Deploy Seaweed Farming for Food involves cultivating seaweed (often called macroalgae) in the ocean for human consumption as a partial replacement for low-protein foods grown on land (e.g., grains, vegetables). This solution considers the emissions avoided by substituting one kilogram of low-protein food with one kilogram of seaweed. Current evidence suggests that farming seaweed for food could result in lower greenhouse gas emissions compared to some terrestrial crops. Advantages include the potential to reduce land-based agricultural impacts, improve water quality, and achieve globally meaningful climate impacts at a smaller spatial scale than growing seaweed for carbon removal by sinking (see Deploy Ocean Biomass Sinking). Disadvantages include potential adverse effects on marine ecosystems, uncertain climate benefits due to limited data on effectiveness, and opportunity costs if seaweed used for food could have delivered a greater climate impact in other emerging uses. Based on our assessment, we will “Keep Watching” this potential solution.
The overall effectiveness of seaweed cultivation for food as a climate solution remains uncertain. It could deliver climate benefits at modest cultivation scales while providing a useful end product. Expansion could also benefit land and food systems by reducing agricultural pressures, but it may introduce environmental trade-offs in the ocean that are not yet well understood. We will “Keep Watching” this solution.
| Plausible | Could it work? | Yes |
|---|---|---|
| Ready | Is it ready? | Yes |
| Evidence | Are there data to evaluate it? | Limited |
| Effective | Does it consistently work? | ? |
| Impact | Is it big enough to matter? | ? |
| Risk | Is it risky or harmful? | ? |
| Cost | Is it cheap? | ? |
This solution involves expanding the cultivation of marine seaweed for human consumption as an alternative to higher-emission, lower-protein foods. These can include grains (e.g., wheat, rye, maize, oats, and rice) and, to a lesser extent, vegetables (e.g., potatoes, cassava, broccoli, and cabbage). By switching part of food production from land to ocean systems, seaweed farming helps avoid some sources of terrestrial agricultural emissions, such as those from fertilizer, irrigation, and soil disturbance. Seaweed cultivation, at modest scales and in suitable locations, does not require additional nutrients or irrigation, which can result in lower emissions. Emissions from physical cultivation activities also differ, with tractor use in land-based agriculture being replaced by emissions from boat operations in seaweed farming. Currently, roughly 80% of cultivated seaweed is consumed by humans in food products.
The climate impact of cultivating seaweed is understudied, but existing estimates suggest that growing a ton of seaweed generates less than a quarter of the emissions from growing a ton of vegetables, such as broccoli and cabbage. The actual climate impact will depend on which types of foods are displaced in diets. Replacing higher-emission, low-protein foods, such as some grain-based staples (e.g., bread or rice), with seaweed could provide even greater climate benefits. More data are needed to assess full cradle-to-grave emissions for seaweed that include transport, processing, and storage prior to consumption. Actual benefits may be lower once full life cycle emissions are considered, or higher if seaweed replaces more emissions-intensive foods.
Unlike terrestrial crops, seaweed cultivation does not require fresh water for irrigation or pesticides for pest management. It is the fastest-growing sector of global aquaculture, and can produce higher biomass yields per area than some land-based crops. Because it grows in the ocean, seaweed farming reduces land demand, which can therefore support terrestrial biodiversity and conservation efforts. If deployed in the right place, seaweed cultivation can also help reduce nutrient pollution in coastal areas.
Compared with other seaweed-based climate solutions, farming seaweed for food could achieve a meaningful global climate impact using far less ocean area (1–2 Mha versus 6–7 Mha estimated for solutions like Deploy Ocean Biomass Sinking), though estimates remain highly uncertain. Cultivation might also provide additional carbon removal benefits by selecting for high-productivity cultivars and strategically placing farms in areas where carbon fixation and burial are naturally high.
Finally, global diets currently overrely on starch-rich grain crops, highlighting a potential opportunity for seaweed, which is a nutritious source of protein, essential fatty acids, and minerals, to replace these foods and diversify diets in many regions. Across commonly consumed species, seaweeds are generally low in fat and calories and can be rich in fiber and micronutrients, including iron, iodine, calcium, and magnesium.
There are several environmental and feasibility concerns associated with seaweed cultivation, especially if it is expanded to use large areas of ocean habitat. Global estimates of ocean area suitable for seaweed cultivation range substantially, from 10 to 4,800 Mha, but often lack consideration for real-world nutrient limitations or ecological impacts. A more recent analysis that considers nitrogen, phosphorus, and iron limitations suggests that the viable ocean seaweed farming area is closer to 400 Mha. If regions are prioritized based on where cultivation is not nutrient-limited, where it can achieve high carbon removal efficiency, and where there are lower risks of adverse ecological impacts, potential seaweed farming areas could be limited to the western North Pacific and North Atlantic. The costs of such an expansion are also poorly understood, with some cost estimates per ton of CO₂
fairly high.
Similarly, it’s unclear how viable seaweed is as a large-scale substitute for low-protein foods in real-world diets. Using vegetables as an example, achieving a climate impact of at least 0.1 Gt CO₂‑eq/yr
could require replacing over 25% of global vegetable production. Assuming productivity typical of subtidal seaweed (6.6 t C/ha/yr), this would translate to an additional ~2.6 Mha of ocean cultivation. An area of 2.6 Mha would equate to a 100-meter-wide continuous belt of seaweed cultivation along 22% of the global coastline. For comparison, seaweed cultivation currently covers less than 400,000 ha.
At large scales, seaweed cultivation could alter food webs through competition with phytoplankton for nutrients and/or requiring external nutrient inputs, raising serious concerns similar to Deploy Ocean Biomass Sinking. Cultivation can have a range of other negative impacts on coastal ecosystems, too. Seaweed farms established in or near seagrass beds, for instance, can displace existing habitats and species. More research is needed to assess these trade-offs, including the spatial scale required for a globally meaningful climate impact and how seaweed cultivation relates to potential land-use benefits. Further work is also needed to evaluate whether seaweed cultivation could deliver greater climate benefits through other emerging products, rather than as a direct food replacement.
Berger, M., Kwiatkowski, L., Bopp, L., & Ho, D. T. (2025). Efficacy of seaweed-based carbon dioxide removal reduced by iron limitation and nutrient competition with phytoplankton. CDRXIV. Link to source: https://doi.org/10.70212/cdrxiv.2025385.v1
Bhuyan, M. S. (2023). Ecological risks associated with seaweed cultivation and identifying risk minimization approaches. Algal Research, 69, 102967. Link to source: https://doi.org/10.1016/j.algal.2022.102967
DeAngelo, J., Saenz, B. T., Arzeno-Soltero, I. B., Frieder, C. A., Long, M. C., Hamman, J., Davis, K. A., & Davis, S. J. (2023). Economic and biophysical limits to seaweed farming for climate change mitigation. Nature Plants, 9(1), 45-57. Link to source: https://doi.org/10.1038/s41477-022-01305-9
EAT-Lancet Commission. (2025). Healthy diets from sustainable food systems: Summary report of the EAT-Lancet Commission. EAT. Link to source: https://eatforum.org/wp-content/uploads/2025/09/EAT-Lancet_Commission_Summary_Report.pdf
Food and Agriculture Organization of the United Nations. (2021). Global seaweeds and microalgae production, 1950–2019: WAPI factsheet. Link to source: https://openknowledge.fao.org/server/api/core/bitstreams/97409d09-2f8e-4712-b11e-60105d89959b/content
Food and Agriculture Organization of the United Nations. (2023). Agricultural production statistics 2000–2022. Link to source: https://openknowledge.fao.org/server/api/core/bitstreams/fba4ef43-422c-4d73-886e-3016ff47df52/content
Froehlich, H. E., Afflerbach, J. C., Frazier, M., & Halpern, B. S. (2019). Blue growth potential to mitigate climate change through seaweed offsetting. Current Biology, 29(18), 3087-3093. Link to source: https://doi.org/10.1016/j.cub.2019.07.041
Hasselström, L., & Thomas, J. B. E. (2022). A critical review of the life cycle climate impact in seaweed value chains to support carbon accounting and blue carbon financing. Cleaner Environmental Systems, 6, 100093. Link to source: https://doi.org/10.1016/j.cesys.2022.100093
Jones, B. L. H., Eklöf, J. S., Unsworth, R. K. F., Coals, L., Christianen, M. J. A., Clifton, J., Cullen-Unsworth, L. C., de la Torre-Castro, M., Esteban, N., Huxham, M., Jiddawi, N. S., McKenzie, L. J., Nakaoka, M., Nordlund, L. M., Ooi, J. L. S., & Prathep, A. (2025). Risks of habitat loss from seaweed cultivation within seagrass. Proceedings of the National Academy of Sciences, 122(8), Article e2426971122. Link to source: https://doi.org/10.1073/pnas.2426971122
Lomartire, S., Marques, J. C., & Gonçalves, A. M. (2021). An overview to the health benefits of seaweeds consumption. Marine Drugs, 19(6), 341. Link to source: https://doi.org/10.3390/md19060341
Lozano Muñoz, I., & Díaz, N. F. (2020). Minerals in edible seaweed: Health benefits and food safety issues. Critical Reviews in Food Science and Nutrition, 62(6), 1592-1607. Link to source: https://doi.org/10.1080/10408398.2020.1844637
Peñalver, R., Lorenzo, J. M., Ros, G., Amarowicz, R., Pateiro, M., & Nieto, G. (2020). Seaweeds as a functional ingredient for a healthy diet. Marine Drugs, 18(6), 301. Link to source: https://doi.org/10.3390/md18060301
Pessarrodona, A., Assis, J., Filbee-Dexter, K., Burrows, M. T., Gattuso, J. P., Duarte, C. M., Krause-Jensen, D., Moore, P. J., Smale, D. A., & Wernberg, T. (2022). Global seaweed productivity. Science Advances, 8(37), eabn2465. Link to source: https://doi.org/10.1126/sciadv.abn2465
Pessarrodona, A., Howard, J., Pidgeon, E., Wernberg, T., & Filbee-Dexter, K. (2024). Carbon removal and climate change mitigation by seaweed farming: A state of knowledge review. Science of the Total Environment, 918, 170525. Link to source: https://doi.org/10.1016/j.scitotenv.2024.170525
Rajapakse, N., & Kim, S. K. (2011). Nutritional and digestive health benefits of seaweed. Advances in Food and Nutrition Research, 64, 17-28. Link to source: https://doi.org/10.1016/B978-0-12-387669-0.00002-8
Ross, F., Tarbuck, P., & Macreadie, P. I. (2022). Seaweed afforestation at large-scales exclusively for carbon sequestration: Critical assessment of risks, viability and the state of knowledge. Frontiers in Marine Science, 9, 1015612. Link to source: https://doi.org/10.3389/fmars.2022.1015612
Spillias, S., Valin, H., Batka, M., Sperling, F., Havlík, P., Leclère, D., Cottrell, R. S., O’Brien, K. R., & McDonald-Madden, E. (2023). Reducing global land-use pressures with seaweed farming. Nature Sustainability, 6(4), 380–390. Link to source: https://doi.org/10.1038/s41893-022-01043-y
Zhang, L., Liao, W., Huang, Y., Wen, Y., Chu, Y., & Zhao, C. (2022). Global seaweed farming and processing in the past 20 years. Food Production, Processing and Nutrition, 4(1), 23. Link to source: https://doi.org/10.1186/s43014-022-00103-2
We define the Deploy Silvopasture solution as the adoption of agroforestry practices that add trees to grazing land, including planted pastures and natural rangelands. (Note that this solution does NOT include creating forested grazing land by thinning existing forest; this is a form of deforestation and not desirable in terms of climate.) Some silvopastures are open savannas, while others are dense, mature tree plantations. The trees may be planted or managed to naturally regenerate. Some silvopasture systems have been practiced for thousands of years, while others have been recently developed. All provide shade to livestock; in some systems, the trees feed livestock, produce timber or crops for human consumption, or provide other benefits. New adoption is estimated from the 2025 level as a baseline which is therefore set to zero.
In silvopasture systems, trees are planted or allowed to naturally regenerate on existing pasture or rangeland. Tree density is generally less than forest, allowing sunlight through for good forage growth.
Silvopasture has multiple climate impacts, though carbon sequestration is the only one which has been thoroughly studied across all climates and sub-practices.
Silvopasture sequesters carbon in both soil and woody biomass. Carbon sequestration rates are among the highest of any farming system (Toensmeier, 2017). The lifetime accumulation of carbon in both soils and biomass is higher than for managed grazing alone (Montagnini et al., 2019; Nair et al., 2012).
Silvopasture can also reduce GHG emissions, though not in every case. We do not include emissions reductions in this analysis.
Conversion from pasture to silvopasture slightly increases capture and storage of methane in soils (Bentrup and Shi, in press). In addition, in fodder subtypes of silvopasture systems, ruminant livestock consume tree leaves or pods. Many, but not all, of the tree species used in these systems have tannin content that reduces emissions of methane from enteric fermentation (Jacobsen et al., 2019).
Some subtypes of silvopasture reduce nitrous oxide emissions from manure and urine, as grasses and trees capture nitrogen that microbes would otherwise convert to nitrous oxide. There are also reductions to nitrous oxide emissions from soils: 76–95% in temperate silvopastures and 16–89% in tropical-intensive silvopastures (Ansari et al., 2023; Murguietio et al., 2016).
Many silvopasture systems increase productivity of milk and meat. Yield increases can reduce emissions from deforestation by growing more food on existing farmland, but in some cases can actually worsen emissions if farmers clear forests to adopt the profitable practice (Intergovernmental Panel on Climate Change [IPCC], 2019). The yield impact of silvopasture varies with tree density, climate, system type, and whether the yields of other products (e.g., timber) are counted as well (Rojas et al., 2022).
