Hydropower is one of the most controversial forms of renewable energy with some authors questioning its environmental compatibility when balancing its “red” and “green” aspects (e.g., Seliger et al., 2016; Geist, 2021). Hydropower can affect aquatic ecosystems and aquatic species in many ways, including impaired habitat quality by the implementation and operation of the power plant, water diversions and abstractions needed for energy production, and direct injuries or mortality of aquatic animals, particularly fish (Anderson et al., 2006; Zarfl et al., 2015; Reid et al., 2019; Algera et al., 2020). Therefore, hydropower often conflicts with the international and national legislation such as the European Water Framework Directive (European Parliament, 2000) and the Habitats Directive (European Parliament, 1992), targeting the protection of highly endangered species and their habitats.

Great challenges arise when it comes to restoring habitat quality in rivers affected by hydropower production, due to the many restrictions and constraints for such action (Pander and Geist, 2018). In free-flowing rivers unaffected by hydropower exploitation, naturally fluctuating discharge, strongly changing water levels, and flow velocities permanently create areas of new habitat. Such succession is an important part of natural river functioning with evolutionary adaptions of all riverine fish to these patterns (Postel and Richter, 2012). These discharge regimes and the associated dynamic formation of diverse habitat conditions are key for the completion of the life cycles of specialized riverine fish. Unfortunately, such dynamic conditions in natural rivers are contrasting the need for a ground-based hydropower exploitation, and many natural rivers were therefore transformed into heavily modified water bodies to control their hydromorphological dynamics. In such systems, alteration of flow and sediment regimes, straightened and rip-rap protected riverbanks, as well as the disconnection with the floodplain have led to a severe loss of critical habitats for spawning and juvenile development of riverine fish species as well as other ecological and socioeconomic impacts (Bunn and Arthington, 2002; Tockner and Stanford, 2002; Aarts et al., 2004; Auerswald et al., 2019). In this context, tributaries, side channels, and fish passes got more and more into the focus of conservation practitioners to restore aquatic habitats beyond the main channel (Pander et al., 2017), with the core intention to promote population development in declining aquatic species (Moyle and Leidy, 1992; Mueller et al., 2020). However, restoration in these waters is also often more restricted than in the main channel, often being limited to small-scale instream restoration measures (Geist and Hawkins, 2016). These measures can comprise the creation of shallow bank habitats (Lorenz et al., 2013; Pander et al., 2017; Pander et al., 2021), the introduction of gravel and boulders (Barlaup et al., 2008; Pander et al., 2015), or the placement of large woody debris (Gurnell et al., 2005; Entrekin et al., 2008). It is known that these restoration measures can create valuable habitats for many aquatic species and harbor large fish diversity, often comprising relict populations of highly protected fish and macroinvertebrate species (Pander and Geist 2016; Pander et al., 2017; Meulenbroek et al., 2018; Pander et al., 2021; Nagel et al., 2022). However, tributaries, side channels, and fish passes are often also limited by hydropower-induced morphological constraints such as damming, bank protection, or maintenance-induced management (Pander et al., 2022, in press). For example, it is a common practice to completely or partly dewater side channels for maintenance reasons to overhaul turbines or to clean the channels from dumped rubbish, typically every year or every few years. This practice of reducing the water level can result in drying out of bank habitats. Since species communities in bank habitats can be particularly prone to swift water level fluctuations (Jepsen, 1997; Pander et al., 2021), it is most likely that aquatic organisms in these habitats are heavily affected by this management practice. This holds even more true when bank habitats have been previously restored with the intention to support fish population development. Despite the high ecological value of restored river banks, especially as nursery habitats for rheophilic fish species, less is known about the direct and indirect effects of regular maintenance such as the dewatering scenarios in practice. This concerns both the direct mortality of fish due to stranding and the indirect effects arising from the disturbance due to the limited accessibility and availability of shallow bank zones that are crucial habitats for the early life history of riverine fish species. In addition, dewatering of channels may also affect other aquatic organisms such as molluscs (Curley et al., 2022), crayfish, or other invertebrates (Muehlbauer et al., 2011). To date, dewatering events and fast changing water levels are only assessed regarding hydropeaking—an operational tool of hydropower production to feed peaks of energy demand and maximize energy production revenue (Venus et al., 2020). Hydropeaking can cause sudden changes in flow and is known to result in the degradation of river habitat quality (Bruder et al., 2016; Greimel et al., 2018), the alteration of stream thermal regimes (Casas-Mulet et al., 2016; Choi and Choi, 2018), the dewatering of spawning grounds (Grabowski and Isely, 2007; Casas-Mulet et al., 2015), and the stranding of fish (Puffer et al., 2015; Schmutz et al., 2015; Auer et al., 2017; Bartoň et al., 2021). Since hydropeaking-induced water level fluctuations strongly differ from dewatering during the maintenance in magnitude and frequency, such results are of limited transferability to channel dewatering for maintenance reasons that typically last several days or even months and can include much greater water level amplitudes (Auer et al., 2017).

