There is no overview of area distribution or changes in distribution of salt marshes at the Greenland scale. However, some local distribution studies exist (e.g., Glooschenko et al., 1993; Lepping and Daniëls, 2007; Bültmann and Daniëls, 2013), and salt marsh sediments have also been studied as archives of recent relative sea level change (Jensen et al., 2006; Woodroffe and Long, 2009; Long et al., 2012).

Salt marshes in Iceland are classified into two habitat types. One is characterized by Puccinellia maritima (Atlantic lower shore communities under the EU habitat type 1330). The other is classified by Carex species (Icelandic Carex lyngbyei salt meadows). The latter is not accepted under EUNIS but proposed as a sub-type of Atlantic black sedge salt meadows that falls under Annex I at a higher level (see A2.5 – https://eunis.eea.europa.eu/habitats/2931). Together, the habitat types cover approximately 65 km², with the majority characterized by P. maritima (Ottósson et al., 2016). The habitat mapping was done mainly by satellite image interpretations so estimation errors can be relatively high, especially for Carex habitats due to indistinct differences between similar habitats. No national monitoring of salt marshes exists in Iceland and long-term changes and dynamics are therefore unknown.

No overviews of distribution area or dynamics are available for the Faroe Islands (https://kort.foroyakort.fo/kort/). Salt marsh species such as Carex lyngbyei, Glyceria fluitans, G. maxima and Puccinellia maritima occur in the Faroe Islands but are all reported as very rare (reported 1–10 times) or rare (reported 11–25 times) (Jóhansen et al., 2000) indicating that salt marsh areas are negligeable in extent.

Norway has no systematic mapping program for salt marshes, but some mapping efforts have been completed, e.g. in connection to mapping RAMSAR areas. Typical species in Norway are Carex paleacea, C. vacillans, Glyceria fluitans, G. maxima, Phragmites australis, Schoenoplectus lacustris, S. tabernaemontani and Bolboschoenus maritimus. Currently, 606 occurrences of salt marshes have been recorded in Norway with a total area of 57 km² (Borgersen et al., 2020). This is most likely an underestimate due to the lack of systematic mapping. The habitat is considered as of “least concern” in Norway’s red list for habitats (Gundersen et al., 2018).

Danish salt marshes are represented by the EU habitat types 1310, 1320 and 1330 ( Table S4 ), with 1330 being the most widely distributed and categorized into two subtypes, salt marsh and salt swamp, depending on whether it is being grazed or not. The grazed plant communities are dominated by halophytic grasses and herbs, whereas the non-grazed ones are dominated by reeds such as Phragmites australis, Schoenoplectus maritimus and Schoenoplectus tabernaemontani. This division is the result of management practices with a long history of grazing preventing overgrowth by reed. The Wadden Sea has the largest tidal range among Danish salt marshes and includes typical tidal Spartina spp. salt marshes (EU habitat type 1320). Many of these marshes were formed as part of land reclamation efforts (Jakobsen, 1954). All habitat types are monitored and mapped every six years under the Danish monitoring program NOVANA since 2007, with the most recent estimate representing a total of 410 km² (sum of the area of the EU habitat types 1310, 1320 and 1330; Fredshavn et al., 2019; Table 1 , https://naturdata.miljoeportal.dk/). The saltmarsh area is estimated to have been twice as large as 150-200 years ago, with the loss caused primarily by expansion of agriculture (Vestergaard, 2000). EU habitat types 1310 and 1330 are categorized as having an unfavorable preservation status, the latter showing decline over the past 12 years, while 1320 has a favorable status (Fredshavn et al., 2019, Table S4 ).

The extent of Swedish salt marshes is estimated at 158 km² based on data and information from the report of conservation status of habitats and species that is provided by the implementation of the EU Habitats Directive (under Article 17), which is synthesized and published every six years by the Swedish Species Information Centre, ArtDatabanken. The estimates include the EU habitat types 1310, 1330 and 1630 ( Table S4 ), which are all categorized as having bad conservation status.