Ansari, J., Udawatta, R. P., & Anderson, S. H. (2022). Soil nitrous oxide emission from agroforestry, rowcrop, grassland and forests in North America: a review. Agroforestry Systems, 97(8), 1465–1479. Link to source: https://doi.org/10.1007/s10457-023-00870-y
Basche, A., Tully, K., Álvarez-Berríos, N. L., Reyes, J., Lengnick, L., Brown, T., Moore, J. M., Schattman, R. E., Johnson, L. K., & Roesch-McNally, G. (2020). Evaluating the untapped potential of US conservation investments to improve soil and environmental health. Frontiers in Sustainable Food Systems, 4, 547876. Link to source: https://doi.org/10.3389/fsufs.2020.547876
Batcheler, M., Smith, M. M., Swanson, M. E., Ostrom, M., & Carpenter-Boggs, L. (2024). Assessing silvopasture management as a strategy to reduce fuel loads and mitigate wildfire risk. Scientific Reports, 14(1), 5954. Link to source: https://doi.org/10.1038/s41598-024-56104-3
Bentrup, G. & Shi, X. (in press). Multifunctional buffers: Design guidelines for buffers, corridors and greenways. USDA Forest Service.
Bostedt, G., Hörnell, A., & Nyberg, G. (2016). Agroforestry extension and dietary diversity–an analysis of the importance of fruit and vegetable consumption in West Pokot, Kenya. Food Security, 8, 271–284. Link to source: https://doi.org/10.1007/s12571-015-0542-x
Briske, D. D., Vetter, S., Coetsee, C., & Turner, M. D. (2024). Rangeland afforestation is not a natural climate solution. Frontiers in Ecology and the Environment. Link to source: https://doi.org/10.1002/fee.2727
Cadavid, Z., & BE, S. T. (2020). Sistemas silvopastoriles: aspectos teóricos y prácticos. CIPAV. Link to source: https://cipav.org.co/sdm_downloads/sistemas-silvopastoriles-aspectos-teoricos-y-practicos/
Cardinael, R., Umulisa, V., Toudert, A., Olivier, A., Bockel, L., & Bernoux, M. (2019). Revisiting IPCC Tier 1 coefficients for soil organic and biomass carbon storage in agroforestry systems. Environmental Research Letters, 13(12), 124020. Link to source: https://doi.org/10.1088/1748-9326/aaeb5f
Chapman, M., Walker, W.S., Cook-Patton, S.C., Ellis, P.W., Farina, M., Griscom, B.W., & Baccani, A. (2019). Large climate mitigation potential from adding trees to agricultural lands Global Change Biology, 26(80), 4357–4365. Link to source: https://doi.org/10.1111/gcb.15121
Chatterjee, N., Nair, P. R., Chakraborty, S., & Nair, V. D. (2018). Changes in soil carbon stocks across the forest-agroforest-agriculture/pasture continuum in various agroecological regions: A meta-analysis. Agriculture, ecosystems & environment, 266, 55–67. Link to source: https://doi.org/10.1016/j.agee.2018.07.014
Damania, Richard; Polasky, Stephen; Ruckelshaus, Mary; Russ, Jason; Amann, Markus; Chaplin-Kramer, Rebecca; Gerber, James; Hawthorne, Peter; Heger, Martin Philipp; Mamun, Saleh; Ruta, Giovanni; Schmitt, Rafael; Smith, Jeffrey; Vogl, Adrian; Wagner, Fabian; Zaveri, Esha. (2023). Nature's Frontiers: Achieving Sustainability, Efficiency, and Prosperity with Natural Capital. Environment and Sustainable Development series. Washington, DC: World Bank Link to source: https://hdl.handle.net/10986/39453
de Sherbinin, A., VanWey, L. K., McSweeney, K., Aggarwal, R., Barbieri, A., Henry, S., Hunter, L. M., Twine, W., & Walker, R. (2008). Rural household demographics, livelihoods and the environment. Global Environmental Change, 18(1), 38–53. Link to source: https://doi.org/10.1016/j.gloenvcha.2007.05.005
Den Herder, M., Moreno, G., Mosquera-Losada, R. M., Palma, J. H., Sidiropoulou, A., Freijanes, J. J. S., & Burgess, P. J. (2017). Current extent and stratification of agroforestry in the European Union. Agriculture, Ecosystems & Environment, 241, 121–132. Link to source: https://doi.org/10.1016/j.agee.2017.03.005
Di Prima, S., Wright, E. P., Sharma, I. K., Syurina, E., & Broerse, J. E. W. (2022). Implementation and scale-up of nutrition-sensitive agriculture in low- and middle-income countries: A systematic review of what works, what doesn’t work and why. Global Food Security, 32, 100595. Link to source: https://doi.org/10.1016/j.gfs.2021.100595
deStefano, A, & Jacobson, M.G. (2018). Soil carbon sequestration in agroforestry systems: A review. Agroforestry Systems, 92, 285–299. Link to source: https://doi.org/10.1007/s13593-014-0212-y
Dudley, N., Eufemia, L., Fleckenstein, M., Periago, M. E., Petersen, I., & Timmers, J. F. (2020). Grasslands and savannahs in the UN Decade on Ecosystem Restoration. Restoration Ecology, 28(6), 1313–1317. Link to source: https://doi.org/10.1111/rec.13272
Dupraz, C, and Liagre, F.(2011). Agroforesterie: Des Arbres et des Cultures. Editions France Agricole. Link to source: https://agroboutique.com/agroecologie-catalogue/12-agroforesterie-des-arbres-et-des-cultures.html
FAO Statistical Service (2024). FAOStat. Link to source: https://www.fao.org/faostat/en/
Feliciano, D., Ledo, A., Hillier, J., & Nayak, D. R. (2018). Which agroforestry options give the greatest soil and above ground carbon benefits in different world regions?. Agriculture, ecosystems & environment, 254, 117–129. Link to source: https://doi.org/10.1016/j.agee.2017.11.032
Frelat, R., Lopez-Ridaura, S., Giller, K. E., Herrero, M., Douxchamps, S., Djurfeldt, A. A., Erenstein, O., Henderson, B., Kassie, M., Paul, B. K., Rigolot, C., Ritzema, R. S., Rodriguez, D., Van Asten, P. J. A., & Van Wijk, M. T. (2016). Drivers of household food availability in sub-Saharan Africa based on big data from small farms. Proceedings of the National Academy of Sciences of the United States of America, 113(2), 458–463. Link to source: https://doi.org/10.1073/pnas.1518384112
Garrett, H. E., Kerley, M. S., Ladyman, K. P., Walter, W. D., Godsey, L. D., Van Sambeek, J. W., & Brauer, D. K. (2004). Hardwood silvopasture management in North America. In New Vistas in Agroforestry: A Compendium for 1st World Congress of Agroforestry, 2004 (pp. 21–33). Springer Netherlands. Link to source: https://doi.org/10.1007/978-94-017-2424-1_2
Goracci, J., & Camilli, F. (2024). Agroforestry and animal husbandry. IntechOpen. Link to source: https://doi.org/10.5772/intechopen.1006711
Government of Colombia (2020). Actualización de la Contribución Determinada a Nivel Nacional de Colombia. Government of Colombia. Link to source: https://unfccc.int/sites/default/files/NDC/2022-06/NDC%20actualizada%20de%20Colombia.pdf
Greene, H., Kazanski, C. E., Kaufman, J., Steinberg, E., Johnson, K., Cook-Patton, S. C., & Fargione, J. (2023). Silvopasture offers climate change mitigation and profit potential for farmers in the eastern United States. Frontiers in Sustainable Food Systems, 7, 1158459. Link to source: https://doi.org/10.3389/fsufs.2023.1158459
Hart, D.R.T, Yeo, S, Almaraz, M, Beillouin, D, Cardinael, R, Garcia, E, Kay, S, Lovell, S.T., Rosenstock, T.S., Sprenkle-Hyppolite, S, Stolle, F, Suber, M, Thapa, B, Wood, S & Cook-Patton, S.C (2023). “Priority science can accelerate agroforestry as a natural climate solution”. Nature Climate Change. Link to source: https://doi.org/10.5281/zenodo.8209212
Husak, A. L., & Grado, S. C. (2002). Monetary benefits in a southern silvopastoral system. Southern Journal of Applied Forestry, 26(3), 159–164. Link to source: https://doi.org/10.1093/sjaf/26.3.159
IPCC (2019). Climate Change and Land: an IPCC special report on climate change, desertification, land degradation, sustainable land management, food security, and greenhouse gas fluxes in terrestrial ecosystems [P.R. Shukla, J. Skea, E. Calvo Buendia, V. Masson-Delmotte, H.-O. Pörtner, D. C. Roberts, P. Zhai, R. Slade, S. Connors, R. van Diemen, M. Ferrat, E. Haughey, S. Luz, S. Neogi, M. Pathak, J. Petzold, J. Portugal Pereira, P. Vyas, E. Huntley, K. Kissick, M. Belkacemi, J. Malley, (eds.)]. Link to source: https://www.ipcc.ch/srccl/
IPCC AR6 WG3 (2022). Climate Change 2022: Mitigation of Climate Change. Contribution of Working Group III to the Sixth Assessment Report of the Intergovernmental Panel on Climate Change [P.R. Shukla, J. Skea, R. Slade, A. Al Khourdajie, R. van Diemen, D. McCollum, M. Pathak, S. Some, P. Vyas, R. Fradera, M. Belkacemi, A. Hasija, G. Lisboa, S. Luz, J. Malley, (eds.)]. Cambridge University Press, Cambridge, UK and New York, NY, USA. Link to source: https://doi.org/10.1017/9781009157926
Jacobsen (2019). Secondary metabolites in leaf hay as a mitigation option for enteric methane production in ruminants. Aarhus University. Link to source: https://pure.au.dk/ws/portalfiles/portal/197235590/Secondary_Metabolites_in_Leaf_Hay_as_a_Mitigation_Option_for_Enteric_Methane_Production_in_Ruminants.pdf
Jose, S., & Dollinger, J. (2019). Silvopasture: a sustainable livestock production system. Agroforestry systems, 93, 1-9. Link to source: https://doi.org/10.1007/s10457-019-00366-8
Lal, R., Smith, P., Jungkunst, H. F., Mitsch, W. J., Lehmann, J., Nair, P. R., & Ravindranath, N. H. (2018). The carbon sequestration potential of terrestrial ecosystems. Journal of soil and water conservation, 73(6), 145A–152A. Link to source: https://doi.org/10.2489/jswc.73.6.145A
Lee, S., Bonatti, M, Löhe, K, Palacios, V., Lana, M.A., and Sieber, S (2020). Adoption potentials and barriers of silvopastoral systems in Colombia: Case of Cundinamarca region. Cogent Environmental Science 6(1). Link to source: https://doi.org/10.1080/23311843.2020.1823632
Lemes, A. P., Garcia, A. R., Pezzopane, J. R. M., Brandão, F. Z., Watanabe, Y. F., Cooke, R. F., Sponchiado, M., Paz, C. C. P., Camplesi, A. C., Binelli, M., & Gimenes, L. U. (2021). Silvopastoral system is an alternative to improve animal welfare and productive performance in meat production systems. Scientific Reports, 11(1), 14092. Link to source: https://doi.org/10.1038/s41598-021-93609-7
Lorenz, K., & Lal, R. (2018). Carbon sequestration in agricultural ecosystems. Springer, Cham. Link to source: https://link.springer.com/book/10.1007/978-3-319-92318-5
Mehrabi, Z., Tong, K., Fortin, J., Stanimirova, R., Friedl, M., & Ramankutty, N. (2024). Global agricultural lands in the year 2015. Earth System Science Data Discussions, 2024, 1–44. Link to source: https://doi.org/10.5194/essd-2024-279
Montagnini, F (2019). Función de los sistemas agroforestales en la adaptación y mitigación del cambio climático. Sistemas agroforestales: Funciones productivas, socioeconómicas y ambientales, 269-299. Link to source: https://cipav.org.co/wp-content/uploads/2020/08/sistemas-agroforestales-funciones-productivas-socioeconomicas-y-ambientales.pdf
Morena, G., & Rolo, V. (2019). Agroforestry practices: Silvopastoralism. In Agroforestry for sustainable agricultura (1st ed.). Burleigh Dodds Science Publishing. Link to source: https://doi.org/10.1201/9780429275500
Murgueitio, E., Uribe, F., Molina, C., Molina, E., Galindo, W., Chará, J., & González, J. (2016). Establecimiento y manejo de sistemas silvopastoriles intensivos con Leucaena. Editorial CIPAV, Cali, Colombia. Link to source: https://www.researchgate.net/profile/Juan-Naranjo-R/publication/310460876_Establecimiento_y_manejo_de_sistemas_silvopastoriles_intensivos_con_leucaena/links/582e30cb08ae138f1c01d8b9/Establecimiento-y-manejo-de-sistemas-silvopastoriles-intensivos-con-leucaena.pdf
Nair, P.K. R. (2012). Climate change mitigation: A low-hanging fruit of agroforestry. Agroforestry: The future of global land use, 31–69. Link to source: https://doi.org/10.1007/978-94-007-4676-3_7
Ortiz, J., Neira, P., Panichini, M., Curaqueo, G., Stolpe, N. B., Zagal, E., & Gupta, S. R. (2023). Silvopastoral systems on degraded lands for soil carbon sequestration and climate change mitigation. Agroforestry for Sustainable Intensification of Agriculture in Asia and Africa, 207–242. Link to source: https://doi.org/10.1007/978-981-19-4602-8_7
Pent, G. J. (2020). Over-yielding in temperate silvopastures: a meta-analysis. Agroforestry Systems, 94(5), 1741–1758. Link to source: https://doi.org/10.1007/s10457-020-00494-6
Pezo, D., Ríos, N., Ibrahim, M., & Gómez, M. (2018). Silvopastoral systems for intensifying cattle production and enhancing forest cover: the case of Costa Rica. Washington, DC: World Bank. Link to source: https://www.profor.info/sites/default/files/Silvopastoral%2520systems_Case%2520Study_LEAVES_2018.pdf
Poudel, S., Pent, G., & Fike, J. (2024). Silvopastures: Benefits, past efforts, challenges, and future prospects in the United States. Agronomy, 14(7), 1369. Link to source: https://doi.org/10.3390/agronomy14071369
Quandt, A, Neufeldt, G, & Gorman, K (2023). Climate change adaptation through agroforestry: Opportunities and gaps. Current Opinion in Environmental Sustainability. 60, 101244. Link to source: https://doi.org/10.1016/j.cosust.2022.101244
Rivera, J. E., Serna, L., Arango, J., Barahona, R., Murgueitio, E., Torres, C. F., & Chará, J. (2023). Silvopastoral systems and their role in climate change mitigation and Nationally Determined Contributions in Latin America. In Silvopastoral systems of Meso America and Northern South America (pp. 25–53). Cham: Springer International Publishing. Link to source: https://doi.org/10.1007/978-3-031-43063-3_2
