Global and European regional assessment: Vulnerable (VU)
EU 27 regional assessment: Endangered (EN)

The Danube Salmon (Hucho hucho) is endemic to the Danube River system in Central and Southeastern Europe, where it largely occurs in tributaries draining the Alps, the Dinaric Alps, and the Carpathian mountain ranges, in addition to the Bohemian-Moravian Highlands, the Bohemian Forest and the Weinsberger Wald plateau. The extant subpopulations are confined to a series of discontinuous river reaches comprising c. 30% of its original range, the majority of which was lost prior to the turn of the century.

Although the population size is understood to be suffering a continued reduction (number of mature individuals), the current rate of decline has not been quantified at the range-wide scale. However, it is suspected that a future decline will meet the threshold for Vulnerable under Criterion A (≥ 30% within the next 24 years = three generation lengths) based on the observed continued decline in some parts of its range (e.g., Austria, Germany, Croatia and Serbia), plus a series of identified threats which include a projected loss of habitat quality (e.g., Austria, Slovenia and the Western Balkans), actual levels of exploitation (e.g., the Western Balkans, Romania and Ukraine) and pollution (e.g., Bosnia and Herzegovina, Slovakia, and Romania). In the EU 27 member states, it is suspected that this future decline will meet the threshold for Endangered (≥ 50% within the next 24 years).

This species was previously assessed as Endangered (B2ab(ii,iii)) globally, but current knowledge indicates that its range does not approach the thresholds required to qualify for a threatened category under Criterion B (extent of occurrence (EOO) <20,000 km², area of occupancy (AOO) <2,000 km²) or D2. The population size is understood to exceed 10,000 mature individuals, and thus does not approach the thresholds for criteria C or D. There exists no comprehensive quantitative analysis which would permit the application of criterion E.

Therefore, the Danube Salmon is assessed as Vulnerable under Criterion A (A3cde), both globally and for Europe, and Endangered under Criterion A (A3cde) for the EU 27 member states.

This species is endemic to the Danube River system in Central and Southeastern Europe, where it primarily inhabits tributaries draining the Alps, Dinaric Alps, and Carpathian mountain ranges, in addition to the Bohemian-Moravian Highlands, the Bohemian Forest and the Weinsberger Wald plateau (Holčík et al. 1988).

The extant subpopulations are confined to a series of fragmented river reaches comprising an estimated 2,853 km of linear length, and representing c. 30% of its historic distribution (see 'Population' and 'Threats'). Most of them inhabit stretches measuring 5–75 km in length, with only seven subpopulations residing in uninterrupted reaches of >100 km (Freyhof et al. 2015).

In Germany, its native distribution extends from the upper Danube in the state of Baden-Württemberg to the Ilz and Inn rivers at the border with Austria. It currently occupies around half of this range (but see 'Population'), including parts of the Iller, Lech, Regen, Isar, Ilz and Inn rivers plus a few short stretches of the Danube main stem (Hanfland et al. 2015, Freyhof et al. 2023; Schmutz et al. 2023).

In Austria, it remains present in the Inn, Traun, Enns, Ybbs and Pielach rivers, plus some associated stretches of  the Danube main stem (but see 'Population'). South of the Alps, it also occurs in the Austrian portion of the Drava River watershed, where the most notable subpopulations inhabit the Mur and Gail tributary systems. Historic records indicate that its range in the Inn River might have sporadically extended upstream to Switzerland, but this has never been unequivocally confirmed (Ratschan 2014, Pinter et al. 2024).

It was extirpated from its native range in the Czechia (Morava River) and Poland (Czadeczka and Czarna Orawa rivers, both of which belong to the Váh River system) during the late 19^th and mid-20^th centuries, respectively (Hanel et al. 2013, Witkowski et al. 2013, Ihuț et al. 2014).

In Slovakia, it occurs in parts of the Váh and Hron rivers, while in Hungary the most recent records suggest that it is currently confined to a short stretch of the upper Tisza River at the border with Ukraine, having formerly occurred in the Danube main stem and Drava River (Sallai and Kontos 2005, Harka 2006).

