Open Access
Issue
Knowl. Manag. Aquat. Ecosyst.
Number 425, 2024
Article Number 15
Number of page(s) 10
DOI https://doi.org/10.1051/kmae/2024013
Published online 04 October 2024

© E. Arevalo et al., Published by EDP Sciences 2024

Licence Creative CommonsThis is an Open Access article distributed under the terms of the Creative Commons Attribution License CC-BY-ND (https://creativecommons.org/licenses/by-nd/4.0/), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. If you remix, transform, or build upon the material, you may not distribute the modified material.

1 Introduction

Pickett and White (1985) defined natural disturbance as a relatively discrete event in time that disrupts ecosystems, including communities, as well as changing the physical environment. In natural ecosystems, communities are subject to disturbances of different frequencies and intensities (Downes et al., 1998). Organisms may adapt to predictable seasonal fluctuations (photoperiod, discharge, temperature, etc.), but adapt with difficulty to disturbances that vary greatly in their frequency and intensity over temporal and spatial scales (Resh et al., 1988). In rivers, hydrological disturbances are major sources of mortality both for macroinvertebrates and fish (Naiman et al., 2008). With climate change, summer droughts and drying events are expected to become increasingly severe and prolonged in European rivers (Trenberth et al., 2014; Dayon et al., 2018; Reid et al., 2019) because of the predicted reduction of precipitation levels together with increased evapotranspiration as both air and water temperatures rise. Additionally, water deficits are expected to be exacerbated by increased water uses for human activities (i.e., agriculture, industry or drinking water, Palmer et al., 2009).

In streams, climate is a strong driver for fish assemblages (Magalhães et al., 2007), including for salmonid populations (Zabel et al., 2006; Walsh and Kilsby, 2007), with severe consequences for development, growth, behaviour, survival and population dynamics (Clews et al., 2010). During droughts and low flow periods, shallow sections (such as riffles and runs) disappear and streams become a series of fragmented pools (Lake, 2003). Fish emigration is prevented (Elliott et al., 1997; Armstrong et al., 2003) and fish are confined to restricted areas with a high risk of mortality from desiccation, predation, starvation or suboptimal environmental conditions, particularly regarding dissolved oxygen concentration (Magalhães et al., 2002). For example, Elliott (2000) demonstrated that severe droughts affected the trout population of Black Brows Beck (UK) by increasing mortality and limiting the growth of both juveniles and adults, which was partly attributed to a decrease in resource availability.

Although macroinvertebrate larvae can use the hyporheic zone as a refuge when facing drying events, their response is highly context dependent (Wood et al., 2010). Several studies reported an increase in the abundance of macroinvertebrates in the hyporheic zone during adverse conditions in surface streams (e.g. floods or drying events), whilst others did not observe any significant changes in macroinvertebrate distribution (Dole-Olivier, 2011; Stubbington, 2012). Part of the reason for having these contradictory observations can be linked to macroinvertebrate taxa showing a very variable set of ecological traits (Fenoglio et al., 2007). The assemblages of aquatic ecosystems regularly subjected to droughts are dominated by small or medium-sized, short-lived macroinvertebrates with aquatic respiration and aerial dispersal such as Chironomidae, Simuliidae and Ephemeroptera, which can very successfully use the hyporheic zone as refuge during streambed drying (Williams and Hynes, 1974; Boulton et al., 1998). Additionally, the physico-chemical characteristics of the hyporheic zone of riffles could be more favourable to macroinvertebrate survival than the conditions in isolated pools created during droughts (e.g. cooler water temperatures, higher dissolved oxygen concentrations; Dewson et al., 2007). Depending on the survival of macroinvertebrates during these hydrological events, the subsequent delivery rate of prey by drift to fish can highly change (Atkinson et al., 2014; Chadd et al., 2017). In this regard, Hakala and Hartman (2004) revealed that limited food availability caused by drought reduced body condition (body weight and fat reserves) and the winter survival of brook trout (Salvelinus fontinalis). The authors highlighted that a 10% decrease in adult weight caused by drought resulted in a reduction of young-of-the-year (YoY) production by 67% the following year. Thus, hydrological events can cause long-term effects and lead to the decline of fish populations.

In the present study, we tested the effect of a surface water drying event (with a continuous flow in the hyporheic zone) on the trophic availability of brown trout juveniles. In 6 experimental channels, a reference discharge was maintained for 48 days to allow macroinvertebrates to colonise. Then, the discharge was diminished by 80% and the water level was reduced to the benthic surface in 3 channels for 2 weeks. Hydraulic conditions in the 3 dry channels returned to reference conditions and juvenile brown trout were added in all 6 experimental channels during 3 weeks of growth. We hypothesized that macroinvertebrates use the hyporheic zone as a refuge, mitigating the decrease in macroinvertebrate abundance caused by the surface water drying event and, consequently, the negative impacts on the survival and growth of brown trout juveniles.

2 Materials and methods

2.1 Experimental design

This experiment was conducted at the National Institute of Agronomic and Environmental Research's (INRAE) experimental facilities authorised for animal experimentation (A640141). The experiment was carried out in strict accordance with the E.U. legal frameworks, specifically those relating to the protection of animals used for scientific purposes (i.e., Directive 2010/63/EU), and under the French legislation governing the ethical treatment of animals (Decret no. 2013-118, February 1st, 2013).

The experiment was conducted in 6 outdoor channels (L: 12 m, W: 52 cm and H: 25 cm), continuously fed by gravity with unfiltered water pumped from an outdoor reservoir (4500 m3), which gets water from the Aitzeguerria and the Lapitxuri brooks in south-western France (43°16' N, 1°28' W; Fig. 1). In each 12-metre-long channel, a layer of 10 cm of coarse substrate (ø = 3 cm) was uniformly distributed over the 10 central metres and kept in place by two 0.5 cm mesh grids on both sides. Seventeen cobbles (ø = 7–8 cm) per channel were added to provide shelter and visual isolation for fish. Channels were covered with a tarpaulin for the experiment to mimic brightness under the shade of trees in the Lapitxuri brook (684 lux on average, SD = 418, Iso-tech digital light meter Lux-1335). Water temperature was recorded every hour using a temperature logger at each channel (mini logger IIT, Vemco; Fig. S1).