Rojas, D, & Rodriguez Anido, N. (2022) Potential of silvopastoral systems for the mitigation of greenhouse gasses generated in the production of bovine meat. In Sistemas silvopastoriles: Hacia una diversificación sostenible. CIPAV. Link to source: https://cipav.org.co/sistemas-silvopastoriles-hacia-una-diversificacion-sostenible/
Riset, J.Å., Tømmervik, H. & Forbes, B.C. (2019). Sustainable and resilient reindeer herding. Reindeer Caribou Health Dis, (23–43). Link to source: https://www.researchgate.net/publication/344787755_Ch13_Sustainable_and_resilient_reindeer_herding
Shelton, M., Dalzell, S., Tomkins, N. and Buck, S. R. (2021). Leucaena: The productive and sustainable forage legume. University of Queensland. Link to source: https://era.dpi.qld.gov.au/id/eprint/9425/
Shi, L., Feng, W., Xu, J., & Kuzyakov, Y. (2018). Agroforestry systems: Meta‐analysis of soil carbon stocks, sequestration processes, and future potentials. Land Degradation & Development, 29(11), 3886–3897. Link to source: https://doi.org/10.1002/ldr.3136
Smith, M. M., Bentrup, G., Kellerman, T., MacFarland, K., Straight, R., Ameyaw, L., & Stein, S. (2022). Silvopasture in the USA: A systematic review of natural resource professional and producer-reported benefits, challenges, and management activities. Agriculture, Ecosystems & Environment, 326, 107818. Link to source: https://doi.org/10.1016/j.agee.2021.107818
Sprenkle-Hyppolite, S. Griscom, B., Griffey, V., Munshi, E., Chapman, M. (2024). Maximizing tree carbon in cropland and grazing lands while sustaining yields. Carbon Balance and Management 19:23. Link to source: https://doi.org/10.1186/s13021-024-00268-y
Toensmeier, E. (2017). Perennial staple crops and agroforestry for climate change mitigation. Integrating landscapes: Agroforestry for biodiversity conservation and food sovereignty, 439-451. Link to source: https://doi.org/10.1007/978-3-319-69371-2_18
U.S. Department of Agriculture Natural Resources Conservation Service [USDA NRCS]. (2025). Conservation Practice Physical Effects [Dataset]. Link to source: https://www.nrcs.usda.gov/resources/guides-and-instructions/conservation-practice-physical-effects
Udawatta, R. P., Walter, D., & Jose, S. (2022). Carbon sequestration by forests and agroforests: A reality check for the United States. Carbon footprints, 1(8). Link to source: https://doi.org/10.20517/cf.2022.06
Zeppetello, L. R. V., Cook-Patton, S. C., Parsons, L. A., Wolff, N. H., Kroeger, T., Battisti, D. S., Bettles, J., Spector, J. T., Balakumar, A., & Masuda, Y. J. (2022). Consistent cooling benefits of silvopasture in the tropics. Nature communications, 13(1), 708. Link to source: https://doi.org/10.1038/s41467-022-28388-4
Zhu, X., Liu, W., Chen, J., Bruijnzeel, L. A., Mao, Z., Yang, X., Cardinael, R., Meng, F.-R., Sidle, R. C., Seitz, S., Nair, V. D., Nanko, K., Zou, X., Chen, C., & Jiang, X. J. (2020). Reductions in water, soil and nutrient losses and pesticide pollution in agroforestry practices: A review of evidence and processes. Plant and Soil, 453(1–2), 45–86. Link to source: https://doi.org/10.1007/s11104-019-04377-3
Eric Toensmeier
Ruthie Burrows, Ph.D.
Yusuf Jameel, Ph.D.
Daniel Jasper
Aiyana Bodi
Hannah Henkin
Ted Otte
Paul C. West, Ph.D.
We found a median carbon sequestration rate of 9.81 t CO₂‑eq /ha/yr (Table 1). This is based on an above-ground biomass (tree trunks and branches) accumulation rate of 6.43 t CO₂‑eq /ha/yr and a below-ground biomass (roots) accumulation rate of 1.61 t CO₂‑eq /ha/yr using a root-to-shoot ratio of 0.25 (Cardinael et al., 2019). These are added to the soil organic carbon sequestration rate of 1.76 t CO₂‑eq /ha/yr to create the combined total.
Table 1. Effectiveness at carbon sequestration.
Unit: t CO₂-eq/ha/yr, 100-yr basis
| 25th percentile | 4.91 |
| Mean | 14.70 |
| Median (50th percentile) | 9.81 |
| 75th percentile | 20.45 |
100-yr basis
Reductions in nitrous oxide and methane and sustainable intensification impacts are not yet quantifiable to the degree that they can be used in climate mitigation projections.
Because baseline grazing systems are already extensive and well established, we assumed there is no cost to establish new baseline grazing land. In the absence of global data sets on costs and revenues of grazing systems, we used a global average profit per hectare of grazing land of US$6.28 from Damania et al. (2023).
Establishment costs of silvopasture vary widely. We found the cost to establish one hectare of silvopasture to be US$1.06–4,825 (Dupraz & Liagre, 2011; Lee et al., 2011). Reasons for this wide range include the low cost of natural regeneration and the broad range in tree density depending on the type of system. We collected costs by region and used a weighted average to obtain a global net net cost value of US$424.20.
Cost and revenue data for silvopasture were insufficient. However, data on the impact on revenues per hectare are abundant. Our analysis found a median 8.7% increase in per-hectare profits from silvopasture compared with conventional grazing, which we applied to the average grazing value to obtain a net profit of US$6.82/ha. This does not reflect the very high revenues of silvopasture systems in some countries.
We calculated cost per t CO₂‑eq sequestered by dividing net net cost/ha by total CO₂‑eq sequestered/ha.
Table 2. Cost per unit of climate impact.
Unit: 2023 US$/t CO₂-eq
| Median | $43.25 |
100-yr basis & 20-yr basis are the same.
There is not enough information available to determine a learning curve for silvopasture. However, anecdotal evidence showed establishment costs decreasing as techniques for broadscale mechanized establishment were developed in Australia and Colombia (Murguietio et al., 2016; Shelton et al., 2021).
Speed of action refers to how quickly a climate solution physically affects the atmosphere after it is deployed. This is different from speed of deployment, which is the pace at which solutions are adopted.
At Project Drawdown, we define the speed of action for each climate solution as emergency brake, gradual, or delayed.
Deploy Silvopasture is a DELAYED climate solution. It works more slowly than gradual or emergency brake solutions. Delayed solutions can be robust climate solutions, but it’s important to recognize that they may not realize their full potential for some time.
Living biomass and soil organic matter only temporarily hold carbon (decades to centuries for soil organic matter, and for the life of the tree or any long-lived products made from its wood in the case of woody biomass). Sequestered carbon in both soils and biomass is vulnerable to fire, drought, long-term shifts to a drier precipitation regime, and other climate change impacts, as well as to a return to the previous farming or grazing practices. Such disturbances can cause carbon to be re-emitted to the atmosphere (Lorenz & Lal, 2018).
Like all upland, terrestrial agricultural systems, over the course of decades, silvopastures reach saturation and net sequestration slows to nearly nothing (Lorenz & Lal, 2018).
Lack of data on the current adoption of silvopasture is a major gap in our understanding of the potential of this solution. One satellite imaging study found 156 million ha of grazing land with more than 10 t C/ha in above-ground biomass, which is the amount that indicates more than grass alone (Chapman et al., 2019). However, this area includes natural savannas, which are not necessarily silvopastures, and undercounts the existing 15.1 million ha of silvopasture known to be present in Europe (den Herder et al., 2017).
Sprenkle-Hippolite et al. (2024) estimated a current adoption of 141.4 Mha, or 6.0% of grazing land (Table 3). We have chosen this more recent figure as the best available estimate of current adoption. Note that in Solution Basics in the dashboard above we set current adoption at zero. This is a conservative assumption to avoid counting carbon sequestration from land that has already ceased to sequester net carbon due to saturation, which takes place after 20–50 years (Lal et al., 2018).
Table 3. Current (2023) adoption level.
Unit: million ha
| Mean | 141.4 |
There is little quantifiable information reported about silvopasture adoption trends.
Grazing is the world’s largest land use at 2,986 Mha (Mehrabi et al., 2024). Much grazing land is too dry for trees, while other grasslands that were not historically forest or savanna should not be planted with trees in order to minimize water use and protect grassland habitat (Dudley et al., 2020). Three studies estimated the total potential area suitable for silvopasture (including current adoption).
Lal et al. (2018) estimated the technical potential for silvopasture adoption at 550 Mha.
Chapman et al. (2019) estimated the suitable area for increased woody biomass on grazing land as 1,890 Mha.
Sprenkle-Hippolite (2024) assessed the maximum area of grazing land to which trees could be added without reducing livestock productivity. They calculated a total of 1,589 Mha, or 67% of global grazing land (Table 4). To our knowledge, this is the most accurate estimate available.
Table 4. Adoption ceiling.
Unit: ha converted
| 25th percentile | 1069000000 |
| Mean | 1343000000 |
| Median (50th percentile) | 1588000000 |
| 75th percentile | 1739000000 |
Unit: % of grazing land
| 25th percentile | 45 |
| Mean | 36 |
| Median (50th percentile) | 53 |
| 75th percentile | 58 |
In our Achievable – High scenario, global silvopasture starts at 141.4 Mha and grows at the Colombian Nationally Determined Contribution growth rate of 6.5%/yr. This would provide the high end of the achievable potential at 206.3 Mha by 2030, of which 64.9 million ha are newly adopted (Table 5). For the Achievable – Low scenario, we chose 1/10 of Colombia’s projected growth rate. This would provide 147.0 Mha of adoption by 2030, of which 5.6 Mha are new.
Few estimates of the global adoption potential of silvopasture are available, and even those for the broader category of agroforestry are rare due to the lack of solid data on current adoption and growth rates (Shi et al., 2018; Hart et al., 2023). The IPCC estimates that, for agroforestry overall, 19.5% of the technical potential is economically achievable (IPCC AR6 WG3, 2022). Applying this rate to Sprenkle-Hippolite’s estimated 1,588 Mha technical potential yields an achievable potential of 310 Mha of convertible grazing land.
Our high adoption rate reaches 13% of the adoption ceiling by 2030. This suggests that silvopasture represents a large but relatively untapped potential that will require aggressive policy action and other incentives to spur scaling.
Table 5. Range of achievable adoption levels.
Unit: Mha
| Current adoption | 141.4 |
| Achievable – low | 147.0 |
| Achievable – high | 206.3 |
| Adoption ceiling | 1,588.0 |
Unit: Mha
| Current adoption | 0.00 |
| Achievable – low | 5.6 |
| Achievable – high | 64.9 |
| Adoption ceiling | 1,447.4 |
Carbon sequestration continues only for a period of decades; because silvopasture is an ancient practice with some plantings centuries old, we could not assume that previously adopted hectares continue to sequester carbon indefinitely. Much of the current adoption of silvopasture has been in place for centuries and sequestration there has presumably already slowed down to almost zero. We apply an adoption adjustment factor of 0.25 to current adoption (see Methodology) to reflect that most current adoption is no longer sequestering significant carbon, yet there is substantial new adoption within the past 20–50 years.
For new adoption the calculation is effectiveness * new adoption = climate impact.
For current adoption the calculation is effectiveness * 0.25 * current adoption = climate impact
Climate impacts shown in Table 6 are the sum of current and new adoption impacts. Carbon sequestration impact is 0.35 Gt CO₂‑eq/yr for current adoption, 0.40 Gt CO₂‑eq/yr for Achievable – Low, 0.98 Gt CO₂‑eq/yr for Achievable – High, and 14.54 Gt CO₂‑eq/yr for our Adoption Ceiling.
Table 6. Climate impact at different levels of adoption.
Unit: Gt CO₂‑eq/yr
| Current adoption | 0.35 |
| Achievable – low | 0.40 |
| Achievable – high | 0.98 |
| Adoption ceiling | 14.54 |
100-yr basis, New adoption only
Lal et al. (2018) estimated a technical global carbon sequestration potential of 0.3–1.0 Gt CO₂‑eq/yr. Sprenkle-Hyppolite et al. (2024) estimated a silvopasture technical potential of 1.4 Gt CO₂‑eq/yr ; this assumes a tree density of 2–6 trees/ha, which is substantially lower than typical silvopasture. For agroforestry overall (including silvopasture and other practices), the IPCC (2022) estimates an achievable potential of 0.8 Gt CO₂‑eq/yr and a technical potential of 4.0 Gt CO₂‑eq/yr.
Silvopasture can also increase and diversify farmer income. Tree fruit and timber often provide income for ranchers. A study in the southern United States showed that silvopasture systems generated 10% more income than standalone cattle production (Husak & Grado, 2002). A more comprehensive analysis across the eastern United States (Greene et al., 2023) found that virtually all silvopasture systems assessed had a positive 20- and 30-yr internal rate of return (IRR). For some systems, the 30-yr IRR can be >15% (Greene et al., 2023).