The majority of Finnish salt marshes belongs to the EU habitat type 1630 (Boreal Baltic coastal meadows). There are no overviews for habitat types 1310, 1320 or 1330 ( Table S4 ), but 1330 is part of a recent classification (Pätsch et al., 2019). Traditionally managed (grazing, mowing) salt marshes have been classified (by IUCN) and their status assessed (improving, declining, stable) in the national red list of threatened habitat types (Lehtomaa et al., 2018). These marshes belong to the habitat type 1630 and are currently classified into six types: seashore meadows Eleocharis parvula-E. acicularis, Eleocharis palustris-Schoenoplectus tabernaemontani-Bolboschoenus maritimus, tall-sedge seashore meadows, low-graminoid seashore meadows, tall seashore meadows and salt patches. Of a total of about 6900 km² of traditionally managed salt meadows in the 18^th century, only about 420 km² exist today (Lehtomaa et al., 2018). Reed belts have been favored by a dramatic decline in grazing by livestock as well as by eutrophication so the area of completely or partly preserved meadows of the habitat type 1630 is only 62 km², of which about 20 km² is overgrown by reed. Hence, the habitat type 1630 is now classified as critically endangered (Schulman et al., 2008). It is estimated that about 121 km² of this habitat type can potentially be restored, and there are ongoing management initiatives (including grazing). Coastal reed belts with Phragmites australis, Schoenoplectus and Bolboschoenus maritimus or Typha are additional potential BC sinks not belonging to the habitat type 1630.

Recent studies of eelgrass in the Nuuk fjord system at 64 °N show that the meadows are confined to inner, protected fjord branches with sandy seafloor where summer temperatures are relatively high (up to 13°C), although a sparse meadow grows in an area with maximum summer temperatures of 8°C (Olesen et al., 2015). These meadows support biomasses as high as those at lower latitudes but have lower productivity (Olesen et al., 2015). There is no assessment of the overall distribution area, but it is relatively small, limited by both availability of suitable sandy/soft seafloor in shallow protected settings and by low water temperatures (Olesen et al., 2015). Dating of sediments in eelgrass meadows in the Nuuk fjord system along with analyses of the origin of organic matter in these sediments, nevertheless, suggest that eelgrass has been expanding in these locations over the past century (Marbà et al., 2018).

Macrovegetation on littoral sediments include eelgrass beds, which are most common along the west coast and in lagoons in the south-east while rare elsewhere (Boström et al., 2014). They cover roughly 1% (11 km²) of the coast (Ottósson et al., 2016), but this is likely an underestimate as the distribution area has only been partially explored. Lack of monitoring and mapping efforts prevent a detailed account of the status and trends.

Eelgrass is very rare in Faroe Islands. It has been reported less than 10 times (Jóhansen et al., 2000) but with no recent observations. The extent is therefore considered negligeable with no overview of the distribution area or trend. Eelgrass occurs in one of the fjords on the southernmost island, Suðuroy, where Ostenfeld (1905-08) already a century ago commented that this fjord was likely the only place where the species occurred.

Eelgrass meadows are the most common and widespread habitat-building species on soft seafloor in Norwegian fjords and bays. The largest occurrences of eelgrass meadows have been systematically mapped by the Norwegian National program for mapping of coastal biodiversity (Bekkby et al., 2013), and are found in sheltered and moderately wave exposed areas all the way from the Swedish border in southern Norway to the Barents Sea region in Finnmark. More than 3000 eelgrass meadows have been identified in Norway, covering 60 km² in total (Bekkby et al., 2013). Following the “red list” protocol (Gundersen et al., 2018), this area was multiplied by a factor of 1.5 to include meadows assumed to be present but not recorded, giving an estimate of 90 km² (Frigstad et al., 2021). The habitat is not considered threatened in Norway’s red list for habitats (from 2018), but the status and trends in the North Sea in particular are largely unknown and highly debated, as the methods used for monitoring do not capture changes in meadow size and lower growth limit (Gundersen et al., 2018). Degradation of the seagrass meadows in the Oslofjord has been documented (Dahl et al., 2008; Espeland and Knutsen, 2014).

