Italian and Corsican Brown Trouts
The systematics of Italian and Corsican trouts within the Brown Trout (Salmo trutta) complex (see discussion on Eurasian and North African Brown Trouts below) have yet to be definitively resolved (Splendiani et al. 2019a). For instance, the insular subpopulations from Corsica and Sardinia, treated here as S. ghigii, exhibit significant genetic differentiation from mainland subpopulations. Together, these subpopulations form a locally endemic sub-lineage that may merit recognition as a distinct taxon (Segherloo et al. 2021, Splendiani et al. 2024).

Mediterranean brown trout represent a distinct evolutionary lineage that has diverged significantly from the Atlantic lineage, which inhabits the river systems of Central Europe and the Atlantic basin (see below). The prolonged historical isolation between Atlantic and Mediterranean brown trout has led to the development of specific mitochondrial lineages and unique genetic variations at nuclear loci, particularly at the lactate dehydrogenase C1 locus (LDH-C1*). Research indicates that the LDH-C1*100 allele is exclusively found in native Mediterranean brown trout populations, whereas the *90 allele is fixed in North European populations from Atlantic drainages (Hamilton et al. 1989, García-Marín et al. 1991, Bernatchez 2001, Caputo et al. 2004, Almodóvar et al. 2006, Splendiani et al. 2016.

S. ghigii subpopulations predominantly exhibit AD and/or ME mitochondrial lineages (see below), with the MA lineage appearing sporadically. Several studies have highlighted exceptionally high genetic diversity among these subpopulations, detected at both mitochondrial and nuclear markers, and observed at inter- and intra-basin levels (Zaccara et al. 2015; Rossi et al. 2019, 2022; Splendiani et al. 2019b, 2024). This genetic divergence is also linked to the loss of anadromous behaviour, a trait shared among brown trout populations across the Mediterranean region.

Eurasian and North African Brown Trouts
At the broader scale, there is currently no general consensus regarding the systematic classification of most Eurasian and North African brown trouts, an assemblage comprising all representatives of the genus Salmo except the well-differentiated Atlantic Salmon (Salmo salar), Marble Trout (Salmo marmoratus), Softmouth Trout (Salmo obtusirostris) and Ohrid Belvica (Salmo ohridanus). While numerous, often range-restricted, members of this grouping have been described based largely on their ecological and morphological diversity, this variability is not consistently reflected by phylogenetic and phylogeographic evidence (Sanz 2018, Whiteley et al. 2019, Segherloo et al. 2021).

Despite a relatively recent diversification history spanning the period 0.5-2.5 Mya, brown trouts exhibit marked ecological and phenotypic variability throughout their large native range, which extends eastward from Europe and Northwest Africa to Russia and the Aral Sea basin. They occupy a wide range of habitats, from mountain streams and larger rivers to lakes and estuaries. Individual subpopulations can exhibit sedentary, anadromous or potamodromous life history strategies. Some freshwater systems are inhabited by multiple sympatric forms which differ in traits associated with foraging and reproductive ecology, and are sometimes referred to as "morphs", "ecomorphs" or "ecotypes" (Klemetsen et al. 2003, Kottelat and Freyhof 2007, Gratton et al. 2014, Ferguson et al. 2019, Segherloo et al. 2021).

Some authorities have viewed this combination of factors to be representative of high species diversity and recognised around 50 nominal taxa, a number of which have been described this century (Kottelat and Freyhof 2007, Snoj et al. 2011, Sanz 2018). Alternatively, their systematics have been viewed from a phylogenetic and phylogeographic perspective based largely on mitochondrial DNA (mtDNA) analyses, with all subpopulations treated as a single polymorphic taxon customarily referred to as the “Brown Trout (Salmo trutta) complex” (Sanz 2018, Whiteley et al. 2019, Segherloo et al. 2021).

The latter approach led to brown trout diversity being defined by ten mtDNA lineages or sublineages corresponding to extensive catchments (the Danube, Atlantic, Mediterranean and Adriatic basins), specific geographic areas (the Balkan Peninsula and North Africa), individual watersheds (the Dades, Duero and Tigris rivers) and a distinctive phenotype (Marble Trout). Subsequent studies revealed that the distribution of some of these mtDNA lineages extends beyond their defined boundaries, e.g., the Adriatic lineage occurs from the Iberian Peninsula to the Republic of Türkiye, and the Marble Trout lineage is present in areas where no marbled phenotype exists, such as Corsica, central and southern Italy, Albania and Greece (Bernatchez et al. 1992, Apostolidis et al. 1997, Bernatchez 2001, Suárez et al. 2001, Cortey et al. 2004, Sušnik et al. 2005, 2007; Splendiani et al. 2006, Martínez et al. 2007, Snoj et al. 2009, 2011; Tougard et al. 2018, Talarico et al. 2021, Rossi et al. 2022, Schöffmann et al. 2022).