While evidence on the impact of silvopasture on yields is mixed, this practice can improve food security by diversifying food production and income sources (Bostedt et al., 2016; Smith et al., 2022). In pastoralists in Kenya, Bostedt et al. (2016) found that agroforestry practices were associated with increased dietary diversity, an important aspect of food and nutrition security. Diverse income streams can mediate household food security during adverse conditions, such as droughts or floods, especially in low- and middle-income countries (de Sherbinin et al., 2008; Di Prima et al., 2022; Frelat et al., 2016).
Trees boost habitat availability, enhance landscape connectivity, and aid in forest regeneration and restoration. In most climates they provide a major boost to biodiversity compared with pasture alone (Smith et al., 2022; Pezo et al., 2018).
By providing shade, silvopasture systems reduce heat stress experienced by livestock. Heat stress for cattle begins at 30 °C or even lower in some circumstances (Garrett et al., 2004). In the tropics, the cooling effect of integrating trees into a pastoral system is 0.32–2.4 °C/t of woody carbon added/ha (Zeppetello et al., 2022). Heifers raised in silvopasture systems had higher body mass and more optimal body temperature than those raised in intensive rotational grazing systems (Lemes et al., 2021). Improvement in livestock physiological conditions probably results from access to additional forage, increased livestock comfort, and reduced heat stress in silvopastoral systems. Silvopasture is highly desirable for its improvements to animal welfare (Goracci & Camilli, 2024).
Silvopasture and agroforestry are important for ensuring soil health (Basche et al., 2020). These practices improve soil health by reducing erosion and may also contribute to soil organic matter retention (U.S. Department of Agriculture Natural Resource Conservation Service ([USDA NRCS], 2025). There is evidence that silvopasture may improve soil biodiversity by preventing soil organism habitat loss and degradation (USDA NRCS, 2025).
Perennials in silvopasture systems could reduce runoff and increase water infiltration rates relative to open rangelands (Smith et al., 2022; Pezo et al., 2018). This increases the resilience of the system during drought and high heat. Silvopasture can improve water quality by retaining soil sediments and filtering pollutants found in runoff (USDA NRCS, 2025). On average, silvopasture and agroforestry practices can reduce runoff of sediments and excess nutrients into water 42–47% (Zhu et al., 2020). The filtering benefits of silvopasture can also mitigate pollution of antibiotics from livestock operations from entering waterways (Moreno & Rolo, 2019).
Some of the tree and forage species used in silvopastures are invasive in certain contexts. For example, river tamarind (Leucaena leucocephala) is a centerpiece in intensive silvopasture in Latin America, where it is native, but also in Australia, where it is not. Australian producers have developed practices to limit or prevent its spread (Shelton et al., 2021).
Livestock can damage or kill young trees during establishment. Protecting trees or excluding grazing animals during this period increases costs (Smith et al., 2022).
Poorly designed tree layout can make herding, haying, fencing, and other management activities more difficult. Tree densities that are too high can reduce livestock productivity (Cadavid et al., 2020).
Silvopasture represents a way to produce some ruminant meat and dairy in a more climate-friendly way. This impact can contribute to addressing emissions from ruminant production, but only as part of a program that strongly emphasizes diet change and food waste reduction.
Forms of silvopasture that increase milk and meat yields can reduce pressure to convert undeveloped land to agriculture.
Silvopasture is a technique for restoring farmland.
Silvopasture is a form of savanna restoration.
Expanding silvopasture could restrict land availability for renewable energy or raw material and food production, since many technologies and practices could be sited on grazing lands. Silvopasture and forest restoration can also compete for the same land.
Silvopasture is a kind of agroforestry, though in this iteration of Project Drawdown “Deploy Agroforestry” refers to crop production systems only. With that said, some agroforestry systems integrate both crops and livestock with the trees, such as the widespread parkland systems of the African Sahel.
ha converted from grazing land to silvopasture
CO₂
Solutions that improve ruminant production could undermine the argument for reducing ruminant protein consumption in wealthy countries.
Certain silvopasture systems reduce per-hectare productivity of meat and milk, even if overall productivity increases when the yields of timber or food from the tree component are included. For example, silvopasture systems that are primarily focused on timber production, with high tree densities, will have lower livestock yields than pasture alone - though they will have high timber yields.
The costs of establishment are much higher than those of managed grazing. There is also a longer payback period (Smith et al., 2022). These limitations mean that secure land tenure is even more important than usual, to make adoption worthwhile (Poudel et al., 2024).
Silvopasture is primarily appropriate for grazing land that receives sufficient rainfall to support tree growth. While it can be implemented on both cropland and grassland, if adopted on cropland, it will reduce food yield because livestock produce much less food per hectare than crops. In the humid tropics, a particularly productive and high-carbon variation called intensive silvopasture is an option. Ideally, graziers will have secure land tenure, though pastoralist commons have been used successfully.
Areas too dry to establish trees (<450 mm annual precipitation) are not suitable for silvopasture by tree planting, but regions that can support natural savanna may be suitable for managed natural regeneration.
Most silvopasture today appears in sub-Saharan Africa (Chapman et al., 2019), though this may reflect grazed natural savannas rather than intentional silvopasture. This finding neglects well-known systems in Latin America and Southern Europe.
Chapman et al. (2019) listed world grasslands by their potential to add woody biomass. According to their analysis, the countries with the greatest potential to increase woody biomass carbon in grazing land are, in order: Australia, Kazakhstan, China, the United States, Mongolia, Iran, Argentina, South Africa, Sudan, Afghanistan, Russia, and Mexico. Tropical grazing land accounts for 73% of the potential in one study. Brazil, China, and Australia have the highest areas, collectively accounting for 37% of the potential area (Sprenkle-Hippolite 2024).
We do not present any maps for the silvopasture solution due to the uncertainties in identifying current areas where silvopasture is practiced, and in identifying current grasslands that were historically forest or savanna.
There is a high level of consensus about the carbon biosequestration impacts of silvopasture, including for the higher per-hectare sequestration rates relative to improved grazing systems alone. A handful of reviews, expert estimations, and meta-analyses have been published on the subject. These include:
Cardinael et al. (2018) assembled data by climate and region for use in the national calculations and reporting.
Chatterjee et al. (2018) found that converting from pasture to silvopasture increases carbon stocks.
Lal et al. ( 2018) estimated the technical adoption and mitigation potential of silvopasture and other practices.
Udawatta et al. (2022) provided an up-to-date meta-analysis for temperate North America.
The results presented in this document summarize findings from two reviews, two meta-analyses, one expert opinion and three original studies reflecting current evidence from a global scale. We recognize this limited geographic scope creates bias, and hope this work inspires research and data sharing on this topic in underrepresented regions.
There is low consensus on the reduction of methane from enteric emissions, nitrous oxide from manure, and CO₂ from avoided deforestation due to increased productivity. We do not include these climate impacts in our calculations.
Until recently there was little understanding of the current adoption of silvopasture. Sprenkle-Hyppolite et al. (2024) used Delphi expert estimation to determine current adoption and technical potential. Rates of adoption and achievable potential are still largely unreported or uninvestigated. See the Adoption section for details.
Alternative grazing refers to approaches to managing ruminant livestock – including multiple fenced paddocks and frequent animal movement, short grazing periods, and enforced rest periods for forage – that aim to increase soil organic carbon (SOC) by allowing vegetation to recover. Robust evidence, including independent, commercial-scale before-and-after control studies that track SOC to at least 30 cm over 5–10+ years, is lacking. The best available research suggests that these practices do not consistently produce additional, durable SOC gains relative to continuous grazing, and that SOC outcomes strongly depend on climate and land-use contexts.
Alternative grazing can deliver localized ecosystem benefits, particularly where previous continuous grazing has caused erosion near waterways or where more strategic movement of ruminant livestock could help. However, these practices require added labor, fencing, and water infrastructure, which can constrain adoption. The potential benefits of alternative grazing do not alter the fact that beef is the highest emitting type of food, and an emphasis on deploying alternative grazing can delay shifts to lower-emitting food options that also allow land restoration. As a result, alternative grazing is not currently supported as an effective climate solution, but remains an area to “Keep Watching.”
Alternative grazing may reduce SOC loss that’s common in continuous overgrazing by allowing areas to rest, but evidence that it delivers additional carbon removal is inconsistent and limited to niche environmental conditions and geographies. Therefore, we will “Keep Watching” this solution.
| Plausible | Could it work? | Yes |
|---|---|---|
| Ready | Is it ready? | Yes |
| Evidence | Are there data to evaluate it? | Limited |
| Effective | Does it consistently work? | No |
| Impact | Is it big enough to matter? | No |
| Risk | Is it risky or harmful? | No |
| Cost | Is it cheap? | No |
Alternative grazing, also called adaptive multi-paddock, rotational, holistic, regenerative, or mob grazing, is a suite of ruminant livestock movement strategies that differ from conventional continuous grazing. Conventional continuous grazing mostly allows unmanaged grazing of a large pasture with limited or no structured rest periods for vegetation recovery. Alternative grazing uses multiple fenced paddocks and frequent animal movement, short grazing periods to redistribute pressure, and enforced rest periods for forage. The proposed carbon removal mechanism is that improved grazing management could increase SOC storage by enhancing plant growth, especially below ground root growth.
The biological mechanism by which recovering vegetation can increase SOC is well understood, but factors such as climate, soil type, prior land use, slope, and stocking rate often have greater influence on SOC than the type of grazing practice. Results from the few higher-quality studies suggest that alternative grazing does not produce consistent, additional, or durable SOC gains relative to continuous grazing. Most positive SOC findings come from wetter regions with higher rainfall, including lands converted from cropland, where gains may occur regardless of grazing style. Positive SOC findings rarely measure changes over time at the same site and often are based on shallow soil samples. A global meta-analysis found that strategic-rest grazing increased vegetation biomass only when stocking rates were reduced, and even then it did not guarantee SOC gains.
Alternative grazing, in limited and context-specific cases, can deliver localized ecosystem benefits, particularly where grazing is already occurring and degradation is severe. Managing grazing away from waterways and heavily compacted areas can reduce overgrazing damage and in some regions may benefit grassland bird habitat, though evidence is limited.
There are several reasons to be concerned about alternative grazing.
Creating areas of rest away from cattle grazing accrues additional management costs related to labor, fencing (physical or virtual), and water infrastructure, which could reduce adoption and scalability. Education gaps limit the ability of livestock managers to adopt the best approach.
Alternative grazing practices can result in reduced animal weight gains and associated increases in enteric methane emissions per unit of beef protein produced, potentially offsetting or outweighing any small, temporary SOC gains relative to continuous grazing. Any claimed SOC gains from grazing management are difficult to measure and verify. Many studies lack baseline SOC measurements, adequate controls, sufficient duration, or adequate soil-depth sampling, making it difficult to attribute carbon gains to grazing practices. In particular, grazing-specific solutions need to be compared with no-grazing or protected native ecosystems nearby, where carbon storage is greater and use does not generate methane.
A last key concern is that alternative grazing is frequently framed as a way to fix the climate impacts of beef production, the highest-GHG-emitting food option, rather than reducing beef demand through improving diets. Like feed additives and genetic efficiency gains, this can overstate production-side fixes, obscure the need for demand-side shifts, and delay more effective carbon removal pathways.
