Alders are deciduous riparian trees distributed mostly in the Mediterranean, temperate and boreal zones of the Northern Hemisphere (Bjelke et al., 2016) up to the Himalayas and the Andes. These trees compose the genus Alnus (Family Betulaceae), represented by more than 40 species. In the Iberian Peninsula, two different species coexist: Alnus glutinosa (L.) Gaertn. (common name: common alder) and Alnus lusitanica Vít, Douda & Mandák (common name: Iberian alder; Vít et al., 2017). Common alder is the most widespread alder species in Europe and Western Asia, playing a significant ecological role as a key component of the riparian vegetation along streams (Claessens et al., 2010; Smeriglio et al., 2022). However, in the Iberian Peninsula, the presence of A. glutinosa is limited to the northeast region. Conversely, A. lusitanica is the most representative and widespread alder species with a large distribution in the northwest region (Gomes Marques et al., 2024a; Martín et al., 2024).

Alders are actinorhizal plants that fix atmospheric nitrogen, therefore contributing very significantly to nitrogen dynamics at the local and landscape scales. In addition, they stabilize streams and riverbanks, functioning as a protective barrier against flooding, preventing waterlogging of crops and surrounding areas, and mitigating the damage caused after periods of widespread rainfall. Also, alders contribute to the rapid colonization of abandoned sites and the maintenance of biodiversity by providing refuge for terrestrial and aquatic organisms (Compton et al., 2003; Wipfli and Musslewhite, 2004; Claessens et al., 2010; Handa et al., 2014). This combination of attributes, along with the provision and cultural ecosystem services that are derived from streams and rivers, makes alder replacement by other species very difficult.

Over the recent decades, different alder species have been severely impacted by decline and mortality events caused by abiotic factors, such as extended periods of drought followed by flooding, as well as biotic factors (Ferreira et al., 2022; Gomes Marques et al., 2022), which accumulate to already existing long-term pressures on rivers (Vörösmarty et al., 2010). Along with global change, an increase in outbreaks of invasive pathogenic fungi and oomycetes has severely affected native plants worldwide (Fisher et al., 2020; Gomes Marques et al., 2024b).

Recently, a strong reduction of alder stands has been detected in Europe due to infection by the Phytophthora alni species complex Brasier & S.A. Kirk (Bjelke et al., 2016). This complex includes a group of pathogens that cause Phytophthora disease of alder, also known as alder dieback, affecting different organs of the aerial part of the tree and roots (Gibbs et al., 1999). Although P. ×alni (Husson et al., 2015) is the most aggressive species within the complex, a possible synergy between Phytophthora species in the damage caused to alders is unknown. The aggressiveness is favored by mild winters and warm, but not too hot, summers (Bjelke et al., 2016; Gomes Marques et al., 2024b; Horta Jung et al., 2024). Also, global environmental changes may promote shifts in the pathogen distribution and impact. Their dispersal occurs primarily along water currents which transport thousands of infective zoospores, thus constituting an important route of dissemination of Phytophthora species. Due to the dendritic structure of river networks, Phytophthora can spread rapidly to new areas, notably downstream (Bjelke et al., 2016). The Phytophthora zoospores usually infect the host through the root system, mainly fine roots, or by wounds at the base of the trunk and ascend through its tissues causing lesions in the cambium (Brasier and Kirk, 2001; Černý and Strnadová, 2012). Thus, alders may be subjected to multiple infections over time due to their proximity to the river and the contact of their tissues with surface runoff. Therefore, changes in phytosanitary state and vigor, as well as the degree of tree survival over time are conditioned by environmental factors, including the concentration of inoculum in the soil, the soil type, the water flow velocity, and the geomorphic position of the tree in relation to water (Gomes Marques et al., 2024b).

The wide distribution of alders, with isolated local populations, has resulted in a high genetic diversity that allows them to respond differently to selection pressures related to stand structure (like canopy composition and density) and edaphic, abiotic and biotic factors. In the same way, it has also resulted in inbreeding and moderate local differentiation, partly associated with the ease of seed dispersal through river channels (Štochlová et al., 2012). Given the local adaptation of alders, the identification/selection of more tolerant or resistant genotypes in distant stretches of the same river should seek to retain the adaptive traits so that they can thrive in different environmental conditions. Even when infected, some alders may look asymptomatic and can remain so for a long time (Elegbede et al., 2010). This would also contribute to broadening the genetic base of resistance, thereby reducing the risk of the pathogen overcoming resistance (Sniezko and Koch, 2017). For these reasons, it is very important to understand and protect the existing diversity and to identify resistant genotypes with as much genetic variation as possible and from suitable sources, so they can be used for riverbank restoration.

