Mitigation of Ecological Impacts on Fish of Large Reservoir

STOTEN-144053; No of Pages 12 Science of the Total Environment xxx (xxxx) xxx

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Science of the Total Environment

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Mitigation of ecological impacts on ﬁsh of large reservoir sediment management through controlled ﬂushing – The case of the Verbois dam (Rhône River, Switzerland)

Franck Cattanéo a,⁎, Jean Guillard b, Seydina Diouf c, Jane O'Rourke a, David Grimardias a a University of Applied Sciences and Arts Western Switzerland – HEPIA Geneva, route de Presinge 150, CH-1254 Jussy, Switzerland b Université Savoie Mont Blanc, INRAE, CARTELL, 74200 Thonon-les-Bains, France c SIG – Services Industriels de Genève, Case postale 2777, 1211 Genève 2, Switzerland

HIGHLIGHTS GRAPHICAL ABSTRACT

• Controlled flushing (CSFO) are new practices to tackle sediment trapping in reservoirs. • Impacts on fish of the first CSFO of the Verbois Dam (Switzerland) were assessed. • From hydroacoustics, fish density in the reservoir did not change after the CSFO. • From radiotelemetry, impacts below the dam were mainly behavioural rather than lethal. • CSFO should be promoted to recover storage volume while lowering ecological impacts.

article info abstract

Article history: Sediment trapping within reservoirs is a worldwide phenomenon which impairs the ecological functioning of up- Received 3 September 2020 stream and downstream ecosystems. It also reduces reservoir water storage volume, which lessens the services Received in revised form 15 November 2020 dams provide such as hydropower production or flood control and questions their sustainability. Hydraulic flush- Accepted 19 November 2020 ing is a widely used operation to recover the reservoir volume, but ecological impacts are massive. Recently, en- Available online xxxx vironmental awareness led dam operators to modify their management practices: ‘Controlled Sediment Flushing ’ Editor: Sergi Sabater Operations (CSFOs) include environmental objectives in their implementation and are designed to be less harmful for aquatic ecosystems by controlling the flow and Suspended Sediment Concentration (SSC) downstream. Keywords: However, CSFOs are not yet widespread, their ecological impacts are poorly documented, and comparisons Siltation with ‘classical’ flushing operations are unreported. Here, we analysed impacts on fish of the first CSFO of the Hydropower dam Verbois reservoir in 2016, both upstream and downstream of the dam, and compared these with those from CSFO the empty flushing of 2012 using the same methodology (Grimardias et al., 2017). Time-series of hydroacoustics Hydroacoustics surveys enabled us to estimate the fish abundance in the reservoir, while radiotelemetry measured movements Radiotelemetry and apparent survival below the dam for four representative species. The 2016 CSFO lasted 10 days, and released Fish a mean Suspended Sediment Concentration (SSC) of 3.47 g·L−1 (peak = 11.98 g·L−1). The fish density as

assessed by the mean acoustic scattering strength SA in the reservoir did not change signiﬁcantly pre- and

post-CSFO, and SA seasonal estimates of year 2016 did not differ from those of 2015 and 2017. The apparent survival estimated from capture-recapture survey of marked fish (N = 118) decreased significantly during the CSFO for all species and differed across species, while the distances moved downstream increased. By comparison with the 2012 empty flushing, the 2016 CSFO allowed fish to remain in the reservoir, while impacts below the dam were mostly behavioural rather than lethal. Overall, despite significant impacts, the CSFO advantageously

⁎ Corresponding author. E-mail address: [email protected] (F. Cattanéo).

Please cite this article as: F. Cattanéo, J. Guillard, S. Diouf, et al., Mitigation of ecological impacts on ﬁsh of large reservoir sediment management through controlled f..., Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2020.144053 F. Cattanéo, J. Guillard, S. Diouf et al. Science of the Total Environment xxx (xxxx) xxx

replaced ‘classical’ ﬂushing from an ecological viewpoint. Provided that an acceptable balance between economical, ecological and technical aspects is found, CSFO can be considered for many reservoirs while accounting for their biological and physical site-speciﬁcity. ©2020ElsevierB.V.Allrightsreserved.

1. Introduction responses can occur, linked with oxygen acquisition, growth, or stress control (Servizi and Martens, 1992, Lake and Hinch, 1999, Henley River flow regulation by dams and reservoirs is a ubiquitous around et al., 2000, Greig et al., 2005, Huenemann et al., 2012). Ultimately, the world, yet severely impairing the natural flux of water and sediment long lasting or acute SSC events affect the survival of individuals, with (Syvitski et al., 2005). This results in severe trans-ecosystem alterations, changes of population abundance and structure, and potentially of com- not only for the ecological functioning of upstream and downstream munity structure or composition (Buermann et al., 1995; Swinkels et al., freshwaters (e.g. channel forms, instream habitats, grain size distribu- 2014; Baoligao et al., 2016). tion…), but also for estuarine and marine coastal ecosystems and the Impacts are dependent on many factors such as sediment particle services they provide (Walling and Fang, 2003, Tang et al., 2016, Yang size, shape, angularity, and biogeochemical characteristics (Bash et al., et al., 2018; but see Chen et al., 2020). Sediment trapping in reservoirs 2001), and vary substantially among family, species and life stages: Sal- is among the most pervading alteration, whereby the sediment flux is monids are more sensitive than other families, and larvae and early ju- drastically reduced for the downstream reaches. In 2003, an estimated veniles are considered more sensitive than adults (Newcombe and 25–30% of the river global sediment flux was trapped within the Jensen, 1996; Crosa et al., 2010). Impact severity correlated with the 45,000 registered reservoirs (Vörösmarty et al., 2003). Today, with ca. SSC and the duration of the event (Newcombe and Jensen, 1996), and 58,000 large dams around the world (ICOLD, 2020), the proportion of heightened by factors such as temperature, presence of pollutants or or- trapped sediment is undoubtedly much larger. Trapped sediment ganic content of the sediment (Garric et al., 1990). However, this knowl- lessens the lifetime of reservoirs (Kondolf, 1997) by reducing their edge is highly biased toward Salmonids, and with most investigations storage capacity, which in turn impairs the flow regulation role of the performed in lab-controlled conditions, their transferability to field con- dam and adversely affects functions such as power production, water ditions is questionable as local factors can mediate impacts of SSC (e.g. supply, flood control or navigation (Brandt, 2000; Schleiss et al., presence of refuges, clear water inputs, water temperature and oxygen 2016). About 1% of the total reservoir volume of the world is lost on av- concentration…). erage every year (White, 2001), which is higher than the volume cre- Due to these variations, predicting impacts of SS release on aquatic ated by the building of new reservoirs, henceforth questioning their biota is difficult. Responses are site-specific(Bilotta et al., 2012), sustainability (Schleiss et al., 2010). A predicted 80% of the total storage which prevents water authorities using a general, widely applicable capacity of reservoirs in Europe and Russia could be lost by 2080 and quantitative model, to set SSC thresholds for sediment manage- (Basson, 2009). Current perspectives of hydropower development (i.e. ment operations. Overall, SSC (or sometimes turbidity, in NTU) thresh- 3700 dams > 1 MW under construction or planned; Zarfl et al., 2015) olds vary largely among regions, and are mostly based on Salmonids suggest that sustainable solutions to sediment trapping have to be tolerance (Berry et al., 2003; Collins et al., 2011). As a guideline for ex- found urgently (Hauer et al., 2018). pected impacts on fish, the severity-of-ill effect of Newcombe and Hydraulic flushing is a commonly used operation to release fine sed- Jensen (1996) combines the intensity and duration of the SS event, iment from small to medium-sized reservoirs (Palmieri et al., 2003; and may be used to help define the operational and technical parame- Petkovšek et al., 2020). It consists in lowering the reservoir water level ters of a flushing (Espa et al., 2019), or to broadly evaluate its effects while opening the bottom outlets of the dam to flush the sediment (Crosa et al., 2010; Grimardias et al., 2017). However, some authors de- downstream (Wang and Hu, 2009). Morris and Fan (1998) distin- noted discrepancies between expected and observed impacts, and sug- guished ‘drawdown, empty or free-flow flushing’, where the reservoir gested that the model should be recalibrated or adapted locally (Xu is emptied to the level of the dam outlets and with a riverine flow et al., 2018; Tritthart et al., 2019). The fact that neither the local context through the reservoir, from ‘pressure flushing’ where bottom outlets nor the possible adaptation of populations to the local sediment regime are fully submerged. While empty flushing is more efficient in were accounted for was criticized (Vondracek et al., 2003). Therefore, dislodging deposited sediment, both methods have significant ecologi- this site-specificity requires that each flushing includes an appropriate cal impacts on downstream ecosystems (Gerster and Rey, 1994), mainly ecological monitoring and stresses the need to better document in situ because sediment releases are outside the range of the natural sediment impacts of sediment management operations. regime (Wohl et al., 2015). During these operations, Suspended Sedi- For a long time, flushing operations consisted of evacuating down- ment Concentration (SSC) downstream of the dam is far higher than stream from reservoirs the largest amount of sediment with the avail- that observed even during major floods, and the timing of sediment re- able water volume (i.e., maximizing the flushing efficiency); there was lease hardly matches the phenology of the natural sediment flux no, or only marginal, environmental considerations. With rising aware- (Lepage et al., 2020). Deposited sediment can lead to instream habitat ness for environmental preservation, ecological objectives are increas- alterations such as pool-filling (Wohl and Rathburn, 2003)andcanse- ingly considered when carrying out a flushing, and now characterize verely impair the quality of spawning grounds (Soulsby et al., 2001). the so-called ‘controlled’ (Espa et al., 2015; Reckendorfer et al., 2019) Impacts of fine suspended and deposited sediment on aquatic biota or ‘environmentally friendly’ (Fruchart and Camenen, 2012; Peteuil have been extensively reviewed (Wilber and Clarke, 2001; Bilotta and et al., 2013) flushing. From here on, we will use the term ‘controlled sed- Brazier, 2008; Kemp et al., 2011) and concern all taxonomic groups iment flushing operation’, noted CSFO, for a flushing which includes ob- (from phytoplankton to fish), trophic positions and biological organiza- jectives of environmental safeguard, as defined by Espa et al. (2019). tion levels (from individuals to communities). On fish, direct impacts The common ways to dampen impacts of flushing are: 1) to limit the range on a timescale from very short to long term effects and grade SSC downstream of the dam, in terms of mean and maximum values; along a severity scale distinguishing behavioural, sub-lethal, para- 2) to limit the duration of the flushing; 3) to combine SSC and duration lethal and lethal effects (Newcombe and Jensen, 1996). Behavioural ef- to fixed thresholds; 4) to choose an adequate time window for the op- fects such as avoidance of the sediment flume are the first signals of eration, i.e. avoid periods of spawning or early-life stages of target spe- impairment (Robertson et al., 2007; Mori et al., 2018). However, if the cies; 5) to synchronize flushing with natural floods. In practice, dam disturbance intensifies, a variety of physiological and metabolic operators fix a sediment volume to be evacuated from the reservoir,

