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\begin{document}

\title{Long-term responses of benthic invertebrates to rotenone treatment}



\author[1]{Gaute Kjærstad}%
\author[1]{Jo Arnekleiv}%
\author[2]{Gaute Velle}%
\author[1]{Anders Gravbrøt Finstad}%
\affil[1]{Norges teknisk-naturvitenskapelige universitet Vitenskapsmuseet Institutt for Naturhistorie}%
\affil[2]{NORCE Norwegian Research Centre AS}%


\vspace{-1em}



  
  \date{May 14, 2024}


\begingroup
\let\center\flushleft
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\selectlanguage{english}
\begin{abstract}
Biological invasions are regarded as one of the largest threats to
native biodiversity. The eradication of non-native parasites by culling
of hosts are a controversial conservation strategy, particularly when
using indiscriminate methods involving whole ecosystem collateral
damage. While short-term effects are abundantly documented, long-term
surveys are needed to detect potential wider ecosystem effects. Here, we
report a six-year study on effects of the piscicide rotenone on
invertebrate communities from a Norwegian water course using a
Before-After-Control-Impact design. Kick-net samples of benthic
invertebrates were collected from three lentic sampling stations and two
lotic stations two to four times per year in both a control and a
treated watershed. In general, only relatively minor short-term effects
immediately after the treatment on species turnover, measured as
temporal beta-diversity, of benthic invertebrates were observed both in
lentic and lotic locations. However, the lotic fauna was temporarily
severely negatively affected following a period of rotenone exposure
from an upstream lake. Species turnover co-varied markedly between
control and treatment locations, indicating that natural environmental
variation override effects of rotenone treatment. Likewise, the
abundance of invertebrate taxa varied considerably both over time and
between control and treatment locations. Our study indicates minor
short-term (i.e. \textless{} one month) or long-term (i.e. four years)
effects of rotenone treatment on benthic invertebrates, but severe
effects on the lotic fauna eight months after treatment. However,
long-term effects are likely to be taxa-specific and vary depending on
habitat connectivity and thus potential for re-colonization and will
differ among locations and among taxa.%
\end{abstract}%



\sloppy


\textbf{Long-term responses of benthic invertebrates to rotenone
treatment}

Running title: Rotenone effects on benthic invertebrates

\selectlanguage{ngerman}\textbf{Gaute Kjærstad\textsuperscript{1} \textbar{} Jo Vegar
Arnekleiv\textsuperscript{1} \textbar{} Gaute
Velle\textsuperscript{2,3}\textbar{} Anders Gravbrøt
Finstad\textsuperscript{1}}

\textsuperscript{1}Department of Natural History, NTNU University
Museum, Norwegian University of Science and Technology, Trondheim,
Norway

\textsuperscript{2}NORCE Norwegian Research Centre, LFI, Bergen, Norway

\textsuperscript{3}Department of Biological Sciences, University of
Bergen, Bergen, Norway

\textbf{Correspondence}

Gaute Kjærstad, Department of Natural History, NTNU University Museum,
Norwegian University of Science and Technology, Trondheim, Norway.

Email: gaute.kjarstad@ntnu.no

Acknowledgements

We thank the Norwegian Environment Agency and the County Governor of
Nordland for financial support. We are grateful to Mark Daverdin, Lars
Rønning and Aslak D. Sjursen for assisting the field work. Antti P.
Eloranta gave valuable comments on an earlier draft of the manuscript.

Abstract

Biological invasions are regarded as one of the largest threats to
native biodiversity. The eradication of non-native parasites by culling
of hosts are a controversial conservation strategy, particularly when
using indiscriminate methods involving whole ecosystem collateral
damage. While short-term effects are abundantly documented, long-term
surveys are needed to detect potential wider ecosystem effects.

Here, we report a six-year study on effects of the piscicide rotenone on
invertebrate communities from a Norwegian water course using a
Before-After-Control-Impact design. Kick-net samples of benthic
invertebrates were collected from three lentic sampling stations and two
lotic stations two to four times per year in both a control and a
treated watershed.

In general, only relatively minor short-term effects immediately after
the treatment on species turnover, measured as temporal beta-diversity,
of benthic invertebrates were observed both in lentic and lotic
locations. However, the lotic fauna was temporarily severely negatively
affected following a period of rotenone exposure from an upstream lake.
Species turnover co-varied markedly between control and treatment
locations, indicating that natural environmental variation override
effects of rotenone treatment. Likewise, the abundance of invertebrate
taxa varied considerably both over time and between control and
treatment locations.

Our study indicates minor short-term (i.e. \textless{} one month) or
long-term (i.e. four years) effects of rotenone treatment on benthic
invertebrates, but severe effects on the lotic fauna eight months after
treatment. However, long-term effects are likely to be taxa-specific and
vary depending on habitat connectivity and thus potential for
re-colonization and will differ among locations and among taxa.

\textbf{Keywords}

alien species, benthic invertebrates, beta diversity, piscicide,
recolonization

\textbf{1 \textbar{} INTRODUCTION}

Biological invasions are regarded as one of the largest threats to
native biodiversity (Mack et al., 2000), including in freshwater (Sala
et al., 2000, Caffrey et al., 2014). Alien species may display invasive
properties inducing serious ecological consequences through both direct
and indirect ecological interactions (Kumschick et al., 2015), as well
as major economic impacts (Dittel \& Epifanio, 2009). Such alien
invasive species may spread rapidly and cause problems that escalate
from local to regional following establishment. Early targeted
eradication actions are often therefore the only effective management
method. This approach may be particularly effective when species are
required to move through environmental corridors, such as in freshwater
ecosystems (Hulme, 2009). However, options for cost-effective and
reliable measures are often limited since it may become virtually
impossible to eradicate invasive aquatic species in anything but small,
enclosed waterbodies (Holdich et al., 1999).

In aquatic ecosystems, one of the most widely used measures for
eradication of invasive species is rotenone. This is a piscicide which
is highly toxic to fish and to certain invertebrates but has low
toxicity to birds and mammals (Ling, 2003). Rotenone is a botanical
compound originating from plant species in the Leguminosae family and
have been used for centuries as a mean of capturing fish in areas where
these plants are naturally occurring (Brooks \& Price, 1961, Meadows,
1973). It has also been used as a fish management tool for almost a
century (Finlayson et al., 2010), either with the purpose of quantifying
fish abundances, manipulating fish populations to maintain sport
fisheries or treatment of rearing ponds (McClay, 2000, 2005). It is one
of the few, or the only, cost-effective and widely accepted measure for
eradicating populations of invasive fish. In addition, rotenone has been
widely used for eradication of fish parasites through culling of host
populations.