In Slovenia, Croatia and the Western Balkans, most extant subpopulations are located within the Sava River system, with only remnants remaining in the upper Drava River (Slovenia) and upper Ibar River in the Great Morava watershed (Serbia). In the Sava, the largest subpopulations inhabit the upper-middle part of the system in Slovenia, plus the Kupa (sl. Kolpa) River in Slovenia and Croatia, the Una River watershed in Croatia and Bosnia and Herzegovina and the Drina River system in Bosnia and Herzegovina, Montenegro and Serbia. In Bosnia and Herzegovina, it is also present in the Vrbas and Bosna rivers (Freyhof et al. 2015, Pliberšek and Tavčar 2024, Schöffmann and Marić 2024).

In Romania, it was formerly present in most left-bank tributaries of the Danube River downstream to the Prut River, but is currently restricted to the Tisza River system, including the upper Mureș River, the Vişeu River, and the stretch of the Tisza main stem that comprises the border with Ukraine, plus the Bistrița River in the Siret River watershed. There is historic evidence indicates that it once occurred in the Danube main stem to the Đerdap (Iron Gate) gorge at the border with Serbia (Holčík 1990, Bănăduc et al. 2013, Witkowski et al. 2013, Ihuț et al. 2014, Năstase and Oţel 2017).

In Ukraine, it remains extant throughout the main stem of the upper Tisza River plus the lower and middle reaches of some of its tributaries, comprising the Rika, Tereblya, Teresva, Shopurka and Black Tisza rivers. It is also present in the upper Prut River downstream to the city of Chernivtsi, including the Cheremosh tributary system (Witkowski et al. 2013, Ihuț et al. 2014, Didenko et al. 2018, Roman et al. 2020).
 
It has been introduced to parts of the Danube system outside of its native range (e.g., the Thaya (cs. Dyje) River in Czechia, the Sajó River in Slovakia, the Grza, Nišava and Mlava rivers in Serbia, the Latorica River in Ukraine), in addition to affluents of the upper Dniester River (Limnytsia, Bystrytsia, and possibly the Zolota Lypa rivers) in Ukraine, some headwater tributaries of the Elbe, Odra and Vistula rivers in Czechia and Poland, Lake Constance in Germany, the Tormes River (an affluent of the Douro (es. Duero) River) in Spain, the Usses River (Rhône River system) in France, and possibly elsewhere (Hanel et al. 2013, Didenko et al. 2018, Afanasyev et al. 2022, Sokolović et al. 2024). These non-native subpopulations are excluded from this assessment, since they were translocated for fisheries rather than conservation purposes.

As a large-bodied obligate predator (see 'Habitat and Ecology'), this species' abundance is naturally low compared with most sympatric fish species (Ratschan 2014, Snoj et al. 2022).

Although it has been extirpated from at least two-thirds of its native range since the early 19^th century, the overall population size is still understood to exceed the minimum threshold for Red List criteria (<10,000 mature individuals) based on a best estimate derived from the consensus view of multiple experts. The current range-wide population trend has not been explicitly quantified, and the total number of extant subpopulations is unclear. Its overall status is described as "Unfavourable-Inadequate" according to the most recent (2013–2018) round of reporting for species covered by Article 17 of the European Union Habitats Directive, while continued population size reductions of > 30% are estimated to have occurred over the past three generations (c. 24 years) in Austria, Germany, Bosnia and Herzegovina and Serbia. Furthermore, it has been estimated that up to 35–40% of its range in southeastern Europe (Slovenia, Croatia and the Western Balkans) could potentially be lost should an extensive series of planned hydroelectric projects be completed (Freyhof et al. 2015, 2020, Pinter et al. 2024, Simić et al. 2024).