Denmark is the Nordic ‘hotspot’ for eelgrass due to the extensive coastline with gently sloping sandy shores in protected settings of intermediate salinity. The national Danish monitoring program assesses annually the cover and depth distribution of eelgrass across the country along transect lines from the shore to the deepest occurrence. Area surveys are not part of the monitoring program, except for the Wadden Sea eelgrass meadows. A first nationwide mapping of the submerged coastal vegetation (eelgrass, other rooted vegetation and macroalgae) has been produced based on optical satellite data from 2018, using advanced radiative transfer modelling and machine learning techniques. The map has 10 m resolution and represents depths down to 4–10 m, depending on local water clarity (http://satlas.dk/marine-vegetation/), but does not distinguish between eelgrass and macroalgae and provides yet no information on the distribution area. However, the potential distribution area of eelgrass in Danish coastal waters has been quantified at 2,204 km² based on a GIS modelling of where habitat requirements are fulfilled (Stæhr et al., 2019). This is about a third of the historic distribution area quantified at 6,726 km² around year 1900 (Petersen, 1914). Overviews for two key distribution areas (the Limfjord and Øresund) suggest that the current actual distribution area of eelgrass is probably only about 20% of the historic area, i.e. about 1,345 km² (Boström et al., 2014). Century-long records reveal shifting challenges to eelgrass over time with the wasting disease decimating the populations in the 1930s, eutrophication peaking in the 1980s causing additional decline, and bottom-trawling and warming further limiting the recovery during the past decades (Krause-Jensen et al., 2021).

The area distribution of eelgrass in Swedish coastal waters is assessed at 400 km² using a model based on satellite remote sensing images (2008 and 2016) and associated ground truthing (>6000 sites) for field validation (the Symphony tool, Hammar et al. 2018) ( Table 1 ). The spatial model described the probability of occurrence of seagrass (in 10 x 10 m pixels, aggregated into 250 x 250 m pixels); the probability level was set at 0.8, which means that there is at least 80% probability of finding the focal habitat in a pixel. The modelled coastal area encompasses primarily eelgrass (Z. marina), but also to some extent widgeon grass (Ruppia spp.) and occasionally some freshwater green algal species (principally along inner margins of Z. marina) (Boström et al., 2003). The model has been used efficiently in recent spatial risk assessment of global change impacts on eelgrass in Sweden (Perry et al., 2020). Satellite remote sensing has been shown useful to assess eelgrass coverage down to about 5-6 meters depth on the Swedish Skagerrak coast (Envall and Isaksson, 2012; Lundén and Gullström, 2013). In general, the cover of soft-bottom vegetation decreases with increasing total nitrogen concentrations and salinity (Wickström et al., 2016). Eelgrass meadows along the Swedish west coast experienced major declines over the period 1980s-2000s (Baden et al., 2003; Nyqvist et al., 2009) due to the combination of eutrophication and overfishing (Moksnes et al., 2008; Baden et al., 2010), and climate change is emerging as an additional stressor with increases in sea surface temperature, ocean acidification and wind-driven turbid conditions likely to cause interactive negative effects in risk areas in southern Sweden by the mid-end century (Perry et al., 2019; Perry et al., 2020).

There is no comprehensive area estimate of eelgrass coverage in Finland, but habitat suitability models suggest a maximum of 30 km² (Downie et al., 2013; Boström et al., 2014). As no national monitoring of eelgrass meadows exists, there are only few examples of long-term dynamics of Finnish eelgrass meadows. The available data suggest that these systems have been stable over time i.e. 1970-1990 (Boström et al., 2002). Finnish eelgrass meadows are often multi-specific, and eelgrass occurs here at the physiological boundary and is thus particularly vulnerable. Individual meadows are small, isolated clones with limited geneflow typically reaching over thousand years of age (Reusch et al., 1999; Reusch and Boström, 2011). The overall low reservoir of genetic diversity in the northern Baltic Sea combined with extremely limited sexual reproduction and inbreeding typically lead to reduced fitness (Reusch et al., 1999; Olsen et al., 2004). Thus, these populations are in urgent need of special conservation efforts as they lack potential for rapid adaptation and are likely to go extinct under extreme climate events.

Other submerged aquatic vegetation habitats that include plants of freshwater and/or marine origin are potentially more important BC habitats in Finland. These have been classified and their status assessed (Lehtomaa et al., 2018). They include habitats dominated by Potamogeton and/or Stuckenia pectinata, Ranunculus spp., Zannichellia spp. and/or Ruppia spp., Myriophyllum spicatum and/or M. sibiricum, and exposed and sheltered habitats characterized by Charales and Najas marina.

