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Effects of a diatom ecosystem engineer (Didymosphenia geminata) on stream food webs: implications for native fishes


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Stream habitat changes affecting primary consumers often indirectly impact secondary consumers such as fishes. Blooms of the benthic algae Didymosphenia geminata (Didymo) are known to affect stream macroinvertebrates, but the potential indirect trophic impacts on fish consumers are poorly understood. In streams of the Kootenai River basin, we quantified the diet, condition, and growth rate of species of trout, char, and sculpin. In 2018, macroinvertebrate taxa composition was different between a stream with Didymo and a stream without, but trout diets, energy demand, and growth rates were similar. Trout abundance was higher in the stream with Didymo, but the amount of drifting invertebrates was higher in the stream without. In 2019, we surveyed 28 streams with a gradient of coverage. Didymo abundance was correlated only with the percentage of aquatic invertebrates in trout diets and was not related to diets of char or sculpin or condition of any species. Thus, we found no evidence for a trophic link between Didymo blooms and the condition or growth of trout, char, or sculpin in mountainous headwater streams.
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Effects of a diatom ecosystem engineer (Didymosphenia geminata)
on stream food webs: implications for native shes
Niall G. Clancy, Janice Brahney, James Dunnigan, and Phaedra Budy
Abstract: Stream habitat changes affecting primary consumers often indirectly impact secondary consumers such as shes.
Blooms of the benthic algae Didymosphenia geminata (Didymo) are known to affect stream macroinvertebrates, but the potential
indirect trophic impacts on sh consumers are poorly understood. In streams of the Kootenai River basin, we quantied the
diet, condition, and growth rate of species of trout, char, and sculpin. In 2018, macroinvertebrate taxa composition was different
between a stream with Didymo and a stream without, but trout diets, energy demand, and growth rates were similar. Trout abun-
dance was higher in the stream with Didymo, but the amount of drifting invertebrates was higher in the stream without. In 2019,
we surveyed 28 streams with a gradient of coverage. Didymo abundance was correlated only with the percentage of aquatic inverte-
brates in trout diets and was not related to diets of char or sculpin or condition of any species. Thus, we found no evidence for a
trophic link between Didymo blooms and the condition or growth of trout, char, or sculpin in mountainous headwater streams.
Résumé : Les modications des habitats de cours deau qui ont une incidence sur les consommateurs primaires ont souvent
deseffetsindirectssurlesconsommateurssecondairescommelespoissons.Sil est établi que les proliférations de lalgue
benthique Didymosphenia geminata (didymo) ont une incidence sur les macroinvertébrés lotiques, leurs impacts trophiques
indirects potentiels sur les poissons consommateurs ne sont pas bien compris. Nous avons quantiélerégimealimentaire,
lembonpoint et le taux de croissance de truites, dombles et de chabots dans des cours deau du bassin de la rivière Koote-
nai. En 2018, un cours deauoùlalgue didymo était présente et un autre dont elle était absente présentaient des composi-
tions taxonomiques de macroinvertébrés différentes, alors que les régimes alimentaires, la demande énergétique et les
taux de croissance des truites y étaient semblables. Labondance de truites était plus grande dans le cours deau contenant
des algues didymo, mais la quantité dinvertébrés à la dérive était plus grande dans le cours deau exempt de lalgue. En
2019, nous avons inventorié 28 cours deau dénissant un gradient de couverture par lalgue didymo. Labondance des
algues était seulement corrélée au pourcentage dinvertébrésaquatiquesdanslerégimealimentairedestruitesetnétait
pas corrélée au régime alimentaire des ombles ou des chabots, ni à lembonpoint daucune espèce. Ainsi, nous navons rel-
evéaucunepreuvedun lien entre les proliférations dalgues didymo et lembonpoint ou la croissance des truites, ombles
ou chabots dans des cours deau supérieurs de montagne. [Traduit par la Rédaction]
Trophic structure determines the availability of food resources
to organisms within a food web. The growth of these organisms
depends not only on the amount of energy in a system, but also
on the proportion of that energy that can be consumed and digested
(Horton 1961;Huryn 1996;Bellmore et al. 2013). In human-
modied landscapes, activities such as logging, mineral extrac-
tion, and river impoundment have altered in-stream habitats and
riparian areas (Hand et al. 2018), often resulting in a lack of habi-
tat complexity and nutrient availability, and thus affecting the
amount of energy available as food for aquatic organisms (Meredith
et al. 2014;Minshall et al. 2014). Such change can alter stream mac-
roinvertebrate assemblages and impact consumers of both larval
and adult life stages of aquatic insects (Power et al. 1996;Nakano
et al. 1999;Baxter et al. 2005). Fo r many sh species, aquatic mac-
roinvertebrates are a primary source of energy (Behnke 2010).
Understanding how specic environmental changes alter the
ow of in-stream energy to sh can thus be of great importance to
conservation and management efforts (Cross et al. 2011;Bellmore
et al. 2012;Scholl et al. 2019).
Organisms that alter physical and biological habitats upon which
other organisms rely are ubiquitous in many environments and have
long been a topic of biological study (Darwin 1881). In fresh waters, a
wide variety of animal ecosystem engineers such as dam-building
beavers, nest-digging shes, pebble-moving craysh, sediment-
disturbing tadpoles, and net-spinning caddisies alter aquatic
habitat through the manipulation of their environment (Moore
2006;Albertson and Daniels 2016;Tumolo et al. 2019). Other spe-
cies, such as corals, act as agents of biogeomorphic change through
the growth of their own physical structures altering or creating
habitats due to their own architectures (Jones et al. 1994). Corre-
sponding ecosystem engineers in streams are almost all autotrophs
(but see Beckett et al. 1996 and Hopper et al. 2019)andincludetrees
that form large woody debris (LWD) jams, macrophytes that shelter
juvenile shes, and algae that harbor invertebrates.
One such autotrophic ecosystem engineer is the diatomaceous
algae Didymosphenia geminata (hereinafter, Didymo). Overgrowths
Received 10 April 2020. Accepted 28 September 2020.
N.G. Clancy* and J. Brahney. Department of Watershed Sciences, Utah State University, 5210 Old Main Hill, Logan, UT 84322, USA.
J. Dunnigan. Montana Fish, Wildlife & Parks, 385 Fish Hatchery Road, Libby, MT 59923, USA.
P. Budy. US Geological Survey, Utah Cooperative Fish and Wildlife Research Unit & Department of Watershed Sciences, Utah State University, 5205 Old
Main Hill, Logan, UT 84322, USA.
Corresponding author: Niall G. Clancy (email:
*Present address: Montana Fish, Wildlife & Parks, 490 N. Meridian Road, Kalispell, MT 59901, USA.
Copyright remains with the author(s) or their institution(s). Permission for reuse (free in most cases) can be obtained from
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(colloquially, blooms) of this North American native are charac-
terized by the production of a long polysaccharide stalk from
individual diatoms, which can cover large areas of the substrate.