Abdalla, M., Hastings, A., Chadwick, D. R., Jones, D. L., Evans, C. D., Jones, M. B., Rees, R. M., & Smith, P. (2018). Critical review of the impacts of grazing intensity on soil organic carbon storage and other soil quality indicators in extensively managed grasslands. Agriculture, Ecosystems & Environment, 253, 62–81. Link to source: https://doi.org/10.1016/j.agee.2017.10.023
Cusack, D. F., Kazanski, C. E., Hedgpeth, A., Chow, K., Cordeiro, A. L., Karpman, J., & Ryals, R. (2021). Reducing climate impacts of beef production: A synthesis of life cycle assessments across management systems and global regions. Global Change Biology, 27(9), 1721-1736. Link to source: https://doi.org/10.1111/gcb.15509
Garnett, T., Godde, C., Muller, A., Röös, E., Smith, P., de Boer, I. J. M., van Zanten, H., Herrero, M., Schader, C., van Middelaar, C., & Thornton, P. (2017). Grazed and confused? Ruminating on cattle, grazing systems, methane, nitrous oxide, the soil carbon sequestration question. Food Climate Research Network, University of Oxford. Link to source: https://www.tabledebates.org/sites/default/files/2022-04/fcrn_gnc_report.pdf
Godde, C. M., Boone, R. B., Ash, A. J., Waha, K., Sloat, L. L., Thornton, P. K., & Herrero, M. (2020). Global rangeland production systems and livelihoods at threat under climate change and variability. Environmental Research Letters, 15(4), 044021. Link to source: https://doi.org/10.1088/1748-9326/ab7395
Hawkins, H. J., Venter, Z. S., & Cramer, M. D. (2022). A holistic view of Holistic Management: What do farm-scale, carbon, and social studies tell us? Agriculture, Ecosystems & Environment, 323, 107702. Link to source: https://doi.org/10.1016/j.agee.2021.107702
Jorns, T. R., Scasta, J. D., Derner, J. D., Augustine, D. J., Porensky, L. M., Raynor, E. J., & CARM Stakeholder Group. (2024). Adaptive multi-paddock grazing management reduces diet quality of yearling cattle in shortgrass steppe. The Rangeland Journal, 45(4), 160–172. Link to source: https://doi.org/10.1071/RJ23047
Lajtha, K., & Silva, L. (2022). Grazing cattle, well-managed or not, is unlikely to increase soil carbon sequestration. Proceedings of the National Academy of Sciences, 119(30), e2203408119. Link to source: https://doi.org/10.1073/pnas.2203408119
Ma, Z., Shrestha, B. M., Bork, E. W., Chang, S. X., Carlyle, C. N., Döbert, T. F., Silva Sobrinho, L., & Boyce, M. S. (2021). Soil greenhouse gas emissions and grazing management in northern temperate grasslands. Science of the Total Environment, 796, 148975. Link to source: https://doi.org/10.1016/j.scitotenv.2021.148975
Poore, J., & Nemecek, T. (2018). Reducing food’s environmental impacts through producers and consumers. Science, 360(6392), 987–992. Link to source: https://doi.org/10.1126/science.aaq0216
Reinhart, K. O., Worogo, H. S. S., & Rinella, M. J. (2022). Ruminating on the science of carbon ranching. Journal of Applied Ecology, 59(3), 642–648. Link to source: https://doi.org/10.1111/1365-2664.14100
Sanderman, J., Partida, C., Xia, Y., Lavallee, J. M., & Bradford, M. A. (2026). Systematic review reveals soil organic carbon benefits of alternative grazing depend on study quality. Communications Earth & Environment. Link to source: https://doi.org/10.1038/s43247-026-03790-8
Schlesinger, W. H. (2022). Biogeochemical constraints on climate change mitigation through regenerative farming. Biogeochemistry, 161, 9–17. Link to source: https://doi.org/10.1007/s10533-022-00942-8
Searchinger, T. D., Wirsenius, S., Beringer, T., & Dumas, P. (2018). Assessing the efficiency of changes in land use for mitigating climate change. Nature, 564(7735), 249–253. Link to source: https://doi.org/10.1038/s41586-018-0757-z
Wang, Y., de Boer, I. J. M., Persson, U. M., Ripoll-Bosch, R., Cederberg, C., Gerber, P. J., Smith, P., & van Middelaar, C. E. (2023). Risk to rely on soil carbon sequestration to offset global ruminant emissions. Nature Communications, 14(1), 7625. Link to source: https://doi.org/10.1038/s41467-023-43452-3
Seaweed ecosystem protection is the long-term protection from degradation of wild subtidal brown and red seaweed ecosystems. Seaweeds, also called macroalgae, are photosynthetic marine organisms that absorb CO₂ from the water and convert it into biomass. This can lower surface-water CO₂ concentrations, allowing additional CO₂ from the atmosphere to be dissolved in the ocean. Some of the fixed carbon can be sequestered through export to the deep sea or burial in the seafloor, while a portion may persist in forms that resist degradation even at the ocean surface.
Protecting seaweed ecosystems can reduce a range of human impacts (wild harvesting, coastal development, overgrazing, and poor water quality) and improve resilience to other stressors (warming), which helps preserve carbon removal by the seaweed and avoid CO₂ emissions from biomass losses.
This solution focuses on legal mechanisms of protection through the establishment of Marine Protected Areas (MPAs), which are managed with the primary goal of conserving nature. This solution does not include cultivated seaweed (see Deploy Seaweed Farming for Food).
Seaweeds are diverse marine photosynthetic organisms composed of three groups: brown (Phaeophyceae), green (Chlorophyta), and red algae (Rhodophyta). They can form ecosystems, such as kelp forests, and contribute to other marine ecosystems by providing habitat and food. Seaweeds are distinguished from other algae, such as phytoplankton, based on their larger size and because most are attached to substrate rather than free-floating. Seaweeds cover an estimated 600 Mha of the ocean (Duarte et al., 2022), an area that is an order of magnitude greater than the area associated with coastal wetlands (~55 Mha, see Protect Coastal Wetlands).
This solution focuses on wild subtidal (always submerged) brown and red seaweed ecosystems, which together account for over 75% of global seaweed extent (Duarte et al., 2022) (Figure 1). We do not include green seaweeds due to their smaller extent and data limitations. We also do not include seaweeds that occur in intertidal zones, as free-floating colonies (e.g., some species of Sargassum) or are cultivated due to data limitations or coverage in other Explorer solutions (e.g., Deploy Seaweed Farming for Food).
Figure 1. Seaweed ecosystem types considered in this solution (left to right): subtidal brown (central California, USA) and subtidal red (Atlantic coast of Spain)
fdastudillo | iStock; Damocean | iStock
Seaweed ecosystems exhibit high net primary productivity (NPP) rates, comparable to those of terrestrial forests (Filbee-Dexter, 2020). Unlike many terrestrial ecosystems, however, nearly all carbon storage in seaweed ecosystems occurs as above-ground biomass, since seaweeds lack below-ground roots. A smaller amount can be buried on site in sediment (Krause-Jensen & Duarte, 2016). Most long-term carbon storage attributable to seaweeds occurs largely outside of seaweed ecosystems, through the export of carbon in dissolved and suspended forms (Figure 2). Some of this carbon reaches the deep sea, where it can persist for more than 100 years (Krause-Jensen & Duarte, 2016; Krause-Jensen et al., 2018; Ortega et al., 2019). Roughly 11.4% (25th quartile, 6.0%; 75th quartile, 13.7%) of NPP from global seaweed ecosystems is estimated to contribute to long-term carbon storage in the deep sea, equivalent to as much as 0.62 Gt CO₂‑eq/yr (173 Tg C/yr, Krause-Jensen & Duarte, 2016). While uncertain and requiring more research, recent modeling efforts support these estimates, suggesting that more than 12.5% of NPP may be removed on 100-yr timescales (Filbee-Dexter et al., 2024b).
Figure 2. Overview of a seaweed ecosystem showing carbon fluxes into and out of the ecosystem (g=gaseous, aq=aqueous) that can result in carbon removal. Some carbon is exported to the shallow sea, where it may be recycled or persist for longer periods depending on its form, some is exported to the deep sea (~1000 m), and some is buried in seafloor sediments.
Adapted from: Hurd, C. L., Gattuso, J.-P., & Boyd, P. W. (2024). Air-sea carbon dioxide equilibrium: Will it be possible to use seaweeds for carbon removal offsets? Journal of Phycology, 60(1), 4–14.
Seaweed ecosystems face growing threats from a range of climate change impacts (Harley et al., 2012), such as increasing sea surface temperatures, marine heat waves, ocean acidification, and extreme storm events, as well as local drivers, such as overfishing, overgrazing, pollution, disease outbreaks, invasive species, and bottom fishing (Corrigan et al., 2025; Filbee-Dexter et al., 2024a; Hanley et al., 2024). For instance, overfishing can deplete top predators in ecosystems, leading to increases in herbivores, such as sea urchins, that overgraze seaweed (Steneck et al., 2002).
In this solution, we calculate how legal protection of seaweed ecosystems via MPAs can reduce CO₂
emissions and preserve carbon removal through avoided ecosystem loss. In addition to preventing direct losses from impacts such as wild harvest, MPAs can help restore predator populations that keep herbivores in balance. For instance, many MPAs include no-take zones that allow predatory fish populations to recover, thereby lessening overgrazing impacts over time. MPAs can also increase the resilience of seaweed ecosystems against climate change stressors, such as marine heat waves (Kumagai et al., 2024; Ortiz-Villa et al., 2025). While some seaweed can release methane, offsetting CO₂
removal (Roth et al., 2023), we exclude this process from our analysis due to existing data limitations. We also do not consider nitrous oxide, though protection might provide additional climate benefits because enhanced nitrous oxide production has been tied to nutrient-polluted seaweed systems (Wong et al., 2021).
We present estimates of climate impact as likely upper bounds under several key assumptions (see Appendix and Caveats), which can be improved upon as more research unfolds. We consider subtidal brown and red seaweed to be protected if they are within designated MPAs based on global datasets from the United Nations Environment Programme World Conservation Monitoring Centre (UNEP-WCMC) and the International Union for Conservation of Nature and Natural Resources (IUCN) (2024). Importantly, protection can help reduce – but will not eliminate – ecosystem loss in MPAs relative to unprotected areas (see Effectiveness).
Akaike, S., & Mizuta, H. (2024). Optimizing the biomass balance of macroalgae and sea urchins in kelp beds by removing the urchins. European Journal of Phycology, 59(2), 218–231. Link to source: https://doi.org/10.1080/09670262.2024.2306397
Balmford, A., Gravestock, P., Hockley, N., McClean, C. J. and Roberts, C. M. (2004). The worldwide costs of marine protected areas. Proceedings of the National Academy of Sciences, 101(26), 9694–9697. Link to source: https://doi.org/10.1073/pnas.0403239101
Bauer, J., Beas-Luna, R., Malpica-Cruz, L., Abadía-Cardoso, A., Filz, P., Bonilla, J. C., & Lorda, J. (2025). Community-led management maintains higher predator biomass supporting kelp forests persistence in Baja California. Scientific Reports, 15(1), Article 23253. Link to source: https://doi.org/10.1038/s41598-025-86140-6
Best, R. J., Chaudoin, A. L., Bracken, M. E. S., Graham, M. H., & Stachowicz, J. J. (2014). Plant–animal diversity relationships in a rocky intertidal system depend on invertebrate body size and algal cover. Ecology, 95(5), 1308–1322. Link to source: https://doi.org/10.1890/13-1480.1
Corrigan, S., Cottier-Cook, E. J., Lim, P. E., & Brodie, J. (2025). The state of the world’s seaweeds 2025 [Data set]. Natural History Museum. Link to source: https://doi.org/10.5519/4ln9oqk7
Cotas, J., Gomes, L., Pacheco, D., & Pereira, L. (2023). Ecosystem services provided by seaweeds. Hydrobiology, 2(1), 75–96. Link to source: https://doi.org/10.3390/hydrobiology2010006
Cottier-Cook, E. J., Lim, P. E., Mallinson, S., Yahya, N., Poong, S. W., Wilbraham, J., Nagabhatla, N., & Brodie, J. (2023). Striking a balance: Wild stock protection and the future of our seaweed industries [Policy brief]. United Nations University Institute on Comparative Regional Integration Studies. Link to source: https://cris.unu.edu/sites/cris.unu.edu/files/UNU-CRIS_Policy-Brief_CottierCook_Et.al_23.06.pdf
Cuba, D., Guardia-Luzon, K., Cevallos, B., Ramos-Larico, S., Neira, E., Pons, A., & Avila-Peltroche, J. (2022). Ecosystem services provided by kelp forests of the Humboldt current system: A comprehensive review. Coasts, 2(4), 259–277. Link to source: https://doi.org/10.3390/coasts2040013
Duarte, C. M., Gattuso, J. P., Hancke, K., Gundersen, H., Filbee-Dexter, K., Pedersen, M. F., Middelburg, J. J., Burrows, M. T., Krumhansl, K. A., Wernberg, T., Moore P., Pessarrodona, A., Ørberg, S. B., Pinto, I. S., Assis, J., Queirós, A. M., Smale, D. A., Bekkby, T., Serrão, E. A., & Krause-Jensen, D. (2022). Global estimates of the extent and production of macroalgal forests. Global Ecology and Biogeography, 31(7), 1422–1439. Link to source: https://doi.org/10.1111/geb.13515
Earp, H. S., Smale, D. A., Almond, P. M., Catherall, H. J. N., Gouraguine, A., Wilding, C., & Moore, P. J. (2024). Temporal variation in the structure, abundance, and composition of Laminaria hyperborea forests and their associated understorey assemblages over an intense storm season. Marine Environmental Research, 200, Article 106652. Link to source: https://doi.org/10.1016/j.marenvres.2024.106652
Eger, A. M., Eddy, N., McHugh, T. A., Arafeh-Dalmau, N., Wernberg, T., Krumhansl, K., Verbeek, J., Branigan, S., Kuwae, T., Caselle, J. E., Ospina, A. G., & Vergés, A. (2024). State of the world’s kelp forests. One Earth, 7(11), 1927–1931. Link to source: https://doi.org/10.1016/j.oneear.2024.10.008
Eger, A. M., Marzinelli, E. M., Christie, H., Fagerli, C. W., Fujita, D., Gonzalez, A. P., Hong, S. W., Kim, J. H., Lee, L. C., McHugh, T. A., Nishihara, G. N., Tatsumi, M., Steinberg, P. D., & Vergés, A. (2022). Global kelp forest restoration: Past lessons, present status, and future directions. Biological Reviews, 97(4), 1449–1475. Link to source: https://doi.org/10.1111/brv.12850