Before the start of the experiment (i.e., from 12/05/2015 to 29/06/2015), the 6 channels received a discharge of 1 L s−1 that created a water height of 4.5 cm above the substrate. These reference conditions were kept for 48 days to allow the natural colonisation of the channels by macroinvertebrates. After this colonisation period, a 100 µm mesh net was placed in the water inlet of each channel to exclude the immigration of macroinvertebrates from the beginning of the drying event to the end of the experiment (Mureithi et al., 2018; Slack et al., 1991). The drying event was simulated in 3 channels and lasted two weeks. The discharge was reduced to 0.2 L s−1 and the water height to 0 cm (benthic surface level). The other 3 channels were kept with the initial discharge. After the 2-week drying event, the 3 channels were brought back to reference discharge conditions.

thumbnail Fig. 1

Experimental design. Water levels were dropped onto the surface of the sediment in some channels (A, left channel), while 4.5 cm of water was maintained in others (A, right channel). (B) Channels were naturally fed by the Aitzeguerria and Lapitxuri brooks. After 48 days of colonisation by macroinvertebrates 100 μm mesh nets were placed at the beginning and end of the channels to study the dynamics of the community in place. A drying event was simulated for 15 days (grey channels) and then fish grew for 21 days (hatched areas).

2.2 Macroinvertebrates

Macroinvertebrates were sampled three times in the upstream, middle and downstream parts of each channel: before the drying event, after it and after the three weeks of fish rearing (see “Fish” section below). To obtain the macroinvertebrate samples, a corer with a 20 cm diameter was buried into the substrate, gravel was removed and macroinvertebrates in the water within the pipe pumped and sieved through a 500 μm mesh and preserved in 70% ethanol. The locations for the first and second sampling corers were randomly chosen, but after the fish rearing period, two samples were collected in the 6-metre upstream channel section without trout and a sample in the 4-metre downstream section that had trout juveniles (see “Fish” section below). A total of fifty-four samples were obtained (three periods × six channels × three samples). Macroinvertebrates were identified and grouped into the following taxonomic groups: Diptera, Coleoptera, Oligochaeta, and “Others”, which included all taxa representing less than 5% of total abundance (i.e., Trichoptera, Ephemeroptera, Mollusca, Plecoptera and Crustacea).

2.3 Fish

Brown trout eggs were obtained through the artificial fertilization of gametes of wild brown trout (1♀ × 1♂) caught in the Nivelle watershed and grew for 4 months under semi-natural conditions (artificial channel fed by a derivation from a headwater stream, the Lapitxuri brook, with natural food and substratum). A total of 102 juveniles were produced for the experiment. After a 24 h fasting period, juveniles were anesthetized (30 mg L−1 benzocaine), individually weighed (wet mass) on a Sartorius CP153 electronic balance to the nearest 0.001 g, measured (total body length) to the nearest 1 mm and photographed for individual identification using melanophore distribution patterns (Garcia de Leaniz et al., 1994). Trout were sorted into 6 similar batches (according to weight and length) of 17 fish with an average individual length of 55.67 mm (SD = 1.88) and an average individual mass of 8.76 g (SD = 1.00). Following recommendations by Grant and Kramer (1990), we estimated that the minimal area for the set of juveniles of this length (17 fish of 55 mm) was 2 m2. Thus, to rear fish under conditions close to habitat saturation, the downstream 4 m of the channels were separated from the top 6 meters by means of a 0.5 cm mesh grid (4 m long × 0.5 m wide = 2 m2), and fish were added only to this part of the channels. The top 6 meters had no fish but contributed to the production of drifting macroinvertebrates. After the drying event, fish were released into the 6 experimental channels for three weeks. They were recovered by electrofishing at the end and after a 24 h fasting period they were anesthetized, individually weighed, measured, and photographed. The performance of fish was estimated by means of survival rate, individual growth and condition factor. After the experiment, juveniles were released into the Lapitxuri brook.

2.4 Statistical analyses

Abundance of macroinvertebrates, survival, growth and condition factor of fish were analysed using a Bayesian modelling approach computed with OpenBUGS®.

2.4.1 Abundance of macroinvertebrates

Macroinvertebrates were sampled at 3 periods (k): before the drying event (k = 1), after 2 weeks from the drying event (k = 2) and after 3 weeks of trout growth (k = 3). The abundance of macroinvertebrates was therefore compared between control and dry channels for these 3 periods. We also evaluated the predation pressure exerted by the fish on macroinvertebrates by comparing the 6-meter upstream section (without trout) with the 4-meter downstream section of the channels (with trout). If Abundancei was the number of individuals counted in the ith sample, we assumed:

μ.log(Abundancei+1)i=μ+αk[i]Dryi+βj[i]+γk[i]+δFishl

where µ was the average of the logarithm for macroinvertebrate abundance, α was a fixed-effect parameter for the differences between the control (Dry = 0) and dry channels (Dry = 1) before the drying event (k1), after 2 weeks of drying (k2) and after 3 weeks of trout growth (k3), β was a random effect corresponding to each j channel, γ was a random effect corresponding to the k steps and δ was a fixed-effect parameter for trout predation pressure (l = 0 without fish and l = 1 with fish). The same modelling treatment was applied to the total abundance of the main macroinvertebrate groups (i.e., Diptera, Coleoptera, Oligochaeta and “Others”).

2.4.2 Survival of trout

Fish found outside the growing area or not recovered were not considered for the analyses (Si = 0). Then, if p.Si was the probability of survival of the ith fish, we assumed:

SiBernoulli(p.Si)

Logit(p.Si)=μ+αDryi+βj[i]

where µ was the mean of the logit for survival probability, α was a fixed-effect parameter for the indirect impact of drying (Dry = 0 for control and Dry = 1 for dry channels) and β was a random effect corresponding to each j channel.

2.4.3 Growth of trout

All the fish were weighed at the beginning of the growing period (W.begi) and all the survivors were weighed at the end (W.endi). Photos allowed the individual recognition of juveniles and growth (Gi) was calculated by dividing the difference between final and initial weight by initial weight (in g). We stated that the growth Gi followed normal distribution with µ.Gi the mean and σ.Gj the standard deviation. We assumed:

μ.Gi=μ+αDryi+βj[i]

where µ was the mean of the individual growth of juveniles, α was a fixed-effect parameter for the indirect impact of drying (Dry = 0 for control and Dry = 1 for dry channels) and β was a random effect corresponding to each j channel.