2 F. Cattanéo, J. Guillard, S. Diouf et al. Science of the Total Environment xxx (xxxx) xxx implying that SSC and duration of flushing are intrinsically correlated the Seujet dam and the water coming from the Arve River (mean (i.e. the higher the SSC, the shorter the duration of the flushing, and flow = 79 m3·s−1; Fig. 1). This important tributary, characterized vice versa). Operators can control the SSC by adjusting the incoming dis- by a very high load of suspended sediment, drains the Mont Blanc al- charge and the water level of the reservoir, provided that the amount of pine massif and annually carries about 0.68 ± 0.29 Mt/year of flysch water upstream is sufficient and can be regulated. Therefore, the possi- and molasse particles into the Rhône River, of which 50% settle in the bility to control effectively a flushing operation is dam-specific, and re- VB reservoir (Launay et al., 2019). Two smaller tributaries (Allondon lies on a trade-off between technical constraints, economical costs, and and Laire) flow into the Rhône River along the study area, but only environmental objectives (Espa et al., 2013; Espa et al., 2016). marginally increase its water and sediment fluxes. The fish commu- In this paper, our goals were twofold: firstly, we analysed impacts on nity in the study area is composed of about 20 species, among the fish community of the first controlled flushing of a large hydro- whichEuropeanperch(Perca fluviatilis), chub (Squalius cephalus), power reservoir on the Swiss Rhône River. Immediate and medium- barbel (Barbus barbus), pike (Esox lucius)androach(Rutilus rutilus) term impacts were evaluated both within the Verbois reservoir and are the most abundant (Teleos, 2018). Limnophilic species such as downstream of the dam. An uncommon combination of hydroacoustics tench (Tinca tinca), bream (Abramis brama)orcarp(Cyprinus carpio) and radio-telemetry allowed us to analyse changes in fish assemblage are present, as well as some coldwater, rheophilic Salmonids (brown density in the reservoir, in movement patterns and in apparent survival trout, Salmo trutta;Europeangrayling,Thymallus thymallus;GREN on marked individuals for four common species. Secondly, we com- Basson, 2009). pared impacts on fishes of two distinct sediment management operations, i.e. the empty flushing of 2012 and the controlled flushing of 2.2. The VB reservoir management and the 2016 CSFO 2016, using a set of operational and ecological indicators. Specifically, we hypothesized that: i) the fish assemblage in the reservoir should Between 1969 and 2003, the VB reservoir was operated by means of be less impaired by the 2016 controlled flushing than by the 2012 empty flushing (full drawdown) every three years (Cohen and Briod, empty flushing; ii) the apparent survival of fishes below the dam should 1989; SIG and SFMCP, 2017a). Vast ecological impacts and growing so- be higher in 2016 than in 2012 due to the controlled SSC. These results cietal discontent from the Swiss and French stakeholders stopped the allowed us to discuss the implementation of general operational mea- three-year flushing cycle. From 2003 on, a ‘passive management’ was sures for an eco-friendlier management of hydroelectric reservoirs. tested, where the dam was operated without accounting for fine sediment, with the expectation that sediment deposition would reach an 2. Material and methods equilibrium state allowing both the reservoir uses and the preservation of public safety. In 2012, due to the filling of the reservoir and increasing 2.1. Study area risk of flooding in Geneva city, dam operators had to perform a new flushing (SIG and SFMCP, 2017a). Meanwhile, a working group includ- A detailed description of the study area can be found in Grimardias ing the Swiss and French Rhône stakeholders elaborated a new manage- et al. (2017). It encompasses a 24 km long section of the Rhône River, ment strategy for fine sediment (COTECH, 2014). It relies on three from the outlet of Lake Geneva to the France-Switzerland border measures: 1) controlled flushing of the reservoir, 2) localised dredging (Fig. 1), along which three run-of-the-river hydropower dams have operations, and 3) dam gates opening during floods of the Arve River. been erected. At the outlet of Lake Geneva, the Seujet dam regulates In comparison to previous empty flushing, the CSFO is characterized the lake water level and the flow in the downstream Rhône River (an- by a partial rather than a full lowering of the reservoir, and by the con- nual mean flow = 251 m3·s−1). The Verbois dam (hereafter noted trol of SSC below the dam according to three ‘concentration – duration’ VB), 15 km downstream, is a 34 m high dam. Its reservoir is 11.4 km thresholds. SSC should not exceed mean values of 5 g·L−1 over the en- in length and has a 13 Mm3 storage capacity. The Chancy-Pougny dam tire operation, 10 g·L−1 during 6 h and 15 g·L−1 during 0.5 h. Because of (noted CP), located 7 km downstream of VB, is 10.7 m high. Its reservoir their longitudinal succession along the Rhône River, dams downstream is 3.7 km long for a 2.7 Mm3 storage capacity. All three dams are of VB were operated simultaneously to allow the transit of the equipped with fish bypasses. The flow regime in this section of the suspended sediment load, and to avoid deposition in downstream res- Rhône River combines the water released from Lake Geneva through ervoirs (Peteuil et al., 2013). Consequently, the CP reservoir, which is

Fig. 1. Map of the study area. The areas of VB and CP reservoirs sampled by hydroacoustics are in white. The thick grey line represents the France – Switzerland boarder.

