Abstract
Kelp forests are experiencing extreme declines globally, yet local social-ecological factors may mediate their responses to climate perturbations. Using a combination of long-term field surveys (2016–2023) and satellite data (2004–2023), we examined the regional effects of the 2014–2016 marine heatwaves regime on kelp forest ecosystems along a 600 km latitudinal gradient in Baja California, Mexico. We documented three distinct geographical subregion trajectories: (1) a remarkable resilience of giant kelp Macrocystis pyrifera, at its thermal limits in the southern subregion, recovering 95% of the historical canopy area despite experiencing the highest temperatures; (2) understory-dominated state in the mid-subregion, with a 96% decrease in M. pyrifera canopy, where understory brown macroalgae filled the ecological niche left by declining giant kelp, maintained by moderate herbivore densities; and (3) a shift to an urchin barren alternative stable state in the northern subregion, characterized by a 2,485% increase in herbivore densities from 2016 to 2023 —dominated by purple sea urchin Strongylocentrotus purpuratus (44%)— and a 95% reduction in giant kelp canopy area compared to pre-warming baseline. The southern subregion’s resilience was associated with high predators’ (10.26 ± 5.63 per 60 m²) and low herbivores’ (25.20 ± 8.89 per 60 m²) densities, dominated by turban snails Megastraea spp. (83%) with S. purpuratus absent. In contrast, the northern subregion’s shift from kelp forests to urchin barrens coincided with low predator abundances (5.05 ± 2.32 per 60 m²), potentially linked to fragmented fisheries management. Our findings challenge the assumption that range-edge populations are inherently more vulnerable to warming and underscore how local biological factors and management approaches can either enhance or compromise ecosystem resilience to climate stress. These divergent trajectories suggest that while temperature acts as an initial driver of change, the persistence of alternative states is governed by ecological feedback involving predator-herbivore dynamics, herbivores’ assemblages, and management contexts, providing insights for developing climate-adaptive conservation strategies under intensifying global environmental change.
Introduction
Coastal ecosystems face mounting pressures from an expanding array of human activities, including coastal development, tourism, pollution, and fishing, which interact synergistically with climate-related stressors (; ; ; ). Among these vulnerable systems, kelp forests are experiencing extreme declines globally, driven by the compounding effects of climate change, overfishing, and intensified herbivore grazing, with cascading consequences for biodiversity and human societies (; ).
Foundational kelp species (order Laminariales) create complex three-dimensional habitats that support high biodiversity and ecosystem services, including coastal protection, marine carbon retention, and small-scale fisheries (; ; ; ). Recent estimates suggest that up to 40% of kelp forests have been lost globally in the past decades, with marine heatwaves (MHWs) —defined as discrete periods of anomalously warm water— emerging as a primary driver of this decline (; ). This vulnerability to thermal stress is particularly concerning as MHWs become more frequent and intense under climate change, potentially triggering widespread ecosystem reorganization (; ).
Kelp forests’ response to climate perturbations operates through multiple interacting mechanisms that can lead to alternative stable states (). These states typically manifest as either productive kelp-dominated or impoverished herbivore-dominated states (known as urchin barrens), maintained through distinct feedback mechanisms (; ; ). The maintenance of kelp-dominated states depends on top-down control, where predators regulate herbivore populations through consumptive and non-consumptive effects (; ; ), and the effectiveness of this predatory control is linked to fisheries management ().
Importantly, when kelp resources become scarce during disturbance events (e.g., MHWs), sea urchins can undergo behavioral shifts from passive detrital to active feeding, creating intense localized grazing that reduces habitat complexity and primary production (; ; ). These behavioral changes, along with reduced predation and substrate modifications, contribute to shifts in ecosystem dynamics. For example, crustose coralline algae can promote urchin recruitment and often inhibit kelp settlement, thereby reinforcing herbivore-dominated states (; ; ). However, these effects vary among coralline species ().
The 2014–2016 MHWs regime in the Northeastern Pacific provides a compelling example of how these mechanisms operate at different spatial scales (; ; ). These warming events triggered widespread kelp forest declines —up to 90% in some regions— and the formation of urchin barrens (). Moreover, the loss of key predators, including the sunflower sea star (Pycnopodia helianthoides, Brandt 1835) due to a massive die off, exacerbated by warming ocean conditions during this period (), amplified ecosystem vulnerability (; ; ; ; ). However, the extent and persistence of these impacts varied significantly across the region; hence, documenting regional and local (10 to 100s km) effects is key to understanding the impacts of extreme warming on the ecosystem.
The Baja California peninsula in Mexico was significantly impacted by these MHWs, with surface temperatures exceeding physiological thresholds for giant kelp (Macrocystis pyrifera, Agardh 1820) —the main species in forming kelp forests— reproduction and survival in many areas (; ). Despite the peninsula’s importance as a biogeographical transition zone with unique ecological characteristics, comprehensive long-term assessment of kelp forest responses in this region remains limited (; ; ). This knowledge gap is particularly critical given Baja California’s position at the southern distribution limit of M. pyrifera in the Northern Hemisphere, making it an ideal system to examine climate-driven ecosystem transitions ().
Notably, Baja California’s small-scale fisheries target key mesopredators such as spiny lobster and sheephead —important urchin predators that maintain top-down control in kelp forest ecosystems (). With nearly a century of operational history, these fisheries offer examples of relatively responsible and adaptive management, providing a valuable counterpoint to the widespread overfishing and destructive practices that have contributed to global kelp forest decline (; ). The confluence of these factors makes Baja California a pivotal location for understanding the complex interplay between climate change, ecological resilience, and human activities in shaping kelp forest ecosystems.
To better understand the geographic variability of factors shaping ecosystem responses, we examined the ecological community trajectories from the Mexico-U.S.A. border (henceforth northern subregion) to its southern thermal limit in the Northern Hemisphere (henceforth southern subregion), about 600 km of coast. Then, to understand the drivers shaping those trajectories, we analyzed the abundance of M. pyrifera and key functional groups across the region. This stretch encompasses different management in small-scale fisheries and species assemblages that create a natural experiment (; ).
For instance, the northern subregion faces challenges in fisheries management, characterized by fragmented oversight and limited surveillance (), while the southern subregion benefits from well-established fishing cooperatives demonstrating effective governance (; ). This management gradient is complemented by biological variation, particularly in herbivores’ assemblages, where the purple sea urchin (Strongylocentrotus purpuratus, Stimpson 1857) —a key species in forming and maintaining barrens (; )— is present in the northern subregion but at lower densities in the southern subregion at its distribution limit ().