Use of rotenone is controversial. This is largely due to its
indiscriminate impacts on the wider aquatic fauna. Short-term impacts on
non-target organisms, like benthic invertebrates is well documented
(e.g. Koksvik \& Aagaard, 1984, Melaas et al., 2001, Dalu et al., 2015).
The observed effects often include reduction in the total abundance of
benthic invertebrates comparted to pre-treatment levels. However, there
are large variations ranging from little or no reduction (Cook \& Moore,
1969, Dudgeon, 1990, Bellingan et al., 2019) to up to 95\% reduction
(Hamilton et al., 2009). Part of the variation may be explained by
taxa-specific tolerance to rotenone. Ephemeroptera, Plecoptera and
Trichoptera, which are abundant in oxygen rich lotic habitats, are
generally reported to be rotenone sensitive (Mangum \& Madrigal, 1999,
Arnekleiv et al., 2001, Eriksen et al., 2009), whereas Coleoptera,
Odonata and Gastropoda, which are often abundant in lentic habitats, are
considered rotenone tolerant (Chandler \& Marking, 1982, Holocombe et
al., 1987, Arnekleiv et al.,1997, Kjærstad \& Arnekleiv, 2011). There
are also species-specific differences in tolerance among closely related
taxa. For example, species of gill breathing Ephemeroptera are
particularly sensitive to rotenone (Arnekleiv et al. 2001). Recovery of
the total benthic invertebrate density can occur within a year (Kjærstad
\& Arnekleiv, 2004, Pham et al., 2018). However, recovery of single taxa
may potentially take several years (Arnekleiv et al., 1997, Mangum \&
Madrigal, 1999).

Long-term consequences of failing to act towards continuous spread of
invasive fish or parasites may be severe to the ecosystem and ecosystem
services. Short-term disturbances to the ecosystem are expected, and to
certain degree accepted, as appropriate collateral damage. However,
long-term consequences of rotenone treatment may be less accepted. It is
currently a lack of studies assessing long-term ecosystem effects caused
by rotenone treatments (Vinson et al., 2010). Therefore, this study
focused on the long-term consequences of an intensive rotenone treatment
on lotic and lentic benthic invertebrate communities. Specifically, we
tested for impact of rotenone treatment on the long-term abundance and
temporal turnover of key taxa by comparing time series from a rotenone
treated watercourse and a nearby control watercourse.

\textbf{2 \textbar{} METHODS}

\textbf{2.1 Study sites}

The Fusta (rotenone treated) and Drevja (control) water courses are
situated in Northern Norway (Figure 1). The Fusta watercourse has a
catchment area of 538 km\textsuperscript{2}. The largest and lowermost
lake of the catchment, Lake Fustvatnet (38 m a.s.l.), has a surface area
of 11 km\textsuperscript{2} and a maximum depth of 68 m. Concentrations
of total nitrogen and total phosphorus from water samples taken in
August 2014 of 70 \selectlanguage{greek}μ\selectlanguage{english}g/L and \textless{}2 \selectlanguage{greek}μ\selectlanguage{english}g/L, respectively, indicate
oligotrophic conditions. River Fusta, which is the outlet of Lake
Fustvatnet, is 8.5 km long and drains to the sea. The river has a mean
annual discharge of 34 m\textsuperscript{3} s\textsuperscript{-1}.

The Drevja watercourse (untreated), which is bordering the Fusta
watercourse, has a catchment area of 178 km\textsuperscript{2}. Lake
Drevvatn (47 m a.s.l.), the largest and lowermost lake in the catchment,
has a surface area of 5 km\textsuperscript{2} and a maximum depth of 40
m. Concentrations of total nitrogen and phosphorus from water samples
taken in August 2014 of 55 \selectlanguage{greek}μ\selectlanguage{english}g/L and \textless{}2 \selectlanguage{greek}μ\selectlanguage{english}g/L, respectively,
indicate oligotrophic conditions. River Drevjaelva, which is the outlet
of Lake Drevvatn is 17.7 km long and drains to the sea. The river has a
mean annual discharge of 12 m\textsuperscript{3} s\textsuperscript{-1}.

Both watercourses are surrounded by sparsely vegetated mountain areas,
spruce dominated forests and scattered farmlands.

\textbf{2.2 \textbar{} Study design}

We performed the study using a Before-After-Control-Impact design
(Josefsson et al., 2020). Rotenone was used in the Fusta watercourse to
eradicate the invasive parasite \emph{Gyrodactylus salaris} through
exterminating its host populations of salmonid fish. Three rotenone
treatments were conducted in the watercourse, with two treatments of the
River Fusta (August 2011 and 2012) and one subsequent treatment of the
upstream Lake Fustvatnet (October 2012), causing the downstream River
Fusta to be treated a third time. The neighboring Drevja watercourse was
used as a control site (Figure 1). Sampling was conducted at two lotic
and three lentic sampling stations in each watercourse during the autumn
from 2011 to 2016, as well as one spring sampling in June 2013, eight
months after the lake treatment. Sampling started in August 2011 before
the first rotenone treatment and ended in Autumn 2016. An overview of
times of sampling and rotenone treatments is given in Table 1.

Figure 1

Riverine stations were situated in riffle areas dominated by cobbles
(particle size of 64-256 mm) and pebbles (particle size 4-64 mm) and
with scattered patches of river moss and alga. Two of the stations in
each lake were situated in exposed areas were pebbles (particle size of
4-64 mm) dominated the substrate. One of the stations in each lake was
situated in sheltered areas with a substrate dominated by finer
inorganic and organic substrate (particle size \textless{} 2mm) and with
little or no aquatic vegetation.

Table 1.

\textbf{2.3 \textbar{} Sampling and sample processing}

Benthic invertebrates were collected using a kick net with a 25x25 cm
frame and a mesh size of 0.25 mm. For each sample, an area of
approximately 0.5 m\textsuperscript{2} was covered with a sampling
duration of one minute. At each combination of sampling occasion and
sampling station, five replicate samples were taken. The samples were
preserved in 96\% ethanol before subsequent subsampling of each sample
(1/10) in the laboratory. All invertebrates in the subsample were sorted
and identified to the lowest possible taxonomic unit using a stereo
microscope and following Engblom (1996) for Ephemeroptera, Lillehammer
(1988) for Plecoptera, Rinne \& Wiberg-Larsen (2017) for Trichoptera,
Holmen (1987) and Nilsson \& Holmen (1995) for Coleoptera and Gl\selectlanguage{ngerman}öer
(2002) for Gastropoda. The number of specimens from the sub sample were
then multiplied by 10. All specimens of the remaining sample that
differed morphologically from the specimens in the sub sample, were
picked out, identified to lowest possible taxon and counted.