3 F. Cattanéo, J. Guillard, S. Diouf et al. Science of the Total Environment xxx (xxxx) xxx immediately downstream of the VB dam, was completely lowered (full Tabary’, a sidearm of the Rhône River, and on 4 May 2016 at the drawdown) and the dam gates fully opened. The application of the CSFO confluence between the Allondon and Rhône rivers. One hundred and of the VB reservoir took place in May 2016. eighteen (N = 118) individuals belonging to four species most representative of the Rhône River ichthyofauna were anaesthetized 2.3. Impact assessment on fish assemblages within reservoirs: with clove oil (‘Eugenol’), measured (total length TL, ±1 mm) and hydroacoustic data weighed (±1 g). Species displayed distinct habitat uses and life history traits (Bruslé and Quignard, 2001). The barbel (Barbus barbus,fam. We used hydroacoustics to investigate changes in mean acoustic Cyprinidae; N = 37, mean TL ± s.d. = 607.1 ± 63.1 mm, mean scattering strength (a proxy for fish density) in the VB and CP reservoirs weight ± s.d. = 2077.6 ± 577.9 g) is a typical benthic and rheophilic during the 2016 flushing operation. A Simrad EK60 echosounder species. The chub (Squalius cephalus, fam. Cyprinidae; N = 17, mean (SIMRAD, Oslo, Norway), 70 kHz, using a 256 ms pulse length and TL ± sd. = 299 ± 67.7 mm, mean weight ± sd. = 360.5 ± 281.2 g) is pinging at 5 pings per second, was used for data acquisition. The circular a eurytopic, water-column species. The roach (Rutilus rutilus,fam. split-beam transducer, 11.16° × 11.46° at −3dB,fixed on the right side Cyprinidae; N = 37, mean TL + s.d. = 330.4 ± 68.8 mm, mean of an aluminium boat, 0.30 m below the water surface emitted vertically weight + s.d. = 440.8 ± 198.6 g) is a gregarious water-column species (Samedy et al., 2013). The transducer was linked to a computer with the which prefers lentic habitats. The brown trout (Salmo trutta, fam. Simrad ER60 software and connected to a GPS to record boat positions. Salmonidae; N = 27, mean TL + s.d. = 335.7 ± 53.6 mm, mean Acoustic data were analysed using Sonar 5-Pro software (v. 6.0.1; Balk weight + s.d. = 427.1 ± 200.1 g) is a typical water-column and and Lindem, 2011). The parameterization (following European recom- rheophilic species. All fish were equipped with externally mounted mendations, CEN, 2014) and data treatment are detailed in Grimardias radio-transmitters fixed on the dorsal muscle, which are preferred to et al. (2017). For each hydroacoustic survey, the sampling protocol internal transmitters when the marking occurs close to, or during, the consisted in a series of consecutive transects, from one bank to the reproductive period (Bridger and Booth, 2003). Two sizes of transmit- other, according to a zigzag trajectory (Guillard and Vergès, 2007)de- ters were used to respect the 2% tag to fish body weight ratio (Winter, fined beforehand. Cruising speed was between 6 and 8 km·h−1. Each 1996; Bridger and Booth, 2003). Fish between 130 g and 240 g, were transect was considered as the statistical unit in data analyses. To obtain marked with the non-coded ATS F1940 and F1970 transmitters (weight representative data, the coverage ratio (i.e. the total length of all tran- in air: 2.6 and 4.8 g, respectively; expected lifetime = 160 days; 148–149 sects in km divided by the square root of the reservoir area in km2) MHz frequency range; Advanced Telemetry System). Bigger fish should be >6 (Aglen, 1983). In our case, the sampling trajectories (minimum weight = 430 g) were tagged with the coded ATS F2020 were composed of 33 transects for both the VB and CP reservoirs, transmitters (weight in air: 8.6 g; expected lifetime = 440 days; which resulted in coverage ratios of 8.00 and 7.37, respectively 148–149 MHz frequency range; Advanced Telemetry System). (Grimardias and Cattanéo, 2016). Overall, 12 hydroacoustics surveys Fish were released on the same day after fully recovered equilib- were conducted between September 2015 and January 2018. Survey rium and spontaneous swimming activity, at three locations. A first timings were decided to reflect potential immediate changes in fish group (N = 14: 4 barbels, 4 roaches, 4 trout and 2 chubs) was re- density due to the flushing operation (i.e. before-after flushing compar- leased upstream of the VB dam at the exit of the fish ladder to isons), and to compare seasonal estimates between years with and check if the flushing could induce downstream movements through without flushing (Fig. 2). the dam gates. A second (N = 64: 18 barbels, 18 roaches, 18 trout

Mean acoustic scattering strength values, expressed in SA and 10 chubs) and a third (N = 40: 15 barbels, 15 roaches, 5 trout (m2·ha−1), were used as proxy for fish density (Boswell et al., 2010; and 5 chubs) group were released in the lotic reaches downstream Yule et al., 2013) for each transect. To test whether the flushing modi- of the VB and CP dam, respectively, to analyse fish movements due fied fish density within the reservoir, mean acoustic scattering strength to the sediment release. SSC was similar downstream of both dams. values (SA) were compared across surveys using the non-parametric Two operators prospected a 14 km river length during daylight on a Kruskal-Wallis rank sum test (data did not meet the hypotheses of nor- weekly basis from 19 April to 11 August 2016, except during the flush- mality and variance homogeneity even after transformation). Consecu- ing operation where the tracking was performed daily (i.e. from 20 May tive pairs of surveys were compared with the Wilcoxon rank sum test to to 28 May). Fish were located according to the ‘homing in’ method identify where differences occurred. For all tests, the significance level (Adams et al., 2012), using a R2000 receiver (Advanced Telemetry Sys- was corrected for multiple comparisons using the Bonferroni method tem Inc., Isanti, Minnesota, USA), and the individual's position was fixed

‘αi’. All analyses were performed with the R statistical software (version where the signal strength was the highest. Given the size of the river, we 3.6.1; R Development Core Team, 2008). estimate that the uncertainty around an individual's position is about 100 m (Grimardias and Cattanéo, unpublished data). 2.4. Impact assessment below the dam: radio-telemetry data 2.4.1. Fish survival Radio-telemetry was used to estimate apparent survival rates and to For each tracking session, the mean (+s.d) distance (in m) of each analyse fish movements during the flushing operation under approval species below the VB dam was calculated from individual positions. from “Direction Générale de l'Eau” of Canton Geneva (Switzerland). Daily apparent survival rates (ϕ) and capture probabilities (ρ) during Six captures took place, including four purges of the fish ladder of the the flushing operation were estimated using Cormack-Jolly-Seber VB dam (on 7 and 21 April 2016, and 4 and 11 May 2016). These proved mark-recapture models (Lebreton et al., 1992), based on the capture to be the most efficient way to capture fish in the Rhône River; two res- history of each marked individual. We tested for effects of i) the species cue electrofishing operations also took place on 23 April 2016 on ‘Bief de (group effect, where ϕ and ρ could vary according to the species), and ii)

Fig. 2. Sampling design set up for the ﬁsh monitoring of the 2016 CSFO of the VB reservoir. The ﬂushing is the black square in late May 2016. Grey arrows indicate the 12 hydroacoustic surveys. The radiotelemetry marking and tracking periods are noted.