Although the overall pattern of declining abundance appears to have slowed during the past 30–40 years, this has been partially driven by intensive management which largely comprises supportive breeding and restocking efforts. For example, in Germany, the majority of extant subpopulations are not considered to be self-sustaining due to their dependence on supplementation, with wild (as defined by Red List Guidelines) subpopulations present only in the Lech (Wertach River), Isar (Ammer River and middle Isar) and Ilz watersheds, which together comprise c. 10% of its original range. Its presence in Austria is also largely maintained through management activities, but viable wild subpopulations occupy short stretches of the Enns and Ybbs rivers, most of the Pielach and Gail systems, and parts of the middle and upper Mur watershed. Its decline has been comparatively moderate in Slovenia, Croatia and the Western Balkans, albeit only 16 of the 43 identified subpopulations are considered to be wild, while 15 are thought to be decreasing and only three demonstrate an increasing trend. In Ukraine, the extant subpopulations are treated as wild although stocking has taken place in the Prut River system since 2017 and the upper Tisza River since 2022 (Hanfland et al. 2015, Freyhof et al. 2015, Didenko et al. 2018, Snoj et al. 2022, Pinter et al. 2024).

Elsewhere, viability analyses of individual subpopulations are largely lacking, and in some cases, accurate appraisals are likely to be hampered by both stocking activities and incomplete knowledge regarding the genetic composition of the broodstock. Moreover, there is no comprehensive information on the extent of historical trade of eggs, e.g., between Germany, Austria, the Czech Republic (Czechia) and Slovenia, or past stocking activities. There are no legal restrictions to such translocations in place throughout the majority of its range, and these activities are thus likely to continue since management units have not yet been clearly defined (Geist et al. 2009, Snoj et al. 2022).

Nevertheless, an overall population size reduction of  ≥ 30% (≥ 50% in the EU 27 member states) is suspected to be met within the next three generations, based on the observed continued decline in some parts of its range (e.g., Austria, Germany, Croatia and Serbia), plus a series of identified threats which include a projected loss of habitat quality (e.g., Austria, Slovenia and the Western Balkans), actual levels of exploitation (e.g., the Western Balkans, Romania and Ukraine) and pollution (e.g., Bosnia and Herzegovina, Slovakia, and Romania).

There is a broad genetic split between subpopulations inhabiting rivers draining the Alps and those present elsewhere, while at the local scale a number of distinct genetic clusters corresponding to individual river systems have been identified. Subpopulations inhabiting rivers with contrasting ecological regimes might also exhibit adaptive differentiation driven by distinctive selection pressures (Weiss et al. 2011; Marić et al. 2014; Snoj et al. 2011, 2022). These findings have significant implications for this species' conservation management (see 'Threats' and 'Conservation').

This species is an exclusive inhabitant of upland (200-800 metres asl) river channels where it demonstrates a general preference for lotic stretches characterised by cool, oxygen-rich water. However, these vary considerably in slope, width, discharge, sediment regime and channel morphology, depending on the location. Adult individuals tend to occupy deeper pools and glides located beneath rapids or waterfalls, in river bends or at confluences, but have extended and variably-sized (from <10 to >100 kilometres of linear river length, depending on the location and individual fish) home ranges within which they undertake seasonal movements related to foraging, reproduction and water temperature. They are somewhat territorial and select positions in dominance hierarchies based on maximising their energy intake, thus larger specimens tend to occupy the upstream portions of favourable reaches. In contrast, juveniles and subadults have no fixed territories and are often observed in riffles and runs. In addition, this species occasionally enters both natural (e.g., Lake Plav, Montenegro) and artificial (e.g., Lake Višegradsko, Bosnia and Herzegovina or Lake Bicaz, Romania) lakes (Holčík et al. 1988, Kottelat and Freyhof 2007, Muhamedagić and Habibović 2013, Ihuț et al. 2014, Pinter et al. 2024).

Adult individuals tend to spend both winter and summer in deep, low-velocity depressions and pools, but often forage in shallower, fast-flowing reaches with appropriate structural heterogeneity provided that the water temperature remains ≤ 20°C. When conditions become warmer, they may move into cool upstream stretches, areas influenced by upwellings of groundwater, or return to deeper environments, although in Alpine rivers strongly influenced by glacier and snow melt the average summer water temperature can remain as low as 11-12°C. Large individuals are somewhat solitary, but aggregations comprising numerous individuals have been observed in larger pools and thermal/hydrological refuge habitats during both warm and cold periods of the year (Holčík et al. 1988, Ihuț et al. 2014, Pinter et al. 2024).