In this review, the impacts of alder dieback in the Iberian Peninsula and the potential breeding strategies for alder resistance to Phytophthora are summarized. New approaches to improve the resistance selection process and breeding are also described.

During the last two decades, several studies have used molecular markers to address the genetic diversity and structure of European alders (Mingeot et al., 2010; De Kort et al., 2014a, De Kort et al., 2014b; Beatty et al., 2015; Cubry et al., 2015; Havrdová et al., 2015; Gryta et al., 2017). Although traditionally considered diploid species, recent studies using cytometry and molecular markers revealed variation in ploidy level (Lepais et al., 2013; Mandák et al., 2016). This way, Mandák et al. (2016) identified three main groups across Europe in what was before considered to be A. glutinosa. In addition to the most common and widely distributed diploid (2n = 2x = 28) A. glutinosa, two tetraploid (2n = 4x = 56) clusters were identified: one in western Balkan Peninsula (A. rohlenae Vít, Douda & Mandák) and the other in the Iberian Peninsula and North Africa (A. lusitanica) (Vít et al., 2017; Gomes Marques et al., 2024a). The proposed post-glacial recolonization of Europe by alders would have taken place from multiple refuges in the north of the Iberian, Apennine and Balkan Peninsulas, composed of diploid populations (A. glutinosa), with no involvement of tetraploid populations, which may have originated later (Comes and Kadereit, 1998; Havrdová et al., 2015; Mandák et al., 2016; Šmíd et al., 2020).

In the Iberian Peninsula, the Iberian alder (A. lusitanica) and the common alder (A. glutinosa) are native tree species occurring in riparian and wetland forest communities. Both species contribute to shaping the characteristics, communities and functioning of stream ecosystems ( Table 1 ). The most representative and widespread alder is the tetraploid A. lusitanica, which is widespread from Morocco, and shows three main genetic groups with a clear geographical distribution in the northern, western and central and southern regions ( Figure 1A ; Martín et al., 2024), under Atlantic and continental climates. The common alder is present in the Ebro basin and some northeast Cantabrian and Catalonian basins ( Figure 1A ). The Ebro River represents the northeast limit for the distribution of this species (Sanna et al., 2023; Martín et al., 2024), but also a contact zone between A. lusitanica and A. glutinosa. Even so, to date, there is no evidence of gene flow between these species, as evidenced by the absence of triploid individuals among the samples prospected in this area (Martín et al., 2024). In turn, although rare, in the Balkans, triploid individuals have been identified in the overlapping distribution areas of A. glutinosa and the tetraploid A. rohlenae (Mandák et al., 2016; Šmíd et al., 2020). Based on these findings, future studies might also identify triploids in the area where A. lusitanica and A. glutinosa meet in the Iberian Peninsula.

According to the European Habitats Directive 92/43/EEC, alders are considered key components of alluvial forests, which are priority habitats for biodiversity conservation (priority habitat 91E0*). Indeed, alders are trees that live close to water and so have high ecological importance ( Table 1 ).

The extensive root system of alders contributes to defining the morphology of the stream channel, thus creating important habitats for other organisms ( Figure 2A ). Moreover, woody adventitious roots provide habitat, refuge and feeding grounds for aquatic organisms (Claessens et al., 2010). Their root system also provides an important ecosystem service by contributing to stabilize riverbanks. Alders are also nitrogen-fixing trees owing to their symbiotic association with the nitrogen-fixing bacteria Frankia alni (Voronin) Von Tubeuf in root nodules, which increases soil nitrogen availability (Teklehaimanot and Mmolotsi, 2007; Claessens et al., 2010). As a result, alder trees in the watershed could significantly increase nitrogen levels in streams. This was verified in the USA, where redder alder (A. rubra Bong.) cover in the watershed was linked to higher nitrogen concentrations in stream water (Compton et al., 2003; Shaftel et al., 2012). Increases in soil nitrogen availability in the presence of alder may also increase the nutrient concentrations of leaf litter of non-nitrogen-fixing species (Shainsky and Rose, 1995; Rhoades et al., 2001).