Prodigious growth of this bifurcating stalk differentiates Didymo
from most other diatoms, and the interwoven aggregate of stalks
from a multitude of cells produces a thick mat that contains
other algae, detritus, and macroinvertebrates sometimes cov-
ering entire streambeds (Gretz 2008). While phosphorus limita-
tion is thought to play a role in Didymo bloom formation, the
precise causes of blooms remain a current topic of investigation
(Taylor and Bothwell 2014). In recent years, increasing reports of
severe Didymo blooms have led to major concern about its conse-
quences for freshwater organisms (Bickel and Closs 2008;James
and Chipps 2016;Jellyman and Harding 2016).
At high Didymo coverage, stream invertebrate assemblages origi-
nally dominated by Ephemeroptera, Plecoptera, and Trichoptera
(EPT taxa) typically shift towards dominance by Chironomidae, Oli-
gochaeta, Nematoda, or Cladocera taxa generally associated with
reduced habitat quality in trout streams (Kilroy et al. 2009;Gillis and
Chalifour 2010;Larned and Kilroy 2014). There has been widespread
concern about the consequences of Didymo blooms for trout (Gillis
and Chalifour 2010;James et al. 2010a;Jellyman and Harding 2016)
because EPT taxa are often assumed to be a primary food source for
salmonid species (Behnke 2010). However, to date, it is unclear
whether Didymo blooms have had any substantial negative or posi-
tive impacts on salmonids via changes to the macroinvertebrate
food base. In several New Zealand rivers, blooms were correlated
with lower trout abundances, dietary percent EPT, and stomach full-
ness (Jellyman and Harding 2016). In contrast, production of Atlantic
salmon (Salmo salar) in Icelandic and Norwegian rivers has remained
high despite the presence of severe Didymo blooms (Jonsson et al.
2008;Lindstrøm and Skulberg 2008), and escapement and juvenile
production of Pacic salmon and steelhead (Oncorhynchus spp.) in
Vancouver Island streams either increased or did not change in rela-
tion to blooms (Bothwell et al. 2008). In four South Dakota streams,
the condition and forage of large brown trout (Salmo trutta)wasnot
correlated with Didymo blooms, while body condition of juveniles
washigher(James and Chipps 2016). However, Didymo growth in
the study was also correlated with drought, making causal inference
difcult (James et al. 2010b). As such, no study has successfully exam-
ined the mechanistic links among Didymo blooms, macroinverte-
brates, and shes necessary to make causal inference. Further, no
studies have addressed the potential effects of blooms on inland
native trout populations or on nongame species.
To better understand the trophic consequences of Didymo
blooms, we assessed the relationship among blooms, sh diet,
condition, and growth over two summers in the mountainous
Kootenai (Kootenay in Canada) basin of British Columbia, Idaho,
and Montana, much of which falls within the globally rare inland
temperate rainforest biome (Dellasala et al. 2011). We employed a
multifaceted research approach in which we examined potential
Didymo bloom impacts on sh via two methods: (i) a detailed,
mechanistic study comparing food webs in a stream with Didymo
blooms against one without and (ii) an observational study survey-
ing sh diet and condition across 28 streams representing a gradi-
ent of Didymo bloom severity. We also rene a novel method for
estimating food consumption by sh that may serve as a viable al-
ternative to traditional bioenergetics, in some cases. The results
of this study increase ecological understanding of the consequen-
ces of Didymo blooms and will help determine whether treat-
ment of Didymo blooms is a necessary strategy to benetshes.
Study location
Twice-monthly through the summer of 2018, we sampled two
streams located in the Cabinet Mountains of northwestern Mon-
tana, Bear Creek and nearby Ramsey Creek (Fig. 1). Both creeks
have similar physical characteristics (Table 1), but Bear Creek
contains obvious Didymo blooms while Ramsey Creek does not.
The two streams thus offer an opportunity to examine potential
effects of blooms on biotic communities in a paired, reference-
impact framework.
Over the course of the summers of 2018 and 2019, we examined
131 locations on 103 individual streams for the presence of Didymo
blooms throughout the Kootenai River basin (Clancy 2020). In 2019,
we surveyed shes in 28 of those streams (Fig. 1), representing large
differences in bloom coverage: 0%80% (Table A1).
Didymo reference-impact study 2018
We selected a 300 m long reach for study in both Bear and Ram-
sey Creeks with similar habitat conditions. The sh assemblages
of both were predominantly composed of Columbia River red-
band trout (Oncorhynchus mykiss gairdneri)andbulltrout(Salvelinus
conuentus). Ramsey Creek also contained Columbia slimy sculpin
(Uranidea cognata syn. Cottus cognatus) at relatively low abun-
dance. We measured ve habitat variables to ensure Bear and
Ramsey creeks were suitable for comparison: mean substrate
size (sensu Wolman 1954), channel wetted width, mesohabitat
composition (percent cascade, rife, and pool), water tempera-
ture (30 min recording interval, Onset HOBO data loggers), and
water chemistry (soluble reactive phosphorus, nitrate, and sul-
fate; Lachat 8500 QuikChem FIA and IC). Every 2 weeks (early
June early September), we systematically estimated percentage
of substrate covered by blooming Didymo using a 19 L bucket
with a clear bottom, making ve evenly spaced estimates along
lateral transects, each 20 paces apart from reach top to bottom.
We then combined twice-monthly estimates to form monthly
Didymo bloom coverage estimates.
Food-web structure was determined by macroinvertebrate and
sh sampling concurrent with Didymo coverage estimation. In
conjunction with Didymo bloom measurements, we collected
drifting macroinvertebrates by placing two separate 25.4 cm
45.72 cm drift nets in the stream for 30 min at the middle-to-end
of the reach and pooling the combined samples in 70% ethanol.
Sampling occurred between the hours of 10:00 and 17:00. The day
following each Didymo and macroinvertebrate sampling event,
we collected shes through single-pass backpack electroshock-
ing (LR-24 Backpack Shocker Smith-Root, Vancouver, Washing-
ton). We completed multiple passes during the nal sampling
event (September) to maximize summer-long recapture. Each
sh was anesthetized with clove oil, weighed, measured, and
marked by clipping a small section of the caudal n. We gastri-
cally lavaged individuals larger than 100 mm to collect diets and,
if captured during June or July, implanted a uniquely coded,
12 mm passive integrated transponder (PIT) tag (Model HDX12,
Biomark, Boise, Idaho). Gut evacuation was assumed to be mini-
mal due to cold temperatures, and processing was generally less
than an hour after capture. Using the average percent growth
between individuals measured in June and July, we back-calculated
June weights for individuals tagged in July. This represented 57%
of redband trout in Bear Creek and 68% in Ramsey Creek. We also
observed no evidence of a tag effect on growth of PIT-tagged sh
(Clancy 2020).