Moreover, several studies have revealed the presence of mosaic distributions of mtDNA haplogroups and mitochondrial-nuclear phylogenetic discordance among wild brown trout populations (Snoj et al. 2009, Gratton et al. 2014, Pustovrh et al. 2014, Splendiani et al. 2020). This suggests the presence of incomplete lineage sorting or asymmetric introgressive hybridization, which are common phenomena in rapidly diverging lineages and indicate that mtDNA genealogies might be generally unsuitable for defining phylogenetic relationships between brown trout taxa (Pustovhr et al. 2011, 2014). In the case of brown trouts, naturally intricate patterns of diversification and secondary contact shaped by repeated glaciations during the Pleistocene have been additionally complicated by widespread anthropogenic translocation and introgressive hybridisation since the Middle Ages (Largiadèr and Scholl 1996, Sanz et al. 2006, Lerceteau-Köhle et al. 2013). The combined use of multiple nuclear (nDNA, e.g., microsatellites, nuclear genes) and mitochondrial markers has already provided better insight into this complex scenario, resulting in progress towards a deeper understanding of evolutionary relationships at particular geographic scales or among subsets of putative taxa (Snoj et al. 2002, 2010, 2011; Sušnik et al. 2006, 2007; Berrebi et al. 2013, 2019; Gratton et al. 2014, Marić et al. 2017).

An integrative taxonomic approach combining morphological and ecological data with next generation sequencing of nDNA to identify genomic clusters may represent the most promising option for resolving brown trout systematics (Guinand et al. 2021, Segherloo et al. 2021). However, no comprehensive morphological or nDNA analyses have yet been completed, and it is plausible that the elaborate genetic and phenotypic diversity demonstrated by these fishes may never be adequately captured by a single accepted taxonomic system (Whiteley et al. 2019).

Pending a definitive outcome to the above, the Red List broadly follows the nomenclature provided by Fricke et al. (2024).

The Italian and Corsican Mediterranean Trout does not meet the range thresholds for Vulnerable under Criterion B1 (extent of occurrence (EOO) <20,000 km²) or D2, and its uncertain area of occupancy (AOO) precludes assessment under Criterion B2. The population is estimated to exceed 10,000 mature individuals, meaning it does not qualify under Criteria C or D. Additionally, no quantitative analysis is available to apply Criterion E.

Although no comprehensive population trend data exists, an ongoing decline in the area of occupancy (AOO) is suspected based on field observations over the past two decades, habitat degradation, widespread introgression from non-native individuals, and the impacts of climate change. These factors suggest the decline may approach or meet the thresholds for Vulnerable to Critically Endangered under Criterion A4 (≥ 30% to ≥ 80% over the period 2019–2031, equivalent to three generations).

Given the uncertainty in available data, the plausible Red List categories range from Vulnerable to Critically Endangered. Consequently, this species is assessed as Endangered.

This species is native to the Italian Peninsula as well as the islands of Corsica and Sardinia (see 'Taxonomic Notes').

In Peninsular Italy, its distribution within the Po River system is limited to former glacial refugia in the upper reaches of the Dora Riparia, Stura di Lanzo, Pellice, and Tanaro rivers in the Cottian and Maritime Alps, along with the upper Enza, Secchia, Panaro, and Bidente di Pietrapazza rivers in the northern Apennines (Splendiani et al. 2020, Talarico et al. 2023).

Beyond the Po watershed, it inhabits headwater tributaries and a few lowland affluents of river systems draining the Apennine range, extending southward from Marche to Molise on the Adriatic slope and from Liguria to Calabria on the Tyrrhenian slope. Subpopulations are also present in rivers flowing into the Ionian Sea in the Basilicata region (Bernatchez 2001, Splendiani et al. 2019b, Palombo et al. 2021).

In Corsica and Sardinia, it inhabits the upper reaches of several small river systems.