This species is a visual predator. Older age classes demonstrate a pronounced tendency towards piscivory, but also prey opportunistically on small mammals, waterbirds, amphibians and reptiles. Cannibalism has been observed on a regular basis (Holčík et al. 1988, Hanfland et al. 2015).

It is the world's second-largest salmonid species, capable of growing to >1.5 m in length with a body mass exceeding 50 kg, although specimens of such size are exceptionally rare. It is relatively long-lived, with a maximum recorded age of 20+ years. Male individuals mature at an age of 3–4+, when they typically weigh 1–2 kg, whereas females mature at age 4–5+ with a body weight of 2–3 kg (Holčík et al. 1988, Jakšić et al. 2024, Pinter et al. 2024).

The annual reproductive period extends for 2–3 weeks and begins between February and early May, with the precise timing dependent on river type and abiotic factors such as water temperature. This species is iteroparous and potamodromous, and nuptial individuals thus migrate relatively short distances (typically 1–25 km) to specific upstream spawning sites comprising well-washed gravel beds in shallow (typically 40–80 cm depth) reaches with moderate flow, which are often located in small tributaries. Nuptial individuals develop a conspicuous epigamic colour pattern comprising pinkish-red pigmentation on the sides of the body (Holčík et al. 1988, Esteve et al. 2013, Hanfland et al. 2015, Pinter et al. 2024).

Unlike most salmonids, this species forms pairs which possibly bond several days or weeks prior to migration. Upon arrival at spawning sites, pairs normally remain together and the male aggressively defends the site to prevent other males from entering. Meanwhile, the female creates an elliptical depression (redd) in the substrate, into which the gametes are deposited. Females produce 1,500–35,000 eggs, depending on the size of the individual. The presence of unclogged, well-oxygenated interstitial spaces within each redd is considered to be crucial for successful incubation and early development. Alevins typically emerge after around one month, and begin to feed in shallow marginal areas close to the spawning site once their yolk sacs are absorbed. Juveniles (parr) usually disperse downstream in riffles or runs, and often aggregate in the vicinity of cover such as undercut banks or submerged wood. They feed on aquatic invertebrates and the fry of other fish species until age 2+, at which point they become almost entirely piscivorous. Individuals inhabiting warmer, more-productive rivers, e.g., in the Alpine foothills, undergo earlier development and more rapid growth than those native to cooler waters with limited prey opportunities, e.g., in the central Alpine region (Holčík et al. 1988, Nikčević et al. 1998, Esteve et al. 2013, Hanfland et al. 2015, Pinter et al. 2024).

This species' decline has largely been driven by river regulation and other hydromorphological alterations, which have resulted in the widespread loss of the heterogeneous, interconnected fluvial habitats required to complete its life cycle. In particular, the widespread construction of dams, weirs and other barriers has altered natural flow and sedimentation regimes, blocked seasonal migration routes and access to spawning sites, fragmented subpopulations (which leads to reduced genetic diversity due to the naturally low abundance of mature individuals), and generally reduced the extent of suitable habitat for all life stages. Moreover, the majority of existing fishways do not favour the passage of large-bodied fish taxa such as Danube Salmon (Freyhof et al. 2015, 2020; Hanfland et al. 2015; Schöffmann and Marić 2024).

Hydroelectric dams have also created unnatural fluctuations in discharge and water temperature (hydropeaking and thermopeaking) which cause dewatering of foraging and spawning sites, the loss of stable nursery habitat for juveniles, and downstream displacement. The combined effect of hydropeaking, dam flushing operations, changes in land use, and the removal of riparian vegetation is also likely to have increased the accumulation of fine sediments at spawning sites, thus impairing the hatching and survival rates of eggs and larvae. Extensive dam construction is ongoing and/or planned for the future in some parts of the Danube Salmon's range, particularly in Austria, Slovenia and the Western Balkans (Freyhof et al. 2015, 2020; Snoj et al. 2022; Pliberšek and Tavčar 2024; Schöffmann and Marić 2024).