Alder trees do not tolerate shade, making them pioneer species, and the choice to help restore degraded riparian habitats. The shade provided by these trees during summer limits instream primary production while contributes to keeping stream water cool (Bjelke et al., 2016). Furthermore, the alder canopy provides habitat, refuge, food and feeding grounds to terrestrial and aquatic species during their terrestrial life stage, which may provide food inputs to the stream ecosystem when falling into the water (Wipfli, 1997).

Moreover, leaf litter of Iberian and common alders ( Figure 2B ) is soft and has high nitrogen concentration and low concentration of recalcitrant carbon, in comparison with leaf litter of other Iberian native tree species (Jabiol et al., 2019; Ferreira et al., 2022). This makes it a very palatable food resource for aquatic microbial decomposers and macroinvertebrate detritivores, which prefer alder leaf litter to more recalcitrant one (Graça and Cressa, 2010; Alonso et al., 2021), resulting in its fast decomposition (Feio et al., 2010; Woodward et al., 2012). Especially in autumn, alder leaf litterfall offers an important food supply to aquatic food webs (Pozo et al., 1997), providing a high-quality food resource for the early stages of detritivores aquatic insects (Molinero and Pozo, 2006). In addition, alder leaf litter often stimulates the decomposition of litter mixtures (Alonso et al., 2021), not only because it is a fast-decomposing leaf litter itself (Rubio-Ríos et al., 2023), but also because it stimulates the decomposition of more recalcitrant leaf litter due to the likely reduction in nutrient limitation (Ferreira et al., 2012; Alonso et al., 2024).

Riparian forests are particularly vulnerable to pathogen invasions, which spread rapidly along rivers, taking advantage of the current and becoming more effective once they reach stagnant waters (Bjelke et al., 2016). Since the 1990s, the principal pathogens associated with alder disease have been the oomycete of the genus Phytophthora. The first observation of the pathogen infecting alder trees in Europe was in the UK in 1993 (Brasier et al., 1995), and it rapidly spread being quickly detected in a large part of Europe (Brasier et al., 2022). Although the mechanisms behind the rapid spread of the pathogen across Europe are not well understood, it is known that water enhances the pathogen’s sporulation, spread and infection via zoospores (Chen et al., 2022). Furthermore, other authors consider that human activities have significantly contributed to the rapid spread of the disease. For instance, plants already infected by Phytophthora from nurseries may have been used for the reforestation of riparian forests, allowing the introduction of new pathogens (Gibbs et al., 2003; Eschen et al., 2015; Jung et al., 2016; Tremblay et al., 2018; Mora-Sala et al., 2022).

In the Iberian Peninsula, the first reports of outbreaks of Phytophthora associated with alders date from 2009 in Spain and from 2016 in Portugal (Pintos-Varela et al., 2010; Solla et al., 2010; Kanoun-Boulé et al., 2016). The main pathogens causing the disease in alder trees are grouped in the P. alni complex, which includes P. ×alni, P. uniformis and P. ×multiformis, described as species by Husson et al. (2015) (Bjelke et al., 2016; Nave et al., 2021; Trzewik et al., 2021; Bregant et al., 2023). Over the years, new species of Phytophthora ( Table 2 ; Figure 1B ) have been detected and isolated from diseased alders presenting symptoms similar to the disease caused by P. alni complex infection ( Figure 3 ). In particular, P. plurivora Jung & Burgess (Jung and Burgess, 2009) has been isolated from bark cankers on several occasions, being the second most isolated species in the Iberian Peninsula (Vieites-Blanco et al., 2023).