This study aimed at understanding how the dewatering of channels as a measure of hydropower-related maintenance affects fish and their habitats, with a focus on restored shallow bank habitats. We hypothesized that 1) the common practice of quick dewatering of shallow bank habitats results in fish stranding and mortality, with smaller and less mobile species being the most strongly affected, and that 2) dewatering can subsequently affect fish community composition, particularly during the period of spawning, larval drift, and juvenile development of sensitive riverine fish species.

Management and restoration of aquatic ecosystems and aquatic species communities promise the greatest success when they are evidence-based (Palmer et al., 2005; Kondolf et al., 2007; Pander and Geist 2013; Geist 2015). This is particularly true for the mitigation of well-known stressors such as hydropower generation, where many of its effects on aquatic ecosystems have been intensively studied. However, even in well-known stressor-effect chains, filling of knowledge gaps such as those presented in this study can help on the way from ecological understanding into policy-making (Palmer et al., 2005; Geist and Hawkins 2016). The findings of the present study are of particular importance for an improved and less ecologically harmful maintenance management of tributaries in the hydropower-affected rivers. This is particularly important since it is well known that heavily modified water bodies and channels for hydropower generation such as the HB can greatly contribute to overall fish diversity (De Leeuw et al., 2007; Pander and Geist, 2010; Davis and Moore 2016). Such rivers do not only harbor many generalist species but also provide habitat for many highly specialized species (e.g., Pander and Geist, 2018), such as B. barbus or C. gobio which are under strict protection by national and international legislation (European Parliament 1992; Bohl et al., 2003). Legislation regarding the protection of species or habitats is strict if it is framed on European basis or transformed into national laws by the member states. For instance, it is strictly forbidden by law to kill protected animals or to destroy habitats of those by human action (European Parliament 1992). The dewatering of the HB caused both direct mortality of fish by stranding and the inaccessibility of important shallow bank habitats during the low water phase. Such areas can be crucial to fulfil critical life stages in the life cycle of many riverine fish, and with a decrease in the availability of these habitats, population effects are a logical consequence (Pander et al., 2017; Hayes et al., 2019b; Moreira et al., 2019; Stoffers et al., 2022).

The results of our study provide evidence that the dewatering of a channel for hydropower generation had multiple effects. First, it affected single individuals of fish directly by their stranding as well as by their habitats falling dry. The observed effects extended over a certain period since the dewatering also affected fish community composition months after its conductance. This was particularly evident in strong abundance declines or even complete absence of species with high conservation value such as C. nasus, B. barbus, and A. bipunctatus in samplings following the dewatering event. The intensity of fish stranding was dependent on the time of the year and dewatering speed, with more stranded fish being detected during the first very fast dewatering event in August 2016 compared to the second event in 2021. In contrast to fish stranding, community effects caused by dewatering were rather strong, most likely due to the restricted habitat availability in 2021 when the dewatering phase lasted several weeks from the middle of May to the beginning of August. This time of the year falls within the main larval drift phase of riverine fish at the River Inn (Nagel et al., 2021), and these types of bank habitats usually serve as a crucial refuge area for the development and growth of fish larvae and young-of-the-year fish (Pander et al., 2017; Stoffers et al., 2022).