Macroalgae occur along most of Greenland’s coastline, but distribution records are relatively sparse, especially in the northernmost regions (Krause-Jensen et al., 2020). The depth limit and width of the belt of submerged macroalgae, dominated by kelps, increase from north towards south as the length of the open water period increases and more of the surface light potentially reaches the seafloor (Krause-Jensen et al., 2012). The deepest occurrence of kelps (> 60 m depth) is reported in clear waters along open coasts and offshore islands (Krause-Jensen et al., 2019). Intertidal macroalgae are also abundant along the Greenland coast, with potential for high biomasses even in northern regions (Høgslund et al., 2014; Ørberg et al., 2018; Sejr et al., 2021; Thyrring et al., 2021). Whereas no detailed mapping of macroalgae in Greenland exists, a first, coarse, spatial distribution model suggests a potential intertidal macroalgal distribution area of about 33,000 km² and a subtidal distribution area of 77,500 km² (Krause-Jensen et al., 2020). Also, Frigstad et al. (2021) estimated the total coverage of dense Saccharina latissima (sugar kelp) forests in Greenland to be 26,700 km² in a Nordic spatial model but emphasized that this is an undocumented estimate with high uncertainties due to unknown effects of sea ice cover, exposure, and substrate type. Macroalgal distribution and production is hypothesized to increase in parallel with increasing water temperatures and declining sea ice cover (Krause-Jensen et al., 2020).

Rocky shores dominated by macroalgae cover 28% (280 km²) of the intertidal in Iceland (Ottósson et al., 2016). Sheltered intertidal areas that are dominated by A. nodosum cover >7% of the total intertidal area and >25% of all low to moderate energy littoral rocky shores. Breiðafjörður in West Iceland accounts for about 70% of intertidal areas (50 out of 70 km²), dominated by A. nodosum according to habitat mapping done in 2012-2014 (vistgerdakort.ni.is, 2018). More recent, in-depth studies expand this estimate to 91-107 km² (Gunnarsson et al., 2019). Frigstad et al. (2021) estimated a total coverage of 495 km² for rockweed and 2,328 km² for kelps, dominated by Laminaria hyperborea (tangle kelp), based on a spatial model covering the Nordic countries. No national monitoring of macroalgae exists so long-term trends and dynamics are unknown.

Kelps are widespread on the rocky shores of the Faroe Islands, and data from the FARCOS project in the 1990s (Bruntse et al., 1999) were recently used to model the potential kelp distribution around the Faroe Islands, predicting a coverage of 10 km² for sugar kelp and 235 km² for tangle kelp ( Table 1 , Frigstad et al., 2021). No estimates exist for the coverage of macroalgae in the littoral zone of the Faroe Islands.

Rockweeds cover the intertidal shores and the shallow subtidal areas along almost the entire Norwegian coast, with a total coverage estimated (modelled) to be 3,090 km² (Frigstad et al., 2021). Tangle kelp and sugar kelp are the two dominant species that form large subtidal kelp forests along the Norwegian coast. Tangle kelp is found in the wave-exposed areas and has been mapped systematically by the National program since 2007 (Bekkby et al., 2013), estimated to cover 3,810 km² (Frigstad et al., 2021). Sugar kelp forests colonize more sheltered areas, often in fjords, bays and in inner parts of archipelagos, and are mapped less systematically. These forests have been estimated to cover 3,607 km² by Frigstad et al. (2021) and 5,355 km² by Gundersen et al. (2021). Both estimates are adjusted for destructive sea urchin grazing in northern Norway (additionally 1592 km² of kelp forest is considered to be grazed by urchins, Gundersen et al., 2021). Tangle kelp forests are in the Norwegian “Red list” for habitats considered to be “near threatened” in northern Norway due to the intensive grazing by the green sea urchin Strongylocentrotus droebachiensis (Rinde et al., 2014). Sugar kelp is listed as endangered due to sea urchin grazing in the north and due to a combination of eutrophication and increasing frequency and intensity of marine heat waves in the south of Norway (Gundersen et al., 2018; Filbee-Dexter et al., 2020). For sugar kelp in southern Norway, a combination of increased water temperature and increased riverine inputs of terrestrial organic matter (increased light attenuation and sedimentation) has negatively affected the lower growth depth limit over the last decades (Frigstad et al., 2018). This trend is leading to a decrease in macroalgal area coverage, decreased primary production, and a subsequent decrease in macroalgal ecosystem services, including macroalgae-driven C-sequestration (Filbee-Dexter and Wernberg, 2018; Wernberg et al., 2019).