\textbf{2.4 \textbar{} Rotenone treatment and fish reestablishment}

The rotenone treatments of River Fusta were performed on August
18\textsuperscript{th} 2011 and August 17\textsuperscript{th} 2012 using
a rotenone formulation CFT-Legumine with a 2.5\% active gradient. In
both treatments the release of rotenone took place over a 7-hour period
from two stations in the river, in the upper and in the middle part
(Stensli \& Wist, 2014). The upper release was situated approximately 1
km upstream of our uppermost sampling station and the release in the
middle part was just downstream of our lowermost sampling station. In
addition, riverbanks of slow flowing areas, adjacent ponds, oxbow lakes
and streams were rotenone treated with local point releases. The total
use of CFT-Legumine in River Fusta was 754~and 573 L in 2011 and 2012,
respectively. The water temperature was approximately 15° C during both
treatments and the water flow was 12
m\textsuperscript{3}s\textsuperscript{-1} in 2011 and 11
m\textsuperscript{3}s\textsuperscript{-1} in 2012. In the lower part of
River Fusta, treatment concentration of CFT-Legumine was higher than
0.5 ppm over a period of several hours, which was the desired minimum
concentration, periodically reaching up to 7 ppm (Adolfsen et al.,
2014).

The rotenone treatment of Lake Fustvatnet took place between October
16\textsuperscript{th} and October 20\textsuperscript{th} 2012. The
total use of CFT-Legumine in Lake Fustvatnet was 139,200 L, of which
96,600 L were used for the surface waters and 42,600 L were used for
deeper areas. A CFT-Legumine formula of 3.3\% rotenone was used, except
for the surface layer where the treatment included 25,800 L with 2.5\%
rotenone (Moen \& Bardal, 2014). The 3.3\% formula did not contain the
synergist piperonylbutoxid and the solvent N-metylpyrrolidone since the
former increases the toxicity to invertebrates, and not to fish
(Finlayson et al., 2010). Measurements of rotenone concentration were
performed over scattered sites in the surface water and at different
depths during and after the treatment. On October
21\textsuperscript{st}2012 no measurements were below 0.5 ppm
CFT-Legumine, and the average concentration was 0.68 ppm. Due to low
water temperatures, the breakdown of rotenone was slow. In April 2013,
the average concentration of CFT-Legumine was 0.1 ppm and 0.3-0.4 ppm in
the surface layers and in the deeper parts (\textgreater{}10m),
respectively. In June 2013, the concentration was below 0.1 ppm at all
depths. No rotenone was detected in mid-October 2013 (Adolfsen et al.,
2014).

Fish was reestablished by stocking of Atlantic salmon (\emph{Salmo
salar} ) in River Fusta and Arctic charr (\emph{Salvelinus alpinus} )
and brown trout (\emph{Salmo trutta} ) in Lake Fustvatnet. Three
thousand Atlantic salmon smolts were released in the river in 2013 (Lo
\& Holthe, 2014) followed by planting of 187,000 Atlantic salmon eggs in
2014. In both 2015 and 2016, 12,000 Atlantic salmon smolts were
released, as well as 384,000 young of the year (YOY) Atlantic salmon in
2015 and 120,00 YOY in 2016 (Holthe et al. 2018). In Lake Fustvatnet
35,000 Arctic charr, mainly 2-year old, were released during the spring
of 2014, as well as 93,000 brown trout YOY (Lo \& Holthe, 2014, Holthe
et al., 2015).

\textbf{2.5\textbar{} Data analyses}

Changes in taxa composition between each subsequent sampling period was
measured as temporal beta-diversity index (TBI). We calculated two
different TBI indices, using the percentage difference form abundance
data and the Sorensen index based on presence-absence data. The
percentage difference (D\textsubscript{\%diff} ) between two sampling
periods (\emph{t\textsubscript{i}} and \emph{t\textsubscript{i+1}} ) was
defined as D\textsubscript{\%diff}= (B+C)/(2A+B+C), where A is the sum
of the abundance of all taxa found in both sampling
period\emph{t\textsubscript{i}} and \emph{t\textsubscript{i+1}} . This
represents the unscaled similarity between the two sampling periods. B
is the part of abundance of the common taxa that is higher in samples
at\emph{t\textsubscript{i}} than at \emph{t\textsubscript{i+1}} and is
the unscaled sum of taxa \emph{losses} between sampling periods. C is
the part of the abundance that is higher \emph{t\textsubscript{i+1}}
than at\emph{t\textsubscript{i}} and is the unscaled sum of taxa
\emph{gains}between the two sampling periods. The equation for the
Sorensen index, using presence-absence data, is the same as for the
D\%\textsubscript{diff} outlined above (see Legendre (2019) for more
details). Differences in abundances and presence-absence data among
subsequent sampling occasions were tested for significance with a paired
t-test. The p\selectlanguage{english}-values (based upon 9999 random permutations) were
controlled for the family-wise error rate using Holm correction. The
calculations of the TBI indices and significance tests were done in the
R package ``adespatial'' (Dray et al., 2019).

We analyzed data from riverine and lake habitat separately. Responses to
rotenone treatment is expected to vary among habitats. The abundance of
each invertebrate taxon is expressed as average number of specimens per
kick-sample for each sampling period at each site.

\textbf{Results}

\textbf{Rivers}

The temporal species turnover appeared to co-vary in the treated and
untreated river, especially for abundance data (Figure 2). The exception
was the period after the lake rotenone treatment in October 2012 to
summer 2013 where benthic invertebrate abundances of the rotenone
treated River Fusta changed significantly (Figure 2a, Appendix 1),
mainly due to losses (Figure 2b). For presence-absence data, there was a
significant dissimilarity of species occurrences in the treated river
from immediately after the lake treatment in October 2012 until August
2015 (Figure 2d, Appendix 2). The species loss-rate of the treated river
was at its highest from October 2012 to June 2013 (Figure 2e) and
gain-rate at its highest from June 2013 to August 2013 (Figure 2f,
Appendix 2). No significant changes in abundance or species occurrence
were found in the untreated river (Figure 2a, b, Appendix 1 and 2).

Figure 2

The total abundance of benthic invertebrates reflects the temporal
turnover indices (Figure 3a). There were no clear short-term effects on
the total abundance or the abundance on most taxa following the two
rotenone treatments of River Fusta (Figure 3). As an exception, the
abundance of Plecoptera decreased strongly in the treated river
immediately after the first and second river treatment (89\% and 92\%,
respectively), whereas the changes were smaller in the untreated river
(+17\% and -37\%, respectively) (figure 3c).

Figure 3

In contrast, there was a marked increase in the abundance of taxa
observed in samples from both the treated and the untreated river in
October 2012, followed by a pronounced decline in the treated river in
June 2013 (Figure 3a). However, the observed response in total abundance
masks distinct differences in response to the lake rotenone treatment of
riverine benthic invertebrate taxa.