To address this knowledge gap, we analyzed the subregional trajectories of kelp forest communities using underwater surveys (2016–2023) and satellite data (2004–2023) focusing on five research questions: (1) What are the temporal trajectories of kelp forest community composition across subregions following the 2014–16 MHWs regime, and how do these trajectories reflect ecosystem stability and potential alternative state shifts? (2) How did M. pyrifera canopy dynamics vary along the latitudinal gradient before, during, and after the MHWs regime, and did sea surface temperature contribute significantly to this variation? (3) How do key functional groups (canopy-forming M. pyrifera, understory brown macroalgae, herbivores, and mesopredators) interact across subregions, and what are their temporal dynamics following the warming event? (4) What are the relative contributions of these functional groups to community structure across subregions, and how has this relationship changed over time? (5) How does herbivore species composition vary across the latitudinal gradient, and what role might these assemblages have in kelp forests’ stability?
Materials and methods
Study area
Our study encompassed the Pacific coast of the Baja California peninsula (henceforth Baja California), spanning ~600 km from Islas Coronados (32.44°N) in Baja California to Puerto Escondido (27.53°N) in Baja California Sur (Figure 1). We defined three sampling geographical subregions based on latitude: northern (32.44–31.46°N; n = 9 sites), mid (30.51–29.69°N; n = 5 sites), and southern (27.92–27.52°N; n = 2 sites).
Figure 1
The northern subregion has two major cities (Tijuana and Ensenada) and is characterized by fragmented management with multiple stakeholders. Limited enforcement capacity in this area may contribute to poaching, particularly for high-value resources such as lobsters —a key sea urchin predator (; ).
The mid-subregion includes towns with different management strategies. San Quintín has multiple stakeholders, including fishing cooperatives, permit holders, sport fishing, and tourism operators. Also in the mid-subregion, we find El Rosario, which benefits from community-based management led by the fishing cooperative “Ensenada” that has ongoing Fishery Improvement Projects (see www.fisheryprogress.org, accessed February 2025) and three community marine reserves established in 2013: Punta Baja (7.15 km2), La Caracolera (5.64 km2), and Sportfish/China Town (2.82 km2) (; ).
The southern subregion features well-managed fishing cooperatives with over 80 years of experience, demonstrating effective governance and community-based conservation practices (; ; ; ; ). For example, the cooperative “Buzos y Pescadores” at Natividad Island, Baja California Sur, established marine reserves since 2006, Punta Prieta (0.7 km2) and La Plana (1.3 km2) with a continuous community-led monitoring program (). Finally, the cooperative “La Purísima” at Punta Eugenia, Baja California Sur, established a marine reserve (Piedra Blanca, 3.11 km2) in 2016 and, since then, has collaborated in a monitoring program led by the consortium Management of Ecosystems across the Californias (MexCal) from the Universidad Autónoma de Baja California (UABC) ().
Ecological community assessment
To analyze the in-situ ecological community structure, we conducted standardized underwater surveys from 2016 to 2023 using MexCal and Reef Check California protocols (; ). Each monitoring site was divided into subsites (Appendix S1: Table S1), where we identified, counted, and measured algae, invertebrates, and fish. We used 30 x 2 m belt transects (60 m2) for algae and invertebrates to estimate relative abundances for the different species. Within each transect, divers recorded all visible invertebrates >3 cm to species level when possible. For M. pyrifera, we quantified abundance through direct counts of all stripes per individual within the transect area. For fish, we surveyed a rectangle 30 m long, 2 m wide, and 2 m high (120 m3). We managed all class data using R on a rbeas/mexcal/github repository.
As part of our long-term monitoring effort, we aimed to visit each site at least once per year, approximately the same time each year (August–November). However, due to logistical, financial, and institutional constraints, the survey frequency differed by subregion, with more frequent (quarterly) visits to northern subregion sites and primarily annual surveys (August–November) at mid and southern subregions sites. Finally, in the mid-subregion, our monitoring program began in 2017, and in 2021, monitoring was impossible due to the COVID-19 pandemic.
Giant kelp canopy cover and sea surface temperature
We integrated satellite-derived measurements across multiple spatial and temporal scales for kelp cover and sea surface temperature (SST). For broad-scale temporal patterns (2004–2023), we obtained quarterly kelp canopy area measurements from Kelpwatch.org database following methodology, with atmospherically corrected 30 m resolution multispectral satellite imagery. For SST, we obtained quarterly satellite data from Copernicus Climate Change Services (C3S; https://climate.copernicus.eu/) following methodology, with values at 0.05° (~5.5 km) spatial resolution.
Kelp forest ecological community dynamics
We combined the in-situ and remote sensing information to assess kelp forest ecological community dynamics. This dual approach enabled the connection of long-term (2004–2023) canopy dynamics from satellite data with detailed community structure from field surveys (2016–2023) while simultaneously assessing kelp forest responses across subregions (satellite) and local (field surveys) spatial scales.
To validate the integration of satellite and field datasets, we tested correlations between field-measured giant kelp stipe counts and satellite-derived canopy area for overlapping years (2016–2023). We calculated Spearman correlations between annual mean stipe densities (per 60 m²) from underwater surveys and annual mean canopy area (km²) from satellite imagery, both aggregated by subregion and year. This cross-validation confirmed a strong positive relationship (r = 0.652, p = 0.001), supporting our integrated analytical approach.
Functional groups
We categorized species into four functional groups based on their cumulative presence, ecological roles, and trophic interactions in the Baja California kelp forest ecosystem ():
We assessed canopy-forming kelp using M. pyrifera stipe counts.
The understory brown macroalgae included six taxa: Southern palm kelp (Eisenia arborea, Areschoug 1876), Woody-stemmed kelp (Pterygophora californica, Ruprecht 1852), Osmunda kelp (Stephanocystis osmundacea, Trevisan 1843), various species of Sugar kelp (Laminaria spp.), Setchell’s rockweed (Stephanocystis setchellii, Gardner 1913), and Feather boa kelp (Egregia menziesii, Areschoug 1876) that form the sub-canopy habitat layer and potentially compete with giant kelp for settlement area (; ).
The herbivores group comprised six taxa known to consume macroalgae. Two temperate sea urchins: purple (S. purpuratus) and red (Mesocentrotus franciscanus, Agassiz 1863). Three tropical/subtropical sea urchins: crowned (Centrostephanus coronatus, Verrill 1867), black (Arbacia stellata, Blainville 1825), and slate pencil (Eucidaris thouarsii, Agassiz and Desor 1846). Moreover, turban snails (Megastraea spp.), included Wavy (M. undosa, Wood 1828) and Chestnut (M. turbanica, Dall 1910).