The exact number of extant subpopulations remains uncertain, as most show some level of genetic introgression with non-native individuals (see 'Threats'). Only a limited number of non-hybrid subpopulations have been identified, which are generally small, patchily distributed, and confined to isolated river stretches. The absence of genetic exchange among these subpopulations has led to a degree of divergence between them (Nonnis Marzano et al. 2003, Querci et al. 2013, Berrebi 2015, Caputo Barucchi et al. 2015, Fabiani et al. 2018, Berrebi et al. 2019, Rossi et al. 2019, 2022, Splendiani et al. 2019b, 2020, Regione Autònoma della Sardegna 2021, Talarico et al., 2021, 2023, Splendiani et al. 2024). Notably, a number of highly differentiated native subpopulations persist in Corsica (Berrebi 2015, Berrebi et al. 2019, Delling et al. 2020).    

Accurately estimating the overall population size of this species is similarly challenging, though it is believed to exceed the minimum threshold for Red List criteria (<10,000 mature individuals). While no quantitative population trend estimate is available, a decline in Area of Occupancy of at least 25% - potentially as high as 80% - is suspected over the period 2019-2031 (equivalent to three generations, with a generation length of four years) due to ongoing hybridisation, habitat degradation and other drivers (see 'Threats').

This species primarily inhabits low nutrient headwater rivers and tributary streams which contain cool, well-oxygenated water and are often characterised by seasonal fluctuations in discharge. Substrata in such habitats tend to comprise a mixture of exposed bedrock, boulders, rocks, cobbles and gravel, with refuges in the form of overhanging riparian vegetation, undercut banks and woody structures such as branches, roots or fallen trees (Splendiani et al. 2013, Lorenzoni et al. 2019, Rossi et al. 2022).

In fluvial habitats, larger individuals occupy deeper pools and glides and are territorial, selecting stream positions in dominance hierarchies based on maximising their energy intake. In contrast, juveniles and subadults are often observed in riffles and runs (Kottelat & Freyhof 2007).

It is a visual predator which feeds on benthic and drifting invertebrates, e.g., Ephemeroptera, Diptera, Plecoptera, Trichoptera, while larger individuals may also consume amphibian larvae and smaller fishes (Fochetti et al. 2003).

Its lifespan is at least eight years, and adults mature at age 2-3+. The annual reproductive period extends from November to February or March, with the precise timing dependent on location. It is characterised by nuptial individuals undertaking short upstream migrations to specific spawning sites comprising well-washed gravel beds in shallow, fast-flowing reaches. After arriving at these sites, individual females create shallow depressions (redds) in the substrate, into which the gametes are deposited. The presence of unclogged, well-oxygenated interstitial spaces within each redd is considered to be crucial for successful incubation and early development (Caputo et al. 2010, Lobón-Cerviá & Sanz 2017).

This species faces significant threats across its range, primarily due to introgressive hybridisation with non-native Brown Trout (Salmo domestic strain), which continues to be widely introduced to sustain recreational fisheries. The large-scale stocking of non-native trout increased dramatically with the advancement of aquaculture techniques in the late 19^th century, resulting in millions of alevins and fingerlings being released annually. Hatchery-produced individuals today are of mixed origin but are typically derived from the Atlantic Brown Trout lineage and are often introduced illegally (Meraner & Gandolfi 2017, Splendiani et al. 2019b). Furthermore, stocking practices frequently involve hatchery-reared individuals from native populations, often without regard for their specific geographic origins. This indiscriminate approach leads to hybridisation between genetically distinct native subpopulations, leading to further losses of local genetic diversity (Fabiani et al. 2018; Talarico et al. 2023, Splendiani et al. 2019b).

Beyond hybridisation, this species is also severely impacted by river regulation and habitat degradation, which have led to the widespread loss of the heterogeneous, interconnected fluvial habitats essential for its life cycle. The construction of dams, weirs, and other barriers has altered natural flow and sedimentation regimes, blocked access to spawning sites, fragmented subpopulations, prevented recolonisation of some river stretches, and reduced the availability of suitable habitat for all life stages. Additional habitat deterioration has resulted from bank stabilisation, channelisation, and flood protection measures, and the industrial extraction of riverine gravel and sediments for urban development, as well as water exploitation for human development (Splendiani et al. 2019b, Carosi et al. 2020a).