The abundance of Ephemeroptera in the treated river decreased shortly
after the two river treatments (Figure 2b). This was also found in the
untreated river after the first river treatment\emph{.} The
Ephemeroptera \emph{Baetis rhodani and Baetis} sp. decreased in the
treated river after both treatments, and in the untreated river after
the first treatment. Abundances of all taxa from the rivers not
presented in the figures are listed in Appendix 3. For other
Ephemeroptera, like \emph{Heptagenia dalecarlica} and
\emph{Ephemerella}sp., the abundance in both rivers was relatively high
shortly after the river treatments. Immediately after the lake treatment
in October 2012,\emph{B. rhodani} was the only Ephemeroptera with
reduced abundance and was not recorded in the treated river. In June
2013, eight months after the lake treatment, there was a dramatic
decline in the Ephemeroptera abundance in the treated river (Figure 3b)
with only a few specimens present of \emph{Siphlonurus} sp., \emph{B.
rhodani} and \emph{H. dalecarlica} . However, a decline was also found
in the untreated river, but here most taxa were present in June 2013. In
the subsequent years, the Ephemeroptera abundances in the treated river
increased, mainly\emph{B. rhodani} and \emph{H. dalecarlica} . A similar
increase occurred in untreated river, in addition to an increase in
\emph{B. muticus.}

There was a strong decrease in the abundance of Plecoptera in the
treated river after both river treatments, whereas there was an increase
in the untreated river. Taxa with reduced abundances
included\emph{Diura} sp., \emph{Isoperla} sp. and \emph{Leuctra} sp.
Shortly after the lake treatment in October 2012, the abundance of
Plecoptera in the treated river decreased with 95\% compared to
September 2012 (Figure 3c), and only three specimens of the genus
\emph{Nemoura} were recorded. Most taxa were present in the untreated
river and with a relatively high abundance of \emph{Amphinemura
borealis} , \emph{Diura} sp. and\emph{Isoperla} sp. In June 2013, eight
months after the lake treatment, no Plecoptera were detected in the
treated river, while at the same time there was an average of more than
200 specimens of Plecoptera in the untreated river (Figure 3c). Contrary
to the river treatments, were the negative effect were minor,
Trichoptera abundances decreased in the treated river immediately after
the lake treatment but did not change in the untreated river (Figure
3d). Eight months after the lake treatment in June 2013, there was a
dramatic decrease (99\%) in the abundance of Trichoptera in the treated
river, and only a few specimens of\emph{Rhyacophila nubila} ,
\emph{Hydropsychidae} sp. and\emph{Micrasema} sp. were present. A
corresponding but smaller decline (41\%) was observed in the untreated
river.

Both rivers had high abundance of Chironomidae shortly after the lake
treatment (Figure 3e). The abundance of Chironomidae co-varied
considerably among rivers. The abundance of Oligochaeta in the treated
river decreased shortly after the first river treatment and increased
subsequent to the second river treatment and the lake treatment (Figure
3f). In June 2013, eight months after the lake treatment, the abundance
of Oligochaeta was at its minimum in the treated river.

There was a low abundance of Hydrachnidia in the treated river and the
taxon was absent in 2013 (Appendix 3). In the untreated river,
Hydrachnidia was recorded in all sampling occasions and with the highest
abundances after the treatments. The abundance of the riffle beetles
(Elmidae) decreased in the treated river following the treatment. In the
untreated River Drevjaelva, the taxon also decreased in abundance,
albeit at a much lower rate than in the treated River Fusta.

\textbf{Lakes}

Temporal species turnover appeared to co-vary in the treated and
untreated lakes (Figure 4). The untreated lake tended to have a greater
temporal turnover, both measured as dissimilarity in taxa abundances and
species occurrences throughout most of the study period (Figure 4a and
4d). The highest losses in terms of species abundance occurred in the
untreated lake in August 2012 and for the treated lake in August 2015
(Figure 4b). Gains were generally highest for the treated lake (Figures
4c and 4f), especially for abundance data. No significant changes in
abundance or species occurrence indices between sampling periods were
found in the treated or untreated lake (Appendix 4 and 5).

Figure 4

The total abundances of benthic invertebrates in the lakes showed major
variations even before the rotenone treatment, i.e., the abundances
decreased by about 50\% from in from August 2011 to August 2012 in both
lakes (Figure 5a). In October 2012, shortly after the treatment, the
total abundance in both lakes decreased, but stronger in the rotenone
treated Lake Fustvatnet. In June 2013, eight months after the treatment,
a minor decrease occurred in both lakes. From August 2013 and onwards
the variations in abundance differed between the lakes, with higher
values in the treated lake in August 2013, August and October 2014,
October 2015 and August and October 2016.

Figure 5

The Ephemeroptera abundance varied over time both within and between
lakes (Figure 5b). There was a high abundance in both lakes in June
2013, compared to pre-treatment abundances mainly caused by the
Ephemeroptera \emph{Siphlonurus} sp. Abundances of all taxa from the
lakes not presented in the figures are listed in Appendix 6. High
abundances were also registered in the treated lake in August 2014 and
August 2016. In October 2012, shortly after the lake treatment the
abundances was lower in both lakes, compared to samples taken just
before the treatment. The abundance of Ephemeroptera taxa decreased
immediately after the lake treatment, especially \emph{Centroptilum
luteolum} in the treated lake and Leptophlebiidae in both the treated
and untreated lakes. Ephemeroptera species with increased abundances in
the treated lake after the treatment included \emph{Centroptilum
luteolum} , \emph{Caenis horaria} and \emph{Ephemera} \emph{vulgata} .
However, the abundance of \emph{E. vulgata} . decreased strongly in
august and October 2015 (73\% and 71\%, respectively), compared to 2014.
Except from \emph{Siphlonurus} sp. there was no marked temporal change
in the abundances of Ephemeroptera in the untreated lake.

The abundance of Plecoptera in the treated lake was generally low except
in October 2012 just before the rotenone treatment (Figure 5c). In the
untreated lake, abundances varied but were generally much higher than in
the treated lake. In the treated lake shortly after the treatment in
October 2012, there was a strong decrease in Plecoptera abundances
(Figure 5c), with only a few specimens of the genus
\emph{Nemoura}present. The decline was much smaller in the untreated
lake. In June 2013, eight months after the treatment, no Plecoptera were
recorded in the treated lake, and only a few specimens were observed in
the untreated lake. The genus \emph{Capnia} had the most marked decrease
in abundance in the treated lake following the treatment, whereas the
abundance of \emph{Capnia} peaked in some of the years after the
treatment in the untreated lake.

For Trichoptera, the abundances in both lakes varied across sampling
occasions (Figure 5d). In October 2012, shortly after the treatment,
abundances decreased markedly in the treated lake but did not change in
the untreated lake. However, a decreased abundance also occurred before
the treatment in both lakes between August 2011 and August 2012. The
highest abundances in the treated lake after the treatment were recorded
in August 2013 and in August 2014. In the untreated lake, the abundances
peaked in 2016. Polycentropodidae had the most pronounced decrease among
Trichoptera after the treatment in the treated lake, whereas the family
increased slightly during the same period in the untreated lake. A few
Trichoptera taxa in the treated lake, all found at low abundances, were
only recorded before or after the treatment.

The abundance of Chironomidae showed only minor changes after the
treatment in both lakes (Figure 5e). In subsequent years, the abundances
of Chironomidae in the treated lake varied compared to pre-treatment
levels, whereas it did not change or were higher in the untreated lake.