Although most of the Danish seafloor is soft, macroalgae occur on scattered stones along the shores as well as on offshore stone reefs. As a component of the national Danish monitoring program, the cover and species composition of macroalgal communities are assessed nationwide along shoreline depth gradients representing hard substratum, as well as on offshore stone reefs, but a map of overall macroalgal distribution in Denmark is lacking. However, recently, the area extent of Fucus sp., which is dominant in the tidal zone, was quantified for the Kattegat coast at about 54 km² based on analysis of aerial photos supported by monitoring data and drone/field surveys in test areas (Riemann et al., 2020). This area was upscaled to an estimated distribution area of intertidal algae in Denmark at 397 km², assuming similar area extent of Fucus per km of coastline as for the Kattegat (Frigstad et al., 2021). The distribution area of subtidal macroalgae in the Kattegat has been assessed at 1438 km² based on monitoring data, seafloor properties and modelling (Öberg, 2006). Large uncertainties in the area of hard substratum have so far prevented a reliable updated modeling of the subtidal macroalgal area (Frigstad et al., 2021).

Several stress factors, varying in space and time, affect Danish macroalgae. Vast macroalgal habitats have been lost due to extraction of stones for construction, a conservative estimate suggests that a total of 40 km² of exposed stone surface was removed from stone reefs in coastal Danish waters in the second half on the 20^th Century (Dahl et al., 2003) before the practice was banned in 2010 (Dahl et al., 2016a). Bottom trawling for mussels and fish has added to the loss of stones. Eutrophication further limits the cover and diversity of both coastal and offshore stone reef macroalgal communities through light limitation and overgrowth by opportunistic species (Riemann et al., 2016).

The area of hard bottom within the photic zone was modelled at 869 km² and used as a proxy for the extent of subtidal macroalgae (Hammar et al. 2018). The hard bottom model is based on geological mapping surveys conducted by the Geological Survey of Sweden (SGU). The model was set to 0.8, which means that there is at least 80% probability of finding the focal habitat in a pixel. In the Nordic model of Frigstad et al. (2021), the total coverage of tangle kelp and sugar kelp in Sweden were estimated at 127 and 858 km², respectively, all at the Swedish west coast. The differences in area estimates for subtidal macroalgae by the two models can e.g. be due to different resolution of the models. For Fucus sp., the total coverage in Sweden was estimated at 1335 km² using the same approach as for Denmark, assuming similar area extent per km of coastline as for the Kattegat (Riemann et al., 2020), excluding the northern east coast of Sweden with known low distributions of Fucus sp. (Frigstad et al., 2021).

Until today, there are no comprehensive overall estimates of macroalgal distribution or change in cover in Swedish marine waters. But long-term studies on depth distribution (based on annual surveys) of macroalgal zonation, cover and community composition are available (e.g., Kautsky et al., 1986; Karlsson, 2007; Hammar et al. 2018) as well as scattered spatial surveys in localized areas (e.g., Blomqvist and Olsson, 2007; Gullström et al., 2009). A harmonized nation-wide macroalgae monitoring program is, however, ongoing since 2019, which aims to assess ecological status and spatiotemporal distribution of macroalgae on the west- and east coasts of Sweden (led by SWaM and following methodology in Lindegarth et al., 2016).

The Finnish Inventory Programme for the Underwater Marine Environment (VELMU) has systematically made inventories of macroalgae since 2011. Macroalgal monitoring started in 1993 and the monitoring of the lower depth limit of Fucus started in 2000. Based on habitat suitability modelling (GAM, EADM), the Finnish coastline (including Åland Islands) today supports about 240-580 km² of Fucus (including F. radicans) compared to >1000 km² in 1900 (Sahla et al., 2020).

The review identified the known area extent of BC habitats in the Nordic region, which was 1441 km² for salt marshes, 1,861 km² (potentially >2,735 km²) for seagrasses (mainly eelgrass) and 16,532 km² (potentially 130,735 km², including coarse Greenland estimates) for macroalgae. This sums up to 19,833 (potentially as much as 134,910 km²) of blue forests in the Nordic region ( Table 1 and Figure 2A ). Although the numbers are associated with considerable uncertainty and contain major knowledge gaps, the area of Nordic marine forests is substantial. The overview reflects the increasing difficulty of mapping the habitats as the water depth increases and conceals the vegetation. Hence, actual distribution maps are very limited and the difference between documented and potential areas is huge for macroalgal habitats, which extend deepest among the BC habitats. Especially for Greenland, Faroe Island and Iceland, there is very limited documentation of area extent of BC habitats. The various estimates also represent major differences in terms of quantification of presence/absence versus dense vegetation, key macroalgal species (e.g. sugar kelp) versus overall macroalgal cover and regarding the pixel size and scale of the survey. Only unified mapping approaches across the Nordic region, such as the one carried out by Frigstad et al. (2021) offer direct comparison across part of the region but depend on further data input for data poor regions, such as Greenland. Whereas there has been very limited Nordic focus on particularly salt marshes in the BC context, Nordic EU member states quantify salt marsh areas as part of the requirements of the EU habitats directive.