Because a shift to a macroinvertebrate assemblage of smaller
and more abundant individuals may favor juvenile shes (James
and Chipps 2016), we identied large and small size classes of red-
band (cutoff at 105 mm) and bull trout (cutoff at 130 mm) using
monthly lengthfrequency histograms (Clancy 2020). We then
calculated size-specic abundances using LincolnPetersen
markrecapture estimation in which the nal sampling date was
the recapture event and all previous sampling events a single
marking event (Lincoln 1930). We determined this approach to
be reasonable because movement of PIT-tagged shes between
the abutting upper and lower halves of Bear Creek was negligible,
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thus meeting the population closure assumption of LincolnPetersen
estimation (Clancy 2020).
We identied and measured drift and diet macroinvertebrates
to the taxonomic level of family and used published length-
to-mass conversions to estimate biomass and caloric content
(Clancy 2020). We compared taxon-specicproportionsof
drifting macroinvertebrates in Bear and Ramsey creeks by cal-
culating the monthly percent similarity (Schoener 1970):
ð1ÞPercent Similarity ¼100 0:5X
where B
is the percentage of aquatic invertebrates of taxa iin
Bear Creek, and R
is the percent of invertebrates of taxa iin
Ramsey Creek. Using the same equation, we compared trout
diets with the availability of invertebrates in the drift (both
aquatic and terrestrial) to determine whether feeding behavior
was different between the two streams. Then, we also compared
trout diets between the two streams using percent energetic
content for each diet taxa. To evaluate how likely observed dif-
ferences between groups were (drift versus drift, diet versus
drift, and diet versus diet), we used Pearsons
tests. We fur-
ther report monthly and summer-long gut fullness and relative
number and energetic content of invertebrates in the drift
between the two streams.
By pairing individual caloric demand with trout diet composi-
tion, we created energy-ow food webs. To calculate individual
caloric demand, we used a novel and simple bioenergetics equa-
tion that combines aspects of the BenkeWallace trophic basis
of production method for calculating energetic demand (Benke
and Wallace 1980) and traditional sh bioenergetics (Hanson
et al. 1997) while accounting for sh thermal growth optima. We
developed this novel thermal optima bioenergetics equation
(TOBE) because traditional sh bioenergetics models require
extensive species-specic parameterization (Chipps and Wahl
2008) and are thus not available for many species. By creating the
TOBE, where the only species-specic parameter is thermal
Fig. 1. Location of study streams (red dots) within the Kootenai River basin (left) and the upper Libby Creek subbasin (right). Inset A
shows the location of the Kootenai basin within the larger Columbia River watershed. Map created using QGIS (2019) and shapeles from
the National Hydrography Dataset (USGS 2019). [Colour online.]
Table 1. Habitat measurements and nutrient
concentrations in Bear Creek (Didymo) and
Ramsey Creek (no Didymo) during summer 2018
Bear Ck.
Ramsey Ck.
(No Didymo)
Temp. (°C) 6SD 9.7962.32 9.7962.40
Cascade 76% 83%
Rife 16% 10%
Pool 8% 7%
Substrate size (mm) 26.7 23.2
Wetted width (m) 7.24 7.17
Nutrients (lg·L
SRP 1.99560.368 1.53060.409
Nitrate 74.5 25
Sulfate 1235 930
N:P ratio 82.5 36.0
Note: The N:P ratio is the molar ratio of nitrate to
soluble reactive phosphorus (SRP).
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growth optima, and comparing consumption estimates of red-
band trout and bull trout with those generated using traditional
bioenergetics models (Deslauriers et al. 2017), we were able to
determine whether the TOBE is adequate for use with species for
which traditional bioenergetics models are not available. Inputs
to the traditional sh bioenergetics models were summer stream
temperatures (taken at 30 min intervals), start and end weights
of sh, and the output was summer-long energetic consumption.
We used species-specic bioenergetics models for redband (rain-
bow) trout (Railsback and Rose 1999)andbulltrout(Mesa et al.
2013) and substituted a model for prickly sculpin (Cottus asper)for
Columbia slimy sculpin (Moss 2001).
The BenkeWallace method, upon which the TOBE is based,
was originally developed for use with benthic macroinverte-
brates and does not account for differential allocation of energy
by organism size and water temperature, factors known to
strongly inuence sh growth (Brown et al. 2004). Thus, we used
two different numbers for the proportion of total assimilated
energy allocated to growth (net production efciency) in large
versus small sh as suggested by Bellmore et al. (2013).Wethen
applied a temperature-specic factor to adjust for the effects of
suboptimum temperatures. Thus, consumption in kilocalories
was calculated as follows:
ð2ÞConsumption ¼X
DietProportioniGrowth EnergyDensity
TempFactor TissueAllocation Digestiblei0:2Digestiblei
where DietProportion
is the average proportion by kilocalories
of food type iin the diet, Growth is the average accrued body
mass in grams of the sh over the summer (JuneSeptember),
EnergyDensity is the energy density (kcal·g
)ofthesh, and Tis-
sueAllocation is the theoretical maximum proportion of assimi-
lated energy allocated to sh tissue growth (net production
efciency), which was set as 0.22 for large size class trout and 0.5
for small size class trout and sculpin. Digestible
is the estimated
digestible proportion of food type i. We used Digestible
for each
food type from Hanson et al. (1997) and subtracted a value of
to account for specic dynamic action (Hanson
et al. 1997 ). Thus, Digestible
is the assimilation ef-
ciency of food type i. TempFactor is the temperature correction
factor calculated according to the following equation:
ð3ÞTempFactor ¼e0:2StreamTempOptimTemp
where StreamTemp is the average stream temperature for the
measurement interval over the growth period, and OptimTemp
is the thermal optimum for each species. This equation is an
approximation of a shs thermal optimum curve (based on equa-
tions in Bear et al. 2007) that asymptotes at an energy-allocation-to-
tissue value of zero (Clancy 2020). We derived thermal optimum
values from previous eld and laboratory studies: 13.1°C for red-
band trout (Bear et al. 2007), 12.0 °C for bull trout (Dunham et al.
2003), and 12.1 °C for Columbia slimy sculpin (Wehrly et al. 2003).
To derive total estimated consumption by each species, we
multiplied estimated summer TOBE consumption values by cal-
culated sh abundances in each stream. Then, we multiplied the
proportion of energy of each prey item in the average diet of each
sh species by the reach-level consumption estimates. Thus, we
obtained estimates of total energy ow from all prey to sh pred-
ators and compared results for Bear and Ramsey creeks.
Multistream Didymo survey 2019
In a 30.5 m reach of each selected stream, we estimated Didymo
coverage using the same method as in 2018. We also recorded six
other habitat variables: wetted width (n= 5), canopy density (n=5
using a densitometer; Strickler 1959), dominant vegetation type,
substrate type (Cummins 1962), Rosgen channel type (Rosgen
1994), number of LWD items (sensu Kershner et al. 2004), and
mean August stream temperature. From reach top to bottom, we
measured wetted width and canopy density, while we qualita-
tively assessed vegetation, substrate, and channel type. Tempera-
ture loggers were placed in three streams: Bear, Outlet, and Trail
creeks. We estimated mean August temperatures of all other
streams by adding the time-specic difference of each streams
temperature (taken with a handheld thermometer) to the mean
August value of one of the three temperature loggers (Bear Creek
for streams owing into the Kootenai River below the Fisher
River conuence, Outlet Creek for those above the Fisher conu-
ence, and Trail Creek for Fisher River tributaries).