2.4.4 Condition factor of trout

Fulton’s condition factor K for each i trout was calculated with empirical data by dividing the final weight (in mg) by length (in mm) cubed multiplied by 100 (Fulton, 1911). Then, we assumed:

μ.Ki=μ+αDryi+βj[i]

where µ was the mean of the condition factor, α was a fixed-effect parameter for the indirect impact of drying (Dry = 0 for control and Dry = 1 for the dry channels) and β was a random effect corresponding to each j channel.

2.4.5 Bayesian computation

Parameters were given independent “weakly informative” priors (Tab. S1; Gelman, 2006; Gelman and Hill, 2007). For each analysis, three independent chains were used. The first 10,000 iterations were discarded as an initial burn-in period. Then, 10,000 further iterations (resulting from one every ten runs) were performed. The convergence of the chains to their ergodic distribution was tested via the Gelman-Rubin (GR) diagnostics integrated in OpenBUGS®. The significance of the parameters was tested by calculating the posterior probability of these parameters to be positive (P(X > 0) > 0.90) or negative (P(X > 0) < 0.10) (Gelman and Hill, 2007).

3 Results

The abundance of all macroinvertebrate groups was similar in the control and dry channels at all periods (before, after the drying event and after fish growth; Fig. 2, Tabs. 1, and S2), except for the abundance of Oligochaeta, which was significantly higher in the dry channels than in the control channels at the end of fish growth (α3 = 1.50). The presence of fish significantly decreased the abundance of Diptera, Coleoptera and “Others” (δ = –0.698, −1.475 and −0.826, respectively) as well as the total abundance (δ = –0.717) while it did not affect the abundance of Oligochaeta. Considering the periods, the abundance of different groups of macroinvertebrates shifted throughout the summer: the abundance of Coleoptera and Oligochaeta increased (γ1 = γ2 and γ2 < γ3) while the abundance of Diptera, “Others” and the total abundance decreased (γ1 > γ2 and γ2 = γ3; Table 1 and S2).

Of the 102 juveniles used for the experiment, 95 juveniles were recovered at the end. The survival probability was similar across treatments, with a 0.92 survival rate in the control and 0.94 in the dry channels (α = 0.418; P(SurvivalDry > SurvivalControl) = 0.65; Fig. 3 and Tab. S3). In the control channels, juveniles weighed on average 1.89 g (± 1.04) at the beginning and 1.88 g (± 1.10) at the end of the experiment. In the dry channels, they weighed on average 1.87 g (± 0.97) at the beginning and 1.86 g (± 1.11) at the end. Juvenile growth in the control channels was not significantly different from growth in the dry channels (α = –0.009; P(GrowthDry > GrowthControl) = 0.43). In both control and dry channels, average growth was negative (–0.03 and −0.04, respectively). The condition factors were also similar in control and dry channels, with average values of 0.93 and 0.92, respectively (α = –0.012; P(Cond.FactorDry > Cond.FactorControl) = 0.31).

thumbnail Fig. 2

Abundance of macroinvertebrates (mean number of individuals by corer) before the drying event, after 2 weeks of drying and after 3 weeks of fish growth in reference discharge conditions in the control and dry channels, shown in white and grey respectively. Each condition represents 9 samples, for a total of 54 samples. Error bars represent standard errors.

Table 1

Probability of a different abundance between control and dry channels (α1, α2 and α3), the 3 periods (γ1 vs. γ2 and γ2 vs. γ3) and the upstream channel sections without fish and the downstream sections with fish (δ). The parameter influence on invertebrate abundance is symbolised by “▲” when the abundance increased significantly (significance probability ≥ 0.90; in bold), “=” when it did not affect the abundance (0.10 < significance probability < 0.90) or “▼” when the abundance decreased significantly (significance probability ≤ 0.10; in bold).

thumbnail Fig. 3

Model estimates of the performance of juvenile trout over the 3-week growth period in the control and dry channels. Performance includes survival probability (left), growth (middle) and condition factors of juvenile trout (right). Boxplots indicate the percentiles 1, 25, 50, 75 and 99 of the posterior distributions.

4 Discussion

Hydrological anomalies are natural phenomena and more than 51–60% of the global fluvial network experiences drying (Messager et al., 2021). However, these events are expected to intensify in the context of climate change (Skoulikidis et al., 2017). Droughts and drying shape instream biological assemblages but the impacts of such events depend on the natural water regime of rivers, timing and severity (duration and intensity) of the drying disturbance, as well as the presence of refuges (Boulton, 2003). These events are extremely challenging to reproduce under experimental conditions because they strongly influence several ecosystem components (both physico-chemistry and aquatic organisms) and spatio-temporal scales. A drought crosses successive biological thresholds (reviewed in Boulton, 2003). In the beginning, the water level drops and the riparian zone is no longer connected to the main river channel. Then, river water flow stops, and the watercourse is fragmented into pools. The disappearance of surface water is the most critical stage for the fauna, leading to massive mortality. The final stage would be the water level declining in the hyporheic zone. Here, we focused on a surface water drying event, but the interstitial flow did not cease. Our results can contribute to our understanding of the impact of surface drying on trophic availability for carnivorous fish in temperate climates, as dry periods in this climatic area are usually brief, highly predictable and mostly restricted to summer (Bogan et al., 2014).

In our study, the two-week surface water drying event did not alter the macroinvertebrate community. Drying events reduce the available surface habitat and connectivity between these, but macroinvertebrates can shelter in the hyporheic zone if the sediment composition allows high hyporheic permeability (Maridet et al., 1992; Schmid and Schmid-Araya, 2010). Coarse gravel could provide enough inhabitable interstices to support high-density and diverse communities (Strayer et al., 1997). The lack of significant effects of a drying event on macroinvertebrate abundance in our experiment could be linked to the use of the hyporheic zone as refuge. To test this idea, we would have needed to create habitats without a hyporheic zone and compare macroinvertebrate communities. We cannot be completely confident that macroinvertebrates sheltered in the hyporheic zone, but our experiment provides circumstantial evidence supporting this hypothesis.