In addition to the above-listed areas, EU member states report the areas and status/trends of the EU habitats directive’s marine habitat types, several of which contain meadows of eelgrass and other rooted vegetation or macroalgae beds. However, in opposition to the directive’s continental habitat types, the marine habitat types are defined based on geomorphology (e.g. “Estuaries, habitat type 1130”, “Coastal lagoons, habitat type 1150” and “Large shallow inlets and bays, habitat type 1160”) and, hence, unfortunately, do not specifically target the vegetated habitats.

It is important to underline that the above estimates of the area of Nordic BC-habitats do not provide information on the extent of manageable areas with regard to restoration or protection against further loss and, hence, the extent of area that is relevant for climate change mitigation policies.

Plans to reduce eutrophication are widely implemented across the Nordic region, e.g. in response to the European directives (WFD, 2000; MFSD, 2008) and the HELCOM Baltic Sea Action Plan (https://helcom.fi/baltic-sea-action-plan/nutrient-reduction-scheme/targets/). In Denmark, for example, a first national Action Plan for the Aquatic Environment was enacted in 1987, followed by additional ones to achieve an overall objective of reducing total nitrogen and total phosphorus discharges by 50 and 80%, respectively. However, because 60% of the country is intensely cultivated, nutrient loadings remain a problem (Riemann et al., 2016).

MPAs in the Nordic region include sites designated under national laws as well as under the Natura 2000 network of nature protection areas. The latter encompasses Special Areas of Conservation under the Habitats Directive and Special Protection Areas under the Birds Directive ( Figure 2C ). Four of the Nordic MPAs are appointed UNESCO Marine World Heritage Sites: Surtsey (Iceland), West Norwegian Fjords – Geirangerfjord and Nærøyfjord (Norway), Wadden Sea (Germany/Netherlands/Denmark) and High Coast/Kvarken Archipelago (Finland) of which the first three support documented BC habitats (UNESCO, 2020).

The current proportion (in %) of “protected” marine area varies across the Nordic region with 4.5% in Norway [Meld. St. 29 (2020–2021) - regjeringen.no], 9.8% in Finland, 12.8% in Sweden and 15.2% in Denmark (https://biodiversity.europa.eu/countries/). However, the 2021 Nordic ministerial declaration on biodiversity, oceans and climate, the EU biodiversity strategy, the first draft of the CBD post-2020 global biodiversity and the High-level panel for a Sustainable Ocean Economy all support regional and global goals of 30% MPAs.

The actual protection of habitats within MPAs also varies considerably as the term “protection” is subject to interpretation. For the Baltic Sea in general, trawling remains a major stressor despite various controls (de Liedekere et al., 2020). In Denmark, fisheries have been allowed in most protected areas except in the Sound (Øresund) and except for a zone of 200 m surrounding stone reefs and bubble reefs, which implies that only about 2% of the marine areas are protected from trawling; however, a new marine plan aims to increase the targets. In NW Sweden, ~7% of the eelgrass area is negatively affected by small-scale construction of docks and marinas, even though habitats located less than 100 m from the shoreline are protected against exploitation by national law, and ~50% of the eelgrass area is located within protected areas (Eriander et al., 2017).

Under the Habitats Directive, salt marshes are designated a habitat type and are therefore more directly protected than eelgrass meadows and kelp forests, which just form part of the habitat types “Large shallow inlets and bays”, “estuaries” or “mudflats”. The Finnish Nature Conservation Act hence protects coastal meadows but not (yet) eelgrass meadows. In Iceland, national laws specify protection of mudflats and, thereby, eelgrass and salt marshes.