The abundance of Oligochaeta (Figure 5f) and Hydrachnidia generally
varied to a similar extent in both lakes. For Corixidae and Coleoptera,
and especially of the genus \emph{Callicorixa} and dytiscid beetles, the
abundances increased in the treated Lake Fustvatnet after the treatment.
No Corixidae was recorded in untreated Lake Drevvatnet, while Dytiscidae
occurred sporadically.

\textbf{Discussion}

The immediate impact of rotenone treatments in August 2011 and 2012 was
minor for benthic invertebrates in the rotenone treated River Fusta.
Only a few taxa showed a marked decrease in abundance in the treated
river compared to the untreated River Drevjaelva, including the
Ephemeroptera \emph{Baetis rhodani} and the Plecoptera
genera\emph{Diura} and \emph{Isoperla} . Because the upstream Lake
Fustvatnet was treated two months after the last river treatment, River
Fusta also received rotenone from the lake. Still, the immediate effect
of the lake rotenone treatment on benthic invertebrate community in the
downstream river was minor. One exception was the Trichoptera
\emph{Rhyacophila nubila,} which occurred at high abundances in both
rivers, but which disappeared from the River Fusta subsequently to the
treatment of the upstream lake. All the above-mentioned taxa are known
to be rotenone sensitive (Engstrom-Heg et al., 1978, Arnekleiv et al.,
2001, Gladso \& Raddum, 2002) and negative impacts were expected.

The rotenone treatment of Lake Fustvatnet took place in October, just
before the lake ice formed. Breakdown of rotenone was thus slowed down
due to low temperatures and limited light. Consequently, the treated
river was exposed to rotenone from the lake for several months. Even
though the toxic effect of rotenone decreases with decreasing water
temperatures (Meadows, 1973, Kjaerstad et al. 2015), the long exposure
apparently had a pronounced negative effect. In June 2013, eight months
after the lake treatment, we recorded the lowest abundance of benthic
invertebrates in the treated river. A similar decrease did not occur in
the untreated river. The benthic invertebrate community in the treated
river at this time were dominated by Chironomidae, Oligochaeta, Nematoda
and elmid beetles. Other taxa occurred at very low abundances and
several taxa were missing, including rotenone tolerant species like the
Ephemeroptera \emph{Ephemerella} . Moreover, despite that the freshwater
pearl mussel (\emph{Margaritifera margaritifera} ) is known to be
rotenone tolerant (Dolmen et al., 1995), all known specimens in the
treated river died during the spring of 2013 (Larsen, 2015). The
temporal beta diversity confirms that the changes in taxa composition
were by far highest, and significant, in the treated river from October
2012 to June 2013. This decline in benthic invertebrates was mainly due
to losses both in terms of taxa abundances and presence-absence. This
indicates that rivers may receive rotenone for several months following
treatment of upstream lakes, which can have major impacts. Most taxa in
the treated river, including rotenone tolerant taxa, decreased to very
low abundances and some disappeared, like the freshwater pearl mussel.

Despite the strong negative impacts recorded in the treated river eight
months after the lake rotenone treatment, recolonization was relatively
fast. The highest gains in terms of species occurrence in the treated
river occurred in August 2013. Already in August 2014 the total
abundances exceeded pre-treatment levels. In August and October 2016,
the change in taxa composition, in terms of abundance and
presence-absence of taxa between sampling occasions, were rather similar
for the treated and the untreated river. This indicates that the benthic
fauna of the treated river had recovered to a great extent. Several
taxa, like the Ephemeroptera \emph{Heptagenia dalecarlica}
and\emph{Ephemerella} , the Plecoptera \emph{Isoperla} and
\emph{Amphinemura borealis} and Oligochaeta reached much higher
abundances in 2013, as compared to pre-treatment levels. Ephemeroptera
and Plecoptera are known to be among the most abundant groups in the
drift of swift-flowing temperate streams (Brittain \& Eikeland, 1988).
These groups likely colonized rapidly from drift in untreated tributary
streams, although some may have survived as eggs. Oligochaeta may have
avoided rotenone exposure by digging deeper into the substratum. It took
almost four years after the last rotenone treatment before most of the
taxa had recolonized. The Plecoptera \emph{Taeniopteryx nebulosa} were
among the slowest to recolonize and appeared in 2016. Also, two rotenone
tolerant taxa, the Ephemeroptera \emph{Ephemerella} and elmid beetles
did not reached pre-treatment abundances in the treated river four years
after the treatments.

According to Vinson et al. (2010), the overall invertebrate abundances
generally return to pre-treatment levels quicker than the biodiversity
and taxonomic composition measures. This is in accordance to our
findings. However, the recolonization will greatly depend on the
taxonomic resolution of the data. Biodiversity measured at family and
genera level will recolonize fast, while it may take longer before all
species to recolonize. We have identified some taxa to species level,
such as Ephemeroptera, Plecoptera and Trichoptera, and some taxa to a
higher taxonomic level, such as the specious Chironomidae and
Hydrachnidia. Thus, changes in abundance and recolonization patterns of
many species in this study remain unknown.

A reduction of total benthic invertebrate abundance was observed both in
the treated lake and untreated lake immediately after the lake
treatment, but to a higher extent in the treated lake. This indicates
both natural changes and a negative effect of rotenone. Taxa with
clearly decreasing abundances shortly after the treatment in Lake
Fustvatnet, but not in the untreated Lake Drevvatnet, included the
Ephemeroptera \emph{Centroptilum luteolum} and the
Plecoptera\emph{Capnia} sp. While \emph{Capnia} is known to be rotenone
sensitive,\emph{C. luteolum} is quite tolerant (Kjaerstad \& Arnekleiv,
2011). The reason for the seemingly high sensitivity of \emph{C.
luteolum} to rotenone in the treated lake is unknown.

The recolonization in terms of abundance of benthic invertebrates in the
treated lake was fast, reaching pre-treatment levels eight months after
the treatment. A few increased in abundance after the treatment in the
treated lake, but not in the untreated lake. This was particularly
evident for the diving beetles (Dytiscidae). Beetles are generally known
to be very rotenone tolerant (Engstrom-Heg et al., 1978). Some will also
benefit from increased prey availability because of the general increase
in invertebrate abundance after the recolonization. Even though most
taxa increased in abundance during the years following the treatments,
the Plecoptera \emph{Capnia} , the Trichoptera Polycentropodidae and the
Gastropoda \emph{Gyraulus acronicus} decreased and had not reached their
pre-treatment levels in the treated lake four years after the treatment.
Considering that Gastropoda are rotenone tolerant (Holocombe et al.,
1987, Arnekleiv et al., 1997), the decline of \emph{G. acronicus} in the
treated lake was surprising. Mangun \& Madrigal (1999) found that of
several taxa missing five years after rotenone treatments, most belonged
to rotenone sensitive groups, like Ephemeroptera, Plecoptera and
Trichoptera. Taxa that were tolerant to rotenone were identified as''
non-missing`` or ''briefly missing`` (Mangun \& Madrigal, 1999).