Eger, A. M., Marzinelli, E. M., Beas-Luna, R., Blain, C. O., Blamey, L. K., Byrnes, J. E. K., Carnell, P. E., Choi, C. G., Hessing-Lewis, M., Kim, K. Y., Kumagai, N. H., Lorda, J., Moore, P., Nakamura, Y., Pérez-Matus, A., Pontier, O., Smale, D., Steinberg, P. D., & Vergés, A. (2023). The value of ecosystem services in global marine kelp forests. Nature Communications, 14(1), Article 1894. Link to source: https://doi.org/10.1038/s41467-023-37385-0
Elsmore, K., Nickols, K. J., Miller, L. P., Ford, T., Denny, M. W., & Gaylord, B. (2024). Wave damping by giant kelp, Macrocystis pyrifera. Annals of Botany, 133(1), 29–40. Link to source: https://doi.org/10.1093/aob/mcad094
Food and Agriculture Organization of the United Nations. (2024). The state of world fisheries and aquaculture 2024: Blue transformation in action [Report]. Link to source: https://doi.org/10.4060/cd0683en
Filbee-Dexter, K. (2020). Ocean forests hold unique solutions to our current environmental crisis. One Earth, 2(5), 398–401. Link to source: https://doi.org/10.1016/j.oneear.2020.05.004
Filbee-Dexter, K., & Wernberg, T. (2020). Substantial blue carbon in overlooked Australian kelp forests. Scientific Reports, 10, Article 12341. Link to source: https://doi.org/10.1038/s41598-020-69258-7
Filbee-Dexter, K., Feehan, C. J., Smale, D. A., Krumhansl, K. A., Augustine, S., de Bettignies, F., Burrows, M. T., Byrnes, J. E. K., Campbell, J., Davoult, D., Dunton, K. H., Franco, J. N., Garrido, I., Grace, S. P., Hancke, K., Johnson, L. E., Konar, B., Moore, P. J., Norderhaug, K. M., … Wernberg, T. (2022). Kelp carbon sink potential decreases with warming due to accelerating decomposition. PLOS Biology, 20(8), Article e3001702. Link to source: https://doi.org/10.1371/journal.pbio.3001702
Filbee‐Dexter, K., Starko, S., Pessarrodona, A., Wood, G., Norderhaug, K. M., Piñeiro‐Corbeira, C., & Wernberg, T. (2024a). Marine protected areas can be useful but are not a silver bullet for kelp conservation. Journal of Phycology, 60(2), 203–213. Link to source: https://doi.org/10.1111/jpy.13446
Filbee-Dexter, K., Pessarrodona, A., Pedersen, M. F., Wernberg, T., Duarte, C. M., Assis, J., Bekkby, T., Burrows, M. T., Carlson, D. F., Gattuso, J.-P., Gundersen, H., Hancke, K., Krumhansl, K. A., Kuwae, T., Middelburg, J. J., Moore, P. J., Queirós, A. M., Smale, D. A., Sousa-Pinto, I., … Krause-Jensen, D. (2024b). Carbon export from seaweed forests to deep ocean sinks. Nature Geoscience, 17(6), 552–559. Link to source: https://doi.org/10.1038/s41561-024-01449-7
Gao, G., Gao, L., Jiang, M., Jian, A., & He, L. (2022). The potential of seaweed cultivation to achieve carbon neutrality and mitigate deoxygenation and eutrophication. Environmental Research Letters, 17(1), Article 014018. Link to source: https://doi.org/10.1088/1748-9326/ac3fd9
Gibbons, E. G., & Quijón, P. A. (2023). Macroalgal features and their influence on associated biodiversity: Implications for conservation and restoration. Frontiers in Marine Science, 10, Article 1304000. Link to source: https://doi.org/10.3389/fmars.2023.1304000
González-Roca, F., Gelcich, S., Pérez-Ruzafa, Á., Vega, J. M. A., & Vásquez, J. A. (2021). Exploring the role of access regimes over an economically important intertidal kelp species. Ocean & Coastal Management, 212, Article 105811. Link to source: https://doi.org/10.1016/j.ocecoaman.2021.105811
Hanley, M. E., Firth, L. B., & Foggo, A. (2024). Victim of changes? Marine macroalgae in a changing world. Annals of Botany, 133(1), 1–16. Link to source: https://doi.org/10.1093/aob/mcad185
Harley, C. D. G., Anderson, K. M., Demes, K. W., Jorve, J. P., Kordas, R. L., Coyle, T. A., & Graham, M. H. (2012). Effects of climate change on global seaweed communities. Journal of Phycology, 48(5), 1064–1078. Link to source: https://doi.org/10.1111/j.1529-8817.2012.01224.x
Heckwolf, M. J., Peterson, A., Jänes, H., Horne, P., Künne, J., Liversage, K., Sajeva, M., Reusch, T. B. H., & Kotta, J. (2021). From ecosystems to socio-economic benefits: A systematic review of coastal ecosystem services in the Baltic Sea. Science of the Total Environment, 755, Article 142565. Link to source: https://doi.org/10.1016/j.scitotenv.2020.142565
Hurd, C. L., Gattuso, J.-P., & Boyd, P. W. (2024). Air-sea carbon dioxide equilibrium: Will it be possible to use seaweeds for carbon removal offsets? Journal of Phycology, 60(1), 4–14. Link to source: https://doi.org/10.1111/jpy.13405
Kelp Forest Alliance. (2024). State of the world’s kelp report [Report]. Link to source: https://kelpforestalliance.com/state-of-the-worlds-kelp-report/
Krause-Jensen, D., & Duarte, C. M. (2016). Substantial role of macroalgae in marine carbon sequestration. Nature Geoscience, 9(10), 737–742. Link to source: https://doi.org/10.1038/ngeo2790
Krause-Jensen, D., Lavery, P., Serrano, O., Marbà, N., Masque, P., & Duarte, C. M. (2018). Sequestration of macroalgal carbon: The elephant in the blue carbon room. Biology Letters, 14(6), Article 20180236. Link to source: https://doi.org/10.1098/rsbl.2018.0236
Krumhansl, K. A., Okamoto, D. K., Rassweiler, A., Novak, M., Bolton, J. J., Cavanaugh, K. C., Connell, S. D., Johnson, C. R., Konar, B., Ling, S. D., Micheli, F., Norderhaug, K. M., Pérez-Matus, A., Sousa-Pinto, I., Reed, D. C., Salomon, A. K., Shears, N. T., Wernberg, T., Anderson, R. J., … Byrnes, J. E. K. (2016). Global patterns of kelp forest change over the past half-century. Proceedings of the National Academy of Sciences, 113(48), 13785–13790. Link to source: https://doi.org/10.1073/pnas.1606102113
Kumagai, J. A., Goodman, M. C., Villaseñor-Derbez, J. C., Schoeman, D. S., Cavanuagh, K. C., Bell, T. W., Micheli, F., De Leo, G., & Arafeh-Dalmau, N. (2024). Marine protected areas that preserve trophic cascades promote resilience of kelp forests to marine heatwaves. Global Change Biology, 30(12), Article e17620. Link to source: https://doi.org/10.1111/gcb.17620
Lindhart, M., Daly, M. A., Walker, H., Arzeno-Soltero, I. B., Yin, J. Z., Bell, T. W., Monismith, S. G., Pawlak, G., & Leichter, J. J. (2024). Short wave attenuation by a kelp forest canopy. Limnology and Oceanography Letters, 9(4), 478–486. Link to source: https://doi.org/10.1002/lol2.10401
McCrea-Strub, A., Zeller, D., Rashid Sumaila, U., Nelson, J., Balmford, A., & Pauly, D. (2011). Understanding the cost of establishing marine protected areas. Marine Policy, 35(1), 1–9. Link to source: https://doi.org/10.1016/j.marpol.2010.07.001
Ortega, A., Geraldi, N. R., Alam, I., Kamau, A. A., Acinas, S. G., Logares, R., Gasol, J. M., Massana, R., Krause-Jensen, D., & Duarte, C. M. (2019). Important contribution of macroalgae to oceanic carbon sequestration. Nature Geoscience, 12(9), 748–754. Link to source: https://doi.org/10.1038/s41561-019-0421-8
Ortiz‐Villa, E. M., Rassweiler, A., Caselle, J. E., Cavanaugh, K. C., Arafeh‐Dalmau, N., Bell, T. W., & Cavanaugh, K. C. (2025). Marine protected areas enhance climate resilience to severe marine heatwaves for kelp forests. Journal of Applied Ecology, 62(9), 2439–2453. Link to source: https://doi.org/10.1111/1365-2664.70112
Pessarrodona, A., Franco-Santos, R. M., Wright, L. S., Vanderklift, M. A., Howard, J., Pidgeon, E., Wernberg, T., & Filbee-Dexter, K. (2023). Carbon sequestration and climate change mitigation using macroalgae: A state of knowledge review. Biological Reviews, 98(6), 1945–1971. Link to source: https://doi.org/10.1111/brv.12990
Rodríguez-Rodríguez, D., & Martínez-Vega, J. (2022). Chapter three—Ecological effectiveness of marine protected areas across the globe in the scientific literature. In C. Sheppard (Ed.), Advances in marine biology (Vol. 92, pp. 129–153). Academic Press. Link to source: https://doi.org/10.1016/bs.amb.2022.07.002
Roth, F., Broman, E., Sun, X., Bonaglia, S., Nascimento, F., Prytherch, J., Brüchert, V., Lundevall Zara, M., Brunberg, M., Geibel, M. C., Humborg, C., & Norkko, A. (2023). Methane emissions offset atmospheric carbon dioxide uptake in coastal macroalgae, mixed vegetation and sediment ecosystems. Nature Communications, 14(1), Article 42. Link to source: https://doi.org/10.1038/s41467-022-35673-9
Steen, H., Moy, F. E., Bodvin, T., & Husa, V. (2016). Regrowth after kelp harvesting in Nord-Trøndelag, Norway. ICES Journal of Marine Science, 73(10), 2708–2720. Link to source: https://doi.org/10.1093/icesjms/fsw130
Steneck, R. S., Graham, M. H., Bourque, B. J., Corbett, D., Erlandson, J. M., Estes, J. A., & Tegner, M. J. (2002). Kelp forest ecosystems: Biodiversity, stability, resilience and future. Environmental Conservation, 29(4), 436–459. Link to source: https://doi.org/10.1017/S0376892902000322
Tano, S., Eggertsen, M., Wikström, S. A., Berkström, C., Buriyo, A. S., & Halling, C. (2016). Tropical seaweed beds are important habitats for mobile invertebrate epifauna. Estuarine, Coastal and Shelf Science, 183, 1–12. Link to source: https://doi.org/10.1016/j.ecss.2016.10.010
Thurstan, R. H., Brittain, Z., Jones, D. S., Cameron, E., Dearnaley, J., & Bellgrove, A. (2018). Aboriginal uses of seaweeds in temperate Australia: An archival assessment. Journal of Applied Phycology, 30(3), 1821–1832. Link to source: https://doi.org/10.1007/s10811-017-1384-z
United Nations Environment Programme World Conservation Monitoring Centre, & International Union for Conservation of Nature. (2024). Protected planet: The world database on protected areas (WDPA) and world database on other effective area-based conservation measures (WD-OECM) [Data set]. Retrieved April 2024, from Link to source: https://www.protectedplanet.net
United Nations Environment Programme. (2023). Into the blue: Securing a sustainable future for kelp forests [Report]. Link to source: https://wedocs.unep.org/handle/20.500.11822/42331
Waldron, A., Adams, V., Allan, J., Arnell, A., Asner, G., Atkinson, S., Baccini, A., Baillie, J. E. M., Balmford, A., Beau, J. A., Brander, L., Brondizio, E., Bruner, A., Burgess, N., Burkart, K., Butchart, S., Button, R., Carraso, R., Cheung, W., … Zhang, Y. P. (2020). Protecting 30% of the planet for nature: Costs, benefits and economic implications [Report]. Campaign for Nature. Link to source: https://www.conservation.cam.ac.uk/files/waldron_report_30_by_30_publish.pdf
Wong, W. W., Greening, C., Shelley, G., Lappan, R., Leung, P. M., Kessler, A., Winfrey, B., Poh, S. C., & Cook, P. (2021). Effects of drift algae accumulation and nitrate loading on nitrogen cycling in a eutrophic coastal sediment. Science of The Total Environment, 790, Article 147749. Link to source: https://doi.org/10.1016/j.scitotenv.2021.147749
Christina Richardson, Ph.D.
Ruthie Burrows, Ph.D.
Avery Driscoll, Ph.D.
James Gerber, Ph.D.
Daniel Jasper
Alex Sweeney
Aiyana Bodi
Avery Driscoll, Ph.D.
Christina Swanson, Ph.D.
Paul C. West, Ph.D.
The globally weighted average effectiveness of seaweed ecosystem protection is 0.32 tCO₂‑eq
/ha/yr. Protecting 1 ha of seaweed ecosystem avoids emissions of 0.043–0.13 tCO₂‑eq
/ha/yr while also sequestering an additional 0.083–0.43 tCO₂‑eq
/ha/yr, with effectiveness higher in subtidal brown than subtidal red seaweed ecosystems (100-yr GWP; Table 1; Appendix).
We estimated effectiveness as the avoided emissions and retained carbon sequestration capacity attributable to the reduction in seaweed ecosystem loss conferred by protection, as detailed in Equation 1. First, we calculated the difference between the rate of seaweed ecosystem loss outside and inside MPAs (Seaweed lossbaseline). We assumed a reduction in loss of 53% (Reduction in loss), which is based on estimates for a range of ecosystems in MPAs (Rodríguez-Rodríguez & Martínez-Vega, 2022). Importantly, this number is highly uncertain and likely to be highly variable, too.
Next, we multiplied this product by the sum of the avoided CO₂
emissions associated with the one-time loss of all above ground biomass carbon in 1 ha of seaweed ecosystem each year over 30 years (Carbonavoided emissions) and the amount of carbon sequestered via long-term storage (on-site or off-site) in 1 ha of protected seaweed ecosystem each year over 30 years (Carbonsequestration).
We based these rates on original analysis of a subset of studies conducted over, at least, 20 years, collated from Krumhansl et al. (2016), that show a median loss rate of 1.2% per year for kelp forests. Due to data limitations, we applied this loss rate to subtidal red seaweed ecosystems as well, but recognize that loss rates are likely to be highly variable. We did this calculation separately for red and brown seaweed ecosystems due to their distinct biomass densities and sequestration capacities, and then averaged the results with accommodations for their relative global areas.
Equation 1.
Table 1. Effectiveness of seaweed ecosystem protection in avoiding emissions and sequestering carbon.
Unit: tCO₂‑eq /ha/yr, 100-year basis
| Avoided emissions, estimate | 0.13 |
| Sequestration | 0.43 |
| Total effectiveness, estimate | 0.56 |
| Total effectiveness, 25th percentile | 0.21 |
| Total effectiveness, 75th percentile | 0.91 |
Unit: tCO₂‑eq /ha/yr, 100-year basis
| Avoided emissions, estimate | 0.043 |
| Sequestration | 0.083 |
| Total effectiveness, estimate | 0.13 |
| Total effectiveness, 25th percentile | 0.034 |
| Total effectiveness, 75th percentile | 0.22 |
Unit: tCO₂‑eq /ha/yr, 100-year basis
| Avoided emissions, estimate | 0.080 |
| Sequestration | 0.24 |
| Total effectiveness, estimate | 0.32 |
| Total effectiveness, 25th percentile | 0.11 |
| Total effectiveness, 75th percentile | 0.52 |
We estimate that seaweed ecosystem protection might save approximately US$72/tCO₂‑eq , but emphasize that these estimates are highly uncertain due to existing data limitations. This is based on protection costs of roughly US$14/ha/yr and revenue of US$43/ha/yr compared with the baseline (Table 2). The costs of seaweed ecosystem protection also include up-front one-time expenditures of US$208 (surveys, administrative setup, legal fees, etc.), estimated from McCrea-Strub et al. (2011). However, data related to the costs of seaweed ecosystem protection are limited, and these estimates are uncertain. For consistency across solutions, we did not include revenue associated with other ecosystem services.