In the same reach, we collected shes through two-pass (one
upstream, one downstream) backpack electroshocking. We anes-
thetized, weighed, and measured all shes and then released leu-
ciscids and catostomids. Using an in-eld assessment in which
we gastrically lavaged sh, we assessed the diets of salmonids
and cottids by spreading the diet contents in a 30 cm 15 cm
white pan and recording the number of individuals of each inver-
tebrate taxa. We identied insects to Order except for Simuliidae
and Chironomidae, which we identied to Family. Other inverte-
brates we identied to Class or Phylum, and vertebrates to the
lowest practical taxonomic level (usually species).
We generated two response metrics of sh condition (Fultons
K(Heincke 1908;Ricker 1975) and residual analysis of observed
versus predicted weights (Fechhelm et al. 1995)) and four metrics
of diet composition (%Diptera, %EPT, %Aquatics, and gut fullness
(number of diet items/sh length]) for each sh. Using weighted,
univariate logistic (%Diptera, %EPT, %Aquatics) and linear regres-
sions (gut fullness and sh condition) in which sh sample size
was the relative weight of each stream in the regression, we ana-
lyzed each response metric compared with Didymo coverage and
the other six habitat variables. We removed four streams (Koka-
nee, Coffee, Mobbs, and Solo Joe creeks) from regressions due to
low sample size or substantially different substrate type. We
grouped sh by genus due to otherwise small sample size if com-
pared only within species (char (Salvelinus)andsculpin(Uranidea))
or substantial hybridization in the basin that made some eld
IDs difcult (trout, Oncorhynchus spp.). For each comparison of a
habitat variable with a diet metric, we calculated an R
value (or
for logistic regression; Nagelkerke 1991)
and pvalue and considered variables with an R
greater than 0.2
and a pvalue less than 0.2 to be a nonspurious correlation.
Didymo versus reference stream study 2018
Differences in all four physical habitat variables were small
between Bear Creek (Didymo) and Ramsey Creek (no Didymo),
giving us condence the two were suitable for comparison (Table
1). Nutrient concentrations differed between the two (Table 1),
likely playing a role in Didymo bloom presence (Capito 2020).
Didymo bloom severity in Bear Creek increased from 10.9% cover-
age in June to 22.6% coverage in August before falling to 18.9% in
September (Fig. 2). The June to August change in Didymo cover-
age was signicant (p<0.01), but the decline from August to Sep-
tember was not (p=0.21).
Composition of drifting aquatic insects (EPT larvae and Diptera
larvae and pupae) between the two streams became less similar
as Didymo coverage increased throughout the summer and more
similar when Didymo coverage decreased in September (June
September: 84.4%, 78.5%, 67.2%, and 92.0% similar; Fig. 2). From
June through August, percent EPT in the drift was higher in Ram-
sey Creek (no Didymo) and percent larval and pupal Diptera was
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higher in Bear Creek (Didymo; Fig. 2). Percentages were similar
in September. Both total drifting invertebrates and total energy
of drifting invertebrates similarly diverged later in the summer
with the streams having similar numbers in June, Ramsey
Creek having higher numbers in July and August, and Bear
Creek having higher numbers in September. The summer-long
amount of total energy of drifting invertebrates was 2.2 times
higher in Ramsey Creek due to greater total biomass in the
Reach abundance estimates for redband and bull trout were
higher in Bear Creek (Didymo; Table 2). Relative growth (size-
dependent growth per unit of time) of redband trout varied by
size class. We estimated summer relative growth of small trout
(<105 mm) to be 0.0292 g of growth per gram of starting weight
per day (g·g
) in Bear Creek and only 0.0033 g·g
Ramsey Creek, but this difference was likely driven by a very
small June sample size (three in Bear Creek and one in Ramsey
Creek). Relative growth of large size class redband trout (>105 mm)
Fig. 2. Monthly percentage of stream substrate covered by Didymo in Bear Creek, 2018 (top). Pie charts show proportions of aquatic life-
stage insect taxa in the drift in Bear and Ramsey creeks and indicate higher proportion of Diptera larvae and pupae in Bear Creek
(Didymo). [Colour online.]
Table 2. Population (reach) abundance, growth, and consumption estimates for each sh
species and size class in Bear and Ramsey creeks during summer 2018.
Redband trout
Small Bear 132 0.0292 33.3 20.9 4 385.6
Ramsey 91 0.0033 31.9 18.8 2 902.9
Large Bear 196 0.0027 74.9 75.3 14 643.0
Ramsey 81 0.0029 53.6 54.2 4 347.0
Bull trout
Small Bear 60 0.0136 38.9 27.1 2 334.0
Ramsey 2 NA NA NA NA
Large Bear 45 0.0011 46.0 24.8 2 083.8
Ramsey 3 NA NA NA NA
Slimy sculpin
Ramsey 20 0.0030 5.0 8.8 100.0
Note: Consumption estimates generated using traditional bioenergetics with Fish Bioenergetics 4.0 (FB4;
Deslauriers et al. 2017) are shown for comparison with the novel thermal optimum bioenergetics equation
(TOBE) estimates. Population-level consumption estimates are the product of the population abundance
estimate and the TOBE consumption estimate.
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was similar between the two streams: 0.002760.0004 g·g
Bear Creek and 0.0029 60.0007 g·g
in Ramsey Creek (mean 6
standard error; Table 2).
Redband trout diets (including both terrestrial and aquatic
insects of all life stages) were 40.7% similar to the drift in Bear
Creek (
test: p<0.01) and 40.1% similar to the drift in Ramsey
Creek (
test: p<0.01). By energetic content, redband trout diets
were 81.2% similar between Bear and Ramsey creeks for the
whole summer (
test: p= 0.84): 55.6% similar in June, 77.5% simi-
lar in July, 99.7% similar in August, and 75.0% similar in Septem-
ber. Monthly gut fullness was not signicantly different between
the two streams. Diets of small individual redband trout in Bear
Creek (Didymo) had more EPT (78.6% 68.4%) than large individu-
als (46.4% 63.0%), while gut fullness and %Diptera were similar.
Individual consumption estimates for large redband trout
were 39% higher in Bear Creek (Didymo), while small size class
estimates were similar between the two streams (Table 2). We
estimated reach-level energetic demand by all redband trout at
19 029 kcal in Bear Creek and 7250 kcal in Ramsey Creek (Table
2). Consumption estimates using the TOBE were similar to those
estimated using traditional, species-specicbioenergeticsmod-
els (Table 2).