Fish may play an important role in structuring benthic invertebrate
communities (e.g. Weyl et al., 2010, Alexiades \& Kraft, 2017). Stocking
of Atlantic salmon (\emph{S. salar} ) took place in the treated river
between 2013 and 2016, whereas Arctic charr (\emph{S. alpinus} ) and
brown trout (\emph{S. trutta} ) were stocked in the treated lake in
2014. A few benthic invertebrate taxa did not reach pre-treatment
abundance. However, if fish was the main driver, we would expect a
decrease in abundance of other common taxa that are eaten by fish in the
years after the treatments. Instead, many taxa increased in abundances
and exceeded pre-treatment levels, indicating a relatively low predation
pressure from fish in the treated river. In the treated lake, contrary
to the untreated lake, we observed a strong decrease in total abundances
of benthic invertebrates in August 2015. This was most likely due to
predation from fish stocked during the previous year. The decrease was
particularly high for Chironomidae and Ephemeroptera. Abundance of large
sized and important fish prey, like the Ephemeroptera \emph{Ephemera
vulgata} , increased to a high abundance in 2013-2014, but decreased
drastically in 2015 in the treated lake, probably due to predation from
fish. The species was not found before the treatment, but occurred in
all sampling events after the treatment, indicating a lower predation
pressure from fish after the treatment. Despite the rotenone treatments
and the fish removal/fish stocking, the abundance of many taxa varied to
a similar extent in the impact and control sites. As an exception, the
effect was more prominent in the impacted treated river than in the
untreated river after the lake treatment and subsequent months of
rotenone exposure. This illustrates the importance of using control
sites when performing time-series experiments.

It is striking to observe the large fluctuations in the abundance and in
the presence-absence of benthic invertebrates over time of both treated
and untreated localities. Even before the lake treatment, the
invertebrate abundances in both lakes were reduced by approximately 50\%
from August 2011 to August 2012. According to Vinson et al. (2010), the
accumulation curve for benthic invertebrate genera showed little
inclination for flattening out during a ten-year monthly sampling of a
Utah river. We did not see the same pattern in our data, perhaps due to
fewer taxa present. Even though it will vary greatly with habitat type,
latitude etc., a rising taxa accumulating curve would certainly be found
elsewhere and shows how random the community composition may be
registered at any given time.

Our study found minor short-term (i.e. \textless{} one month) effects of
the river rotenone treatments, whereas the lake treatment caused major
temporary negative effects on the lotic fauna eight months after the
treatment. No major short- or long-term (i.e. four years) effects on
lake benthic invertebrate communities where detected. Recovery time of
invertebrate abundance after the lake treatment was fast whereas the
recovery in terms of presence-absence of species were comparably much
longer, lasting for several years. The current study underpins the
importance of performing long-terms studies during investigation of the
effects of rotenone on benthic invertebrates.

\textbf{References}

Adolfsen, P., Sandvik, M. \& Waaler, T. (2014). Rotenonanalyser og
resultater. In: Bekjempelse av \emph{Gyrodactylus salaris} i
Vefsnaregionen. (Eds: J. H. Stensli \& H. Bardal). Veterinaerinstituttets
rapportserie 2-2014.

Alexiades, A.V. \& Kraft, C. E. (2017). Effects of stocked trout on
stream invertebrate communities. \emph{Journal of Freshwater Ecology,
32} , 95-102. https://doi.org/10.1080/02705060.2016.1248502

Arnekleiv, J. V., Dolmen, D., Aagaard, K., Bongard, T. \& Hanssen, O.
(1997). Rotenonbehandlingens effekt pa bunndyr i Rauma og
Hensvassdraget, More \& Romsdal. Del I: Kvalitative undersokelser.
Vitenskapsmuseet Rapport Zoologisk Serie 1997-8.

Arnekleiv J.V., Dolmen D. \& Ronning L. (2001). Effects of rotenone
treatment on mayfly drift and standing stocks in two Norwegian rivers.
In: E. Dominguez (ed). Trends in Research in Ephemeroptera and
Plecoptera. Kluwer Academic/Plenum Publishers.

Bellingan, T.A., Hugo, S., Woodford, D.J., Gouws, J., Villet, M.H. \&
Weyl, O.L.F. (2019). Rapid recovery of macroinvertebrates in a South
African stream treated with rotenone. \emph{Hydrobiologia, 834} , 1--11.
https://doi.org/10.1007/s10750-019-3885-z

Brittain, J.E. \& Eikeland, T.J. (1988). Invertebrate drift -- A
review.\emph{Hydrobiologia 166} , 77-93.
https://doi.org/10.1007/BF00017485

Brooks, I. C. \& Price, R. W. (1961). Studies on the chronic toxicity of
Pro\selectlanguage{english}-Noxfish(r), a proprietary synergized rotenone
fish\selectlanguage{english}-toxicant.\emph{Toxicology and Applied Pharmacology, 3,} 49--56.
https://doi.org/10.1016/0041-008X(61)90007-2

Caffrey J.M., Baars J.R., Barbour J.H., Boets P., Boon P. \& Davenport
K. (2014). Tackling invasive alien species in Europe: the top 20 issues.
Management of Biological Invasions, 5, 1-20.
http://dx.doi.org/10.3391/mbi.2014.5.1.01

Chandler, J. H. \& Marking, L.L. (1982). Toxicity of Rotenone to
Selected Aquatic Invertebrates and Frog Larvae. \emph{The Progressive
Fish-Culturist 44} , 78-80.
https://doi.org/10.1577/1548-8659(1982)44{[}78:TORTSA{]}2.0.CO;2

Cook, S.F. jr. \& Moore, R.L. (1969). The effects of rotenone treatment
on the insect fauna of a California stream. \emph{Transactions of the
American Fisheries Society, 98} , 539-544.
https://doi.org/10.1577/1548-8659(1969)98{[}539:TEOART{]}2.0.CO;2

Dalu, T., Wasserman, R.J., Jordaan, M., Froneman, W.P. \& Weyl, O.L.F.
(2015). An Assessment of the Effect of Rotenone on Selected Non-Target
Aquatic Fauna. \emph{-- PloS ONE}

\emph{(11)} : e0142140. https://doi.org/10.1371/journal.pone.0142140

Dar P.A. \& Reshi Z.A. (2014). Components, processes and consequences of
biotic homogenization: a review. \emph{Contemporary Problems of Ecology
7,} 123--136. https://doi.org/10.1134/S1995425514020103

Dittel, A. \& Epifanio, C. (2009). Invasion biology of the Chinese
mitten crab \emph{Eriochier sinensis} : A brief review. \emph{Journal of
Experimental Marine Biology and Ecology, 374} , 79-92.
https://doi.org/10.1016/j.jembe.2009.04.012

Dolmen, D., J.V. Arnekleiv, \& T. Haukebo. (1995). Rotenone tolerance in
the freshwater pearl mussel \emph{Margaritifera margaritifera}
.\emph{Nordic Journal of Freshwater Research 70} , 1-30.