We estimated costs of MPA maintenance at US$14/ha/yr based on data from existing MPAs, though only 16% of MPAs surveyed reported their current funding was sufficient (Balmford et al., 2004). Maintenance is critical for seaweed ecosystems, especially those prone to overgrazing. Tourism revenues directly attributable to protection were estimated to be $43/ha/yr (Waldron et al., 2020) based on estimates for all MPAs (and PAs) and not including downstream revenues. However, estimates of tourism revenues are highly uncertain for seaweed ecosystems. In some seaweed ecosystems, such as kelp forests, tourism is likely a real revenue generator through diving or other recreational activities, but the financial contribution is generally unclear and poorly documented across all seaweed ecosystems.
Table 2. Cost per unit of climate impact. Negative values indicate cost savings.
Unit: 2023 US$/tCO₂‑eq , 100-yr basis
| Estimate | -72 |
A learning curve is defined here as falling costs with increased adoption. The costs of seaweed ecosystem protection do not fall with increasing adoption, so there is no learning curve for this solution.
Speed of action refers to how quickly a climate solution physically affects the atmosphere after it is deployed. This is different from speed of deployment, which is the pace at which solutions are adopted.
At Project Drawdown, we define the speed of action for each climate solution as emergency brake, gradual, or delayed.
Protect Seaweed Ecosystems is an EMERGENCY BRAKE climate solution. It has the potential to deliver a more rapid impact than gradual and delayed solutions. Because emergency brake solutions can deliver their climate benefits quickly, they can help accelerate our efforts to address dangerous levels of climate change. For this reason, they are a high priority.
Additionality is an important caveat for ecosystem protection. In our analysis, we used baseline rates of seaweed ecosystem loss to calculate the effectiveness of protection, which are highly uncertain and understudied. This assumes that seaweed ecosystems would continue to be lost at these rates in the absence of protection and thus that protection provides additional carbon benefits from the ecosystems whose loss is avoided.
Importantly, effective protection depends on adequate funding and management. Poorly managed MPAs can fail to prevent key stressors, such as urchin overgrazing, from increasing and undermine the viability of seaweed ecosystems. Similar dynamics have been documented in kelp restoration efforts, where inadequate management has led to overgrazing and project failure (Eger et al., 2022).
The permanence of ecosystem carbon benefits is another key caveat. While seaweed ecosystems are expanding or expected to expand with climate change, in some regions many will contract (Corrigan et al., 2025). Protection may increase resilience to some climate change stressors, but it will not fully prevent ecosystem loss in many regions. Additionally, because seaweed ecosystems sequester carbon both on-site and off-site, the effectiveness of protection partly depends on downstream activities. For instance, carbon at the seafloor is threatened by disturbances such as bottom fishing and mining (see Protect Seafloors). Protection of seaweed ecosystems does not prevent loss of downstream stored carbon, some of which is contributed by seaweed ecosystems (Ortega et al., 2019). Additionally, seaweed biomass extent can change dramatically from year to year, which could result in substantial variability in carbon removal rates despite protection.
Another caveat in this solution lies in our assumptions about carbon dynamics at the ocean surface. We assume that seaweed NPP results in an equivalent removal of CO₂ from the atmosphere. In reality, this influx may not be fully efficient (Hurd et al., 2024). In some regions of the ocean, water carrying a CO₂ deficit from seaweed photosynthesis might be subducted before it reaches equilibrium with the atmosphere, which would reduce the atmospheric removal attributed to seaweed productivity in our calculations.
In our analysis, avoided emissions are calculated under the assumption that destruction of a seaweed ecosystem results in the loss of all biomass carbon. This likely overestimates near-term emissions, as some carbon may remain in the ocean for long periods. However, this fraction is expected to be small given that an estimated 6.0–13.7% (average: 11.4%) of NPP is thought to be stored long term (Krause-Jensen & Duarte, 2016).
Finally, the relative fraction of NPP removed and durably stored (>100 years) is also uncertain (Pessarrodona et al., 2023). Despite this uncertainty, our use of 11.4% is supported by recent modeling of particulate carbon fluxes that suggest ~12.5% of NPP could be sequestered on a 100-year timescale (based on 44 Tg C of particulate organic carbon export to 1,000 m, where carbon is less likely to return to the atmosphere within a century, and ~353 Tg C as NPP; Filbee-Dexter et al., 2024b), but requires more research.
A total of 78.80 Mha of seaweed ecosystems are currently within MPAs (Table 3). Cumulatively, roughly 18% of seaweed ecosystems are under some form of protection, with 4% located in strictly protected MPAs, 6% in nonstrict MPAs, and 8% in other IUCN protection categories. Subtidal brown and red seaweed ecosystems have similar rates of existing protection in all protection categories (Figure 3).
Table 3. Current (circa 2024) extent of seaweed ecosystems under legal protection. “Strict protection” includes land within IUCN categories I–II Marine Protected Areas (MPAs). “Nonstrict protection” includes land within IUCN Categories III–VI MPAs. “Other” includes land within all remaining IUCN MPA categories. Values may not sum to global totals due to rounding.
Unit: Mha protected
| Strict protection | 8.43 |
| Nonstrict protection | 11.4 |
| Other | 15.5 |
| Total | 35.3 |
Unit: Mha protected
| Strict protection | 9.28 |
| Nonstrict protection | 16.3 |
| Other | 18.0 |
| Total | 43.5 |
Unit: Mha protected
| Strict protection | 17.7 |
| Nonstrict protection | 27.6 |
| Other | 33.4 |
| Total | 78.8 |
We calculated the rate of MPA expansion in seaweed ecosystems based on recorded year of establishment (UNEP-WCMC & IUCN, 2024). Protection expanded by a median of 0.74 Mha/yr in subtidal brown seaweed ecosystems and 0.97 Mha/yr in subtidal red seaweed ecosystems (Table 4; Figure 3a). The global average rate of expansion was roughly 2.13 Mha/yr, with a median of 1.71 Mha/yr. The adoption trend for subtidal brown and red seaweed was relatively similar, with both expanding 0.46–0.55%/yr, on average (median of 0.39–0.40%/yr) (Figure 3b).
Table 4. 2000–2024 adoption trend. Global totals reflect independent statistics, not sums of subtidal brown and red values.
Unit: Mha/yr
| 25th percentile | 0.40 |
| Median (50th percentile) | 0.74 |
| Mean | 1.01 |
| 75th percentile | 1.31 |
Unit: Mha/yr
| 25th percentile | 0.62 |
| Median (50th percentile) | 0.97 |
| Mean | 1.12 |
| 75th percentile | 1.45 |
Unit: Mha/yr
| 25th percentile | 1.02 |
| Median (50th percentile) | 1.71 |
| Mean | 2.13 |
| 75th percentile | 2.76 |
Figure 3. Trend in seaweed ecosystem protection (2000–2024) in terms of (A) total hectares protected and (B) the percent of the current adoption ceiling that is currently protected. These values reflect only the area located within Marine Protected Areas. Units: million hectares protected and percent protected relative to the adoption ceiling.
We estimated that approximately 430 Mha of wild seaweed ecosystems are available for protection (Table 5). Subtidal red seaweeds compose ~240 Mha, with subtidal brown seaweeds occupying the remaining ~190 Mha. These adoption areas do not include other types of seaweed habitats/ecosystems, such as those found in the intertidal zone, rhodolith beds, Halimeda bioherms, coral reefs, and pelagic, free-floating seaweed, which could account for an additional ~150 Mha (Duarte et al., 2022). These adoption areas are highly uncertain due to data limitations and are also likely to shift with climate change.
Table 5. Adoption ceiling: upper limit for the adoption of legal protection of seaweed ecosystems.
Unit: Mha
| Estimate | 189.6 |
Unit: Mha
| Estimate | 243.0 |
Unit: Mha
| Estimate | 432.6 |
We defined the lower end of the achievable range for seaweed ecosystem protection (across all IUCN categories) as 50% of the adoption ceiling and the upper end of the achievable range as 70% of the adoption ceiling (Table 6). These adoption levels are ambitious relative to existing levels of protection (~18%), but align with targets to protect 30% of ecosystems by 2030 (Eger et al., 2024) and serve as an optimistic benchmark for the 30-year time horizon considered in our analysis. Several countries already protect more than 30% of subtidal brown seaweed ecosystems, such as kelp forests (Kelp Forest Alliance, 2024). For example, the United Kingdom, Japan, China, and France protect over 41%, 68%, 68%, and 47% of their kelp beds, respectively.
Table 6: Range of achievable adoption levels for seaweed ecosystems.
Unit: Mha
| Current adoption | 78.8 |
| Achievable – low | 216.3 |
| Achievable – high | 302.9 |
| Adoption ceiling | 432.6 |
We estimated that MPAs currently avoid emissions of 0.03 GtCO₂‑eq/yr in seaweed ecosystems, with potential impacts of 0.14 GtCO₂‑eq/yr at the adoption ceiling (Table 7). Achievable levels of seaweed ecosystem protection could safeguard 0.07 to 0.10 GtCO₂‑eq/yr by reducing emissions from biomass loss and retaining sequestration fluxes (Table 7). However, these estimates are highly uncertain and will benefit from more research (see Caveats).
Limited data exist on the potential climate impacts of seaweed ecosystem protection for comparison. However, a rough estimate of the benefits of conservation, restoration, and afforestation interventions of seaweeds suggests carbon benefits of at least 0.04 GtCO₂‑eq/yr (Pessarrodona et al., 2023). Other estimates suggest that total carbon sequestration in seaweed ecosystems could be on the order of 0.22–0.98 GtCO₂‑eq/yr (Krause-Jensen & Duarte, 2016). This is higher than our estimates because we account only for the carbon benefits of protection in seaweed ecosystems at risk of loss.
Table 7. Climate impact at different levels of adoption. Values may not sum to global totals due to rounding.
Unit: GtCO₂‑eq/yr, 100-year basis
| Current adoption | 0.02 |
| Achievable – low | 0.05 |
| Achievable – high | 0.07 |
| Adoption ceiling | 0.11 |
Unit: GtCO₂‑eq/yr, 100-year basis
| Current adoption | 0.01 |
| Achievable – low | 0.02 |
| Achievable – high | 0.02 |
| Adoption ceiling | 0.03 |
Unit: GtCO₂‑eq/yr, 100-year basis
| Current adoption | 0.03 |
| Achievable – low | 0.07 |
| Achievable – high | 0.10 |
| Adoption ceiling | 0.14 |
Seaweeds can provide coastal resilience to the impacts of storms by lowering wave heights before they reach shorelines (Corrigan et al., 2025; Cotas et al., 2023). The magnitude of this benefit can vary based on the species and location of seaweed, and some evidence suggests that severe storms can harm seaweed habitats (Earp et al., 2024). Evidence suggests that kelp forests can attenuate wave heights locally, especially in the summer at peak kelp growth, but protection varies at larger spatial scales (Elsmore et al., 2024; Lindhart et al., 2024). Emerging research has found that protected seaweed ecosystems show more resilience to marine heat waves than unprotected areas (Kumagai et al., 2024). During heat waves, protected ecosystems maintain a habitat for species such as sea urchins that consume species that might degrade kelp ecosystems (Bauer et al., 2025; Kumagai et al., 2024).
Seaweeds support species that are important for tourism and fishing (Cuba et al., 2022; Eger et al., 2023). Many species that are supported by seaweeds have high economic value for fishing, such as crabs, lobsters, and abalones (Corrigan et al., 2025). For example, Eger et al. (2023) estimated that 1 ha of kelp forest where about 900 kg of fish biomass is harvested could yield about US$29,900 a year. The same study estimated that the global value of kelp forests that support fisheries is about US$465–562 billion (Eger et al., 2023). Seaweed habitats can also be tourist destinations for snorkeling and diving (UNEP, 2023), providing income-earning opportunities for nearby communities.
The contribution of seaweeds to fisheries production can play a role in global food security (Cottier-Cook et al., 2023; Eger et al., 2023). Additionally, seaweeds are an essential part of many diets, especially in East Asia (FAO, 2024). Because seaweeds are a culturally important food in many geographies, protecting seaweeds can play an important role in equitably improving global food security (FAO, 2024).
For some cultures, seaweeds and their habitats shape shared identities and livelihoods (Cotas et al., 2023). For example, seaweeds are a source of traditional foods, medicines, art, and knowledge for many coastal communities and Indigenous peoples (Thurstan et al., 2018). Protecting seaweeds can preserve the cultural identities, practices, and knowledge of Indigenous communities that are often vulnerable (Corrigan et al., 2025).
Seaweeds support biodiversity by providing habitat for a variety of marine species (Best et al., 2014; Cuba et al., 2022; Gibbons & Quijón, 2023; Tano et al., 2016). Literature reviews of the ecosystem services of seaweeds find that they contribute to increases in biodiversity (Gibbons & Quijón, 2023). Seaweeds can provide habitat and refuge from large predators (Best et al., 2014; Gibbons & Quijón, 2023). Invertebrates, detritivores, and other small species found in seaweed forests are essential food sources for other marine species (Cuba et al., 2022; Tano et al., 2016).