The primary sources of energy (>5% of demand) for redband trout
in Bear Creek (Didymo) were Ephemeroptera (38.0% of energy
intake), Hymenoptera (15.1%), Trichoptera (14.4%), Plecoptera (9.5%),
and Diptera (7.6%) (Fig. 3; also refer to online Supplementary Table
). Primary energy sources for Ramsey Creek (no Didymo) redband
trout were Ephemeroptera (45.8%), Hymenoptera (15.7%), Diptera
(9.8%), Trichoptera (9.0%), and Plecoptera (6.3%) (Fig. 3). Primary sour-
ces of energy for bull trout in Bear Creek were Ephemeroptera
(48.0%), Trichoptera (13.1%), Nematoda (7.2%), Plecoptera (6.3%), and
Hymenoptera (5.1%) (Fig. 3). We collected only four bull trout diets
during the entire summer in Ramsey Creek, and we did not consider
this sufcient to calculate average diet compositions. Columbia
slimy sculpin were only captured in Ramsey Creek and thus had no
Didymo comparison.
Multistream Didymo survey 2019
Between-site variation in FultonsKwas too low to assess possi-
ble explanatory variables (coefcients of variation (CV) 0.1;
Fig. 3. Energy-ow food web for shes in Bear Creek (Didymo) and Ramsey Creek (no Didymo) indicating that similar energy sources
sustained sh growth in both streams. Line thickness represents proportion of total energy demand by the given sh species met by each
invertebrate taxon. Only taxa representing at least 5% of energy demand are shown. Left to right: Hymenoptera, Diptera, Ephemeroptera,
Plecoptera, Trichoptera, and Nematoda. [Colour online.]
Supplementary data are available with the article at
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Supplementary Table S2
). Between-site variation in sh relative
condition, calculated as a shs observed weight compared with
its predicted weight, was similarly low for trout and sculpin (CV
of 0.12 and 0.04, respectively) and moderately low for char (CV =
0.28). Despite slightly more variation in char relative condition
between sites, there was no relationship between condition and
Didymo coverage (R
=0.03,p= 0.46).
For all diet metrics across all three sh taxa, percent Didymo
cover was only correlated with percentage of aquatic life-stage
invertebrates in Oncorhynchus diets (Fig. 4). Canopy cover, LWD, ri-
parian vegetation type, and stream temperature were also corre-
lated with percent aquatic invertebrates in Oncorhynchus diets,
with LWD having the highest pseudo-R
(Supplementary Table
). In fact, few sh diet metrics were correlated with habitat
variables (Supplementary Table S3
During the summers of 2018 and 2019, we examined the
response of trout, char, and sculpin species to Didymo blooms at
reach and regional scales. Similar to other studies, Didymo in
Bear Creek was positively correlated with an increase in the
proportion of larval Diptera and a decrease in the proportion of
EPT taxa (Marshall 2007;James et al. 2010a;Anderson et al. 2014).
Despite changes to their macroinvertebrate food base, redband
trout diets and growth rates with Didymo were similar to those
in the no-Didymo reference stream. In fact, diets were most simi-
lar in August when Didymo coverage was at its peak. While a no-
Didymo comparison was not available for bull trout because so
few were captured in Ramsey Creek, bull trout diets from Bear
Creek also contained relatively few Diptera. Corroborating these
results, we similarly did not observe major differences in the diets
or condition of trout, char, or sculpin species across a gradient of
Didymo bloom coverages in 2019. This held true even in Outlet
Creek, British Columbia, where Didymo blooms covered over 80% of
the streambed, but rainbow trout diets remained similar to those in
streams with little to no Didymo.
Stream-resident trout are considered generalist invertivores
(Behnke 1992), but strong selection by redband trout in Bear
Creek (Didymo) and Ramsey Creek (no Didymo) for the same taxa
indicates this subspecies may show a strong preference for may-
ies (Ephemeroptera; Fig. 3). Despite this strong selection, the
similarity in trout growth rates between the two streams in 2018
Fig. 4. Correlations of Didymo coverage to each sh taxons diet and condition response metrics from 2019. Each dot represents the
average value for sh in a single stream.
is Nagelkerkespseudo-R
value. Asterisks (***) indicate p0.05. Results indicate few
relationships between Didymo coverage and sh diet metrics.
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Clancy et al. 7
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and condition across streams in 2019 may indicate summer trout
growth in Kootenai River headwaters is more limited by factors in-
dependent of forage such as nontrophic interspecic competition.
Consumption estimates using the novel TOBE were similar to
those generated using traditional bioenergetics models, and the
small discrepancies between bull trout estimates from the two
approaches were probably due to the fact we used a thermal opti-
mum va lue of 12.0 °C ( Dunham et al. 2003) that was likely more
appropriate for resident Kootenai basin bull trout than the 16.0 °C
optimum (Mesa et al. 2013) used by Fish Bioenergetics 4.0
(Deslauriers et al. 2017). While further renement of TOBE (espe-
cially of the size-specic tissue allocation) could certainly make
estimates more accurate, the relative similarities between the
two estimates demonstrate the potential utility of TOBE as a
means to generate consumption estimates when bioenergetics
models are not available or otherwise appropriate (Donner 2011).
Although not the impetus of our study, we observed interest-
ing differences in percentage of aquatic invertebrates in trout
diets in streams with differing riparian vegetation. We observed
higher proportions of riparian invertebrates in trout diets in
alder-dominated (Alnus spp.) streams than in pine-dominated
streams (largely lodgepole pine, Pinus contorta), a pattern similar
to that observed in juvenile salmon (Oncorhynchus kisutch)in
Alaska coastal temperate rainforests (Picea spp.; Allan et al.
2003). However, unlike the Alaska juvenile salmon, trout in our
inland temperate rainforest streams dominated by cedar (Thuja
plicata)andhemlock(Tsuga spp.) had similar aquaticterrestrial
ratios to alder-dominated streams.
Our 2018 study was limited to two streams, and while Didymo
was associated with differences in the macroinvertebrate com-
munity, it is possible differences would exist independent of
Didymo. Although not collected for this study, analysis of
benthic invertebrate samples would likely be more sensitive to
variation in Didymo coverage because they are less affected by
potential drift from upstream sources. Further, we examined the
impacts of Didymo blooms only from June to September during
both years, a time when terrestrial invertebrate inputs, and trout
reliance upon them, are high (Nakano and Murakami 2001). It is
possible some negative or positive consequence of Didymo can
only be observed by studying shes across seasons (e.g., winter).
In fact, some studies have reported severe Didymo blooms during
winter months (e.g., Kolmakov et al. 2008), and we observed
severe blooms in the Lardeau River during April of 2018, prior to
the freshet. Trout growth in headwater streams is higher in summer
months, but foraging (Thurow 1997)andgrowth(Al-Chokhachy
et al. 2019) still occur over winter. Further, our study did not
examine the potential effects of Didymo on sh spawning sites or
early life-stage survival (Bickel and Closs 2008). We therefore sug-
gest future work examine the relationship of Didymo blooms to
sh diet and growth in winter, as well as egg and larval sh sur-
vival. Nonetheless, we demonstrate at both reach and regional
scales that Didymo blooms do not seem to affect sh diets, condi-
tion, or growth in mountainous headwaters during summer.