The predators group included three mesopredator species: California spiny lobster (Panulirus interruptus, Randall 1840), California sheephead (Bodianus pulcher, Ayres 1854), and horn shark (Heterodontus francisci, Girard 1854). These predators were selected based on their documented role in controlling herbivore populations through consumptive and non-consumptive effects (; ; ; ). The sunflower star, P. helianthoides, was historically present but disappeared from the region due to sea star wasting disease in 2014 and was not observed during our study period (2016–2023).
We also included a fifth category encompassing all remaining fish, invertebrates, and algae not classified in the four main functional groups, referred to as “other species,” to capture the whole community composition in our relative contribution analyses.
Statistical analysis
To capture vertical zonation patterns, at each site we surveyed at least two transects at three depth-stratified subsites: shallow (4–8 m), mid-depth (9–13 m), and deep (14–18 m). To account for this spatial structure in the analyses, we first calculated densities at the finest scale (individual transects) and then aggregated these to subsite means. We used these subsite-level estimates to calculate site-level means and final subregion estimates.
For community composition trajectories analysis, we employed non-metric multidimensional scaling (nMDS) to visualize spatiotemporal trajectories of community structure across subregions and years. We assessed the significance of temporal and spatial patterns in community composition using PERMANOVA with 999 permutations based on Bray-Curtis dissimilarity matrices calculated from square-root-transformed abundance data. Our model included year and subregion as fixed factors with their interaction term with year as random to test whether temporal trajectories differed among subregions. To account for the nested sampling structure where transects were repeatedly sampled within sites, we constrained permutations within sites. This approach properly accounts for non-independence of repeated sampling while testing for spatiotemporal differences in kelp forest community structure.
We employed a hierarchical analytical approach to examine temporal and spatial patterns in kelp forest communities. We defined 2014–16 as the MHWs regime, a period previously characterized in detail for this region (; ). We first conducted Kruskal-Wallis tests for kelp cover and SST to determine significant differences across the three time periods (before, during, and after the MHWs regime) and subregions (northern, mid, and southern). Following Kruskal-Wallis’s significant results, we performed post-hoc Dunn tests with Benjamini-Hochberg adjustment to account for multiple comparisons. This non-parametric approach was selected due to the non-normal data distribution and unequal sample sizes across periods (before n = 40, during n = 12, after n = 28). We conducted linear and quadratic trend analyses for satellite kelp cover data within each subregion to characterize temporal patterns across the study period.
To analyze the effect of functional groups on canopy-forming M. pyrifera densities, we performed Generalized Linear Mixed Models (GLMMs) with Poisson distribution. We ran three separate models examining the density effects of (1) predators, (2) herbivores, and (3) understory brown macroalgae on canopy-forming kelp. We included the subregion as a fixed effect in each model and tested its interaction with the predictor variable. We initially assessed the year as a potential factor, but it did not significantly affect any models and was therefore excluded from the final analysis. We included subsite as a random effect to account for spatial autocorrelation. We fitted the models using a log link function and assessed model fit using generalized chi-square statistics and examination of residuals.
To analyze community structure and the relative contribution of key functional groups, we used the four functional groups and the “other species” group. To determine the relative contribution of each group to the community, we first calculated total community density by adding the densities of all species observed within each transect. We then computed the percentage contribution of each group relative to this total community density at the transect level. These percentages were averaged across transects within subsites and, finally, across subregions and years.
We conducted statistical analyses using JMP Pro 18 (Statistical Discovery LLC, 2022) and R version 4.2.1 (R Core Team, 2024), with the package tydiverse for data management, vegan for multivariate analyses, and FSA for post hoc tests. We set significance at α = 0.05. We used the ggplot2 package for all figures.
Results
Subregional ecological community trajectories
Ecological communities showed clear spatial structuring along a north-south biogeographic gradient, with notable temporal variability on subregional patterns (Figure 2). Multivariate analyses revealed strong spatial differentiation, as evidenced by PERMANOVA (Subregion: R² = 18.0%, p < 0.001) and ANOSIM (Global R = 0.29, p < 0.001). Pairwise contrasts confirmed the strongest dissimilarity between the northern and southern subregions (R = 0.39, p = 0.001), moderate differentiation between the southern and mid subregions (R = 0.33, p = 0.002), and non-significant differences between the northern and mid subregions (R = 0.16, p = 0.065). Interannual variability explained a comparable amount of variance (Year: R² = 19.2%, p < 0.001), highlighting the dynamic responses of communities to environmental drivers.
Figure 2
The significant year × subregion interaction (F = 1.11, R² = 17.2%, p = 0.049) confirmed that subregions followed distinct temporal trajectories (Figure 2). Subregion-specific temporal analyses revealed that the southern subregion exhibited the strongest temporal dynamics (R² = 77.4%, p = 0.001), consistent with its initial decline and subsequent recovery pattern. The mid subregion showed moderate temporal change (R² = 73.1%, p = 0.045), reflecting its transition to understory dominance, while the northern subregion showed comparatively weaker temporal structuring (R² = 26.5%, p = 0.081), indicating stabilization in an alternative urchin barren state following the initial transition.
The northern subregion transitioned significantly from a kelp-dominated to an herbivore-dominated assemblage driven by S. purpuratus (24.1% contribution). This change is illustrated by nMDS trajectories indicating a gradual divergence from 2016 to 2023 (Figure 2). In contrast, the mid-subregion exhibited transitional dynamics, stabilizing around a fish-understory brown macroalgae equilibrium (Blacksmith fish (Chromis punctipinnis, Cooper 1863): 14.6%, E. arborea: 9.5%), and showing central clustering in nMDS plots, suggesting buffered variability.
The southern subregion maintained resilient kelp-dominated communities (M. pyrifera stipes: 22.1%), evidenced by cyclical nMDS patterns diverging from 2016 —the year without M. pyrifera (Figures 3, 4A)— and compositional distinctness from both mid and northern subregions. Despite varied dispersion among subregions (p = 0.024), the congruence across multivariate tests, species contributions, and nMDS trajectories highlighted strong spatial-temporal structuring (Appendix S1: Table S2). These patterns collectively illustrate ecosystem-specific responses and transitions to different alternative stable states driven by the interplay between biogeographic gradients and interannual environmental fluctuations: the north’s shift to urchin barrens, the mid’s brown understory macroalgae-dominated state, while the south’s recovered and maintained giant kelp persistence.