The constant renewal of water in our experiment maintained optimum oxygen conditions in the control and dry channels (above 11 mg L−1 at all times; Tab. S4), which could promote the colonisation of the hyporheic zone by macroinvertebrates. In natural conditions, droughts generally increase water temperature and decrease the concentration of dissolved oxygen. For example, the temperature rose from 14 to 25 °C and dissolved oxygen decreased from 12 to 4 mg L−1 during a 2-week drought in three pools in France's Albarine river (Datry, 2017), creating unsuitable conditions for many aquatic species (Vander Vorste et al., 2020). Similar decreases in oxygen could be expected in the hyporheic zone, but groundwater inputs could also be involved in the physico-chemical processes occurring in the hyporheic zone. Changes in terms of dissolved oxygen in the hyporheic zone during drying events, and the consequences for the biota, are difficult to predict and site specific.

Macroinvertebrate access to the hyporheic zone depends on the permeability of the substrate. The water from the Lapitxuri and Aitzeguerria brooks does not carry large amounts of fine particles (particles < 2 mm in size, following Wood and Armitage, 1997). These two brooks feed the Nivelle River and point measurements over 4 consecutive years near the confluence between the Lapitxuri brook and the Nivelle River revealed between 11 and 20% of fine particles in the gravel bed (Arevalo, 2014). Although we did not measure it during the experiment, the inputs of fine particles into the experimental channels were observed to be negligible. Thus, the very high permeability of the gravel bed was behind the successful colonisation of the hyporheic zone by macroinvertebrates (Stubbington, 2012). The level of fine sediments in the river bed reflects differences in land-use, geology and catchment stability (Boulton, 2003), and is increasingly recognised as a key threat to the ecological integrity of riverine ecosystems globally (Mathers et al., 2017). Previous studies have demonstrated reductions in macroinvertebrate density and diversity directly induced by streambed colmation (Descloux et al., 2013; Vadher et al., 2015, 2018; Mathers and Wood, 2016). Marked changes in the abundance and composition of benthic assemblages are noticeable when levels of fine particles in the streambed exceed 10% (Harrison, 2010), 20% (Zweig and Rabeni, 2001) or 30% (Bo et al., 2007; Descloux et al., 2013; Relyea et al., 2000). Clogging homogenises benthic habitats and decreases available spaces between larger particles, which limits permeability for most macroinvertebrates (Descloux et al., 2013; Mathers et al., 2014, 2019). As the hyporheic zone can contribute to resilience capacity of populations (Milner et al., 2022), guaranteeing its ecological integrity is crucial with ongoing global changes (Jones et al., 2012).

The persistence of macroinvertebrates during the drying event could also be related to the ecological traits of the dominant taxa. Chironomidae (Diptera) and Oligochaeta have a vermiform shape and are adapted to life in the hyporheic zone (Williams and Hynes, 1974; Stubbington, 2012). Thus, the main taxa on our macroinvertebrate communities, with taxa adapted to respond to natural flow variations, could have contributed to the limited effect of the surface water drying event observed (Naiman et al., 2008). However, we also observed that the total abundance of macroinvertebrates decreased during summer, due to a large reduction in the abundance of Diptera (mainly Chironomidae). The abundance of Coleoptera and Oligochaeta increased but not enough to compensate for the loss of Chironomidae (Georgian and Wallace, 1983; Boulton et al., 1992; Milner et al., 2018). For macroinvertebrates, summer is often a stressful period due to low flows associated with high temperatures, diminishing total macroinvertebrate biomass both in the benthos and in the drift (Rashidabadi et al., 2022). The phenology of aquatic insects, with the progressive emergence of adults of many species throughout summer could explain this reduction in macroinvertebrate abundance (Hynes, 1970; Shearer et al., 2003; Baxter et al., 2017). Consequently, habitat quality for stream carnivorous fish, including salmonids, could drop due to a reduction in prey availability (Rashidabadi et al., 2022).

The supply of macroinvertebrates in our experimental channels was controlled with 100-μm mesh nets and tarpaulins to focus on the dynamics of the established community. This experimental choice may also explain the gradual decline in macroinvertebrate abundance throughout summer. In natura, community turnover involves a range of strategies, including the drift of macroinvertebrates from upstream reaches (Datry et al., 2017) and aerial dispersal (both active and passive) from perennial nearby sources (Bogan et al., 2014; Sarremejane et al., 2017). After droughts, 80% of the community can recover after 60 days (Pařil et al., 2019), 73 days (Doretto et al., 2019) or 150 days (Di Sabatino et al., 2023) through these dispersal strategies. However, in a changing global context, a growing number of streams are experiencing dry periods, including perennial streams. Consequently, the effects of global change on network connectivity and the resilience capacity of systems are crucial issues. This experiment provides some evidence demonstrating that a rupture in connectivity leads to a reduction in the abundance of macroinvertebrates in isolated reaches, potentially impeding post-disturbance recolonisation.

The substrate and water velocity configuration of our channels was comparable to riffles in the Lapitxuri brook. However, riffles are suboptimal habitats for fish development in summer because of the limited delivery rate of prey (Atkinson et al., 2014; Chadd et al., 2017) and the energetically demanding conditions (Elliott, 2000; Lennox et al., 2019). Deep pools provide more energetically favourable habitats than riffles for young-of-the-year salmonids in streams (Rosenfeld and Boss, 2001; Harvey et al., 2005; Kahler et al., 2011) by limiting energy expenditure associated with swimming and station holding, while prey capture and detection rates are increased due to the concentration of prey (Nislow et al., 2004). Thus, the reduction in the abundance of macroinvertebrates during summer together with the suboptimal flow conditions in the experimental channels could explain the negative performances of trout juveniles in our experiment (Sweka and Hartman, 2008).

Finally, the presence of fish significantly reduced invertebrate abundance. This phenomenon could be explained by two mechanisms. Firstly, fish can have a direct impact on invertebrates through predation. Diptera can constitute over 80% of their energy inputs (Vignes and Heland, 1995; Sánchez-Hernández et al., 2012; Arevalo et al., 2020), although they are opportunistic and may ingest large quantities of less favourable prey, such as Ephemeroptera and Trichoptera (Johansson, 1991). However, some groups like Coleoptera are less vulnerable to predation because of their chemical defences and unpalatability (Otto and Svensson, 1980; Keeley and Grant, 1997). Secondly, the reduction in macroinvertebrate abundance induced by the presence of fish can be explained by invertebrate predator-avoidance behaviour. Macroinvertebrates detect the chemical cues of predators and actively drift to “safe” areas with low predation pressure to minimize mortality (Muotka et al., 1999; Huhta et al., 2000; Naman et al., 2016), which in our channel meant leaving the experimental channels.