The quality of habitat has been further diminished by bank stabilisation, channelisation and other efforts to enhance flood protection or exploit water for human development. The industrial extraction of riverine gravel and other sediments for urban development has contributed to an observed reduction in the extent and quality of available spawning sites (Curtean-Bănăduc et al. 2019, Hanfland et al. 2015, Schöffmann and Marić 2024).

This species is also susceptible to declining water quality caused by diffuse and point source agricultural, domestic and industrial pollution, which has resulted in eutrophication or discharge of toxic substances at some locations, e.g., the Váh River downstream of Ružomberok and Hron River downstream of Dubová in Slovakia, the Bosna River in Bosnia and Herzegovina, the Vișeu River in Romania (Holčík et al. 1988, Muhamedagić and Habibović 2013, Curtean-Bănăduc et al. 2019, Jakšić et al. 2024).

In some cases, the threats described above have also driven declines in its prey species, which may have further exacerbated local declines (Schmutz et al. 2023, Pinter et al. 2024).

Some of this species' biological traits, such as large body size, relatively low abundance and late sexual maturity, render it susceptible to recruitment overfishing. Licensed and unregulated harvesting and angling have thus contributed to its decline, with the negative impact in some cases exacerbated by insufficient fisheries regulations (e.g., a lack of catch and release practices, minimum size limits which permit the removal of sexually immature individuals, continued offtake of adults during the annual reproductive period) and/or a reduction in the size of local subpopulations due to habitat fragmentation and other factors. Licensed angling is no longer considered to represent a widespread threat due to the implementation of fisheries' regulations and an ongoing increase in catch-and-release policies (see 'Conservation'), but unregulated harvesting using nets, spears, poisons and even explosives continues to occur in the Western Balkans, Romania and Ukraine (Holčík et al. 1988, Witkowski et al. 2013, Didenko et al. 2011, 2018, Curtean-Bănăduc et al. 2019).

The impact of supportive breeding and stocking with hatchery-reared individuals derived either from captive broodstocks or through the stripping of wild adults is increasingly viewed as a threat to the integrity of natural gene pools. Individuals continue to be translocated between different river systems (see 'Population'), stocking is often not precisely-documented, and widespread admixture between different genetic clusters has occurred. For example, in Austria, all rivers which retain extant subpopulations are regularly stocked with individuals derived from the Mur River, and it is believed that the original wild gene pools have been completely lost. In addition, the repeated use of captive, semi-domesticated broodstocks for the production of gametes, plus the practice of mixing gametes obtained from multiple male and female individuals within a single container ("mixed milt fertilisation") may be creating artificial selection processes which are likely to drive genetic drift, reduce the scale of differentiation between locally-adapted subpopulations, and lead to a loss of hereditary behaviours such as spawning site fidelity. Stocking with hatchery-reared individuals may also lead to increased competition, predation pressure and the transfer of parasites and pathogens (Kuciński and Fopp-Bayat 2021, Snoj et al. 2022, Jakšić et al. 2024).

Rising water temperatures due to climate change also represent a plausible threat, since they may interfere with food availability, lifespan, and the timing of reproductive processes (Jakšić et al. 2024, Pinter et al. 2024).

Subpopulations inhabiting some rivers draining the Alps may also be threatened by increasing abundance of the piscivorous Eurasian Otter (Lutra lutra), Great Cormorant (Phalacrocorax carbo) and Goosander (Mergus merganser), which consume both Danube Salmon individuals and its prey species (Hanfland et al. 2015, Schmutz et al. 2023).

This species has been viewed as an iconic sports fish since at least the late 17^th century, and is considered to be of high value in recreational fisheries across most of its extant range. It is particularly favoured by fly fishing enthusiasts, and supports a thriving international tourist industry in some countries, e.g., Slovenia, Slovakia and Montenegro (Witkowski et al. 2013, Dekić et al. 2024, Schöffmann and Marić 2024).