Figure 3
Figure 4
Subregional kelp canopy and SST
To assess differences in kelp canopy and SST, we divided the analyses into three periods: before (2004–2013), during (2014–2016), and after (2017–2023) the MHWs regime. Kelp canopy response varied across subregions (Figure 3; Table 1). The mid and southern subregions experienced sharp initial declines during the MHWs, while the northern subregion showed less pronounced immediate impacts.
Table 1
| Variable | Measure | Subregions | Combined analysis | ||
|---|---|---|---|---|---|
| Kelp Cover | northern | mid | southern | ||
| Historical mean ± SE (millions m²) | 2004–2013 | 2.43 ± 0.31 | 2.46 ± 1.04 | 2.73 ± 0.34 | – |
| 2023 mean ± SE (millions m²) | 2023 | 0.12 ± 0.10 | 0.09 ± 0.06 | 2.58 ± 1.02 | – |
| Net change from historical baseline | 2004–2013 vs 2023 | -94.9% | -96.4% | -5.4% | – |
| Period effects (K-W) | χ² (df = 2) | 13.71*** | 20.45*** | 16.15*** | 28.33*** |
| Trend pattern | Quadratic decline | Linear decline | Recovery | – | |
| SST | |||||
| Before period (mean °C ± SD) | 2004–2013 | 16.6 ± 1.8 | 16.2 ± 1.7 | 17.8 ± 1.6 | – |
| During period (mean °C ± SD) | 2014–2016 | 18.3 ± 0.9 | 18.1 ± 0.8 | 19.9 ± 0.7 | – |
| After period (mean °C ± SD) | 2017–2023 | 17.2 ± 1.2 | 16.9 ± 1.1 | 18.5 ± 1.0 | – |
| Period effects (K-W) | χ² (df = 2) | 7.60* | 9.46** | 7.39* | 21.17*** |
| Combined effects | |||||
| Subregional differences | χ² (df = 2) | Kelp: 12.13** SST: 24.75*** | |||
Statistical summary and quantitative responses of kelp cover and sea surface temperature across subregions and periods.
K-W, Kruskal-Wallis; SST, Sea Surface Temperature; df, degrees of freedom; SE, Standard Error; SD, Standard Deviation; Statistical significance: *p < 0.05, **p < 0.01, ***p < 0.001.
Recovery patterns diverged markedly after 2017. The southern subregion recovered rapidly, surpassing pre-warming coverage by 2018 and maintaining high levels through 2023, resulting in only a 5.4% net decline from historical baselines. In stark contrast, the mid and northern subregions experienced persistent declines through 2023, with 96.4% and 94.9% decreases from historical levels, respectively (Table 1).
SST patterns confirmed the regional extent of the 2014–16 MHWs regime, with all subregions experiencing significant warming during this period compared to pre-warming baselines (Table 1). The thermal impacts were most pronounced in the southern subregion, which peaked at 25.2°C, compared to 22.9°C and 22.8°C in the mid and northern subregions, respectively. Following the MHWs regime, temperatures returned to levels that were statistically indistinguishable from pre-warming conditions across all subregions, indicating that the extreme thermal stress was transient rather than a permanent climate shift (Figure 3; Table 1). Detailed statistical comparisons for both kelp cover and SST responses are provided in Appendix S1: Table S3.
Relationships between canopy-forming M. pyrifera and functional groups
We analyzed yearly mean (± SE) abundances in canopy-forming M. pyrifera (Figure 4A), herbivores (Figure 4B), understory brown macroalgae (Figure 4C), and predators (Figure 4D) across subregions. Our focus on detecting multi-year trajectories and alternative state transitions suggests that annual sampling is sufficient to capture the significant major community shifts documented in this study.
Herbivores
Herbivores’ densities showed a significant relationship with M. pyrifera stipes across subregions (β = 0.337, SE = 0.1466, p = 0.022), but this relationship varied by subregion (p <0.001). In the northern subregion, where herbivores’ densities increased by 2,485% from 2016 (25.25 ± 10.35 per 60 m²) to 2023 (625.67 ± 355.06 per 60 m²), we observed a negative association with M. pyrifera densities of 99.94 ± 43.31 stipes per 60 m2 (Figures 4A, B; Table 2). In contrast, the southern subregion maintained the lowest herbivores’ densities (25.20 ± 8.89 per 60 m²) and the highest M. pyrifera densities with 339.57 ± 95.33 stipes per 60 m2 (Figures 4A, B). The mid-subregion had intermediate herbivores (68.60 ± 20.73 per 60 m2) and the lowest M. pyrifera densities (52 ± 21.09 stipes per 60 m2) throughout the study period (Figures 4A, B). Herbivores showed a weak positive effect on understory brown macroalgae density (β = 0.284, SE = 0.5446, p = 0.603), with no significant regional variation in this relationship (p = 0.859) (Figures 4B, C).
Table 2
| Response | Predictor | Overall effect | P-value | Subregional effects (β ± SE) | Subregion × predictor | ||
|---|---|---|---|---|---|---|---|
| β ± SE | Northern | Mid | Southern | P-value | |||
| Canopy-forming M. pyrifera | Predators | 0.885 ± 0.266 | 0.001** | -0.574 ± 0.263* | -0.553 ± 0.321 | 1.306 ± 0.817 | < 0.001*** |
| Herbivores | 0.337 ± 0.147 | 0.022* | -0.760 ± 0.233** | 0.017 ± 0.407 | 1.690 ± 0.525** | <0.001*** | |
| Understory brown macroalgae | 0.566 ± 0.144 | <0.001*** | -0.789 ± 0.257** | -0.413 ± 0.313 | 0.935 ± 0.477* | <0.001*** | |
| Understory brown macroalgae | Predators | -0.114 ± 1.128 | 0.919 | -0.115 ± 0.181 | 0.382 ± 0.207 | -1.041 ± 0.357** | 0.032* |
| Herbivores | 0.284 ± 0.545 | 0.603 | -0.329 ± 0.340 | 0.269 ± 0.365 | -0.834 ± 0.167*** | 0.859 | |
The Generalized Linear Mixed Model results in subregional effects and relationships among functional groups in kelp forest communities.
All models included subsite as a random effect, variance = 0.455 ± 0.180; Significance levels: *p < 0.05, **p < 0.01, ***p < 0.001; SE, Standard Error.