All these Phytophthora taxa ( Table 2 ) generate common symptoms in alders ( Figure 3 ) including small and yellowish leaves, increased cone production but cones smaller than the ones from healthy alders, canopy dieback, growth reduction, bark necrosis, bleeding cankers, exudations in the collar and basal part of the stem, root and collar rot, and tree mortality (Bjelke et al., 2016; Corcobado et al., 2023; Handa et al., 2014; Redondo et al., 2015a). Nonetheless, the pathogenicity of the different Phytophthora species varies. Different methods have been described to study the pathogenicity of Phytophthora species under controlled conditions (Haque et al., 2015; Chandelier et al., 2016). Within the P. alni complex, P. ×alni is considered the most aggressive (Haque et al., 2015). However, studies comparing the pathogenicity between P. plurivora and P. ×alni (the most pathogenic species isolated from alders in the Iberian Peninsula; Jung et al., 2013; Haque et al., 2015; Ferreira et al., 2022; Horta Jung et al., 2024) suggest that the oomycete with the highest pathogenicity on alder is P. plurivora (Zamora-Ballesteros et al., 2017; Corcobado et al., 2023; Vieites-Blanco et al., 2023). This difference in pathogenicity may depend on many factors, such as the alder species and their defense mechanisms, the environmental factors of different riparian forests and the isolates of the Phytophthora pathogen tested and their infection capacity. It is important to note that the higher pathogenicity of P. plurivora compared to P. ×alni may be due to two factors. First, infection of P. plurivora causes a low response in alder, which implies a low defense against this pathogen. Secondly, P. plurivora colonizes the xylem and phloem, in contrast to P. ×alni which primarily occurs in the phloem. The ability to invade the xylem may provide P. plurivora with a competitive advantage over P. ×alni (Vieites-Blanco et al., 2023). Phytophthora lacustris Brasier, Cacciola, Nechwatal, Jung & Bakonyi ( Figure 1B , Table 2 , Rial-Martínez et al., 2023) and P. hydropathica Hong & Gallegly ( Table 2 , Pintos-Varela et al., 2016) were also detected in alder trees or river water associated with riparian alder, but their pathogenicity remains to be demonstrated.

Given the ecological importance of alders ( Table 1 ), their disappearance from riparian forests, due to Phytophthora occurrence, will likely alter the plant community diversity and structure, and affect the characteristics of streams, because of modified channel morphology, decreased bank stability and increased water temperature ( Figure 4A ). Root rot caused by the pathogen will result in reduced tree stability, and fallen trees in riverbanks will cause accelerated erosion, affecting river geomorphological conditions and increasing damages caused by flooding in crops and farms ( Figure 4B ). Also, with alder death, the loss of woody adventitious roots will reduce the instream habitat available to aquatic organisms, and the loss of wide-canopy trees will reduce the habitat availability to terrestrial organisms. Moreover, the disappearance of alder leaf litter from autumn leaf fall will reduce the supply of high-quality organic matter, disrupting the stream detrital pathway (Alonso et al., 2021). Consequently, instream nutrient cycling will be impaired due to the decreased decomposition rate of leaf litter, caused by the loss of high-quality alder leaf litter (Rubio-Ríos et al., 2023). Additionally, the absence of alder will stop contributing to the increase of the quality of leaf litter from non-nitrogen-fixing species (Rhoades et al., 2001), which may decompose slower due to the loss of the stimulatory effect of alder litter presence on the decomposition of other low-quality leaf litter in mixtures (Ferreira et al., 2012; Alonso et al., 2024).

Even before alder trees completely disappear from streamside forests due to the disease, sick alder trees might already be affecting the functioning of the stream ecosystem. For instance, diseased alder trees generally have sparse and small-sized leaves (Jung et al., 2018), and their leaves have lower nitrogen and higher phosphorus concentrations compared with leaves from healthy trees (Ferreira et al., 2022). As aforementioned, these changes will reduce the amount of leaf litter inputs to streams and likely affect the instream cycling of litter-derived nutrients. Microbial-mediated leaf litter decomposition was faster for Iberian alder trees infected with P. alni complex than for healthy trees, probably because of the higher litter phosphorus concentration in diseased trees (Ferreira et al., 2022).

On the other hand, the mortality of alder trees by Phytophthora often generates landscape gaps, which trigger germination and proliferation of some exotic invasive species ( Figure 4C ), such as Acacia dealbata Link and A. melanoxylon R. Br. in specific areas (Portela-Pereira et al., 2022), leading to biodiversity homogenization. Dieback of mother trees induced by Phytophthora may hamper forest succession by reducing the rate of successful recruitment events, thus compromising the long-term sustainability of the community (Rodríguez-González et al., 2010). Moreover, soil and water infestation by Phytophthora species may impede the successful regeneration of alders, by producing damage to the embryo and radicle of seeds during germination, as reported in other Phytophthora-infested ecosystems (Martín-García et al., 2015; Figure 4D ). Thus, direct mortality along with indirect effects on the alder life cycle might lead to non-linear changes in community composition threatening biodiversity hosted by alder-dominated forests (Biurrun et al., 2021).