5 Conclusions

In our experiment, the surface water drying event did not reduce macroinvertebrate abundance. The coarse substrate and the high water quality of the hyporheic zone probably provided sufficient refuge for macroinvertebrates. The survival and growth of salmonid juveniles were unaffected by the drying event. Our study suggests that preserving the quality of the hyporheic zone (e.g. by minimizing the inputs of fine sediment transport from the catchment to the watercourses and maintaining optimal oxygen conditions) would be key to mitigating the effects of droughts and drying on macroinvertebrates and fish.

Supplementary materials

Figure S1. Water temperature (°C) recorded after the drying during trout growth (from the 16/07/2015 to the 06/08/2015) in the 3 control (black lines) and the 3 dry channels (grey dotted lines).

Table S1. Prior distributions assigned to parameters in the models used. E, VAR and CV correspond to mean, variance and coefficient of variation, respectively.

Table S2. Main statistics of the posterior probability distribution functions of the α (differences between control and dry channels), β (all the other sources of variations, corresponding to each channel), γ (period) and δ (fish predation) parameters affecting the abundance of macroinvertebrates.

Table S3. Main statistics of the posterior probability distribution functions of the α (dry effect) and β (all the other sources of variations, corresponding to each channel) parameters affecting the survival, the growth and the condition factor of juveniles.

Table S4. Punctual measurements of dissolved oxygen concentration (in mg L–1) and water temperature (°C) during a dry event in control and dry channels.

Access here

Acknowledgements

EA was funded by a cross-border grant (Univ Pau & Pays Adour/Univ of the Basque Country) and the French Embassy's Mérimée program. Experiments were carried out using IE ECP facilities and financial support by Conseil Général 64. We are extremely grateful to students for their assistance with macroinvertebrate identification and Phillip Basterra for professional proofreading of the English text. The authors also thank the anonymous reviewers and Editors for their relevant and helpful comments that allowed to significantly improve this work.

Data availability statement

All files are available from the Zenodo database (doi https://doi.org/10.5281/zenodo.10842477).