Predators
Predators’ densities exhibited a positive relationship with M. pyrifera (β = 0.8848, SE = 0.2658, p = 0.001), with significant regional variation in this effect (p <0.001). The southern subregion maintained the highest predator densities (10.26 ± 5.63 per 60 m²) with peaks in 2019 (18.0 ± 6.76 per 60 m²), while the northern and mid subregions showed consistently lower predator abundances, with 5.05 ± 2.32 and 5.82 ± 3.89 per 60 m², respectively (Figure 3D). Predators did not influence understory algae density (β = -0.1144, SE = 1.1279, p = 0.919), and this relationship varied significantly across subregions (p = 0.032) (Figures 4A, D; Table 2).
Understory brown macroalgae
Understory brown macroalgae densities showed a significant positive relationship with M. pyrifera (Table 2; β = 0.5655, SE = 0.1444, p < 0.001), with regional variation in this effect (p < 0.001). This relationship was most pronounced in the mid-subregion, which maintained the highest understory densities (71.76 ± 14.28 per 60 m²) and peaked in 2022 (105.75 ± 9.06 per 60m²) before declining sharply in 2023 (20.94 ± 6.43 per 60 m²). The northern and southern subregions maintained comparatively lower understory densities throughout the study period, with 40.07 ± 30.54 and 28.98 ± 14.35 per 60 m², respectively (Figures 4A, C; Table 2).
Ecological community composition changes
Our analysis revealed distinct subregional patterns in community structure and temporal shifts in functional group dominance across the study period from 2016 to 2023 (Figure 5). In the southern subregion, M. pyrifera dominated the community (58.2 ± 8.6%), with particularly high dominance in 2020, reaching 83.7%. The other species group comprised 27.3 ± 7.4% of the community, while herbivores maintained low contributions (5.5 ± 0.7%). Predators maintained a modest presence (3.0 ± 0.6%) throughout the study.
Figure 5
In the mid-subregion, understory brown macroalgae and other species groups contributed the largest proportions (29.1 ± 5.8% and 34.2 ± 3.2%, respectively), while M. pyrifera averaged 15.0 ± 4.9%. Herbivores’ contribution showed significant temporal variation, averaging 19.1 ± 5.3% but increasing from 6% in 2017 to 28% by 2023. Predator populations remained stable but low (2.6 ± 0.3%).
The northern subregion exhibited the most dramatic temporal shift in community structure (Figure 5). While other species group averaged 34.4 ± 6.1% overall, herbivores’ dominance increased substantially from 6.4% in 2016 to 71.6% in 2023 (mean: 38.4 ± 8.6%). This shift was accompanied by variable M. pyrifera abundances, from 44% in 2017 to 9.1% in 2023, with an overall contribution of 17.2 ± 4.9%. Understory brown macroalgae presence was consistently low (7.8 ± 2.3%). Predators maintained a low overall abundance (2.2 ± 0.9%).
Herbivores’ assemblage composition
The herbivores’ assemblages (mean % ± SE) exhibited distinct subregional patterns in species composition and dominance (Figure 5). In the southern subregion, Megastraea spp. dominated the herbivores’ community (82.9 ± 4.12%), followed by C. coronatus (11.7 ± 2.59%). The warm-water species E. thouarsii was exclusively present in the south, contributing 3.93 ± 3.35%. The temperate sea urchin species, M. franciscanus and S. purpuratus, accounted for less than 0.4% (Figure 6).
Figure 6
The mid-subregion was also dominated by Megastraea spp. (71.6 ± 10.2%), followed by S. purpuratus and M. franciscanus, which contributed similarly to the herbivores’ assemblage (12.7 ± 5.24% and 11.1 ± 3.70%, respectively). Finally, C. coronatus remained a minor component of the community (4.54 ± 2.30%), while A. stellata was rarely observed (0.07 ± 0.07%) (Figure 6).
In the northern subregion, S. purpuratus was the dominant herbivore (43.6 ± 2.62%), followed by Megastraea spp. (34.3 ± 2.76%). Mesocentrotus franciscanus also maintained a notable presence (17.9 ± 3.18%), while C. coronatus and A. stellata contributed minimally (2.37 ± 0.71% and 1.84 ± 0.90%, respectively). The tropical species, E. thouarsii, was absent from the mid and northern subregions (Figure 6).
1 Discussion
Our study demonstrates how coastal ecosystems, such as kelp forests, exhibit diverse ecological community trajectories following extreme disturbances. Kelp forest community dynamics in Baja California presented distinct subregional trajectories following the 2014–16 MHWs regime (Figure 2). We used an ecological long-term, large-scale dataset from 2016 to 2023 across a 600 km latitudinal gradient. We identified three contrasting ecosystem responses: (1) the resilient kelp-dominated southern subregion, exhibiting remarkable post-disturbance recovery and persistence of M. pyrifera, despite facing the highest absolute temperatures at the onset of oceanographic changes; (2) the understory-dominated mid-subregion, showing overall little change in fish, macroalgae, and invertebrates; and (3) the herbivore-dominated northern subregion, demonstrating a clear phase shift to urchin barrens without kelp cover recovery despite the return of favorable temperature conditions.
These divergent trajectories suggest that while temperature acted as the initial driver of change, local biological factors played decisive roles in determining post-disturbance ecosystem states. The interaction among these drivers —herbivores’ densities and assemblages, predators’ abundances, and competition of understory brown macroalgae with canopy-forming M. pyrifera— created distinct kelp forest trajectories with unique structures and likely functions (; ). Alternative stable states are critical in understanding marine ecosystems (; ), and kelp forests have intricate dynamics of alternative stable states commonly characterized by lush forests or urchin barrens (; ). In this work, we analyzed the community assemblages that foster different kelp forest states in Baja California. Specifically, we investigated recovery trajectories post-disturbance, aiming to inform ecological restoration.
Synergistic drivers of change
While temperature was the initial driver of kelp decline during the 2014–16 MHWs regime, our results demonstrate that local biological factors influenced the post-disturbance trajectories. Despite all subregions facing temperatures exceeding known physiological thresholds for M. pyrifera reproduction (>23.6°C; ), subsequent recovery patterns varied dramatically. This divergence is particularly pronounced in the southern subregion, which experienced an average of 204.38 ± 68.18 days of MHWs per year during 2014–16 (), leading to the nearly complete loss of M. pyrifera during that period (Figures 3, 4A). This pattern aligns with previous findings that habitats and biological populations near tropical-temperate transition zones are more susceptible to warm-water anomalies (; ).