The harvest of wild kelp (Laminaria hyperborea) in Norway amounts to ~150,000 metric tons per year (Steen, 2018), representing ~0.3% of the estimated standing stock of living biomass (55 million metric tons, Frigstad et al., 2021; Gundersen et al., 2021). Usually, the Norwegian harvest area is ~6% of a regional kelp resource (Norderhaug et al., 2020), but can exceed 75% at monitoring points (Steen et al., 2018). The harvested yield has been stable for the past 30 years but is foreseen to increase as the number of commercial actors have recently doubled (from 1 to 2). Kelp harvest is regulated by law and regulations, and the Norwegian coast is divided into harvesting zones by latitude minutes (1 nautical mile), where each zone is open for harvest every 5-6 years, to allow for kelp regrowth. There is no upper limit on biomass harvest by regulations, and each zone can theoretically be depleted every 5 years. In practice, commercial actors prioritize areas with a high biomass per area and bottom conditions that fit the harvesting equipment. This often results in a patchy harvest. Harvest may be restricted based on the outcome of monitoring conducted half a year prior to planned harvesting. If the commercial demand is increasing in the future, and more efficient harvesting equipment is developed, macroalgae biomass, habitats and ecosystems may be severely impacted without further legislation.

The rockweed Ascophyllum nodosum is also harvested commercially in Norway, with yearly landings around 20,000 metric tons (Directorate of Fisheries, Norway, www.fiskeridir.no/tall-og-analyse/aapne-data). The Marine Resource Act does not regulate this harvesting, as it largely takes place in the intertidal, which is private property, and can be conducted with permission from the landowners.

In Iceland, the harvest of macroalgae is mainly confined to Breiðafjörður. The harvest of A. nodosum amounts to 15,000-20,000 metric tons per year and 4,000-7,000 metric tons for L. digitata and L. hyperborea (Ingólfsson, 2010; Gunnarsson et al., 2019). The Marine & Freshwater Research Institute has advised that the annual harvest 2018-2022 of A. nodosum in Breiðafjörður should not exceed 40,000 metric tons, or around 3% annually of estimated rockweed biomass in the area (Hafro.is, 2018).

There is limited tradition for commercial exploitation of wild macroalgae in Denmark, as well as in the Baltic Sea in general, but a significant harvest of pristine communities of the drifting red algae Furcellaria lumbricalis took place in the central Kattegat in the 1940-1960s and decimated the population (Schramm, 1998; Weinberger et al., 2020).

For salt marshes, overgrowth and shadowing by tall grass (e.g. reeds) is being managed by grazing and hay harvest to increase biodiversity. In Denmark, around two thirds of the total area of mapped salt marshes (2016-2019) is managed by such practices. In Finland, abandonment of traditional agricultural activities (Lehikoinen et al., 2017) has rendered boreal coastal meadows endangered (Schulman et al., 2008), and the conservation status under the EU Habitats Directive was assessed as unfavorable or bad in 2007 and 2013. However, due to new management efforts, the total area of Finnish salt marshes has now increased from 60 to 62 km², and further supporting initiatives are in place via national funding (strategic nature conservation, restoration and management programme HELMI, 2020–2030) and EU funding (LIFE-project CoastNet LIFE, 2018-2025).

As many former salt marsh areas were transformed to agriculture land by dikes, drainage and pumping, removal of these structures to reestablish the natural hydrology forms direct restoration measures. Such restoration links to “managed coastal realignment”, which is a new coastal protection strategy integrating coastal and nature protection by removing or abandoning coastal protection measures to reestablish natural processes and dynamics (Schernewski et al., 2018). There are examples from e.g. the southern Baltic Sea (Germany) and Denmark (Tryggelev Nor) (Karnauskaitė et al., 2018) citing the database “Our Coast” and the European Maritime Spatial Planning (MSP) platform database (www.msp-platform.eu). However, as saltmarshes are not necessarily positively conceived by the public, and engineering solutions to coastal protection are probably more generally acknowledged, there is a need for thorough public communication and involvement to build confidence and develop best practices (Stewart-Sinclair et al., 2020).