It was viewed as a delicacy and harvested commercially in the past, although few detailed records are available. Today, it is occasionally landed as bycatch in nets used by professional fishers working on artificial accumulation lakes, and is targeted by unlicensed fishers for consumption at the local scale, including restaurants, in some parts of its range (Holčík et al. 1988, Didenko et al. 2011, 2018, Curtean-Bănăduc et al. 2019, Schöffmann and Marić 2024, Simić et al. 2024).

Danube Salmon farming has been ongoing since declines in abundance were first noted during the late 19^th century. The practice was first developed in Slovakia and later shifted to Austria, the former Socialist Federal Republic of Yugoslavia (SFRY) and Czechia. The stocking of fisheries with hatchery-reared alevins and juveniles began shortly afterwards, and today, it is most often carried out using fingerlings (age 0+) to age 2+ individuals derived from captive broodstocks. In some countries, e.g., Poland, and Romania, hatchery-production has declined in recent decades. During the 20^th century, eggs and fry were regularly translocated between different countries, e.g., from the Czech Republic (Czechia) or SFRY to Austria, and from Slovakia to various countries including Austria, Germany, Switzerland and Poland. Fertilised eggs were also exported from the Czech Republic (Czechia) and Slovakia to countries outside of their native range, including Finland, Sweden, Denmark, Belgium, Spain, Morocco and Canada (Holčík et al. 1988, Ihuț et al. 2014).

This species is included in Appendix III of the Bern Convention and Annexes II and V of the European Union Habitats Directive.

It is nationally-protected and listed as Endangered in the national Red Books or Red Lists of most countries within its native range, and is currently treated as Extinct in the Wild in Czechia and Poland (Mrakovčić et al. 2006, Witkowski et al. 2013, Freyhof et al. 2023).

It is present within the boundaries of various protected areas, many of which are included in the European Union's Natura 2000 network or the Emerald Network of Areas of Special Conservation Interest. A dedicated Danube Salmon protected area is established at Bosanska Krupa in Bosnia and Herzegovina, where fishing is not permitted anywhere within a 600-metre stretch of the Una River (Dekić et al. 2024, Jakšić et al. 2024).

Angling regulations are in place throughout its range, but differ at both local (e.g., provincial) and national scales depending on country. For example, minimum landing sizes vary from 65 centimetres in the Czechia to 100 centimetres in Serbia, while the extent of closed fishing seasons and the number of individuals which each angler may catch per season are also somewhat variable. Only artificial bait can be used in some countries, while night fishing may or may not be permitted. Catch and release practices have become increasingly common in recreational fisheries since the turn of the century, while all forms of fishing are prohibited in Romania and Ukraine (Witkowski et al. 2013, Didenko et al. 2018, Dekić et al. 2024, Jakšić et al. 2024, Pliberšek and Tavčar 2024).

This species has been included within the framework of several conservation activities co-funded by the European Union's LIFE programme. From 2012–2016, the 'Ljubljanica Connects' project (LIFE10 NAT/SI/000142) reconstructed two dysfunctional fish passes and removed other migratory barriers along the Ljubljanica River in Slovenia. The 'Life in Upper Drau River' project (LIFE06 NAT/A/000127) ran from 2006–2011, and led to improved habitat conditions in a short stretch of the upper Drava River in Austria. The LIFE 'Living Space of Danube Salmon' scheme (LIFE99 NAT/A/006054) focused on habitat rehabilitation and monitoring in the Pielach River and adjacent stretch of the Danube in Austria and took place from 1999–2003 (Schmutz et al. 2002, Unterlercher 2011, Sapač and Zabret 2016).

In some European Union member states, stocks may have benefitted from efforts to improve the ecological status of rivers within the framework of the European Water Framework Directive 2000/60/EC. This legislation was implemented in 2003 and has alongside a general decline in heavy industry resulted in improving water quality across much of its range. Pollution has thus been reduced in many areas, while management actions focussed on freshwater fishes have included measures to improve habitat quality, e.g., restoration of spawning sites or stream banks, or connectivity, e.g., dam-removal projects or installation of nature-like fishways to reestablish migration routes. Such activities have facilitated recolonisation of some formerly occupied river reaches in Austria and Germany, although no significant increases in abundance have yet been detected (Ratschan 2014; Hanfland et al. 2015, Schmutz et al. 2023, Pinter et al. 2024).