Contrary to expectations, the southern subregion exhibited the most pronounced post-disturbance recovery and long-term persistence of M. pyrifera despite initial declines (Figure 3, 4A). In contrast, the mid and northern subregions showed a continued decline from 2018 onward, with M. pyrifera canopy area decreasing by ~95% in 2023 compared to pre-warming baselines (2004–2013), even after temperature conditions returned to normal (Figure 3). These results contrast with previous findings indicating that many kelp species have been detrimentally affected at their warm distribution limits (; ), while populations at the range center or the cold-water range edge have been less affected ().
The persistence of alternative stable states long after the return of favorable temperature conditions provides evidence for self-reinforcing feedback mechanisms that extend beyond simple temperature effects. Our observations of regional variation in post-disturbance trajectories support the “stressor versus ecosystem-type hypothesis” (), suggesting that local biological and physical characteristics strongly influence ecosystem responses to perturbations. The contrasting outcomes we observed —rapid recovery in the south versus persistent degradation in the north— exemplify how similar climatic perturbations can trigger divergent ecosystem states depending on local biological (predation and competition) and social (fisheries management) contexts ().
2 Mechanisms driving kelp forest alternative stable states in Baja California
Our findings highlight four distinct mechanisms driving kelp forests’ alternative stable states in Baja California:
Herbivores’ densities and M. pyrifera decline
First, herbivores’ effects showed strong regional context-dependency, with the northern subregion’s 2,485% increase in herbivores’ densities from 2016 (25.25 ± 10.35 per 60 m²) to 2023 (625.67 ± 355.06 per 60 m²) (Figure 4B) coinciding with the 95% M. pyrifera cover decline (Figure 3). This pattern suggests a threshold effect in which intense herbivore pressure surpasses kelp resilience, aligning with global observations of tipping point-driven phase shifts and suggesting a transition to an alternative urchin barren stable state (; ). In contrast, the southern subregion maintained the lowest herbivore’ densities (25.20 ± 8.89 per 60 m²) and the highest M. pyrifera stipes densities with 339.57 ± 95.33 per 60 m2 (Figures 4A, B).
Predators’ densities and M. pyrifera abundances
Second, the positive relationship between predators’ densities and M. pyrifera abundances (β = 0.8848, p = 0.001) provides potential evidence for the role of top-down control in maintaining kelp forest stability. The significant regional variation in this relationship suggests that predatory effects are modulated by local social-ecological contexts, particularly fisheries management (; ). Similar patterns were documented in Tasmania and Aotearoa New Zealand, where predator abundances are impacted by fisheries management and strongly influence kelp-dominated or urchin barrens phase stability (; ; ; ).
Recent findings in our previous work () provide further insight into the relationship between predator biomass and fisheries management in Baja California. We reported exceptionally low abundances of lobster P. interruptus and horn sharks H. francisci in the northern subregion (0.006 ± 0.01 and 0 per 60 m2, respectively) across seven years of underwater surveys (2016–2022). Additionally, sheepheads B. pulcher had smaller sizes (20.44 ± 1.21 cm) in the northern subregion compared to the southern subregion (30.28 ± 1.28 cm). The severe depletion of key predators in the northern subregion appears to be associated with historically fragmented management approaches and persistent enforcement challenges (). This has likely contributed to sea urchin population expansion, as larger sheepheads and lobsters are particularly effective at controlling urchin populations (). Moreover, the documented decline in sheephead size is concerning, as larger individuals (>450 mm) consume significantly more urchins (38.4% of diet) compared to smaller fish (<200 mm, 2.8% of diet) (). Additionally, reported in southeastern Australia that Packhorse rock lobster (Sagmariasus verreauxi, Edwards 1851) had only a 4% predation rate on urchins, while crested horn sharks (Heterodontus galeatus, Günther 1870) had a 45%. This highlights the importance of maintaining diverse predator assemblages and the ecological consequences of the absence of horn sharks in the northern subregion. Consequently, the reduction in kelp (particularly drift kelp) availability during the 2014–16 MHWs regime, combined with the functional loss of predation linked to overfishing and disappearance of the sunflower star P. helianthoides, may have facilitated both the initial formation and subsequent persistence of urchin barrens in the northern subregion, even after environmental conditions improved (see. ; ).
3 Herbivores’ composition
Third, herbivores’ composition was crucial in shaping subregional trajectories (Figure 6). The northern subregion’s transition to urchin dominance was primarily driven by S. purpuratus (43.6 ± 2.62% SE of herbivore assemblage). Contrarily, the stability in the southern subregion was associated with the dominance of Megastraea spp. (82.9 ± 4.12% SE) and the absence of S. purpuratus. This difference in herbivores’ composition likely contributed to the southern subregion’s resilience and persistence of M. pyrifera, as Megastraea spp. exhibit less destructive feeding behaviors compared to S. purpuratus, which actively consumes M. pyrifera holdfasts and can detach even an adult (; ).
Importantly, our findings show that northern Baja California exhibits a more complex ecological configuration, where multiple herbivore species –not just S. purpuratus– interact synergistically to maintain the barren state (Figures 5, 6). For example, sea urchin species with tropical affinities, such as C. coronatus and A. stellata, increased in abundance in the northern subregion following the 2014–16 MHWs regime (Figure 6; ), a process known as tropicalization.
Lastly, the reduction in kelp forests is intrinsically linked to the sea urchin fisheries in Baja California (), as urchins depend on kelp availability to produce high-quality gonads. This reduction in fishing pressure aligns with findings and could have contributed to the increased densities of M. franciscanus and S. purpuratus in the northern subregion after 2018, corresponding to the last year of substantial giant kelp presence (Figures 3, 4B, 6). Nonetheless, further investigation is needed to fully assess this relationship, as M. franciscanus experienced mass mortality events during 2014–16 (), and this population may be undergoing natural recovery.
4 Understory brown macroalgae and M. pyrifera competition
Fourth, understory brown macroalgae had negative associations with M. pyrifera in the mid-subregion (β = -0.413 ± 0.313, p < 0.001), suggesting a shift in dominance following the 2014–16 MHWs regime disturbance. documented this suppression of competition and/or increased facilitation among understory algae following the 2014–16 MHWs regime in the Northeastern Pacific. Nonetheless, the high densities observed in the mid-subregion (69.9 ± 6.9 per 60 m²) suggest that understory algae may serve as ecological buffers, temporarily maintaining some ecosystem function during periods of reduced M. pyrifera cover (Figures 3, 4C). This pattern is consistent with the findings of , who described understory algae as a form of “ecological insurance” during canopy disturbance. However, understory algae compete with M. pyrifera for settlement substrate and may affect recovery (; ).