Implications for management
Because Didymo blooms are visually obvious and macroinver-
tebrates assemblages are altered, it has been routinely hypothe-
sized Didymo negatively impacts sheries (Bickel and Closs 2008;
Beville et al. 2012;Klauda and Hanna 2016). Authors of previous
studies have suggested nutrient amendments (James et al. 2015;
Coyle 2016) and dam releases (Cullis et al. 2015)asviablemeansto
manage nuisance Didymo blooms. Indeed, both methods show
promise for reduction of blooms at local scales. The impetus for
this bloom reduction may be independent of concern for shes,
including aesthetics, fouling of infrastructure, or to prevent hy-
poxia. However, we did not observe any large impacts of Didymo
blooms on the diet, condition, or growth of trout in Kootenai ba-
sin headwaters. This overall result is similar to those for brown
trout in a South Dakota stream (James and Chipps 2016). There-
fore, clearly identifying objectives of Didymo control may be
useful. At present, it is not clear that objectives to increase
growth or condition of sh would be met by reducing Didymo
blooms in headwaters. Where Didymo control is implemented,
additional monitoring of invertebrates, sh diets, and growth
will lead to greater understanding of specic impacts. While less
often an objective for major stream restoration efforts, consider-
ing potential impacts of Didymo blooms to imperiled inverte-
brates (especially sedentary freshwater mussels that may not be
able to avoid blooms) may identify conservation opportunities of
Didymo control.
This work was supported by the British Columbia Ministry of
the Environment; the US Forest Service; Montana Fish, Wildlife &
Parks (in-kind); the US Geological Survey Utah Cooperative Fish
and Wildlife Research Unit (in-kind); and the Utah State University
School of Graduate Studies, Department of Watershed Sciences,
and Ecology Center. We thank Chuck Hawkins for helpful reviews
of previous versions of this manuscript. Jon McFarland, Ryan West,
Chris Clancy, and Marshall Wolf provided valuable assistance in
the eld. Thank you to Jay DeShazer, Ryan Sylvester, Jared
Lampton, Jordan Frye, Brian Stephens, Monty Benner, and Mike
Hensler for help at the Libby Field Ofce and to Jeff Burrows, Greg
Andrusak, Joe Thorley, Murray Pearson, and Jeff Curtis in British
Columbia. Gary Thiede at Utah State University (USU) and Brett
Roper and Mike Young of the US Forest Service provided additional
support. The National Aquatic Monitoring Center in Logan, Utah,
provided generous help with identication of sh diet samples.
Thanks also to the associate editor and two anonymous reviewers,
whose comments greatly improved the manuscript. Any use of
trade, rm, or product names is for descriptive purposes only and
does not imply endorsement by the United States Government.
This study was performed under the auspices of the USU IACUC
protocol No. 10006, The University of British Columbia animal
care protocol No. A19-0171, Montana scientic collectors permits
07-2018 and 16-2019, and British Columbia sh collection permit
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Appendix Table A1 appears on the following page.
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Appendix A
Table A1. Steams surveyed in 2019.
Subbasin (state
or prov.)
% Didymo
No. of
Fish spp.
48.95124 115.54154 Yaak R. (Mont.) 1.7 82.6 3.4 Pine B 12 10.1 RB
48.82052 115.29097 Koocanusa
33.3 80.2 4.7 Cedar A 6 12.5 WCT
Bear Cr. 48.16886 115.58853 Kootenai R.
30.6 88.7 8.3 Cedar B 19 11.9 BULL, RB
Big Cherry
48.20577 115.59134 Kootenai R.
29.0 74.5 5.9 Cedar A 75 12.5 BULL, RB,
Burnt Cr. 48.72936 115.87002 Yaak R. (Mont.) 31.6 64.5 7.3 Cedar A 12 14.4 EB, LN DC, MWF,
Coffee Cr. 49.69663 116.91781 Kootenay L. (B.C.) 2.0 14.3 Cedar A 12.9 BULL, WCT
Davis Cr. 50.14236 116.95520 Kootenay L. (B.C.) 20.5 60.6 9.8 Cedar A 5 11.8 BULL, MWF, RB,
E. Fork
Pipe Cr.
48.61675 115.61885 Kootenai R.
24.6 83.7 3.2 Alder B 12 10.9 EB, RB, RBCT,
E. Fork
Yaak R.
48.94885 115.53378 Yaak R. (Mont.) 52.6 31.5 7.6 Pine B 4 11.4 RB
Granite Cr. 48.29544 115.62011 Kootenai R.
21.3 68.1 10.3 Cedar B 7 11.6 BULL, EB, RB,
Hope Cr. 50.45751 117.19081 Lardeau R. (B.C.) 16.6 67.3 4.8 Cedar A 16 12.7 BULL, MWF, RB
49.60506 117.12635 Kootenay L. (B.C.) 0.0 32.0 15.9 Alder A 13.9 RB
Lake Cr. 48.44899 115.87918 Kootenai R.
27.0 6.3 19.0 Alder B 0 RB, TCOT
Leigh Cr. 48.22127 115.60603 Kootenai R.
3.2 88.0 4.6 Cedar A 20 9.9 EB, RBCT
Lizard Cr. 49.48972 115.10481 Elk R. (B.C.) 15.4 26.0 6.8 Pine B 5 11.5 EB, RBCT
49.50892 116.78628 Kootenay L. (B.C.) 40.6 80.4 5.0 Cedar A 11.5 BULL, LNDC,
Mobbs Cr. 50.50673 117.27104 Lardeau R. (B.C.) 48.3 28.1 3.2 Alder Side
8 10.6 BULL, RB, SLCOT
Lost Ledge
49.90834 116.90733 Kootenay L. (B.C.) 5.7 69.2 4.9 Cedar A 9 12.8 RB
N. Fork 17
Mile Cr.
48.66022 115.76750 Yaak R. (Mont.) 17.7 66.0 3.8 Cedar A 15 11.9 EB, RB, SLCOT,
Outlet Cr. 50.16812 115.46405 White R. (B.C.) 80.5 64.0 6.3 Pine B 8 18.2 RB
48.37814 115.62908 Kootenai R.
0.9 83.2 7.7 Cedar B 40 10.7 EB, RB, SLCOT
48.82799 115.24295 Koocanusa
39.4 53.4 4.5 Alder A 11.3 EB, RB
Solo Joe
48.92425 115.53963 Yaak R. (Mont.) 0.0 81.1 2.8 Pine A 21 11.7 RB
Trail Cr. 48.03825 115.46030 Fisher R. (Mont.) 13.0 63.7 4.9 Pine B 10 13.9 EB, TCOT, WCT
W. Fisher
48.04253 115.47351 Fisher R. (Mont.) 1.3 26.9 5.6 Pine B 2 11.1 BULL, TCOT,
Wolf Cr. 48.23393 115.28529 Fisher R. (Mont.) 0.0 32.0 9.8 Alder B 3 20.2 LNDC, LSSU,
49.80625 117.02835 Kootenay L. (B.C.) 0.2 37.9 10.1 Cedar B 8.6 BULL, WCT
Weasel Cr. 48.94904 114.73400 Wigwam R.