However, active conservation strategies in most countries are limited to stocking programs and angling regulations, with these measures often managed by local angling associations or NGOs. In most of central and eastern Europe, Danube Salmon management depends solely on broodstocks, while stocking is less frequent in southeastern Europe, where an absence of comprehensive genetic research also hampers the development of appropriate conservation actions. In Serbia, it has been successfully reintroduced to the Đetinja and Moravica rivers in the West Morava River (Great Morava system), using stock from the Drina watershed (Witkowski et al. 2013, Ihuț et al. 2014, Snoj et al. 2022).

A number of authors have recommended that stocking should be avoided wherever possible, or that it should be discontinued once there is evidence of viable natural reproduction. When it is considered necessary, the use of high-resolution genotyping to identify current and potential broodstock, strengthen supportive breeding activities and define genetic clusters which represent meaningful management units at the evolutionary microgeographic scale is strongly recommended. Where possible (i.e., in the case of subpopulations supporting a viable number of mature individuals), the practice of translocating individuals between different genetic clusters should be discontinued in order to preserve what remains of the species' natural genetic structure. In addition, gametes should be obtained from wild individuals rather than captive broodstock and analysed rapidly to ensure their viability. Restocking with eyed eggs rather than hatchery-reared juveniles should also be explored, and the practice of mixing gametes from multiple adult individuals should be discontinued. Genetic samples should be retained for all broodstocks in order to facilitate future subpopulation analyses, and a comparable protocol should ideally be applied throughout the species' range. Furthermore, improved restocking techniques should be combined with habitat restoration efforts in areas where the latter is a realistic target, since stocking may have little long-term impact if ecological conditions remain unsuitable. Some of these recommendations have already been adopted or trialled, e.g., in Slovenia and parts of Bosnia and Herzegovina (Geist et al. 2009, Ihuț et al. 2014, Marić et al. 2014, Weiss and Schenekar 2016, Kuciński and Fopp-Bayat 2021, Snoj et al. 2022, Pliberšek and Tavčar 2024, Schöffmann and Marić 2024).

The abundance, genetic structure and location of important spawning and nursery sites for all subpopulations that have not been analysed to date should ideally be determined in order to facilitate informed management decisions. Actions to reduce overharvesting by both humans and predators could include the enforcement of catch and release angling regulations (see below), an increase in minimum take size limits, the establishment of additional dedicated protected areas and no-fishing zones (the latter targeting particularly small subpopulations), the development and installation of effective fishways, and the removal of obsolete barriers such as dams and weirs. A repeatable protocol for assessing the viability of individual subpopulations should ideally be adopted at the range-wide scale, and a promising integrated approach combining genetic analyses, field observations and sampling data has already been developed and trialled in Austria (Weiss and Schenekar 2016, Jakšić et al. 2024, Pinter et al. 2024).

The prevention of further hydropower development in free-flowing river stretches should be treated as a conservation priority, particularly in Austria, Slovenia and the Western Balkans where major expansion is ongoing. The restoration of connectivity to fragmented rivers is also of prime concern, and further research regarding the provision of both up- and downstream migration pathways for Danube Salmon is needed, especially with regard to efficient fishways bypassing larger dams. Further habitat restoration efforts should focus on the reestablishment of natural river channel morphology, flow regimes and bed load dynamics, including suitable spawning areas for adult individuals and heterogeneous riparian zones for juveniles. Enhancing the abundance of prey may also be required at some locations. These management activities should initially focus on river stretches which currently support the largest viable subpopulations in order to increase the extent of suitable habitat and connect them to less-resilient stocks (Ihuț et al. 2014, Pinter et al. 2024).

There is also an urgent need for a regional inventory and threat analysis plus further research focusing on this species' habitat requirements in the Western Balkans region (Pinter et al. 2024).