The persistence of this alternative stable state appears facilitated by moderate herbivores’ densities (72.4 ± 8.9 per 60 m²), likely maintained by moderate predator densities (5.82 ± 3.89 per 60 m²) preventing the system from transitioning to an herbivore-dominated state (Figure 4B). Our results add nuance to the traditional kelp-urchin dichotomy, indicating that alternative ecosystem stable states may exist along a continuum rather than as discrete conditions (). For example, documented a similar phenomenon on the East coast of the U.S.A., where reducing sea urchin abundances to almost zero individuals m-2 did not immediately result in kelp forest recovery due to the persistence of alternative algal assemblages.
In contrast, the northern subregion demonstrates how high herbivore pressure can prevent understory dominance even after canopy loss. The 2,485% increase in herbivores’ densities from 2016 to 2023 has created a similar scenario described by , where grazing intensity exceeds the recovery capacity of both canopy-forming and understory algae. The consistently low understory algae densities in this subregion (Figure 4C), despite reduced competition from M. pyrifera, suggest that herbivore control, rather than competitive release, is the dominant structuring force. Finally, the southern subregion’s stable but lower understory densities (Figure 4C) likely reflect competitive suppression by persistent M. pyrifera canopy, as documented in other stable kelp forest systems (; ).
Our findings should be interpreted within the context of certain methodological constraints. The varying sampling frequency across subregions (quarterly in the northern subregion versus annually in the mid and southern subregions) may have influenced our ability to detect short-term fluctuations in community structure, particularly in the mid and southern subregions. While logistical constraints necessitated this sampling design, it could have masked fine-scale temporal dynamics in these areas. However, our subregional differences persist at annual scales, suggesting that our primary conclusions about alternative stable state trajectories remain robust despite these sampling limitations.
Implications for conservation and management
Fisheries management
The contrasting subregional fisheries management contexts influence kelp forest trajectories through several mechanisms. In the northern subregion, historically fragmented management and limited enforcement capacity, leading to poaching (), coincide with extremely low biomass of sea urchin key predators such as lobster, sheephead, and horn shark compared to the southern subregion (). These patterns suggest that enhancing management approaches could result in higher predator biomass with cascading effects on trophic interactions. However, establishing direct causal links between management approaches and ecological outcomes requires additional long-term data on enforcement efforts, fishing pressure, and population dynamics (see ).
Community-led no-take marine reserves
Marine reserves (MRs) are an interesting tool for increasing predators’ abundance and other important species. For example, community-led MRs in Baja California have increased important economic resources such as abalone, lobster, sheephead, and horn sharks (; ; ). Moreover, marine reserves can serve as a buffer against climate change (). Importantly, we should identify and protect sites with persistent kelp forests or seabed topography favorable for retaining drift kelp () and climate microrefugia where species can survive during warm-water events (; ). However, some predators without economic importance may not benefit from MRs, and other tools should be explored. The apparent declining trend in predator densities across subregions in 2022–2023 (Figure 3) warrants continued monitoring, as further predator declines could compromise the stability of recovered kelp forests.
Enhancing predator populations
Not all predators are decreasing because of fisheries management. For example, the sunflower star P. helianthoides, which disappeared from Baja California due to a disease in 2014 and has not returned until today, may not directly benefit from MRs (; ; ). As P. helianthoides is key in predation on sea urchins and can potentially facilitate kelp forest recovery, we should focus on efforts to recover this top predator (). In that sense, additional conservation tools such as captive breeding and reintroduction may be interesting to explore. Enhancing and maintaining predator populations will ultimately be crucial to herbivore population control.
Herbivore control
If predators’ abundances are not high enough and management scenarios are too complex, human intervention in herbivore control could be explored. Herbivore control has proved effective in short-term kelp forest recovery (). Moreover, documented that in densities above 11 urchins m-2 —typical in urchin barrens— the primary cause of mortality was density-dependent and not predation. Thus, controlling urchin populations will be necessary for future management. Nonetheless, such control/removals’ scalability and economic feasibility require innovative biotechnological solutions to add value to sea urchins obtained from barrens.
Biotechnology innovation
Aquaculture represents an opportunity to aid in removing sea urchins from barrens and simultaneously address market demand for high-quality uni. Gonad enhancement of wild-caught adult sea urchins (known as sea urchin ranching) has been proposed as a restorative aquaculture strategy to generate economic benefits for coastal communities and induce the production of sustainable food, while macroalgae are restored in the local removal areas (; ; ; ). In that sense, implementing sustainable sea urchin harvesting programs in historical kelp forests that have switched to urchin barrens could serve dual purposes —controlling herbivore populations while creating sustainable blue economy opportunities through restorative aquaculture for coastal communities (). Nonetheless, developing environmental and economically viable feed formulations is the most critical factor for developing the sea urchin ranching industry ().
Active kelp forest restoration
Beyond predator enhancement and herbivore control, active restoration efforts represent a complementary approach for degraded kelp forest systems. Global experiences with kelp forest restoration have demonstrated the importance of addressing underlying stressors before implementing restoration activities (). In the Mediterranean, successful macroalgal forest restoration has highlighted the critical role of foundation species in ecosystem recovery (). For Baja California’s kelp forests, active restoration approaches such as kelp transplantation or spore seeding could be particularly valuable in the northern subregion, where natural recovery appears limited by persistent herbivore pressure. However, such efforts would need to be coupled with herbivore control and predator enhancement to address the underlying drivers maintaining the urchin barren state.
Integral marine zones/marine prosperity areas
Finally, “integral marine zones” that combine all the tools above could be explored (). Defining no-take MRs complemented with restorative aquaculture, ecotourism, active restoration, and predator reintroduction could enhance ecosystem resilience while supporting local economies. Similarly, the concept of Marine Prosperity Areas (MPpAs) offers a framework that integrates area-based conservation with human well-being through targeted socio-economic investments during ecological recovery periods (). This approach addresses the temporal mismatch between ecosystem recovery timelines and socio-economic needs by implementing strategic interventions across community engagement, capacity building, and collaborative management phases. Such integral approaches demonstrate how combining traditional marine protection with systematic investment in local prosperity can enhance both ecosystem resilience and community sustainability.