52.5 58.0 3.5 Pine B 5 15.3 RBCT
Note: Species codes: EB, brook trout; BULL, bull trout; LNDC, longnose dace; LSSU, largescale sucker; MWF, mountain whitesh; RB, rainbow trout; RBCT,
rainbowcutthroat hybrid; RSSH, redside shiner; SLCOT, slimy sculpin; TCOT, torrent sculpin; WCT, westslope cutthroat trout.
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... This study was completed as part of a larger project that examined the influence of benthic algae on fish food webs (Clancy et al. 2021). Bear and Ramsey creeks are tributaries to Libby Creek in the Kootenai River basin of northwestern Montana (Figure 1). ...
... The estimated mean June-August discharge at the Bear Creek site is 30.1 ft 3 /s (0.85 m 3 /s), and 30.2 ft 3 /s (0.86 m 3 /s) at the Ramsey Creek site (McCarthy et al. 2016). The mean summer water temperature at both study sites during 2018 was 9.79°C (Clancy et al. 2021). The fish assemblage was dominated by native Columbia River Redband Trout (hereafter, "Redband Trout") and federally threatened Bull Trout Salvelinus confluentus. ...
... Upon initial capture (during the first three sampling events), we anesthetized each trout that was larger than 100 mm with clove oil, implanted each with a 12-mm passive integrated transponder (PIT) tag (Model HDX12, Biomark, Boise, Idaho), and marked them by clipping a small piece of the caudal fin. We measured the total length and wet weights of the trout at each capture, recorded the PIT tag number, and gastrically lavaged a large proportion (91%) to obtain diets for a separate study (Clancy et al. 2021). We made multiple passes during the final sampling event to maximize the recapture of tagged fish. ...
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Research on fishes sometimes requires that individual fish be captured and subjected to invasive procedures multiple times over a relatively short time span. Electrofishing is one of the most common techniques used to capture fish, and it is known to cause injury to fish under certain circumstances. We evaluated the relationship of growth rates in Columbia River Redband Trout Oncorhynchus mykiss gairdneri to the number of times that they were captured via electrofishing and gastrically lavaged during the summer of 2018 in a mountainous, headwater stream. We captured fish between two and seven times over the course of 86 d using continuous (smooth) DC backpack electrofishing. We observed no relationship between the growth rate of Columbia River Redband Trout and the number of times that they were captured or gastrically lavaged. Although these findings contrast with hatchery electrofishing experiments, they may represent the greater resiliency of wild fish. It appears that researchers can use electrofishing and gastric lavage in cold waters at least once per month, and potentially up to twice per month, without greatly affecting the growth of wild Columbia River Redband Trout.
... Though D. geminata. Blooms do not appear to influence the condition of food resources for resident trout populations (Clancy et al., 2020), questions remain on whether severe blooms may interfere with spawning or redds, particularly if oxygen concentrations are depleted. Glacier loss is extensive in southern Canada , and these rivers have already experienced reduced seasonal streamflow (Brahney et al., 2017b). ...
Glaciers provide cold, turbid runoff to many mountain streams in the late summer and buffer against years with low snowfall. The input of glacial meltwater to streams maintains unique habitats and support a diversity of stream flora and fauna. In western Canada, glaciers are anticipated to retreat by 60-80% by the end of the century, and this retreat will invoke widespread changes in mountain ecosystems. We used a space-for-time substitution along a gradient of glacierization in western Canada to develop insights into changes that may occur in glaciated regions over the coming decades. Here we report on observed changes in physical (temperature, turbidity), and chemical (dissolved and total nutrients) characteristics of mountain streams and the associated shifts in their di-atom communities during de-glacierization. Shifts in habitat characteristics across gradients include changes in nutrient concentrations, light penetration, temperatures, and flow, all of which have led to distinct changes in di-atom community composition. Importantly, glacial-fed rivers were 3-5°C cooler than rivers without glacial contributions. Declines in glacial meltwater contribution to streams resulted in shifts in the timing of nutrient fluxes and lower concentrations of total phosphorus (TP), soluble reactive phosphorus (SRP), and higher dissolved inorganic nitrogen (DIN) and light penetration. The above set of conditions were linked to the overgrowth of the benthic diatom Didymosphenia geminata. These changes in stream condition and D. geminata colony development primarily occurred in streams with marginal (2-5%) to no glacier cover. Our data support a hypothesis that climate-induced changes in river hydrochemistry and physical condition lead to a phenological mismatch that favors D. geminata bloom development.
Full-text available
Didymosphenia geminata (Didymo) is a nuisance algae that can cover entire streambeds under certain environmental conditions. Numerous studies have shown that it changes the composition of stream invertebrates. Fishes in many headwaters are known to feed almost exclusively on invertebrates. Thus, there is concern changes to the amount or type of invertebrates caused by Didymo blooms will impact fishes such as trout, charr, and sculpin. In the Kootenai River basin of Montana and British Columbia, we examined stream invertebrates and fish diets, condition, and growth across 25 streams during the summers of 2018 and 2019. The severity of Didymo blooms in these streams ranged from 0 – 80% coverage of the entire streambed. In 2018, we observed significant shifts in the types of stream invertebrates available to trout in Didymo-affected streams. However, trout diets and growth rate were not affected. In 2019, trout, charr, and sculpin diets in streams with severe Didymo blooms were similar to streams with little to no Didymo. Condition of all three types of fish were unaffected. We therefore conclude that summer Didymo blooms have no obvious impacts on the diet, condition, or growth of these fishes. We suggest further studies document potential impacts during winter months and on sensitive invertebrates such as freshwater mussels.
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Aggregations of freshwater mussels create patches that can benefit other organisms through direct habitat alterations or indirect stimulation of trophic resources via nutrient excretion and biodeposition. Spent shells and the shells of living mussels add complexity to benthic environments by providing shelter from predators and increasing habitat heterogeneity. Combined, these factors can increase primary productivity and macroinvertebrate abundance in patches where mussel biomass is high, providing valuable subsidies for some fishes and influencing their distributions. We performed a 12-wk field experiment to test whether fish distributions within mussel beds were most influenced by the presence of subsidies associated with live mussels or the biogenic habitat of shells. We used remote underwater video recordings to quantify fish occurrences at fifty 0.25-m2 experimental enclosures stocked with either live mussels (2-species assemblages), sham mussels (shells filled with sand), or sediment only. The biomass of algae and benthic macroinvertebrates increased over time but were uninfluenced by treatment. We detected more fish in live mussel and sham treatments than in the sediment-only treatment but found no difference between live mussel and sham treatments. Thus, habitat provided by mussel shells may be the primary benefit to fishes that co-occur with mussels. Increased spatiotemporal overlap between fish and mussels might strengthen ecosystem effects, such as nutrient cycling, and the role of both fish and mussels in freshwater ecosystems.