References

  • Atkinson CL, Julian JP, Vaughn CC. 2014. Species and function lost: Role of drought in structuring stream communities. Biol Conserv 176: 30–38. [CrossRef] [Google Scholar]
  • Arevalo E, Larrañaga A, Bardonnet A. 2020. Comparison of food availability and performances of first-feeding alevins through spring with the occurrence of a flood. Ecol Freshw Fish 29: 693–704. [CrossRef] [Google Scholar]
  • Arevalo E. 2014. Impact du barrage de Lurberria sur l'efficacité des frayères de salmonidés de la Nivelle. Université de Pau et des Pays de l'Adour (UPPA), FRA. [Google Scholar]
  • Armstrong JD, Kemp PS, Kennedy GJA, Ladle M, Milner NJ. 2003. Habitat requirements of Atlantic salmon and brown trout in rivers and streams. Fish Res 62: 143–170. [CrossRef] [Google Scholar]
  • Baxter CV, Kennedy TA, Miller SW, Muehlbauer JD, Smock LA. 2017. Macroinvertebrate drift, adult insect emergence and oviposition, in Methods in Stream Ecology Ecosystem Structure, edited by F.R. Hauer, G.A. Lamberti. Vol. 1. Elsevier, Academic Press, pp. 435–456. [Google Scholar]
  • Bo T, Fenoglio S, Malacarne G, Pessino M, Sgariboldi F. 2007. Effects of clogging on stream macroinvertebrates: an experimental approach. Limnologica 37: 186–192. [CrossRef] [Google Scholar]
  • Bogan MT, Boersma KS, Lytle DA. 2014. Resistance and resilience of invertebrate communities to seasonal and supraseasonal drought in arid-land headwater streams. Freshw Biol 60: 2547–2558. [Google Scholar]
  • Boulton AJ, Peterson CG, Grimm NB, Fisher SG. 1992. Stability of an aquatic macroinvertebrate community in a multiyear hydrologic disturbance regime. Ecology 73: 2192–2207. [CrossRef] [Google Scholar]
  • Boulton AJA, Findlay S, Marmonier P, Stanley EH, Valett HM. 1998. The functional significance of the hyporheic zone in streams and rivers. Annu Rev Ecol Syst 29: 59–81. [CrossRef] [Google Scholar]
  • Boulton AJ. 2003. Parallels and contrasts in the effects of drought on stream macroinvertebrate assemblages. Freshw Biol 48: 1173–1185. [Google Scholar]
  • Chadd RP, England JA, Constable D, Dunbar MJ, Extence CA, Leeming DJ, Murray-Bligh JA, Wood PJ. 2017. An index to track the ecological effects of drought development and recovery on riverine invertebrate communities. Ecol Indic 82: 344–356. [CrossRef] [Google Scholar]
  • Clews E, Durance I, Vaughan IP, Ormerod SJ. 2010. Juvenile salmonid populations in a temperate river system track synoptic trends in climate. Glob Chang Biol 16: 3271–3283. [CrossRef] [Google Scholar]
  • Datry T, Bonada N, Boulton AJ. 2017. General introduction, in: Intermittent Rivers and Ephemeral Streams: Ecology and Management, edited by Datry T, Bonada N, Boulton AJ. Elservier, 2017, pp. 1–20. [Google Scholar]
  • Datry T. 2017. Ecological Effects of Flow Intermittence in Gravel-Bed Rivers. In Gravel-Bed Rivers. . John Wiley and Sons, Ltd. pp. 261–297. [Google Scholar]
  • Datry T, Bonada N, Boulton A. 2017. Intermittent Rivers and Ephemeral Streams, Ecology and Management. London: Elsevier, Academic Press. [Google Scholar]
  • Dayon G, Boé J, Martin É, Gailhard J. 2018. Impacts of climate change on the hydrological cycle over France and associated uncertainties. Comptes Rendus − Geosci 350: 141–153. [CrossRef] [Google Scholar]
  • Descloux S, Datry T, Marmonier P. 2013. Benthic and hyporheic invertebrate assemblages along a gradient of increasing streambed colmation by fine sediment. Aquat Sci 75: 493–507. [CrossRef] [Google Scholar]
  • Dewson ZS, James ABW, Death RG. 2007. A review of the consequences of decreased flow for instream habitat and macroinvertebrates. J North Am Benthol Soc 26: 401–415. [CrossRef] [Google Scholar]
  • Di Sabatino A, Coscieme L, Cristiano G. 2023. No post-drought recovery of the macroinvertebrate community after five months upon rewetting of an irregularly intermittent Apennine River (Aterno River). Ecohydrol Hydrobiol 23: 141–151. [CrossRef] [Google Scholar]
  • Dole-Olivier MJ. 2011. The hyporheic refuge hypothesis reconsidered: A review of hydrological aspects. Mar Freshw Res 62: 1281–1302. [CrossRef] [Google Scholar]
  • Doretto A, Bona F, Falasco E, Morandini D, Piano E, Fenoglio S. 2019. Stay with the flow: How macroinvertebrate communities recover during the rewetting phase in Alpine streams affected by an exceptional drought. River Res Appl 36: 91–101. [Google Scholar]
  • Downes BJ, Lake PS, Glaister A, Angus Webb J. 1998. Scales and frequencies of disturbances: Rock size, bed packing and variation among upland streams. Freshw Biol 40: 625–639. [CrossRef] [Google Scholar]
  • Elliott JM. 2000. Pools as refugia for brown trout during two summer droughts: trout responses to thermal and oxygen stress. J Fish Biol 56: 938–948. [CrossRef] [Google Scholar]
  • Elliott JMA, Hurley MA, Elliott JMA. 1997. Variable effects of droughts on the density of a sea-trout Salmo trutta population over 30 years. J Appl Ecol 34: 1229–1238. [CrossRef] [Google Scholar]
  • Fenoglio S, Bo T, Cucco M, Malacarne G. 2007. Response of benthic invertebrate assemblages to varying drought conditions in the Po river (NW Italy). Ital J Zool 74: 191–201. [CrossRef] [Google Scholar]
  • Fulton TW. 1911. In: The sovereignty of the sea: An historical account of the claims of England to the dominion of the British seas, and of the evolution of the territorial waters W. Blackwood, Edinburgh, London. p. 799. [Google Scholar]
  • Garcia de Leaniz C, Fraser N, Mikheev V, Huntingford F. 1994. Individual recognition of juvenile salmonids using melanophore patterns. J Fish Biol 45: 417–422. [CrossRef] [Google Scholar]
  • Gelman A, Hill J. 2007. Data analysis using regression and multilevel/hierarchical models. Cambridge University Press, 651 p. [Google Scholar]
  • Gelman A. 2006. Prior distributions for variance parameters in hierarchical models (Comment on Article by Browne and Draper). Bayesian Anal 1: 515–534. [MathSciNet] [Google Scholar]
  • Georgian T, Wallace BJ. 1983. Seasonal production dynamics in a guild of periphyton-grazing insects in a Southern Appalachian stream. Ecology 64: 1236–1248. [CrossRef] [Google Scholar]
  • Grant JW, Kramer DL. 1990. Territory size as a predictor of the upper limit to population density of juvenile salmonids in streams. Can J Fish Aquat Sci 47: 1724–1737. [CrossRef] [Google Scholar]
  • Hakala JP, Hartman KJ. 2004. Drought effect on stream morphology and brook trout (Salvenus fontinalis) populations in forested headwater streams. Hydrobiologia 515: 203–213. [CrossRef] [Google Scholar]
  • Harrison E. 2010. Fine sediment in rivers: scale of ecological outcomes. University of Canberra. [Google Scholar]
  • Harvey BC, White JL, Nakamoto RJ. 2005. Habitat-specific biomass, survival, and growth of rainbow trout (Oncorhynchus mykiss) during summer in a small coastal stream. Can J Fish Aquat Sci 62: 650–658. [CrossRef] [Google Scholar]
  • Huhta A, Muotka T, Tikkanen P. 2000. Nocturnal drift of mayfly nymphs as a post-contact antipredator mechanism. Freshwater biology 45: 33–42. [Google Scholar]
  • Hynes HBN. 1970. The Ecology of Running Waters. Liverpool: Liverpool University Press. [Google Scholar]
  • Johansson A. 1991. Caddis larvae cases (Trichoptera, Limnephilidae) as anti-predatory devices against brown trout and sculpin. Hydrobiologia 211: 185–194. [CrossRef] [Google Scholar]
  • Jones JI, Murphy JF, Collins AL, Sear DA, Naden PS, Armitage PD. 2012. The impact of fine sediment on macro-invertebrates. River Res Appl 28: 1055–1071. [CrossRef] [Google Scholar]
  • Kahler TH, Roni P, Quinn TP. 2011. Summer movement and growth of juvenile anadromous salmonids in small western Washington streams. Can J Fish Aquat Sci 58: 1947–1956. [Google Scholar]
  • Keeley ER, Grant JWA. 1997. Allometry of diet selectivity in juvenile Atlantic salmon (Salmo salar). Can J Fish Aquat Sci 54: 1894–1902. [CrossRef] [Google Scholar]
  • Lake PS. 2003. Drought and aquatic ecosystems: an introduction. Freshw Biol 48: 1141–1146. [CrossRef] [Google Scholar]
  • Lennox RJ, Crook DA, Moyle PB, Struthers DP, Cooke SJ. 2019. Toward a better understanding of freshwater fish responses to an increasingly drought-stricken world. Rev Fish Biol Fish 29: 71–92. [CrossRef] [Google Scholar]
  • Magalhães MF, Beja P, Canas C, Collares-Pereira MJ. 2002. Functional heterogeneity of dry-season fish refugia across a Mediterranean catchment: The role of habitat and predation. Freshw Biol 47: 1919–1934. [CrossRef] [Google Scholar]
  • Magalhães MF, Beja P, Schlosser IJ, Collares-Pereira MJ. 2007. Effects of multi-year droughts on fish assemblages of seasonally drying Mediterranean streams. Freshw Biol 52: 1494–1510. [Google Scholar]
  • Maridet L, Wasson J-G, Philippe M. 1992. Vertical distribution of fauna in the bed sediment of three running water sites: influence of physical and trophic factors. Regul Rivers Res Manag 7: 45–55. [CrossRef] [Google Scholar]