Dray, S., Bauman, D., Blanchet, G., Borchard, D., Clappe, S., Guenard,
G., \ldots{}Wagner, H.H. (2019). Adespatial: Multivariate multiscale
spatial analsysis. R package version 0.3-3. Retrieved from
https://cran.r-project.org/package=adespatial

Dudgeon D. (1990). Benthic community structure and the effect of
rotenone piscicide on invertebrate drift and standing stocks in two
Papua New Guinea streams. \emph{Archives of Hydrobiology 119} , 35--53.

Engblom, E. 1996. Ephemeroptera, Mayflies. In: Aquatic insects of North
Europe. A taxonomic handbook. Vol 1. (Ed: A. Nilsson). Apollo Books,
Stenstrup.

Engstrom-Heg R., Colesante R.T. \& Silco, E. (1978). Rotenone tolerances
of stream-bottom insects. \emph{New York Fish Game Journal 25} , 31--41.

Eriksen, T. E., Arnekleiv, J. V. \& Kjaerstad, G. (2009). Short-Term
Effects on Riverine Ephemeroptera, Plecoptera, and Trichoptera of
Rotenone and Aluminum Sulfate Treatment to Eradicate \emph{Gyrodactylus
salaris. Journal of freshwater ecology, 24} , 507-607.
https://doi.org/10.1080/02705060.2009.9664337

Finlayson, B., Somer, W. L. \& Vinson, M. R. (2010). Rotenone toxicity
to rainbow trout and several mountain stream insects. \emph{North
American Journal of Fisheries Management 30} , 102--111.
https://doi.org/10.1577/M09-078.1

Finlayson, B., R. Schnick, D. Skaar, J. Anderson, L. Demong, D.
Duffield, W., \ldots{}Steinkjer, J. (2010). Planning and standard
operating procedures for the use of rotenone in fish
management---rotenone SOP manual. American Fisheries Society, Bethesda,
Maryland.

Gladso, J. \& Raddum, G.G. (2002). Rotenone treatment of a west
Norwegian river: Effects on drift of invertebrates. \emph{Internationale
Vereinigung fur Theoretische und Angewandte Limnologie:
Verhandlungen,28} ,764-769.
https://doi.org/10.1080/03680770.2001.11901816

Gloer, P. (2002). Susswassergastropoden Nord- und Mitteleuropas.
Bestimmungsschlussel, Lebenweise, Verbreitung. ConchBooks.

Hamilton, B.T., Moore, S.E., Williams T.B., Darby, N. \& Vinson, M.R.
(2009). Comparative effects of rotenone and antimycin on
macroinvertebrate diversity in two streams in Great Basin National Park,
Nevada. \emph{N. Am. J. Fish Management 29} , 1620--1635.
https://doi.org/10.1577/M08-178.1

Holdich D., Reynolds J., Souty-Grosset C. \& Sibley P. (2009). A review
of the ever-increasing threat to European crayfish from non-indigenous
crayfish species. \emph{Knowledge and Management of Aquatic Ecosystems,
394-395} , 11. https://doi.org/10.1051/kmae/2009025

Holmen, M. (1987). The Aquatic Adephaga (Coleoptera) of Fennoscanida and
Denmark, Volume 1. Gyrinidae, Haliplidae, Hygrobiidae and Noteridae.
Fauna entomologic Scandinavican 20.

Holocombe, G. W., Phipps, G. L., Sulaiman, A. H. \& Hoffman, A. D.
(1987). Simultaneous multiple species testing: Acute toxicity of 13
chemicals to 12 diverse freshwater amphibian, fish, and invertebrate
families. \emph{Archives of Environmental Contamination and Toxicology
16} ,697-710. https://doi.org/10.1007/BF01055420

Holthe, E., Bjorna, T. \& Lo, H. (2015). Reetableringsprosjektet i
Vefsnaregionen - Arsrapport for aktiviteten i 2014.
Veterinaerinstituttets rapportserie 14-2015.

Holthe, E. Bjorna, T. \& Lo, H. (2018). Reetableringsprosjektet i
Vefsnaregionen. Arsrapport 2017. Veterinaerinstituttets rapportserie
11-2018.

Hulme, P. E. (2009). Trade, transport and trouble: managing invasive
species pathways in an era of globalization. \emph{Journal of Applied
Ecology, 46} , 10--18. https://doi.org/10.1111/j.1365-2664.2008.01600.x

Josefsson, J., Hiron, M., Arlt, D., Auffret, A.G., Berg, A., Chevalier,
M, Glimskar, A., \ldots{}Part, T. 2020. Improving scientific rigor in
conservation evaluations and a plea deal for transparency on potential
biases. Conservation Letters. 2020; 13:e12726.
https://doi.org/10.1111/conl.12726

Kumschick S., Gaertner M., Vila M., Essl, F., Jeschke, J.M., Pysek, P.,
Ricciardi, A., \ldots{}Winter, M. (2015). Ecological impacts of alien
species: quantification, scope, caveats, and
recommendations.\emph{BioScience, 65} , 55--63.
https://doi.org/10.1093/biosci/biu193

Kjaerstad, G. \& Arnekleiv, J. V. (2004). Rotenonbehandling av elver i
Rana-regionen i 2003 og 2004: effekter pa bunndyr. NTNU Vitenskapsmuseet
zoologisk notat 2004-4.

Kjaerstad, G. \& Arnekleiv, J.V. (2011). Effects of Rotenone Treatment on
Lotic Invertebrates. \emph{International Review of Hydrobiology 96} ,
58-71. https://doi.org/10.1002/iroh.201111261

Kjaerstad, G., Arnekleiv \& Speed, J.D.M. (2015). Effects of three
consecutive rotenone treatments on the benthic macroinvertebrate fauna
of the River Ogna, Central Norway. \emph{Rivers Research and
Applications.} https://doi.org/10.1002/rra.2873

Koksvik, J.I. \& Aagaard, K. (1984). Effects of rotenone on the benthic
fauna of a small eutrophic lake. \emph{Internationale Vereinigung fur
Theoretische und Angewandte Limnologie: Verhandlungen, 22} , 658-665.
https://doi.org/10.1080/03680770.1983.11897361

Larsen, B.M. (2015). Elvemusling i Fusta, Nordland -- konsekvenser av
rotenonbehandling i vassdraget og tiltak for a sikre bestanden av
muslinger. NINA Rapport 1189.

Legendre, P. (2019). A temporal beta-diversity index to identify sites
that have changed in exceptional ways in space-times
surveys.\emph{Ecology and Evolution, 9} , 3500-3514.
https://doi.org/10.1002/ece3.4984

Lillehammer, A. (1988.) Stoneflies (Plecoptera) of Fennoscandia and
Denmark. E. J. Brill/Scandinavian Science Press Ltd. Leiden, The
Netherlands. Vol. 21.