As evident from the comparison of relative abundances, the risk for stranding was very different among species. Small and least mobile specimens of C. gobio, a bottom-oriented, cold stenothermic fish without a swim bladder (Utzinger et al., 1998), had an overproportioned high risk of stranding compared to the larger and open-water-oriented species such as C. nasus. Due to the timely sampling in this study, effects of predation on the stranded fish by piscivorous birds or mammals can be widely excluded which could be expected otherwise. In combination with the fact that macrophytes were slumped into dense pads making it difficult to find stranded fish (Bell et al., 2008), our data may actually underestimate real mortality associated with such dewatering events. Furthermore, in isolated puddles with no connectivity to the main channel, it is possible that a successively progressing oxygen depletion (Branco et al., 2016; Hayes et al., 2019a) would have caused much higher mortality rates than detected in the timely sampling after the dewatering in this study. It needs to be noted that partial dewatering of bank habitats can also be a natural process in free-flowing rivers not affected by hydropower. In natural rivers, this effect can contribute to the lateral nutrient exchange. However, in a natural environment, the dewatering of bank habitats is often much slower and only parts of the bank habitats are affected with the possibility for fish to move to equivalent substitute areas. This is particularly the case when the bank angles are flat and the dropping speed of the water is low. Under such conditions, a clear cutoff between bank habitats and the remaining main channel is almost impossible. The lowering of the water level for the maintenance of the hydropower system is usually realized within a very short time span of several hours. In this study, the dewatering speed did not allow all the individuals to escape into other areas or the main stem. Channels are often built as monotonous river courses with steep bank angles not comprising shallow habitats for fish as refuge as would be required. When small fish or fish larvae that are known as weak swimmers (Lechner, et al., 2014) are drifted into the main stem with faster current, they are more likely to get dislocated downstream without the potential to migrate back (Baras and Lucas, 2001; Pavlov et al., 2019). This can lead to shifts in population recruitment patterns with a particular decline of sensitive rheophilic species (Zauner and Eberstaller, 1999) as detected in this study for C. nasus, B. barbus, and A. bipunctatus in the HB. In contrast, S. cephalus, a generalist species of the HB, was able to cope better with the dewatering, potentially due to its ability to spawn several times throughout spring and summer. It can also use the bank habitats for the larval development after the dewatering ended. Species that are known to have a very limited time phase for spawning are much more susceptible to population effects when the dewatering of habitats fall within their spawning season, which may ultimately result in a loss of an entire generation. In addition, shallow areas rich of fine sediment and organic matter, like in the assessed bank habitats, can harbor highly specialized species such as lampreys or freshwater mussels. These species have a distinct life cycle that can comprise a very long phase in a very restricted habitat area. Lampreys such as Eudontomyzon sp. are known to have a sediment phase lasting five to seven years before they eventually emerge, transform to adults, and spawn (e.g., Dawson et al., 2015). In this case, population effects can even affect multiple age classes if these individuals are not able to find other substitute habitats during the dewatering phase.

The evidence about the negative ecological effects detected herein can be transferred into the policy to reduce the negative effects of dewatering to a minimum, and to contribute to the fulfilment of national and international legal requirements. At first, the usefulness of this measure should be seriously checked. If not avoidable, it should be accompanied by monitoring to determine the least harmful time window and duration for such maintenance measures. Generally, if the dewatering is not avoidable, the draining period should be kept as short as possible under the consideration of potential negative effects due to the stranding of fish and subsequently potential negative population effects due to the loss of habitat for critical life stages. To avoid stranding of fish larvae and fish, the dewatering should be carried out slowly over a time span of several days as also suggested in the study by Hayes et al. (2019a). Several general principles of flow management in relation to the reshaped channel morphologies have already been proposed, for example, in the “river in river concept” (Forseth and Harby, 2014). However, it would be very advantageous if the construction design of bank habitats prevented fish or fish larvae from being caught in the remaining puddles (Figure 8). In case fish are stranded, they should be collected and transferred into habitats where their survival is most likely. The dewatering should allow fish larvae and subsequent life stages (particularly weak swimmers) as well as lampreys and invertebrates (e.g., mollusks) to reach adequate substitute habitats during the dewatering phase. These habitats should ensure survival of these species until the dewatering ends and should allow them to migrate back. To avoid negative population effects, the dewatering should be applied outside the seasons of spawning, larval drift, and juvenile development. In this study, the main season for larval drift was spring to summer and the most dense fish population in bank habitats was detected in summer to autumn, matching previous findings of the River Inn system (Nagel et al., 2021). Consequently, the least harmful time span for dewatering in this project area would be late autumn to early spring when least fish are present in the bank habitats. However, it has to be noted that fish in overwintering habitats can also be negatively affected if those get disconnected from the main stem or fall dry. If the implementation of the abovementioned suggestions is impossible due to the many restrictions in channels and heavily modified water bodies, the responsible parties of the dewatering should compensate for the loss of species and negative effects on populations. In this context, a monitoring as mentioned previously could be used to assess the direct and indirect effects on fish and the fish community, and to determine adequate compensation. This can be achieved by a monetary habitat compensation (Venus et al., 2020) or by stocking fish eggs, larvae, or fish from the genetically adapted broodstock (Gum et al., 2009).