4 F. Cattanéo, J. Guillard, S. Diouf et al. Science of the Total Environment xxx (xxxx) xxx the tracking session. We tested this last effect in two ways, first consid- 3. Results ering each tracking session separately (temporal effect, where ϕ and ρ could vary among tracking sessions), then considering an overall effect 3.1. Suspended sediment and flow during the flushing of the flushing operation (period effect, where ϕ could vary according to whether the session occurred during or outside of the flushing opera- The CSFO started on 21 May 2016 12:00 a.m. with a gradual low- tion, not tested for ρ). All nested models were tested, from the model ering of the reservoir level (about 0.15 m·h−1) during the two first with species x tracking session interaction to the null model, where ϕ days, until reaching a plateau where the lowering was stopped dur- and ρ were considered as constant (no effect of other variables). ing 24 h (J + 2 12:00 to J + 3 12:00; Fig. 3). Meanwhile, the flow Model comparison and selection were performed according to the was significantly increased from the Seujet dam from 265 m3·s−1 corrected Akaike Information Criterion (AICc, for small samples). Daily to ca. 500 m3·s−1 to initiate sediment resuspension, and then survival rates were estimated from the best adjusted model considering lowered to about 300 m3·s−1 when the plateau was reached. SSC the lowest AICc (Δ AICc >2;Burnham and Anderson, 2002). Models then ranged between 0 and 3.07 g·L−1. From J + 3 12:00 to J + 6 were computed using the MARK software (version 6.1; White and 0:00, the water level was slowly lowered from the height of 362 to Burnham, 1999). At the end of the flushing, an overall apparent survival 357.5 m (ca. 0.08 m·h−1), together with a gradual increase in flow from rate Φ was estimated per species from fish still displaying swimming ac- 380 to 470 m3·s−1which triggered the sediment resuspension. This ele- tivity or upstream movement over all fish still tracked at the beginning vation was maintained until the beginning of the reservoir filling at J + of the reservoir flushing. 8 12:00. SSC ranged between 2.5 and 8 g·L−1, with only some short peaks above, and reached a maximum of 11.98 g·L−1on J + 6 10:20 2.4.2. Fish movements pm. Throughout the CSFO, the mean SSC was 3.47 g·L−1. The 10 g·L−1 A linear mixed model was used to relate the daily speed of individ- threshold was exceeded 4 times, for a total duration of 3 h 50 min. The uals (i.e. distance covered over the number of days from previous 15 g·L−1 threshold was never exceeded. The reservoir was refilled from detection; in m·day−1, cubic root transformed values) to the period J + 8 12:00 to ca. J + 10 12:00, with flow values ranging between 460 (outside/during the flushing), the species, the individual size, and and 680 m3·s−1. first-order interactions (period × species and species × size), considered as fixed effects. The model was fitted on repeated daily speed estimates 3.2. Fish density within reservoirs on the individuals that remained in the study area and that were still tracked on 9th June 2016. Individuals were thus considered as random In the VB reservoir, the mean acoustic scattering strength signifi- effect. Because of differences in mean and range size across species, cantly changed between surveys (K = 142.4, df = 11, p < 0.001; the individual size was normalized per species. Fixed effects were tested Fig. 4). However, when comparing surveys performed just before and using Type III Anova with Wald F tests, using the R statistical software after the CSFO (May vs June and July 2016), no statistical difference ap- (version 3.6.1; R Development Core Team, 2008), with ‘lmer’ function peared (p > 0.05). The mean acoustic scattering strength decreased by from ‘lme4’ package (Bates et al., 2015). 48.6% between the first two post-CSFO surveys, i.e. June and July 2016 (W = 758, p = 0.006). In an attempt to account for the seasonal pattern 2.5. Assessment of ‘severity of ill’ effects in fish density within the reservoir, we compared seasonal surveys of the year 2016 with the CSFO with corresponding surveys in years The expected impacts of the CSFO on fishes,aswellasthecomparison 2015 and 2017/2018 (without flushing operation). The mean acoustic with the 2012 flushing, were assessed by means of the Newcombe and scattering strength did not differ between March surveys in 2016 and Jensen (1996) severity-of-ill effect (SEV) model. This model combines 2017 (K = 0.248, p = 0.619). In June, no difference appeared between the mean intensity of the sediment disturbance (mean SSC, in mg·L−1) acoustic densities in 2016 and 2017 (K = 3.457, p =0.063).Thethree and the exposure duration (ED, in h): surveys of September 2015, 2016, and 2017 did not differ across years (K = 3.631, p = 0.163). Significant differences were noted across the winter (December or January surveys; K = 41.528, p < 0.001), with SEV ¼ a þ b lnðÞþ ED c lnðÞ mean SSC mean acoustic scattering strength in December 2015 being higher than in December 2016 (W = 964, p < 0.001) and in January 2018 with a, b, c being regression coefficients estimated by Newcombe and (W = 1811.5, p < 0.001). No difference appeared between acoustic Jensen (1996), depending on taxonomic groups or ontogenetic stage. densities of December 2016 and January 2018 (W = 1117, p =0.552). The Model 2 (adult freshwater Salmonids) was used for brown trout, In the CP reservoir, the mean acoustic scattering strength sharply and the Model 6 (adult freshwater non-Salmonids) for chub, barbel decreased between the pre- and post-CSFO surveys (May vs July and roach. The SEV scales from 0 to 14 and is associated either to nil 2016; W = 808.5, p < 0.001). By the end of the year, mean acoustic scat- (SEV = 0), behavioural (1–3), sub-lethal (4–8) or lethal/para-lethal ef- tering strength values in October and December 2016 were similar both fects (9–14). to those observed in September and December 2015, and in September and January 2018, respectively (p >0.05). 2.6. Comparison of the 2016 CSFO vs 2012 empty flushing 3.3. Apparent survival below the VB dam A set of operational and biological metrics was used to compare both flushing operations and their impacts on fish. Operational metrics in- Out of the 118 marked individuals, 69.5% (N =82)weredetectedat cluded the timing and duration of the flushing, mean and peak SSC, es- least once during the study. Among those, 73 individuals were detected timated volumes of sediment released and of water used for the in the study area at the onset of the CSFO: 29 barbels, 15 chubs, 11 roaches flushing, and flushing efficiency (i.e. the ratio of amount of sediment re- and 18 trout (Table 1).AttheendoftheCSFO,48fish were still present in leased over the volume of water used; Morris and Fan, 1998). Biological the study area among the 73 individuals present at the beginning metrics included the SEV, the observed trend in fish density before and (Table 1). Out of those, 39 individuals still displayed activity, while the 9 after the flushing, an assessment of assemblage resilience, the mean dis- others were detected motionless and thus may have been dead. The over- tance travelled downstream by fishes below the dam, the apparent sur- all apparent survival rate Φ was 53.4% (Table 1). The highest rate was ob- vival rate of marked fishes, and an empirical evaluation of fish mortality served for the barbel (Φ = 65.5%). The Chub and the trout displayed below the dam. Because biological monitoring relied on the same medium Φ rates (respectively 53.3 and 55.6%), while the roach was the methods, both events are comparable using the same metrics. most affected, with an overall apparent survival rate Φ of 18.2%.