Over the past decade, several eelgrass restoration efforts have been undertaken in the Nordic region ( Figure 2D and Table S7 ). The activities have benefitted from international experience, which highlights (1) the need for careful site selection so that restoration takes place where eelgrass eelgrass used to grow and where habitat requirements are presently fulfilled, (2) the importance of sufficient spatial scale of the restoration to increase the chance of building resilience in the new patches, and (3) ensuring sufficient time perspective (years) for monitoring the effects of the restoration (Bayraktarov et al., 2016; van Katwijk et al., 2016; Orth et al., 2020). Persistent Nordic restoration efforts confirm these recommendations and provide best practices for site selection and full-scale restoration (e.g. Moksnes et al., 2016; Lange et al., 2022). The most successful Nordic eelgrass restorations have involved eelgrass transplants rather than seeds, and have sometimes involved anchoring of shoots, stabilization of sediments by protective structures, sandcapping, enclosures to avoid bioturbation, traps to reduce crabs, or co-restoration with blue mussels in order to facilitate eelgrass survival and growth (Moksnes et al., 2016; Gagnon et al., 2021; van der Heide et al., 2021, Flindt et al., 2022; Lange et al., 2022) ( Table S7 ). Target areas for restoration have been identified by modelling where eelgrass habitat requirements are fulfilled (e.g. Canal-Vergés et al., 2016; Flindt et al., 2016) supplemented with further site inspection and transplantation trials (Lange et al., 2022). However, because eelgrass restoration is labor intensive, costly and with no guarantee of success, protection of existing meadows is a management priority (Moksnes et al., 2016).

The main threats to macroalgae include but are not limited to removal of anchoring stones, aggressive fishery practices, and destructive grazing by sea urchins. In the cases of macroalgal habitat loss due to removal of stones, restoration of the habitat requires the establishment of new reefs. The Blue Reef project in the Kattegat (www.bluereef.dk) represents a successful example of macroalgal restoration. The experience gained from this and other projects led to a manual on best practices (Dahl et al., 2016a) and has inspired additional projects in Danish coastal waters (e.g. in Als Fjord and Limfjorden, and planned projects).

Restoration of macroalgae on barren grounds after destructive grazing by sea urchins has been carried out along the Norwegian coast by removing sea urchins, transplanting Laminaria hyperborea and Saccharina latissima kelp, planning sites of action taking interactions with other species, such as crabs, into account (Christie et al., 2019 and references therein). Based on this work, a strategy on how to recover kelp ecosystems from urchin barrens has been developed (Verbeek et al., 2021). “Green gravel” (i.e. small rocks seeded with kelp) has also recently been applied to restore sugar kelp (Fredriksen et al., 2020). The project MERCES has restored macroalgal habitats across Europe and concluded that active intervention (such as sea urchin removal) is required if the cause behind the habitat degradation and loss is not dealt with at a broader overall scale (Bekkby et al., 2020), e.g. by ensuring that intact fish populations control the sea urchin populations (Norderhaug et al., 2021). Internationally, Japan has the longest experience on macroalgal restoration including transplantation of kelp and removal of sea urchins, and their practice with country-wide restoration teams including fishermen (Watanuki et al., 2010; see also annex of Duarte et al., 2020) may inspire similar action in the Nordic region.

Macroalgal farming could potentially benefit wild macroalgal forests by leading to reduced harvest of wild macroalgae. Moreover, sustainable macroalgal farming may contribute associated ecosystem services, including uptake of excess nutrients, and supporting C-sequestration and reduced emissions, provided that the farming follows best practice sustainability standards (Duarte et al., 2022). Sustainability standards are needed to avoid potential negative effects of the farming such as competition of resources or physically damaging native habitats, and/or spreading of non-native strains (Campbell et al., 2019). Verified standards for C-accounting and C-offsetting connected with seaweed farming are also lacking and would require challenging and careful documentation (Hurd et al., 2022). The interest in the exploitation of macroalgae through harvest of living biomass and beach cast as well as via seaweed farming is on the rise both in the Nordic Atlantic region and the Baltic region, although the low salinity of the Baltic Sea presents suboptimal growth conditions for most species (Weinberger et al., 2020; Araújo et al., 2021). Among the Nordic countries, Norway has the largest number of seaweed farms (>27) and a large potential for macroalgal cultivation both inshore and offshore (Broch et al., 2019). The Norwegian seaweed cultivation industry is predicted to reach 4 million metric tons annually by 2030 and 20 million metric tons annually by 2050 (Olafsen et al., 2012). Macroalgal cultivation is also taking place in Denmark (6 farms), Faroe Island (2 farms), Greenland, Iceland and Sweden (one farm in each area) (Araújo et al., 2021).