Hydroelectric dams have exacerbated these challenges through hydropeaking and thermopeaking, creating fluctuations in discharge and water temperature that dewater spawning sites, eliminate stable juvenile nursery habitat, and cause downstream displacement and stranding of fish. Moreover, the combined effects of hydropeaking, dam flushing operations, land-use changes, and riparian vegetation removal have increased the accumulation of fine sediments at spawning sites, impairing the hatching and survival of eggs and larvae (Freyhof et al. 2020).

Additionally, the species is affected by diffuse and point-source pollution from agricultural, domestic, and industrial activities, leading to eutrophication and the discharge of toxic substances at some locations (Splendiani et al. 2013, Gumiero et al. 2022).

Overexploitation by recreational anglers is considered to be a threat in areas not governed by strict fishing regulations (Carosi et al. 2022). The introduction of non-native species such as Rainbow Trout (Oncorhynchus mykiss) may also exert a negative impact on native salmonids as a result of increased competition, predation and pathogen transmission (Stanković et al. 2015).

Finally, climate change and rising water temperatures pose a growing threat, as they may disrupt food availability, lifespan, and the timing of reproductive processes (Carosi et al. 2020a). For instance, recent droughts have caused some Sardinian subpopulations to become isolated in small remnant pools as their typically perennial habitats dried up, only to be subsequently impacted by severe autumn floods.

This species plays a significant role in recreational fisheries across its range, where angling for both native and non-native salmonids supports a lucrative industry largely dependent on the stocking of hatchery-produced individuals with mixed genetic origins (Carosi et al. 2022, Polgar et al. 2022).

This species occurs within numerous protected areas, some of which are designated National Parks and/or included in the European Union's Natura 2000 network. However, many of these subpopulations demonstrate genetic admixture with non-native trout, and stocking of hatchery-produced individuals (see 'Threats') continues to take place within protected area boundaries (Lorenzoni et al. 2019, Splendiani et al. 2019, Talarico et al. 2023).

Clarifying its taxonomic status remains a critical research priority, particularly regarding the subpopulations in Corsica and Sardinia, where conservation concerns are especially pressing (Splendiani et al., 2024).

Some subpopulations have been included in European Union co-funded LIFE projects (LIFE03 NAT/F/000101, LIFE12 NAT/IT/0000940, LIFE17 NAT/IT/000547, and LIFE18 NAT/IT/000931). Actions undertaken within the framework of these projects included the identification of genetically intact native subpopulations, targeted removal of non-native trout, and supportive breeding of non-hybrid individuals (Lorenzoni et al. 2019, Carosi et al. 2020b, Talarico et al. 2023).

Regulations governing recreational fisheries include daily bag limits, minimum catch sizes, the application of catch and release policies, and closed fishing seasons, with fishing completely prohibited along certain river stretches. The stocking of non-native trout in Italy is currently regulated under European and national legislation, but enforcement of these regulations remains inconsistent. Stocking activities are predominantly carried out by private fishing associations and local fisheries authorities, and rehabilitation efforts for native trout subpopulations vary widely (Splendiani et al. 2019b, Carosi et al. 2022).

Since the 2010s, the production of Brown Trout from the putatively native Mediterranean genetic lineage has been increasingly promoted to circumvent restrictions on stocking non-native freshwater fishes. These individuals, marketed as "Mediterranean Trout," are in fact hybrids originating from various locations in central and southern Italy, often exhibiting introgression with Atlantic and domestic lineages. Furthermore, stocking frequently occurs in areas where the Mediterranean lineage is not native, meaning that this approach fails to prevent the spread of non-native alleles (Splendiani et al. 2019a, 2019b). Comprehensive scientific analyses of these stocking practices are still required.

It has been widely recommended that the conservation management of European trouts should be independent of their systematic classification, given the lack of taxonomic consensus and the presence of significant microgeographic genetic and phenotypic diversity (see 'Taxonomic Notes'). Each subpopulation should therefore be assessed individually, considering its evolutionary and genetic significance, population trends, and specific threats, to facilitate priority ranking and the effective allocation of conservation resources through site-specific, catchment-scale management plans (Meraner & Gandolfi 2017, Berrebi et al. 2019, Splendiani et al. 2019a, 2019b).

The abundance trends and genetic integrity of many subpopulations remain unknown, and their assessment should be a key focus of future research to ensure proper conservation prioritisation. In practice, local, national, and regional coordination will be essential for the efficient implementation of these efforts.