5 F. Cattanéo, J. Guillard, S. Diouf et al. Science of the Total Environment xxx (xxxx) xxx

Fig. 3. Characteristics of the 2016 CSFO of the VB reservoir (black curves) from the onset (day J-0) to the end of operation (+10), and of the 2012 empty ﬂushing for comparison (grey curves). Are shown: the reservoir water level (A), the suspended solid concentration below the dam (B), and the ﬂow (C). Data: Services Industriels de Genève (SIG).

Within the set of mark-recapture models (N = 24; Appendix The average daily speed of fish depended on the period (p = 0.024; 1) fitted to test the effects of the species, the period (within/outside GLMM, Wald F-test; Appendix 2). During the flushing, daily speeds CSFO) and the tracking session, the most likely model indicated that were larger, and movements were in a downstream direction. For all the apparent survival probability ϕ depended on the species and on species, mean daily speeds (±s.e.) were −91.09 ± 23.95 m·day−1 dur- an overall effect of the CSFO (i.e. period). This model performed much ing the CSFO, while they were 2.31 ± 11.77 m·day−1 outside the CSFO. better than the second one (ΔAICc = 27.8), and was given an Akaike weight of 1.000, while no support was attributed to other models 3.5. Assessment of ‘severity of ill’ effects (AICc weight = 0.000). Overall, ϕ was lower during the CSFO than outside for all species, but the magnitude of change depended on the spe- The 2016 CSFO lasted 244 h, and the mean SSC was 3.47 g·L−1 (SIG cies. The barbel exhibited a slight decrease in ϕ during the flushing and SFMCP, 2017b). According to the Newcombe and Jensen (1996) (from 0.988 ± 003 to 0.972 ± 0.013), the chub and the trout a moderate model, this led to SEV values of 10.5 for trout, and of 10.3 for barbel, decline (from 0.995 ± 0.002 and 0.992 ± 0.003, respectively, to 0.941 ± chub and roach, corresponding to lethal and para-lethal effects with 0.026 and 0.937 ± 0.024), and the roach the sharpest drop (from 0.946 ± 0–20% mortality and severe habitat degradations. 0.013 to 0.761 ± 0.058). Note that considering each tracking session separately instead of an overall CSFO yielded to less supported models, 3.6. Comparison of the 2016 CSFO vs 2012 empty flushing emphasizing that ϕ decreased gently and graduallyamongtrackingses- sions throughout the CSFO. Results presented in Table 2 are a synthesis of operational metrics and observed impacts on fishes of the 2012 and 2016 flushings. The first most 3.4. Fish movement below the VB dam noticeable operational difference is that a significant water volume was maintained in the VB reservoir in 2016 (lowering = −12 m; Fig. 3), Significant downstream movements were observed below the dam whereas it was fully emptied in 2012 (lowering = −17 m). Sec- during the CSFO, for all species (Fig. 5). Among the marked fishes, ondly the mean and peak SSC below the dam were much lower only three roaches were regularly monitored, hence distance and daily in 2016 (mean = 3.47 g·L−1,peak=11.98g·L−1 in 2016 vs. speed were not computed for this species. Individuals (N = 37) from mean = 11 g·L−1,peak=48g·L−1 in 2012) due to the control other species travelled on average 1732 m for trout, 784 m for barbel, measures. Finally, the flushing was about 3 days shorter in 2016. and 95 m for chub. A general tendency to move upstream was observed Asaresult,thevolumeofevacuatedsedimentwasabout2.5 just after the flushing (J + 11; Fig. 5). times less in 2016 than in 2012 (1.3 Mm3 vs 3.2 Mm3), and the

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Fig. 4. Mean (±s.e.) acoustic scattering strength SA value for each hydroacoustic surveys in the VB (N = 12; upper graph) and CP reservoirs (N = 11; lower graph). The thick grey line represents the CSFO.

flushing efficiency much lower (0.33% vs 0.87%). As a logical conse- the water level in two steps (−7 m during the first 48 h, stable water quence, the SEV was slightly lower in 2016, especially for Salmonids level for the next 24 h, then −5 m in 60 h; Fig. 3) caused a gradual in- (10.5 vs 11.4; Newcombe and Jensen, 1996). Overall, the 2016 CSFO crease in water velocities in the main channel, and a progressive waslessdetrimentalforthefish assemblage, both in the VB reservoir dewatering of marginal habitats. Fish were able to cope with these hy- and below the dam. In the reservoir, the mean acoustic scattering draulic conditions by seeking refuge close to the substrate or along the strength SA did not change after the CSFO, whereas it dramatically de- banks. In addition, even when the water level reached −12 m on 26 creased in 2012. Below the dam, despite an overall apparent survival May evening, a 5 m water depth was preserved. In the upstream part of slightly higher in 2012, there were many dead fishes observed along the reservoir the fish community is increasingly composed of rheophilic the Rhône riverbanks during and just after the 2012 flushing, a fact species (GREN, 2009; Teleos, 2018)whichareabletocopewithhighcur- that did not occur in 2016. rent velocities and were little affected by the lowering of the reservoir (Grimardias et al., 2017). 4. Discussion Comparing fish densities at the same season between 2016 and two years without flushing (2015, 2017–18) revealed no difference, except Improving the management of fine sediment accumulated in reser- in the winter surveys. If the CSFO detrimentally affected fish densities voirs is an important issue dam managers and water authorities must on the short or medium term (a few days to ca. 1.5 year after the oper- deal with. The new strategy experimented on the upper-Rhône river ation), we expected lower seasonal densities after the CSFO. However, at the VB dam in 2016 affected the fish community. The overall apparent the sampling design was not fully robust to test the effect of the flushing survival rate of fish decreased due to high SSC downstream of the dam on the seasonal pattern of fish density. A BACI design would have been and duration of flushing. However, impacts were mainly behavioural, preferred (Christie et al., 2019), but in the absence of a control site, we with marked downstream movements. The fish density was preserved used surveys performed the year before and after the flushing for com- in the VB reservoir after the CSFO, while that of the CP reservoir, located parison. Our approach may be questionable as the natural variations of immediately downstream, drastically declined though it recovered on seasonal mean acoustic scattering strength were unknown in this eco- the medium term (i.e. 3 and 6 months later). system and factors other than the flushing could have yielded changes in fish density (temperature, flow regime, recruitment …). The decrease 4.1. Fish density in the VB and CP reservoirs in acoustic densities observed during the winter surveys of 2016 and 2017–18 could not be attributed to the flushing, or else we would also In the VB reservoir, the fish density (assessed by the mean acoustic have lower density values for the June and September 2016 surveys. scattering strength) did not change between the pre- and post-CSFO sur- However, mild hydroclimatic conditions during the autumn and winter veys (May vs June 2016). This suggested that the partial lowering allowed 2015 may have been favourable to fish survival, and thus explain a fish to remain during the flushing, as opposed to a full lowering higher mean acoustic scattering strength in December 2015. Another (Hofmann et al., 2001; Grimardias et al., 2017). Progressively lowering hypothesis is an increased availability after the flushing of marginal

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Table 1 Numbers and percentages of individuals tracked by radiotelemetry, following their status at the end of the flushing operation (30th May 2016; still present, displaying swimming activity, motionless, or lost), detailed per species and release location. “N tracked” represents the number of individuals detected in the study reach at the onset of the flushing operation. The “overall apparent survival rate (%)” was defined as the number of individuals still active at the end of flushing over the number of individuals detected at the onset of the flushing (“Ntracked”), so as to provide an estimate of the “worst” possible fate for fishes.