Identifying natural resistant genotypes and understanding the underlying mechanisms of resistance are essential for developing effective management and conservation strategies (Redondo et al., 2020). Resistance and tolerance responses of alders to Phytophthora infection are complex and multifaceted, involving a combination of genetic, physiological, biochemical and environmental factors (Gomes Marques et al., 2022; Avila-Quezada and Rai, 2023; Macháčová et al., 2024). Alders have physical barriers as structural defenses, like bark and lignified cell walls (Schmitt et al., 2021). However, Phytophthora can penetrate the host through its fine roots and then spread upwards to the trunk, or directly infect the trunk through existing wounds during flooding (Oßwald et al., 2014). This renders the physical barriers of the alder trees ineffective against infection by Phytophthora.

Alders have also developed defense mechanisms that define their tolerance or resistance and the level of pathogenicity of the oomycete. For example, trees can use tylose production, lignin deposition and/or callose production and deposition around sieve plates to prevent the infection of vascular systems (Redondo et al., 2015a; van den Berg et al., 2018). On alders, Phytophthora was found to affect the epidermis, cortex and vascular cylinder of roots, thus altering fibers and vessels, which leads to a detectable increase in tylose production (Vieites-Blanco et al., 2023). Indeed, some authors relate the presence of tyloses to a defense mechanism to hamper the advance of mycelium (Narayan et al., 2022; Martín and López, 2023). However, Vieites-Blanco et al. (2023) considered that the production of tyloses could be an indicator of damage rather than of plant resistance. In turn, callose formation in the crib plates is associated with plant resistance to Phytophthora. Thus, van den Berg et al. (2018) observed a lack of callose formation on susceptible roots with invasive hyphae, and Vieites-Blanco et al. (2023) detected differences in callose formation in alders infected with less aggressive pathogens vs. absence of callose with the more aggressive pathogens.

A direct relationship has been observed between the geographical distribution of Phytophthora species and subspecies and the climatic conditions. Efforts to create a comprehensive and reliable map of the global spread of various Phytophthora species have revealed that their distribution is related with the abiotic conditions of the studied areas. For example, in southern Sweden, a temperature-related barrier separates the survival of two species: P. uniformis, which can withstand lower temperatures, and P. ×alni which is only found in southern regions with milder climatic conditions (Redondo et al., 2015b, Redondo et al., 2020; Teshome et al., 2020).

In addition, biological barriers, such as microbial communities, are very important in enhancing plant defenses. Forest trees maintain close relationships with a wide variety of microorganisms that are essential for maintaining tree health, optimizing nutrient availability and supporting overall ecosystem functions (Fuller et al., 2023). Compared to crops planted yearly, the microbes that live around trees, both helpful and harmful, usually have more steady and stable interactions. This stability is largely due to the deep root systems of trees, which create a more stable environment for microbial communities (Mercado-Blanco et al., 2018). Beneficial microbiota, including plant growth-promoting rhizobacteria (PGPR), plant growth-promoting fungi (PGPF) and biocontrol agents, can significantly influence the metabolic processes of alder trees. These microorganisms enhance growth, improve performance and increase the trees’ resistance to various stresses in an economically efficient manner (Tian et al., 2020). Furthermore, in the search for resistant or tolerant genotypes to the pathogen Phytophthora, it was observed that progenies of alders from an area invaded by P. uniformis were less susceptible to P. ×alni than progenies from a pathogen-free zone. This could suggest an epigenetic regulation in some of the mechanisms of alder resistance to Phytophthora, as recently described in citrus (Rodrigues da Silva et al., 2021). Also, these studies provide valuable information, as they confirm that responses to infection can be inherited and highlight the importance of studying both surviving and healthy trees (not just the diseased ones), since they provide adaptation potential to the future generation. However, they also stress the need to study the stability of these responses over time and generations.