  • Mathers KL, Hill MJ, Wood CD, Wood PJ. 2019. The role of fine sediment characteristics and body size on the vertical movement of a freshwater amphipod. Freshw Biol 64: 152–163. [CrossRef] [Google Scholar]
  • Mathers KL, Millett J, Robertson AL, Stubbington R, Wood PJ. 2014. Faunal response to benthic and hyporheic sedimentation varies with direction of vertical hydrological exchange. Freshw Biol 59: 2278–2289. [CrossRef] [Google Scholar]
  • Mathers KL, Rice SP, Wood PJ. 2017. Temporal effects of enhanced fine sediment loading on macroinvertebrate community structure and functional traits. Sci Total Environ 599–600: 513–522. [CrossRef] [PubMed] [Google Scholar]
  • Mathers KL, Wood PJ. 2016. Fine sediment deposition and interstitial flow effects on macroinvertebrate community composition within riffle heads and tails. Hydrobiologia 776: 147–160. [CrossRef] [Google Scholar]
  • Messager ML, Lehner B, Cockburn C, Lamouroux N, Pella H, Snelder T, Tockner K, Trautmann T, Watt C, Datry T. 2021. Global prevalence of non-perennial rivers and streams. Nature 594: 391–397. [CrossRef] [PubMed] [Google Scholar]
  • Milner AM, Picken JL, Klaar MJ, Robertson AL, Clitherow LR, Eagle L, Brown LE. 2018. River ecosystem resilience to extreme flood events. Int J Bus Innov Res 17: 8354–8363. [Google Scholar]
  • Milner VS, Jones JI, Maddock IP, Bunting GC. 2022. The hyporheic zone as an invertebrate refuge during a fine sediment disturbance event. Ecohydrology 15: 1–14. [Google Scholar]
  • Mureithi PW, Mbaka JG, M'Erimba CM, Mathooko JM, 2018. Effect of drift sampler exposure time and net mesh size on invertebrate drift density in the Njoro River, Kenya. Afr J Aquat Sci 43: 163–168. [CrossRef] [Google Scholar]
  • Muotka T, Huhta A, Tikkanen P. 1999. Diel vertical movements by lotic mayfly nymphs under variable predation risk. Ecological Entomology 24: 443–449. [CrossRef] [Google Scholar]
  • Naiman RJ, Latterell JJ, Pettit NE, Olden JD. 2008. Flow variability and the biophysical vitality of river systems. Comptes Rendus − Geosci 340: 629–643. [CrossRef] [Google Scholar]
  • Naman SM, Rosenfeld JS, Richardson JS. 2016. Causes and consequences of invertebrate drift in running waters: from individuals to populations and trophic fluxes. Can J Fish Aquat Sci 73: 1292–1305. [CrossRef] [Google Scholar]
  • Nislow KH, Sepulveda AJ, Folt CL. 2004. Mechanistic linkage of hydrologic regime to summer growth of age-0 Atlantic salmon. Trans Am Fish Soc 133: 79–88. [CrossRef] [Google Scholar]
  • Otto C, Svensson BS. 1980. The significance of case material selection for the survival of caddis larvae. J Anim Ecol 855–865. [CrossRef] [Google Scholar]
  • Palmer M, Lettenmaier DP, Poff NL, Postel SL, Richter B, Warner R. 2009. Climate change and river ecosystems: protection and adaptation options. Environ Manag 44: 1053–1068. [CrossRef] [PubMed] [Google Scholar]
  • Pařil P, Polášek M, Loskotová B, Straka M, Crabot J, Datry T. 2019. An unexpected source of invertebrate community recovery in intermittent streams from a humid continental climate. Freshw Biol 64: 1971–1983. [CrossRef] [Google Scholar]
  • Pickett ST, White PS. 1985. The ecology of natural disturbance and patch dynamics. Elsevier. [Google Scholar]
  • Rashidabadi F, Rosenfeld JS, Abdoli A, Naman SM, Nicolas A. 2022. Seasonal changes in invertebrate drift: effects of declining summer flows on prey abundance for drift-feeding fishes. Hydrobiologia 849: 1855–1869. [CrossRef] [Google Scholar]
  • Reid AJ, Carlson AK, Creed IF, Eliason EJ, Gell PA, Johnson PTJ, Kidd KA, MacCormack TJ, Olden JD, Ormerod SJ, Smol JP, Taylor WW, Tockner K, Vermaire JC, Dudgeon D, Cooke SJ. 2019. Emerging threats and persistent conservation challenges for freshwater biodiversity. Biol Rev 94: 849–873. [CrossRef] [PubMed] [Google Scholar]
  • Relyea CD, Minshall GW, Danehy RJ, 2000. Stream insects as bioindicators of fine sediment. Proc Water Environ Federat 2000: 663–686. [CrossRef] [Google Scholar]
  • Resh VH, Brown AV, Covich AP, Gurtz ME, Li HW, Minshall GW, Reice SR, Sheldon AL, Wallace JB, Wissmar RC. 1988. The role of disturbance in stream ecology. J North Am Benthol Soc 7: 433–455. [CrossRef] [Google Scholar]
  • Rosenfeld JS, Boss S. 2001. Fitness consequences of habitat use for juvenile cutthroat trout: energetic costs and benefits in pools and riffles. Can J Fish Aquat Sci 58: 585–593. [CrossRef] [Google Scholar]
  • Sánchez-Hernández J, Servia MJ, Vieira-Lanero R, Cobo F. 2012. Ontogenetic dietary shifts in a predatory freshwater fish species: the brown trout as an example of a dynamic fish species. New Adv Contrib to Fish Biol 271–298. [Google Scholar]
  • Sarremejane R, Cañedo-Argüelles M, Prat N, Mykrä, H, Muotka T, Bonada N, 2017. Do metacommunities vary through time? Intermittent rivers as model systems. J Biogeogr 44: 2752–2763. [Google Scholar]
  • Schmid PE, Schmid-Araya JM. 2010. Scale-dependent relations between bacteria, organic matter and invertebrates in a headwater stream. Fundam Appl Limnol 176: 365–375. [CrossRef] [Google Scholar]
  • Shearer KA, Stark JD, Hayes JW, Young RG. 2003. Relationships between drifting and benthic invertebrates in three New Zealand rivers: implications for drift-feeding fish. New Zeal J Mar Freshw Res 37: 809–820. [CrossRef] [Google Scholar]
  • Skoulikidis NT, Sabater S, Datry T, Morais MM, Buffagni A, Dorflinger G, Zogaris S, Del Mar Sanchez-Montoya M, Bonada N, Kalogianni E, Rosado J, Vardakas L, De Girolamo AM, Tockner K. 2017. Non-perennial Mediterranean rivers in Europe: status, pressures, and challenges for research and management. Sci Total Environ 577: 1–18. [CrossRef] [PubMed] [Google Scholar]
  • Slack KV, Tilley LJ, Kennelly SS, 1991. Mesh-size effects on drift sample composition as determined with a triple net sampler. Hydrobiologia 209: 215–226. [CrossRef] [Google Scholar]
  • Strayer DL, May SE, Nielsen P, Wollheim W, Hausam S. 1997. Oxygen, organic matter, and sediment granulometry as controls on hyporheic animal communities. Arch fur Hydrobiol 140: 131–144. [CrossRef] [Google Scholar]
  • Stubbington R. 2012. The hyporheic zone as an invertebrate refuge: a review of variability in space, time, taxa and behaviour. Mar Freshw Res 63: 293–311. [CrossRef] [Google Scholar]
  • Sweka JA, Hartman KJ. 2008. Contribution of terrestrial invertebrates to yearly brook trout prey consumption and growth. Trans Am Fish Soc 137: 224–235. [CrossRef] [Google Scholar]
  • Trenberth KE, Dai A, Van Der Schrier F G, Jones PD, Barichivich J, Briffa KR, Sheffield J. 2014. Global warming and changes in drought. Nat Clim Chang 4: 17–22. [CrossRef] [Google Scholar]
  • Vadher AN, Millett J, Wood PJ. 2018. Direct observations of the effect of fine sediment deposition on the vertical movement of Gammarus pulex (Amphipoda: Gammaridae) during substratum drying. Hydrobiologia 815: 73–82. [CrossRef] [Google Scholar]
  • Vadher AN, Stubbington R, Wood PJ. 2015. Fine sediment reduces vertical migrations of Gammarus pulex (Crustacea: Amphipoda) in response to surface water loss. Hydrobiologia 753: 61–71. [CrossRef] [Google Scholar]
  • Vander Vorste F R, S Sarremejane R, Datry T. 2020. Intermittent rivers and ephemeral streams: a unique biome with important contributions to biodiversity and ecosystem services. Encyclopedia of the World's biomes, 419–429. [CrossRef] [Google Scholar]
  • Vignes JC, Heland M. 1995. Comportement alimentaire au cours du changement d'habitat lié à l'émergence chez le saumon atlantique Salmo salar L. et la truite commune Salmo trutta L., en conditions semi-naturelles. Bull Français la Pêche la Piscic 207–214. [CrossRef] [EDP Sciences] [Google Scholar]
  • Walsh CL, Kilsby CG. 2007. Implications of climate change on flow regime affecting Atlantic salmon. Hydrol Earth Syst Sci Discuss 11: 1127–1143. [CrossRef] [Google Scholar]
  • Williams DD, Hynes HBN. 1974. The occurrence of benthos deep in the substratum of a stream. Freshw Biol 4: 233–256. [CrossRef] [Google Scholar]
  • Winkowski JJ, Zimmerman MS. 2017. Summer habitat and movements of juvenile salmonids in a coastal river of Washington State. Ecol Freshw Fish 27: 255–269. [Google Scholar]
  • Wood P, Armitage PD. 1997. Biological effects of fine sediment in the lotic environment. Environ Manage 21: 203–217. [CrossRef] [PubMed] [Google Scholar]
  • Wood PJ, Boulton AJ, Little S, Stubbington R. 2010. Is the hyporheic zone a refugium for aquatic macroinvertebrates during severe low flow conditions? Fundam Appl Limnol / Arch für Hydrobiol 176: 377–390. [CrossRef] [Google Scholar]
  • Zabel R, Scheuerell M, McClure M, Williams J. 2006. The interplay between climate variability and density dependence in the population viability of Chinook salmon. Conserv Biol 20: 190–200. [CrossRef] [PubMed] [Google Scholar]
  • Zweig LD, Rabeni CF, 2001. Biomonitoring for deposited sediment using benthic invertebrates: a test on 4 Missouri streams. J North Am Bentholog Soc 20: 643–657. [CrossRef] [Google Scholar]