Species/group N tracked Number of radiotracked individuals Percentage of radiotracked individuals Overall apparent survival rate (%)

Present Active Motionless Lost Present Active Motionless Lost

Barbus barbus Up ≥ down VB 1 1 1 0 0 100.0 100.0 0.0 0.0 Down VB 18 15 13 2 3 83.3 86.7 13.3 16.7 Down CP 10 6 5 1 4 60.0 83.3 16.7 40.0 Total 29 22 19 3 7 75.9 86.4 13.6 24.1 65.5

Squalius cephalus Up ≥ down VB 1 1 1 0 0 100.0 100.0 0.0 0.0 Down VB 9 8 7 1 1 88.9 87.5 12.5 11.1 Down CP 5 1 0 1 4 20.0 0.0 100.0 80.0 Total 15 10 8 2 5 66.7 80.0 20.0 33.3 53.3

Rutilus rutilus Up ≥ down VB 0 ––– –––– – Down VB 10 3 2 1 7 30.0 66.7 33.3 70.0 Down CP 1 0 –– 1 0.0 –– 100.0 Total 11 3 2 1 8 27.3 66.7 33.3 72.7 18.2

Salmo trutta Up ≥ down VB 1 1 1 0 0 100 100.0 0.0 0.0 Down VB 15 12 9 3 3 80.0 75.0 25.0 20.0 Down CP 2 0 –– 2 0.0 –– 100.0 Total 18 13 10 3 5 72.2 76.9 23.1 27.8 55.6

All species Up ≥ down VB 3 3 3 0 0 100.0 100.0 0.0 0.0 Down VB 52 38 31 7 14 73.1 81.6 18.4 26.9 Down CP 18 7 5 2 11 38.9 71.4 28.6 61.1 Total 73 48 39 9 25 65.8 81.3 18.8 34.2 53.4

habitats such as reeds, where fish overwinter but remain undetectable conditional on a particularly high estimate in May 2016 (pre-CSFO), by hydroacoustics. Less fish using the main channel and reservoir which may reflect adults migrating for reproduction (FISHLAB/GE, could explain lower acoustic densities in December 2016 and 2017, 2020). This indicates that the full lowering of this reservoir did not and therefore point out the limitation of our sampling. Overall, results allow fish to withstand the SSC, water velocities, and habitat volume showed that the seasonal pattern of fish density did not change across contraction, and that the CSFO may have impaired the migration pro- years and hinted that the flushing did not alter this pattern. cess. Fish were forced to drift downstream, as observed by the radio- On the contrary, the mean acoustic scattering strength in the CP res- telemetry study (see below) during previous drawdown flushing of ervoir drastically dropped after the flushing. However, this result is the VB reservoir (Hofmann et al., 2001, Grimardias et al., 2017), or in

Fig. 5. Mean distance (thick trait; represented as the distance below the VB dam, in m) and mean daily speed (m·day−1; grey bars) of tracked individuals from the onset of ﬂushing to 10 days after (J-0 to +9) for barbel (A), chub (B), trout (C), and for all species pooled (D, excluded roach). Only the individuals present at the onset of ﬂushing and that were detected until 9th June were considered (Ntotal =37;N = 20 for barbel, N =8forchub,N = 9 for trout).

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Table 2 This could be explained by the death of fish due to high and lengthy Synthesis of the biological effects of the 2012 empty flushing and of the 2016 CSFO, with SSC, high flows or low dissolved oxygen concentrations (DO) during their operational characteristics. SEV values are calculated from Newcombe and Jensen flushing, or by escape outside the study area. DO can be inversely (1996). Note that data from hydroacoustics (SA) only refer to the VB reservoir. correlated to SSC during flushing and can reach levels lethal for Flushing event fish (Hesse and Newcomb, 1982; Buermann et al., 1995; Baoligao 2012 2016 et al., 2016). In our study, DO remained at high levels throughout −1 −1 Operational characteristics the CSFO (>10.6 mg·L ; mean DO = 11.7 mg·L ; SIG and 3 −1 Reservoir emptying Full Partial SFMCP, 2017b). Flow reached an hourly maximum of 653 m ·s , SSC below the dam Not controlled Controlled which is a frequent occurrence (e.g. the 2-year return period flood Period 9–21 June 20–30 May is 951 m3·s−1) and cannot explain fish mortality. High SSC can Duration (hours) 278 244 − cause fish mortality in the case of flushing, for very distinct SSC - du- Mean (peak) SSC (g·L 1) 11 (48) 3.47 (11.98) Estimated released sediment volume (Mm3)3.21.3 ration conditions. Espa et al. (2013) reported a reduction in juvenile − Estimated water volume used (Mm3) 368 400 brown trout abundance after 2 weeks at a mean SSC of 3–4g·L 1. Flushing efficiency 0.87% 0.33% Grimardias et al. (2017) observed lower survival rates for three spe- fi fl Metrics for fish impact assessment cies and many dead sh after a 13-day ushing with a mean SSC of −1 −1 SEV - severity of ill effect (Salmonids/non-Salmonids) 11.4/10.7 10.5/10.3 11 g·L and a peak at 48 g·L . Baoligao et al. (2016) observed 2 −1 ↘→ Mean acoustic scattering strength (SA,m ·ha ) many dead fish from 10 species in 5 families after a 58 h flushing after/before flushing −1 −1 fl with a mean SSC of 90.8 g·L (peak = 302 g·L ), but there was Return of SA to pre- ushing level (resilience) No Yes −1 Displacement distances of marked fish below the dam <1500 <2000 an interaction with a strongly decreasing DO (<1.1 mg·L for 1 h) (m) and a high temperature (26–28 °C). As expected by the SEV model, Overall apparent survival rate Φ of marked fish (%) 73.8 69.6a which predicted 0–20% mortality, the SSC during the 2016 CSFO Observed mortality along the riverbanks high no could have caused the death of fish, especially the younger life a Calculated from individuals present only in the downstream section of VB, without stages (Crosa et al., 2010). However, from our field observations, considering roaches, as established in Grimardias et al. (2017) for the 2012 flushing for a the flushing caused little direct mortality, and most of the fish pre- better comparison. sumably left the study area and escaped downstream. The gradual movement of fish downstream during the CSFO con- other studies (Bergstedt and Bergersen, 1997). In the autumn and win- firms this idea, with mean displacements of 95 m for chub, 784 m for ter 2016, acoustic densities did not differ from those observed the pre- barbel, and 1732 m for trout. This contrasted with movements observed vious year at the same seasons. This suggested that the medium-term the month before, and during the 2 months after the flushing, where the impacts of the 2016 flushing were not discernible from the natural pat- mean position of fish hardly changed. At the time of flushing (May), tern of fish density variations in the absence of disturbance. mature individuals from many species, including chub, barbel and roach, were supposed to move upstream to search suitable 4.2. Fish apparent survival and movements below the VB dam spawning habitats (FISHLAB/GE, 2020). Downstream movements are unexpected and reflected a behavioural response to high SSC and Assessing the impacts of flushing using capture-recapture data to water level dropping, with most fish seeking refuge (Bisson and Bilby, analyse the behaviour of individual fish and their apparent survival is 1982; Berg and Northcote, 1985). In a field experiment, Mori et al. uncommon. Most studies deal with population density estimates (e.g. (2018) showed avoidance response of Plecoglossus altivelis when the CPUE, Capture per Unit Effort) and interpret reduced densities as mor- SSC exceeded 200 mg·L−1,withmostfish moving into the non-turbid tality. However, because density estimates may be uncertain, especially stream. In the case of high SSC, refuges can be as diverse as clear in large rivers, and population densities vary naturally over time, a den- water tributaries (Grimardias et al., 2017), vegetated areas, woody de- sity decrease does not imply that fishes are dead; uncontrolled factors bris along the banks, or coarse substrate particles in the riverbed. The such as stocking and fishing may blur the effects of flushing, and these spatio-temporal heterogeneity of the SSC below the dam, both in the studies may be inconclusive (Brignoli et al., 2015; Espa et al., 2015; longitudinal and cross-sectional dimensions, can also reveal areas near Espa et al., 2016). Analysing individual behaviour is a valuable alterna- the banks with lower concentrations that can help mitigate the harmful tive as movement patterns show an early signal of impairment, along impacts of the flushing (Tritthart et al., 2019). with possible physiological damages, but not necessarily implies the The mean overall 53.4% apparent survival rate masked strong dis- death of fish. Our approach however suffered some limitations, espe- parities between species, with high rates for trout, chub and barbel cially concerning the survival estimates. We had no means to ascertain (55.6, 53.3 and 65.5%, respectively), but a much lower rate for roach that a “lost” or “motionless” fish was dead, thus we could only estimate (18.2%). The general tendency that Salmonids are more sensitive to ele- an apparent survival. Most probably, “lost” fish left the study area, either vated SSC than Cyprinids (Newcombe and Jensen, 1996) must be nu- dead or alive. We categorized fish as “motionless” if they were immobile anced and was not verified here with the measured SSC range. We for several successive tracking sessions and never recovered movement showed that different Cyprinid species could exhibit contrasted re- until the end of the survey, and/or if we picked up the transmitter's sponses. Sutherland and Meyer (2007) also found distinct responses mortality signal; we hypothesized that those fish were dead. This to SSC between two Cyprinids (Cyprinella galactura vs Eryomona allowed us to refine estimates and assess an overall apparent survival monachus), and a higher sensitivity than Salmonids to low and moder- rate” (Table 1). Thus, we were confident that fishes were alive and ate SSC. Various responses to SSC were found within genera of five still present in the study area, reflecting a true minimal survival rate Notropis sp. (Cyprinidae; Gray et al., 2014). Our study showed that to the CSFO. In addition, we only considered adult individuals large roach was the most sensitive to the CSFO. Confirming our results, a enough to be marked, but less sensitive to the CSFO than young-of- video-counting monitoring at the CP fish-pass allowed the observation the-year or juvenile fish. of ca. 11,000 roach individuals just before the flushing (5–19 May As a whole, the daily apparent survival rate decreased during the 2016), while only 40 were recorded after (1 June–13 July 2016; CSFO for all species and differed across species. Models including the FISHLAB/GE, 2020). This suggests that the flushing abruptly stopped time effect were not supported, indicating that the rate did not the migration dynamics and forced individuals to move downstream. change significantly between days. The impact was gradual, with Overall, these results stress that impact assessment based on a single the apparent survival rate diminishing slightly day after day and model per family (e.g. in Newcombe and Jensen, 1996) may overlook being significantly lower during the flushing period than outside. species-specific sensitivity.