Through research, several potential biological control agents (BCAs) have been identified to reduce pathogenic spread and associated symptoms including dieback and root rot. The most commonly used BCAs belong to the genera Pseudomonas, Bacillus and Trichoderma. For instance, a study by Zaspel et al. (2014) demonstrated the advantages of root treatment with Pseudomonas veronii Coroler, Elomari, Hoste, Gillis, Izard, Leclerc for enhancing alder rooting. The study found that this bacterium can induce tolerance in some P. ×alni-infected alders, allowing them to survive without exhibiting symptoms caused by the pathogen. Moreover, while studies of BCAs on alder are limited, there are many registered microorganisms with antagonistic activity on other Phytophthora pathogens whose benefits have been successfully demonstrated. For example, Pseudomonas putida (Trevisan) Migula and P. chlororaphis (Guignard and Sauvageau, 1894; Bergey et al., 1930) were used as BCA of Phytophthora root rot in citrus orchids (Steddom et al., 2002) and in P. palmivora E.J. Butler infected-cacao plants, respectively (Acebo-Guerrero et al., 2015). In the same way, Bacillus amyloliquefaciens (Fukomoto) Priest, Goodfellow, Shute & Berkeley induced systemic resistance against P. cactorum (Lebert & Cohn) J. Schroüt (Lee et al., 2015). In addition, Trichoderma virens (J.H. Miller, Giddens & A.A. Foster) Arx, T. harzianum Rifai, T. asperellum Samuels, Lieckfeldt & Nirenberg and T. spirale Bissett showed antagonistic effects against P. palmivora in cacao plants (Mpika et al., 2009). T. saturnisporum Hammill also showed antagonistic effects against several Phytophthora spp (Diánez Martínez et al., 2016; Mercado-Blanco et al., 2018). These findings suggest a promising avenue for biocontrol against Phytophthora species, underscoring the importance of investigating the alder microbiome. Such research aims to identify potential microorganisms capable of inhibiting or parasitizing the pathogen decreasing its pathogenicity, thereby enhancing the host chances of survival and resilience (Redondo et al., 2020).

Forest tree breeding is a laborious and time-consuming process, strongly limited by the long breeding cycle of most tree species. An important factor in maintaining the long-term viability of the alder populations in Europe is the development or maintenance of natural resistance to Phytophthora. Additionally, the presence of asymptomatic alders surviving in highly affected areas could suggest the potential resistance to alder dieback in some common alder genotypes (Jung and Blaschke, 2004). Traditional approaches for breeding are based on the selection and mating of elite trees carrying the desirable traits. It is crucial to identify asymptomatic genotypes, which may be resistant, within natural populations affected by the pathogen. These genotypes should be tested for resistance under controlled conditions, typically using a combination of short- and long-term assessments. This includes planting the selected genotypes in areas with high disease pressure and varying environmental conditions (Sniezko, 2006; Sniezko and Koch, 2017). The confirmed resistant trees can then be used to develop a pool of resistant germplasm, for reforestation or restoration purposes, by clonal propagation and/or by crossing with other resistant trees (Keriö et al., 2019).

Forest biotechnology provides tools for conserving and sustainably managing of natural genetic resources, as well as for optimizing and speeding up genetic improvement programs. The development and use of biotechnological techniques provide, on the one hand, information for better management of natural resources, through the study of the genetic variability of forest tree resources and their functional characterization, employing molecular biology techniques (Díaz-Sala, 2014, 2019). On the other hand, in vitro culture techniques make possible large-scale plant propagation within breeding programs to, among others, preserve selected outstanding genotypes that are difficult to conserve by other methods. The application of tissue culture techniques has also led to the development of cryopreservation, as an additional strategy for conservation widely used in the agricultural sector. Vegetative propagation and in vitro propagation allow the development of clonal trials for phenotypic evaluation, the multiplication and maintenance of the genotypes of interest, the regeneration of a high number of plants in breeding programs and guarantee the health and availability of forest material quickly at any time and season of the year (Díaz-Sala, 2016).

Alders play significant ecological, commercial and recreational roles. However, a strong reduction of alder stands has been detected in Europe due to infection by Phytophthora spp. Continuous tree germplasm screening, effective procedures for the evaluation of the disease, accurate identification of Phytophthora species and identification of tolerant genotypes for large-scale propagation are crucial for the development of breeding programs that will preserve tree genetic diversity. The advances in genetic screening and genomic technologies, reference genomes and bioinformatic tools, including the promising application of Artificial Intelligence techniques in molecular biology (e.g. AlphaFold; Tunyasuvunakool et al., 2021; Abramson et al., 2024), as well as the genomics of Phytophtora-alder interactions and the novel high-throughput phenotyping techniques, will help to identify resistance gene(s), QTLs and pathogen effectors. This knowledge will allow the application of genomic-assisted breeding, gene silencing and gene editing to improve Phytophthora resistance in alder.