Cite this article as: Arevalo E, Bardonnet A, Glise S, Gueraud F, Huchet E, Lange F, Rives J, Larrañaga A. 2024. A permeable hyporheic zone may contribute to buffer the effects of a drying event on prey availability for salmonid juveniles. Knowl. Manag. Aquat. Ecosyst., 425. 15

All Tables

Table 1

Probability of a different abundance between control and dry channels (α1, α2 and α3), the 3 periods (γ1 vs. γ2 and γ2 vs. γ3) and the upstream channel sections without fish and the downstream sections with fish (δ). The parameter influence on invertebrate abundance is symbolised by “▲” when the abundance increased significantly (significance probability ≥ 0.90; in bold), “=” when it did not affect the abundance (0.10 < significance probability < 0.90) or “▼” when the abundance decreased significantly (significance probability ≤ 0.10; in bold).

All Figures

thumbnail Fig. 1

Experimental design. Water levels were dropped onto the surface of the sediment in some channels (A, left channel), while 4.5 cm of water was maintained in others (A, right channel). (B) Channels were naturally fed by the Aitzeguerria and Lapitxuri brooks. After 48 days of colonisation by macroinvertebrates 100 μm mesh nets were placed at the beginning and end of the channels to study the dynamics of the community in place. A drying event was simulated for 15 days (grey channels) and then fish grew for 21 days (hatched areas).

In the text
thumbnail Fig. 2

Abundance of macroinvertebrates (mean number of individuals by corer) before the drying event, after 2 weeks of drying and after 3 weeks of fish growth in reference discharge conditions in the control and dry channels, shown in white and grey respectively. Each condition represents 9 samples, for a total of 54 samples. Error bars represent standard errors.

In the text
thumbnail Fig. 3

Model estimates of the performance of juvenile trout over the 3-week growth period in the control and dry channels. Performance includes survival probability (left), growth (middle) and condition factors of juvenile trout (right). Boxplots indicate the percentiles 1, 25, 50, 75 and 99 of the posterior distributions.

In the text

Current usage metrics show cumulative count of Article Views (full-text article views including HTML views, PDF and ePub downloads, according to the available data) and Abstracts Views on Vision4Press platform.

Data correspond to usage on the plateform after 2015. The current usage metrics is available 48-96 hours after online publication and is updated daily on week days.

Initial download of the metrics may take a while.