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4.3. Comparison of two sediment management operations CSFOs with an accepted balance between the released sediment load and the ecological impacts are described with lower mean Overall, the 2016 CSFO was less detrimental to the fish community SSC (Espa et al., 2015, Reckendorfer et al., 2019). than the 2012 full drawdown flushing. Two reasons may explain this re- The limitation of peak SSC's can help mitigate ecological impacts sult: first, both operations were different in nature, 2012 being an of CSFO (Crosa et al., 2010). Peak SSC values during flushing com- empty flushing while 2016 was a pressure flushing, where the reservoir monly reach between 10's and 100's of g·L−1, generally for less was not fully lowered (Fig. 3). This explains the clear differences ob- than one to a few hours. Such SSC provoke gill damages and block ox- served in fish remaining in the reservoir: the residual water volume, ygen supplies, and usually lead to the death of fish (Kemp et al., the gradual changes in habitat and the smoother hydraulic conditions 2011). Garric et al. (1990) showed that 100% of trout juveniles died induced by the partial lowering allowed fish to maintain their position after 5 h at ca. 15 g·L−1. Newcombe and MacDonald (1991) reported in the reservoir, whereas they were forced to move downstream of 60% mortality in juveniles Salmonids after 6 h at a mean of 82 g·L−1. the dam when the lowering was full. As a result, the fish density and dy- Grimardias et al. (2017) showed that the survival for three fish spe- namics post-flushing showed no signs of impairment in 2016, suggest- cies dropped during the firstthreedaysaftertheonsetofflushing, ing a good resistance to the disturbance; the resilience was weaker in when the mean SSC approximated 18 g·L−1, with regular peaks 2012, where recovery was not achieved 17 months after the flushing >30 g·L−1. For the Valgrosina Reservoir (Italia), Crosa et al. (2010) (Grimardias et al., 2017). It is also likely that lower trophic levels such recommended not to exceed a daily average of 10 g·L−1.Thethresh- as phyto- and zooplankton, and macroinvertebrates were more severely olds of 15 g·L−1 for 0.5 h, and 10 g·L−1 for 6 h looks like an accept- depleted following the empty flushing (see also Quadroni et al., 2016, able compromise in the case of VB reservoir, as it allows an Espa et al., 2019), with a longer recovery time (Peter et al., 2014). Fish efficient flushing while limiting the fish mortality. may have suffered indirectly for several weeks of this food shortage in Sediment flushing operations are pulse events (e.g. Jentsch and White, 2012. Second, the control of SSC below the dam in 2016 drastically 2019) affecting a mainstem river reach within a hydrographic network. helped reducing fish mortality. Though we did not quantitatively assess Referring to metapopulation theory (e.g. Hanski and Gilpin, 1991)and this indicator, there is no doubt that, from our field observations and disturbance ecology (Resh et al., 1988), assemblage resilience can be im- from those of the many people involved, the 2012 flushing resulted in proved when the disturbed reach is connected with undisturbed patches, severe fish mortality (GREN, 2012) while there were no observed i.e. tributaries, allowing stressed fish to find a refuge and new individuals dead fish after the 2016 CSFO. This is partially in agreement with the to recolonize the reach after the perturbation (Yount and Niemi, 1990). SEV indices, which predicted ranges of 0–20% (for 10 ≤ SEV < 11) and The “tributary refuge hypothesis” has been validated in the case of flood 20–40% (for 11 ≤ SEV < 12) mortality, and which were higher in 2012 disturbances (Koizumi et al., 2013), but also with an experimental in- than in 2016 (11.4 vs 10.5 for Salmonids, and 10.7 vs 10.3 for Cyprinids, crease of SSC (Mori et al., 2018) and in the case of the 2012 VB empty respectively). Correlated with the controlled SSC, and because the dura- flushing (Grimardias et al., 2017). Provided that tributaries are present, tion of flushing couldn't be too extended for economic purposes, the management operations that facilitate access to refuges and increase volume of evacuated sediment and the flushing efficiency were much their carrying capacity, such as pool digging at their confluence, may be lower in 2016 (Table 2). valuable. Translocation of individuals before the onset of the flushing may also help. In addition, in the case of the VB reservoir, the presence 4.4. Implications for sediment management of Lake Geneva upstream constitutes a source patch that offers a unique opportunity to promote the assemblage resilience. As long as the lake The implementation of multiple SSC - duration thresholds for the VB water level is still high enough at the end of flushing, opening the Seujet reservoir since 2016 is a clear improvement of flushing practices toward dam at the outlet of Lake Geneva (cf. Fig. 1) will increase the water flow ecological considerations. Using the same thresholds as for the Génissiat and facilitate the recolonization of the VB reservoir. Dam in France (Peteuil et al., 2013) located 40 km downstream, empha- The definition of SSC – duration thresholds should consider the natu- sizes the international coordination between Swiss and French stake- ral sediment regime of the reach. Organisms have evolved with, and are holders and allows a uniform and ecologically friendly sediment adapted to, spatiotemporal variations of the natural flow and sediment management over the whole Upper-Rhône River. Defining the SSC - du- regimes of the river (Poff et al., 1997; Wohl et al., 2015), so that only ‘ex- ration thresholds is difficult, as it results in a trade-off between the high treme’ events cause significant changes in assemblages' structure and SSC needed to release a substantial volume of sediment over a short abundance. Disturbance regimes (in terms of intensity, duration and fre- time at an acceptable economic cost, and the much lower SSC required quency) shape the functional and life history attributes of organisms and to preserve the environment downstream. Accounting for the mean provide assemblages with the resistance and resilience capacities to face and peak (over 6 h and 0.5 h, respectively) values in threshold def- environmental stochasticity. Reservoir flushing releases vast sediment inition is relevant as both can be very different and can trigger dif- loads in short time periods and generates high SSCs. Lepage et al. ferent responses on fish. Peak flows are often necessary to initiate (2020) estimated that the SSC during the VB 2012 flushing was 6 to 8 the mobilization of fine cohesive sediment, and sudden sediment times higher than during major floods, and that about 1/3 of the annual releases generating high SSC peaks during flushing may be unavoid- suspended sediment load transited over the 19 days (5% of time) of the able and unpredictable. The mean SSC value over the whole flushing operation. The SEV model of Newcombe and Jensen (1996) provides a of 5 g·L−1 is similar or close to recommendations from other studies rough - though useful as a guideline – estimate of fish impacts based on in an Alpine context (Schneider et al., 2006; Crosa et al., 2010; the SSC and duration but does not account for the adaptation of the Fruchart and Camenen, 2012). These showed significantly lower local fauna to the sediment regime of the river. This may explain discrep- values (e.g. 1.5 g·L−1 for the Sernio reservoir on the Adda River, Ita- ancies with empirical observations (e.g. Xu et al., 2018). The development lia: Espa et al., 2015;1g·L−1 for the Gepatsch reservoir, Tyrol, of models including the departure of the flushing characteristics from Austria: Hauer et al., 2020). Impacts of sediment releases are site- those of the river sediment regime may improve impact assessments. specific, and it is risky to extrapolate these threshold values to This would however require more extensive field monitoring data, other contexts. Salmonids are not dominant in the Rhône River, which are still scarce concerning reservoir flushing. but two threatened species are present (Salmotruttaf.lacustris and T. thymallus). Some Cyprinids may be more sensitive to SSC 5. Conclusions than Salmonids, as shown here for roach. As such, we must consider this mean value as an upper limit, not a target value, and we would The CSFO implemented for the first time in the VB dam in 2016 is the recommend not exceeding this threshold. Reports of ‘successful’ pivotal measure of a new sediment management strategy. From an