Future research priorities
Understanding the mechanisms of kelp forest resilience and vulnerability requires targeted research in several areas. First, the apparent thermal resilience of southern M. pyrifera populations warrants investigation through comparative genomic analyses and experimental thermal tolerance tests across the latitudinal gradient. Second, studying population connectivity patterns, such as spore dispersal of M. pyrifera and larval dynamics of herbivores and predators, is essential. Additionally, gametophyte banking and recruitment enhancement could support population recovery efforts. This knowledge is essential for predicting and managing ecosystem responses to disturbances, especially given the substantial subregional differences in community composition we documented.
Third, the success of present and future research initiatives should consider integrating multiple knowledge systems (). While our scientific monitoring has revealed subregional patterns in kelp forest dynamics, the successful implementation of conservation and/or economic strategies requires a transdisciplinary approach that bridges the gap between research and action. Integrating monitoring, local ecological knowledge, traditional management practices, and blue economy development can provide complementary insights into long-term ecological changes, while fostering stakeholder engagement necessary for successful conservation outcomes. This collaborative framework, combining rigorous scientific research with community-based knowledge and management, could create a path for preserving kelp forests and the ecosystem services they provide.
Conclusions
Our analysis of kelp forest dynamics along Baja California’s 600 km latitudinal gradient reveals complex, spatially variable responses to climate perturbations. Through integrated field surveys (2016–2023) and satellite data (2004–2023), we documented three distinct ecosystem trajectories following the 2014–2016 MHWs regime: (1) remarkable giant kelp M. pyrifera resilience in the southern subregion despite highest thermal stress, facilitated by higher predators’ abundances and herbivores’ assemblage dominated Megastraea spp.; (2) an understory-dominated alternative state in the mid-subregion; and (3) a persistent shift to urchin barrens in the northern subregion, characterized by a 95% decrease in kelp cover and a 2,485% increase in herbivores’ densities from 2016 to 2023, dominated by S. purpuratus.
These divergent trajectories suggest that while temperature initiated the ecosystem changes, local biological contexts were key in determining post-disturbance trajectories. The southern subregion’s well-established cooperative fisheries management coincided with higher predator biomass, while the northern subregion’s fragmented management aligned with depleted predator communities. These findings emphasize that local ecological interactions and management contexts strongly mediate kelp forest responses to climate perturbations.
Our research highlights the need for regionally tailored conservation and management approaches that account for biological factors (herbivores’ assemblages and predators’ densities) and social contexts (fisheries management). By identifying mechanisms that enhance or compromise ecosystem resilience to climate stress, we provide information for developing effective conservation strategies to maintain ecosystem function under increasing climate perturbations.
Statements
Data availability statement
The original contributions presented in the study are included in the article/Supplementary Material. Further inquiries can be directed to the corresponding author/s.
Author contributions
JB: Writing – original draft, Formal Analysis, Visualization, Methodology, Writing – review & editing, Investigation, Data curation, Conceptualization. JL: Investigation, Supervision, Funding acquisition, Conceptualization, Writing – review & editing, Resources, Data curation, Methodology, Project administration. RB-L: Funding acquisition, Resources, Writing – review & editing, Conceptualization, Project administration, Methodology, Supervision, Investigation. LM-C: Writing – review & editing, Funding acquisition, Supervision, Resources, Investigation, Project administration. AA-C: Funding acquisition, Investigation, Project administration, Writing – review & editing. AP-L: Investigation, Writing – review & editing. JO: Writing – review & editing, Investigation, Supervision, Funding acquisition, Conceptualization, Validation.
Funding
The author(s) declare that financial support was received for the research and/or publication of this article. This work was supported by a postdoctoral grant to JB from SECIHTI (CVU #860611). Part of the fieldwork for this work was supported by a UABC Research Grant to LM-C and AA-C (UABC – XXIII Convocatoria Interna de Apoyo a Proyectos de Investigación). Finally, we would like to extend our sincere gratitude to all the funding agencies that have partially supported this large-scale MexCal monitoring program. This includes UABC internal funds, UC-Mexus, Alianza MX, and San Diego Foundation.
Acknowledgments
We extend our sincere gratitude to the fishing communities who supported this research. We especially thank the Fishing Cooperative “Ensenada” at El Rosario, particularly Miguel Bracamontes Peralta, for his unconditional support. Also, the Fishing Cooperative “La Purísima” at Punta Eugenia, especially Juan Carlos Bonilla. We gratefully acknowledge the Managing Ecosystems Across the Californias (MExCAL) group and the Reef Check California team for their dedication to underwater surveys; your excellent teams of volunteers made this research possible. Special thanks to Daniel Díaz Guzmán for his leadership in field operations and data management, and Alejandro Castillo for data curation. We thank Elliot De la Cruz, Carolina Félix Loaiza, Fernanda Aldaco, Héctor Hernández, Carmen Valdez-Rojas, Daniela Amador, Julia Lara, Mónica Peralta, Alfonso Ferreira, Lucia Rodríguez, Nur Arafeh Dalmau, Guillermo Torres Moye, Ainoa Vilalta, and Paulina Filz for their invaluable contributions to underwater surveys. We acknowledge the reviewers who substantially contributed to the enhancement of this work.
Conflict of interest
The authors declare that the research was conducted in the absence of any commercial or financial relationships that could be construed as a potential conflict of interest.
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Supplementary material
The Supplementary Material for this article can be found online at: https://www.frontiersin.org/articles/10.3389/fmars.2025.1691156/full#supplementary-material
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Summary
Keywords
kelp forest resilience, marine heatwaves, Macrocystis pyrifera, Strongylocentrotus purpuratus, trophic cascades, ecological community, herbivores, ecological trajectories
Citation
Bauer J, Lorda J, Beas-Luna R, Malpica-Cruz L, Abadía-Cardoso A, Paz-Lacavex A and Olmos J (2026) Extreme marine heatwaves drive divergent kelp forest trajectories and alternative stable states. Front. Mar. Sci. 12:1691156. doi: 10.3389/fmars.2025.1691156
Received
22 August 2025
Accepted
15 October 2025
Published
11 March 2026
Volume
12 - 2025
Edited by
Yuqi Shan, Sichuan University, China
Reviewed by
Ettore Nepote, Marche Polytechnic University, Italy
Lianna Gendall, University of Victoria, Canada
Updates
Copyright
© 2026 Bauer, Lorda, Beas-Luna, Malpica-Cruz, Abadía-Cardoso, Paz-Lacavex and Olmos.
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*Correspondence: Rodrigo Beas-Luna, rbeas@uabc.edu.mx; Jorge Olmos, jolmos@cicese.mx
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