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Compensatory growth—when individuals in poor condition grow rapidly to catch up to conspecifics—may be a mechanism that allows individuals to tolerate stressful environmental conditions, both abiotic and biotic. This phenomenon has been documented fairly widely in laboratory and field experiments, but evidence for compensatory growth in the wild is scarce. Cutthroat trout (Oncorhynchus clarkii subsp.) are cold‐water specialists that inhabit montane streams in western North America where seasonal conditions can be harsh and growth rates vary greatly among seasons. Understanding if individuals compensate for periods of reduced growth and body condition will improve understanding of the requirements of fish throughout their life‐cycle and across freshwater habitats. We quantified compensatory growth of juvenile cutthroat trout using extensive mark–recapture data from 11 stream populations (1,125 individuals) and two subspecies inhabiting a wide range of ecological settings in the northern Rocky Mountains, U.S.A. Our objectives were to determine how growth was linked across seasons and whether individuals behaviourally compensated for depressed body condition via emigration. Fish in relatively poor condition consistently demonstrated compensatory growth in mass during subsequent seasons. In contrast, fish in relatively better condition responded with positive growth in length during the summer signalling these fish may be better suited to headwater environments; no compensatory growth in length was found during the winter. Furthermore, there was no evidence that individual condition mediated migration tendencies of fish to seek more favourable habitat. Across a wide range of environmental conditions, we found consistent empirical support for compensatory growth in mass in the wild. A critical next step is to quantify how changing abiotic and biotic conditions influence the ability of stream fishes to compensate for locally or seasonally challenging conditions, thereby affecting long‐term resiliency, viability, and adaptation in the face of changing environmental conditions.
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Ecosystem engineers transform habitats in ways that facilitate a diversity of species; however, few investigations have isolated short-term effects of engineers from the longer-term legacy effects of their engineered structures. We investigated how initial presence of net-spinning caddisflies (Hydropsychidae) and their structures that provide and modify habitat differentially influence benthic community coloniza-tion in a headwater stream by conducting an in situ experiment that included three treatments: (1) initial engineering organism with its habitat modification structure occupied (hereafter caddisfly); (2) initial habitat modification structure alone (hereafter silk); and (3) a control with the initial absence of both engineer and habitat modification structure (hereafter control). Total invertebrate colonization density and biomass was higher in caddisfly and silk treatments compared to controls (~25% and 35%, respectively). However, finer-scale patterns of taxonomy revealed that density for one of the taxa, Chironomidae, was ~19% higher in caddisfly compared to silk treatments. Additionally, conspecific biomass was higher by an average of 50% in silk treatments compared to controls; however, no differences in Hydropsyche sp. biomass were detected between caddisfly treatments and controls, indicating initially abandoned silk structures elevated conspecific biomass. These findings suggest that the positive effects of the habitat modification structures that were occupied for the entirety of the experiment may outweigh any potential negative impacts from the engineer, which is known to be territorial. Importantly, these results reveal that the initial presence of the engineer itself may be important in maintaining the ecological significance of habitat modifications. Furthermore, the habitat modifications that were initially abandoned (silk) had similar positive effects on conspecific biomass compared to caddisfly treatments, suggesting legacy effects of these engineering structures may have pertinent intraspecific feedbacks of the same magnitude to that of occupied habitat modifications. Elucidating how engineers and their habitat modifications differentially facilitate organisms will allow for a clearer mechanistic understanding of the extent to which animal engineers and their actions influence aspects of community organization such as colonization.
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Riverscapes are complex, landscape-scale mosaics of connected river and stream habitats embedded in diverse ecological and socioeconomic settings. Social–ecological interactions among stakeholders often complicate natural-resource conservation and management of riverscapes. The management challenges posed by the conservation and restoration of wild salmonid populations in the Columbia River Basin (CRB) of western North America are one such example. Because of their ecological, cultural, and socioeconomic importance, salmonids present a complex management landscape due to interacting environmental factors (eg climate change, invasive species) as well as socioeconomic and political factors (eg dams, hatcheries, land-use change, transboundary agreements). Many of the problems in the CRB can be linked to social–ecological interactions occurring within integrated ecological, human–social, and regional–climatic spheres. Future management and conservation of salmonid populations therefore depends on how well the issues are understood and whether they can be resolved through effective communication and collaboration among ecologists, social scientists, stakeholders, and policy makers.
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Bioenergetics modeling is a widely used tool in fisheries management and research. Although popular, currently available software (i.e., Fish Bioenergetics 3.0) has not been updated in over 20 years and is incompatible with newer operating systems (i.e., 64-bit). Moreover, since the release of Fish Bioenergetics 3.0 in 1997, the number of published bioenergetics models has increased appreciably from 56 to 105 models representing 73 species. In this article, we provide an overview of Fish Bioenergetics 4.0 (FB4), a newly developed modeling application that consists of a graphical user interface (Shiny by RStudio) combined with a modeling package used in the R computing environment. While including the same capabilities as previous versions, Fish Bioenergetics 4.0 allows for timely updates and bug fixes and can be continuously improved based on feedback from users. In addition, users can add new or modified parameter sets for additional species and formulate and incorporate modifications such as habitat-dependent functions (e.g., dissolved oxygen, salinity) that are not part of the default package. We hope that advances in the new modeling platform will attract a broad range of users while facilitating continued application of bioenergetics modeling to a wide spectrum of questions in fish biology, ecology, and management.
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Human activities frequently result in either intentional or unintentional introductions of species to new locations, and freshwater environments worldwide are particularly vulnerable to species invasions. An introduced freshwater diatom, Didymosphenia geminata, was first discovered in New Zealand in 2004 but there was limited research available to predict the drivers of D. geminata biomass and how biomass variability might influence higher trophic levels (e.g. invertebrates and fish). We examined the effect of D. geminata biomass on benthic invertebrates, invertebrate drift and fish communities in 20 rivers in New Zealand with variable hydrology, physical habitat and water chemistry. Variation in D. geminata biomass was best explained by a model that showed D. geminata biomass increased with time since the last flow event exceeding three times the median annual discharge and decreasing concentration of dissolved reactive phosphorus. Analyses of biotic responses showed that high D. geminata biomass did not affect either invertebrate or fish diversity but altered the structure of benthic communities, changed the composition of drifting invertebrate communities and reduced fish biomass by 90 %, particularly trout. A partial least squares path model was used to disentangle both direct and indirect effects of D. geminata on fish communities and showed D. geminata had a significant negative direct effect on fish communities. This is the first study to show how the potential effects of the introduced diatom D. geminata can impact fish communities and has shown that D. geminata impacts fish both directly and indirectly through changes in their invertebrate prey community.