10 F. Cattanéo, J. Guillard, S. Diouf et al. Science of the Total Environment xxx (xxxx) xxx ecological viewpoint, it advantageously replaces the empty flushing of References 2012 that occurred after a 9-year period without major sediment man- fi Adams, N.S., Beeman, J.W., Eiler, J.H. (Eds.), 2012. Telemetry Techniques: A User Guide for agement event, but impacts were still signi cant. Main improvements Fisheries Research. American Fisheries Society, Bethesda, Maryland. are that the CSFO preserved the fish density in the VB reservoir while Aglen, A., 1983. Random Errors of Acoustic Fish Abundance Estimates in Relation to the that of the CP reservoir recovered in less than a year, therefore enabling Survey Grid Density Applied. FAO Fisheries Report. Food and Agriculture Organiza- a better resilience of the ecosystem. Except for roach, which seemed to tion of the United Nations. Balk, H., Lindem, T., 2011. Sonar4 and Sonar5-pro Post Processing Systems. Operator man- be especially sensitive to SSC increase, apparent survival estimates for ual, version 6.0.1. three representative species were close to those reported in 2012, but Baoligao, B., Xu, F., Chen, X., Wang, X., Chen, W., 2016. Acute impacts of reservoir sedi- fl fi – emigration, rather than direct mortality, most likely accounted for this. ment ushing on shes in the Yellow River. J. Hydro Environ. Res. 13, 26 35. Bash, J., Berman, C.H., Bolton, S., 2001. Effects of Turbidity and Suspended Solids on Sal- It is noteworthy that the shift toward including CSFOs in the sediment monids. University of Washington Water Center. management strategy with the defined SSC – duration thresholds im- Basson, G.R., 2009. Management of siltation in existing and new reservoirs. 23rd Congress plies that such operations should be scheduled every 3–5 years to main- of the International Commission on Large Dams, Brasilia. Bates, D., Mächler, M., Bolker, B., Walker, S., 2015. 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Comité technique franco-suisse, rapport de synthèse (38 p). Crosa, G., Castelli, E., Gentili, G., Espa, P., 2010. Effects of suspended sediments from reser- The authors declare that they have no known competing financial voir flushing on fish and macroinvertebrates in an alpine stream. Aquat. Sci. 72 (1), – fl 85 95. https://doi.org/10.1007/s00027-009-0117-z. interests or personal relationships that could have appeared to in u- Espa, P., Castelli, E., Crosa, G., Gentili, G., 2013. Environmental effects of storage preserva- ence the work reported in this paper. tion practices: controlled flushing of fine sediment from a small hydropower reservoir. Environ. Manag. 52 (1), 261–276. Acknowledgements Espa, P., Crosa, G., Gentili, G., Quadroni, S., Petts, G., 2015. Downstream ecological impacts of controlled sediment flushing in an Alpine valley river: a case study. River Res. Appl. 31 (8), 931–942. This study was funded by the Services Industriels de Genève (SIG, Ge- Espa, P., Brignoli, M.L., Crosa, G., Gentili, G., Quadroni, S., 2016. Controlled sediment flush- neva, Switzerland) and the Société des Forces Motrices de Chancy-Pougny ing at the Cancano Reservoir (Italian Alps): management of the operation and downstream environmental impact. J. Environ. Manag. 182, 1–12. (SFMCP, Geneva, Switzerland) under contract n°45320234, and by the Espa, P., Batalla, R.J., Brignoli, M.L., Crosa, G., Gentili, G., Quadroni, S., 2019. Tackling reser- COGEFé (Geneva, Switzerland; convention 2015-M05). We thank Carole voir siltation by controlled sediment flushing: impact on downstream fauna and re- Nawratil de Bono (SIG), Estelle Lecomte (SIG) and Fabio Heer (COGEFé) lated management issues. PLoS One 14 (6), e0218822. FISHLAB/GE, 2020. http://www.fishlab.ch/pointinfo-ndeg2-abaissement-2016.html Vis- for their kind support. We are indebted to the people of the cantonal de- ited on 24 June 2020. partment ‘Direction Générale de l'Eau’ for study approval and for their Fruchart, F., Camenen, B., 2012. Reservoir sedimentation: different type of flushing - precious help in the field, namely Dimitri Jaquet, Jean-Louis Delabays, friendly flushing example of Genissiat dam flushing. ICOLD International Symposium on Dams for a Changing World, 05/06/2012, Kyoto, Japan. and Christophe Reymond. We are also grateful to the assistants and stu- Garric, J., Migeon, B., Vindimian, E., 1990. Lethal effects of draining on brown trout. A pre- dents from Hepia Geneva who helped in many occasions. dictive model based on field and laboratory studies. Water Res. 24 (1), 59–65.

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