Ling, N. (2003). Rotenone - review of its toxicity and use for fisheries
management. Science for conservation 211.

Lo, H. \& Holthe, E. (2014). Bevaring av fiskebestander. In: Bekjempelse
av \emph{Gyrodactylus salaris} i Vefsnaregionen. (Eds: J.H. Stensli \&
H. Bardal). Veterinaerinstituttets rapportserie 2-2014.

Mack, R.N., Simberloff, D., Mark Lonsdale, W., Evans, H., Clout, M. and
Bazzaz, F.A. (2000). \emph{Biotic invasions: causes, epidemiology,
global consequences, and control. Ecological Applications, 10} ,
689-710.
https://doi.org/10.1890/1051-0761(2000)010{[}0689:BICEGC{]}2.0.CO;2

Mangum, M.A \& Madrigal, J.L. (1999). Rotenone Effects on Aquatic
Macroinvertebrates of the Strawberry River, Utah: A Five-Year
Summary.\emph{Journal of Freshwater Ecology 14} , 125-135.
https://doi.org/10.1080/02705060.1999.9663661

Meadows, B. S. (1973). Toxicity of rotenone to some species of coarse
fish and invertebrates. \emph{Journal of Fish Biology, 5} , 155-- 163.
https://doi.org/10.1111/j.1095-8649.1973.tb04444.x

Melaas, C.L., Zimmer, K.D., Butler, M. G. \& Hanson, M.A. (2001).
Effects of rotenone on aquatic invertebrate communities in prairie
wetlands. \emph{Hydrobiologia 459} , 177-186.
https://doi.org/10.1023/A:1012514124430

Moen, A. \& Bardal, H. (2014). Innsjobehandlingen. In: Bekjempelse
av\emph{Gyrodactylus salaris} i Vefsnaregionen. (Eds: J. H. Stensli \&
H. Bardal). Veterinaerinstituttets rapportserie 2-2014.

Nilsson and Holmen (1995). The Aquatic Adephaga (Coleoptera) of
Fennoscandia and Denmark, Volume II. Dytiscidea. Fauna entomologica
Scandinavia 32.

McClay, W. (2000). Rotenone use in North America (1988-1997). Fisheries
20, 15-21.
https://doi.org/10.1577/1548-8446(2000)025\%3C0015:RUINA\%3E2.0.CO;2

McClay, W. (2005). Rotenone use in North America (1998-2002). Fisheries
30, 29-31.
https://doi.org/10.1577/1548-8446(2005)30{[}29:RUINA{]}2.0.CO;2

Pham, L., Jarvis, M. G., West, D \& Closs, G. P. (2018). Rotenone
treatment has a short-term effect on New Zealand stream
macroinvertebrate communities. \emph{New Zealand Journal of Marine and
Freshwater Research, 52} , 42-54.
https://doi.org/10.1080/00288330.2017.1330273

Rinne, A. \& Wiberg-Larsen, P. (2017). Trichoptera Larvae of Finland:
Identification Key to the Caddis Larvae of Finland and Nearby Countries.
Trificon books.

Sala O.E., Chapin III F.S., Armesto J.J. \& Berlow, E. (2000).
Biodiversity: global biodiversity scenarios for the year
2100.\emph{Science 287} , 1770-1774.
https://doi.org/10.1126/science.287.5459.1770

Stensli, J.H. \& Wist, A.N. (2014). Behandlingene i sone Indre
Vefsnfjorden. In: Stensli, J.H. \& Bardal, H. (eds.). Bekjempelse
av\emph{Gyrodactylus salaris} i Vefsnaregionen. (Eds: J. H. Stensli \&
H. Bardal). Veterinaerinstituttetsrapportserie 2-2014.

Vinson, M.R., Dinger, E.C. \& Vinson, D.K. (2010). Piscicides and
Invertebrates: After 70 Years, Does Anyone Really Know? \emph{Fisheries
35} , 61-71. https://doi.org/10.1577/1548-8446-35.2.61

Weyl, P.S.R., de Moor, F. C., Hill, M.P. \& Weyl, O.L.F. (2010). The
effect of largemouth bass \emph{Micropterus salmoides} on aquatic
macroinvertebrate communities in the Wit River, Eastern Cape, South
Africa. \emph{African Journal of Aquatic Science 2010} , 35, 273--281.
https://doi.org/10.2989/16085914.2010.540776

\textbf{Tables}

TABLE 1 Time of sampling (grey) and rotenone treatments (black) in River
Fusta, River Drevjaelva, Lake Fustvatnet and Lake Drevvatnet

\textbf{Hosted file}

\verb`image1.emf`  available at \url{https://authorea.com/users/426555/articles/531027-long-term-responses-of-benthic-invertebrates-to-rotenone-treatment}

\textbf{Figure captions}

FIGURE 1 Overview of the two studied watercourses Fusta and Drevja.
Triangles indicate the position of sampling stations. A double black bar
indicates the upstream limit of the treatment of River Fusta in August
2011 and August 2012, and a single black bar indicates the upstream
limit of the treatment of the watercourse in October 2012.

FIGURE 2 Mean changes in dissimilarity (\% difference) for abundance
data (a -- c) and for presence-absence data (d -- f) between sampling
occasions (a and d) and its components; losses (b and e) and gains (c
and f). Grey: River Fusta (rotenone treated), black: River Drevjaelva
(untreated). Dashed vertical lines indicate the timing of the rotenone
treatments. Error bars indicate maximum and minimum values of the
temporal biodiversity indices at individual sampling stations.
Significant differences in diversity indices between sampling periods (p
\textless{} 0.05) in River Fusta are marked with grey asterisk (no
signifcant differences detected in River Drevjaelva).

FIGURE 3 Average abundance (with standard deviation) of benthic
invertebrates of River Fusta (rotenone treated) and River Drevjaelva
(untreated) in the period 2011-2016. a) total abundance, b)
Epemeroptera, c) Plecoptera, d) Trichoptera, e) Chironomidae and f)
Oligochaeta. Timing of rotenone treatments are marked with arrows.

FIGURE 4 Mean changes in dissimilarity (\% difference) for abundance
data (a -- c) and for presence-

absence data (d -- f) between sampling occasions (a and d) and its
components; losses (b and e) and gains (c and f). grey: Lake Fustvatnet
(rotenone treated), black: Lake Drevvatnet (untreated). Dashed vertical
lines indicate the timing of the rotenone treatment of the lake. Error
bars indicate maximum and minimum values of the temporal biodiversity
indices at individual sampling stations.

FIGURE 5 Average abundance (with standard deviation) of benthic
invertebrates of Lake Fustvatnet (rotenone treated) and Lake Drevvatnet
(untreated) in the period 2011-2016. A) Total abundance, b)
Ephemeroptera, c) Plecoptera, d) Trichoptera, e) Chironomidae and f)
Oligochaeta. Timing of the rotenone treatment is marked with an arrow.\selectlanguage{english}
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