained equally low during the next two years. Currently, the conditions in the Ebro River seem to allow a higher survival rate for P. auricularius than those in the CIA; nevertheless, future monitoring should confirm their long-term success.
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The fishes of Champaign County Illinois have received probably as intensive and prolonged study as those in any area of equal size in the New World. The long period of observation has furnished an unusual opportunity to evaluate the ecological changes that have occurred in a highly developed agricultural and urban region and to relate these changes to the distribution and abundance of stream fishes. Throughout this study, emphasis has been placed on changes—changes in the county resulting from agricultural development and population increase, changes in the streams resulting from natural and human modifications, changes in aquatic habitats resulting from new developments in land use practices, and changes in the fishes as these adaptable animals adjusted to new conditions in their naturally unstable aquatic environment.
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Freshwater mussels have undergone dramatic population declines due largely to habitat alteration. A commonly employed measure to minimize the effects of anthropogenic habitat disturbance on mussels is short-distance relocations of individuals. However, quantified survival data are lacking to gauge the success of relocations. To evaluate the suitability of short-distance relocations as a conservation tool for freshwater mussels, we experimentally relocated two common species, Mucket (Actinonaias ligamentina) and Plain Pocketbook (Lampsilis cardium), in an active construction zone. We marked 100 mussels with passive integrated transponders, released them ~200 m upstream of the construction site, and monitored them monthly throughout the spring and summer 2013-2015. We used Cormack-Jolly-Seber models to estimate apparent survival rates and found survival was lowest the first two months after relocation but increased and stabilized thereafter. Our models predict 93% of the relocated A. ligamentina and 71% of the L. cardium remained alive three years post-relocation. We conclude short-distance relocations are a viable minimization tool for protecting freshwater mussels at bridge construction sites, but further study is needed examine the factors driving the initial mortality.
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The bluebreast darter Etheostoma camurum (Cope) has a disjunct distribution within the Ohio River drainage. I examined the distribution and life history characteristics of E. camurum in the Vermilion River basin (Wabash River drainage), Vermilion County, Illi-nois, during the summer of 2007. The darter commonly was collected in areas of moderate to swift currents over cobble and boulder. It spawned in these areas during early summer, and consumed predominately aquatic insect larvae for its diet. Because of its small range, it is recommended that E. camurum remain listed as state-endangered.
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From 2005–2011, the federally endangered freshwater mussel Epioblasma capsaeformis (oyster mussel) was reintroduced at three sites in the upper Clinch River, Virginia, using four release techniques. These release techniques were: 1) translocation of adults (Site 1, n = 1,418), 2) release of laboratory-propagated sub-adults (Site 1, n = 2,851), 3) release of 8-week old laboratory-propagated juveniles (Site 2, n = 9,501), and 4) release of stream-side infested host fishes (Site 3, n = 1,116 host fishes). These restoration efforts provided a unique research opportunity to compare the effectiveness of techniques used to reestablish populations of extirpated and declining species. We evaluated the relative success of these four population restoration techniques via monitoring at each release site (2011–2012) using systematic 0.25-m2 quadrat sampling to estimate abundance and post-release survival. Abundances of translocated and laboratory-propagated E. capsaeformis at Site 1 ranged from 577–645 and 1,678–1,700 individuals, respectively, signifying successful settlement and high post-release survival. Two untagged individuals (29.1 and 27.3 mm) were observed, indicating that recruitment is occurring at Site 1. No E. capsaeformis were found at sites where 8-week old laboratory-propagated juveniles (Site 2) and stream-side infested host fishes (Site 3) were released. Our results indicate that translocations of adults and releases of laboratory-propagated sub-adults were the most effective population restoration techniques for E. capsaeformis. We recommend that restoration efforts focus on release of larger (>20 mm) individuals to accelerate augmenting and reintroducing populations and increase the probability for recovery of imperiled species.
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Power analysis after study completion has been suggested to interpret study results. We present 3 methods of estimating power and discuss their limitations. We use simulation studies to show that estimated power can be biased, extremely variable, and severely bounded. We endorse the practice of computing power to detect a biologically meaningful difference as a tool for study planning but suggest that calculation of confidence intervals on the parameter of interest is the appropriate way to gauge the strength and biological meaning of study results.
Book
Nearly 200 years ago, a naturalist named Rafinesque stood on the banks of the Ohio River and began to describe the freshwater mussels he found there. Since that time these animals have become the most imperiled animals in North America. Dozens of species have become extinct, and it is estimated that two-thirds of the remaining freshwater mussels face a similar fate. Yet, despite their importance, the mussels of Ohio remain a poorly documented and largely mysterious fauna. The Freshwater Mussels of Ohio by G. Thomas Watters, Michael A. Hoggarth, and David H. Stansbery brings together, for the first time, the most up-to-date research on Ohio’s mussels. Designed for the weekend naturalist and scientist alike, it synthesizes recent work on genetics, biology, and systematics into one book. Each species is illustrated to a degree not found in any other work. Full-page color plates depict shell variation, hinge detail, and beak sculpture. Full-page maps show the distribution of each species based upon the collections of numerous museums (with historical distributions dating from the 1800s). In addition to species accounts, the book has a substantive introduction that includes information on basic biology, human use, and conservation issues. Extensive synonymies, a key to all species, and an illustrated glossary are included as well.
Article
This well-illustrated book highlights freshwater mussels fabulous diversity, amazing array of often bizarre ecological adaptations, and their dire conservation plight. Summarizing and synthesizing historical and contemporary information as well as original research and analysis, the book describes the diverse array of mussel life history strategies and builds a cohesive narrative culminating in the development of explicit frameworks to explain pervasive patterns in mussel ecology. The fascinating and colorful role of mussels in human society is also described in detail, including the little-known pearl button industry of the early 1900s and the wild and often violent shell harvest of the 1990s. The final chapter details humans efforts to save these fascinating animals and gives a prognosis for the future of the North American fauna. The book provides the first comprehensive review of mussel ecology and conservation for scientists, natural resource professionals, students, and natural history enthusiasts.
Article
The use of the Cormack-Jolly-Seber model under a standard sampling scheme of one sample per time period, when the Jolly-Seber assumption that all emigration is permanent does not hold, leads to the confounding of temporary emigration probabilities with capture probabilities. This biases the estimates of capture probability when temporary emigration is a completely random process, and both capture and survival probabilities when there is a temporary trap response in temporary emigration, or it is Markovian. The use of secondary capture samples over a shorter interval within each period, during which the population is assumed to be closed (Pollock's robust design), provides a second source of information on capture probabilities. This solves the confounding problem, and thus temporary emigration probabilities can be estimated. This process can be accomplished in an ad hoc fashion for completely random temporary emigration and to some extent in the temporary trap response case, but modelling the complete sampling process provides more flexibility and permits direct estimation of variances. For the case of Markovian temporary emigration, a full likelihood is required.