Seaweeds improve water quality by supporting nutrient cycling and reducing pollutants (Cotas et al., 2023; Heckwolf et al., 2021). Evidence suggests that seaweeds can reduce eutrophication by filtering excess nutrients from the water (Corrigan et al., 2025; Gao et al., 2022; Heckwolf et al., 2021).
Leakage, in which protecting one ecosystem results in the degradation of another, could offset the climate impact of seaweed ecosystem protection. For instance, restricting wild harvesting through the establishment of an MPA could shift pressure to other unprotected areas. Another key risk is weakly enforced or poorly managed MPAs. This is a real concern with existing MPAs due to a lack of funding, and can result in low protection effectiveness. Finally, climate change stressors, such as ocean warming and marine heat waves, are a major risk to permanence because they could lead to widespread mortality, even in protected areas.
Intact and healthy seaweed ecosystems can enhance fish stocks, biodiversity, and habitat quality, which benefits all connected coastal and marine ecosystems.
Protecting seaweed ecosystems can help ensure the underlying areas of the seafloor remain intact.
Protection of seaweed ecosystems could potentially reduce the adoption of offshore wind in some regions.
ha of seaweed ecosystem protected
CO₂
Seaweed ecosystems can release methane, which could reduce the climate benefits of protection estimated in this solution. While data are scarce, a recent study suggests that methane emissions could offset 28–35% of the carbon sink capacity in some seaweed ecosystems (Roth et al., 2023) if they escape to the atmosphere, which may be unlikely if methane production occurs at depth in sediments (Pessarrodona et al., 2023).
Seaweed ecosystems cover approximately 600 Mha of the ocean, an area twice the size of Argentina (Duarte et al., 2022). Here we show the estimated global distribution of subtidal red and brown seaweeds, though large uncertainty exists in their spatial distribution and extent.
Duarte, C. M., Gattuso, J. P., Hancke, K., Gundersen, H., Filbee-Dexter, K., Pedersen, M. F., Middelburg, J. J., Burrows, M. T., Krumhansl, K. A., Wernberg, T., Moore P., Pessarrodona, A., Ørberg, S. B., Pinto, I. S., Assis, J., Queirós, A. M., Smale, D. A., Bekkby, T., Serrão, E. A., & Krause-Jensen, D. (2022). Global estimates of the extent and production of macroalgal forests. Global Ecology and Biogeography, 31(7), 1422–1439. Link to source: https://doi.org/10.1111/geb.13515
Seaweed ecosystems cover approximately 600 Mha of the ocean, an area twice the size of Argentina (Duarte et al., 2022). Here we show the estimated global distribution of subtidal red and brown seaweeds, though large uncertainty exists in their spatial distribution and extent.
Duarte, C. M., Gattuso, J. P., Hancke, K., Gundersen, H., Filbee-Dexter, K., Pedersen, M. F., Middelburg, J. J., Burrows, M. T., Krumhansl, K. A., Wernberg, T., Moore P., Pessarrodona, A., Ørberg, S. B., Pinto, I. S., Assis, J., Queirós, A. M., Smale, D. A., Bekkby, T., Serrão, E. A., & Krause-Jensen, D. (2022). Global estimates of the extent and production of macroalgal forests. Global Ecology and Biogeography, 31(7), 1422–1439. Link to source: https://doi.org/10.1111/geb.13515
While seaweed ecosystems are widely distributed across the global ocean, substantial uncertainty exists surrounding their distribution and extent (Kelp Forest Alliance, 2024; Duarte et al., 2022). The effectiveness of protection varies geographically and is likely to be highest in regions where the resource demands of management and enforcement are relatively low and chronic ecological threats are minimal or easy to address. However, because datasets needed to identify such priority regions are limited, we focus here on broad geographic patterns in seaweed extent and distribution.
Based on the most recent estimates of global extent, subtidal red and brown seaweeds cover nearly an order of magnitude greater area (~430 Mha) of the world’s oceans than do coastal wetlands (~55 Mha, see Protect Coastal Wetlands). Areas with subtidal brown seaweeds, such as kelp forests, are more effective per hectare than those with subtidal red seaweeds at mitigating climate change due to their generally higher biomass densities. Kelp forests are common in 25 countries. Canada, the United States, Russia, the United Kingdom, and Denmark contain more than 60% of the world’s kelp forests (~106.8 Mha), particularly in their high polar regions (Kelp Forest Alliance, 2024). In the Southern Hemisphere, Australia, New Zealand, Chile, and Argentina contain the greatest area of kelp forests (~27.8 Mha). Kelp forests in other countries are also critically important for sequestering carbon. Sixteen of the 25 countries with kelp forests have less than 5 Mha, and the median country-level extent is 4.2 Mha, highlighting the need for protection in these areas to achieve globally relevant levels of adoption. No country-level analyses exist for subtidal red seaweed ecosystems, and more research is needed to understand potential priority regions for subtidal red seaweed ecosystem protection.
Consensus of effectiveness at reducing emissions and maintaining carbon removal: Mixed
There is mixed scientific consensus that protection prevents the degradation of seaweed ecosystems, but high consensus that degradation leads to losses in biomass carbon stocks and sequestration capacity. Seaweed ecosystems can be degraded by diverse stressors that directly or indirectly affect biomass stocks. Management actions, such as establishment of MPAs, can help prevent both direct and indirect habitat loss and thereby maintain the carbon removal capacity of seaweed ecosystems with relatively high certainty against stressors such as wild harvesting, coastal development, overgrazing, and poor water quality (Pessarrodona et al., 2023). However, some stressors, such as marine heat waves and ocean warming, are less effectively addressed by protection alone (Filbee-Dexter et al., 2024a). Benefits are still expected in some systems because MPAs can enhance resilience and recovery by reducing co-occurring stressors common that contribute to seaweed ecosystem degradation (Krumhansl et al., 2016; Ortiz-Villa et al., 2025). Moreover, MPAs, even when established in areas with addressable stressors, are typically not fully effective. Here, we applied a protection effectiveness of 53%, based on aggregated estimates from MPAs beyond seaweed ecosystems (Rodríguez-Rodríguez & Martínez-Vega, 2022). If the effectiveness of protection is lower (higher), climate impacts could likewise be lower (higher).
There is high scientific consensus that degradation of seaweed ecosystems leads to losses in biomass carbon stocks and sequestration capacity. While direct estimates of CO₂ emissions from biomass are limited, degradation has been shown to remove biomass carbon and reduce sequestration. For instance, drivers of habitat loss and degradation, such as overharvesting (González-Roca et al., 2021; Steen et al., 2016), overgrazing (Akaike & Mizuta, 2024), and poor water quality (Filbee-Dexter & Wernberg, 2020), reduce standing biomass and therefore associated carbon export from seaweed ecosystems (Pessarrodona et al., 2023).
The carbon sink capacity of seaweed ecosystems, such as kelp forests, is also expected to decline with climate change stressors such as warming, which can increase rates of decomposition by 9–42% (Filbee-Dexter et al., 2022) and drive habitat loss, both of which reduce the likelihood that carbon makes its way to the deep sea for long-term storage. Off the coast of Australia, over 140,000 ha of subtidal brown seaweed forests have already been lost to warming over two decades, representing a decline of 2–4% of regional seaweed biomass carbon stocks and sequestration capacity (Filbee-Dexter & Wernberg, 2020).
The results presented in this assessment synthesize findings from 5 global datasets. We recognize that geographic bias in the information underlying global data products creates bias, and hope this work inspires research and data sharing on this topic in underrepresented regions and on understudied aspects of these ecosystems.
This analysis quantifies emissions that can be avoided by protecting seaweed ecosystems via the establishment of Marine Protected Areas (MPAs). We leveraged two global seaweed distribution maps alongside a shapefile of MPAs, available data on rates of avoided ecosystem loss attributable to MPA establishment, and global data on biomass carbon stores and carbon sequestration rates to calculate climate impacts. This appendix describes the source data products and how they were integrated.
Seaweed Ecosystem Extent
We relied on the global maps of seaweed extent developed by Duarte et al. (2022), which classify subtidal brown and red seaweeds (among others). We used the “LT2 Brown Algae Benthic” raster to calculate subtidal brown seaweed extent and the “LT2 Red Algae Benthic” raster to calculate subtidal red seaweed extent. We did not consider red seaweed in subtidal brown-dominant environments, such as kelp forests, due to existing limitations with the global maps.
Protected Seaweed Ecosystem Areas
We identified protected seaweed ecosystem areas using the World Database on Protected Areas (UNEP-WCMC & IUCN, 2024), which contains boundaries for each MPA and additional information, including the establishment year and IUCN management category (Ia to VI, not applicable, not reported, or not assigned). In this analysis, we considered all categories. While some MPA categories likely allow for wild harvest, which can be unsustainably conducted, wild seaweed harvest is currently estimated at 1.3 Mt/yr (wet weight) (FAO, 2024), which represents a relatively small portion of the global loss rate used (<0.2%/yr). We converted the MPA boundary data to a raster and used them to calculate the seaweed area within MPA boundaries for each seaweed type analyzed (subtidal brown and red) and each MPA category. To evaluate trends in adoption over time, we also aggregated protected areas by establishment year as reported in the WDPA.
Calculation of Effectiveness
The following equations show a detailed breakdown of the stepwise set of calculations used to implement Equation 1, including estimation of avoided seaweed loss and of emissions and retained sequestration across the 30-year time horizon considered.
Avoided Seaweed Ecosystem Conversion
We compiled baseline estimates of seaweed ecosystem loss (%/yr) from existing literature and used them in conjunction with an estimate of reductions in loss associated with protection of 53% (derived from Rodríguez-Rodríguez & Martínez-Vega, 2022) to calculate the rate of avoidable macroalgae loss (Seaweed lossavoided). Seaweed ecosystem loss rates were based on the original analysis of data aggregated from Krumhansl et al. (2016) for studies over 20 years long (Seaweed lossbaseline; median loss rate of 1.2%/yr).
Equation A1.
We then used the avoidable seaweed loss rates to calculate avoided CO₂ emissions and additional carbon sequestration for each adoption unit. Specifically, we estimated the carbon benefits of avoided seaweed ecosystem loss by multiplying avoided seaweed ecosystem loss by avoided CO₂ emissions (Equation A2) and by applying carbon sequestration rates over 30 years (Equation A3) for each seaweed type.
We estimated avoided CO₂ emissions by assuming a one-time release of all aboveground biomass carbon upon loss. We derived our estimates of retained carbon sequestration from global databases on NPP for each seaweed type from Duarte et al. (2022) and a global estimate of NPP-derived sequestration (11.4%) from NPP based on Krause-Jensen and Duarte (2016).
Equation A2.
Equation A3.
We then estimated effectiveness (Equation A4) as the avoided CO₂ emissions and retained carbon sequestration capacity attributable to the reduction in seaweed ecosystem loss conferred by protection estimated in Equations A1–3.
Equation A4.
Restore Salt Marsh Ecosystems involves actively reestablishing salt marshes in areas where they were previously lost to conversion or other disturbance, allowing vegetation to regrow and carbon to accumulate in biomass and sediments. Advantages include salt marshes’ ability to durably store substantial quantities of carbon over long time periods and their numerous co-benefits for the environment and humans. Disadvantages include variable but potentially low effectiveness due to site-to-site differences in carbon removal rates and potential emissions of other GHGs, such as methane and nitrous oxide, as well as costs that might exceed US$500/t CO₂‑eq in some areas. Salt marsh restoration is not expected to have a globally meaningful climate impact (>0.1 Gt CO₂‑eq/yr ), primarily because the adoption ceiling is constrained by the limited area available for restoration, but there are no major environmental risks associated with the solution. Therefore, Restore Salt Marsh Ecosystems is “Worthwhile.”
Based on our analysis, restoring salt marsh ecosystems is a “Worthwhile” carbon removal technique that is ready for large-scale deployment. While the capacity for adoption is limited, limiting climate impact, this solution has no major risks and provides widespread added benefits for people and the environment.
| Plausible | Could it work? | Yes |
|---|---|---|
| Ready | Is it ready? | Yes |
| Evidence | Are there data to evaluate it? | Yes |
| Effective | Does it consistently work? | Yes |
| Impact | Is it big enough to matter? | No |
| Risk | Is it risky or harmful? | No |
| Cost | Is it cheap? | Yes |
Restore Salt Marsh Ecosystems removes carbon from the air by reestablishing salt marshes in areas where they were previously drained, filled, or otherwise degraded and lost. As plants take up CO₂ through photosynthesis and vegetation traps sediments, some of this carbon is stored long term in waterlogged soils with slow decomposition rates. Restoration typically reconnects land to tidal exchange and rebuilds marsh elevation and vegetation, which promotes plant growth and sediment accumulation. Active restoration can include breaching levees or removing barriers to restore tidal flow, regrading or adding sediment to raise elevations, planting native marsh vegetation, and controlling invasive species. In many cases, restoration can also reduce GHG emissions by replacing land uses, such as drained agriculture, that emit CO₂.
The fundamental idea of restoring salt marsh ecosystems is scientifically sound, and, on average globally, restored salt marshes have been shown to remove carbon over long timescales through vegetation recovery and sustained carbon burial in waterlogged soils, even after accounting for methane and nitrous oxide emissions. This solution has been in practice worldwide for many decades, and global assessments suggest it could expand to roughly 2 million hectares because ~67% of salt marshes have been destroyed since the early 1900s. Restoration success rates are high relative to those of many other marine habitats. However, its potential adoption ceiling is still low relative to other nature-based solutions (e.g., Restore Forests) because restoration is limited to suitable coastal areas, which are constrained by coastal development and other human stressors. As a result, its climate impact is likely well below 0.1 Gt CO₂‑eq /year.
Restoration of salt marsh ecosystems is a well-established, scalable practice with many benefits for the environment. Restored salt marshes can reduce shoreline erosion and costal flooding, improve water quality by retaining nutrients and sediments, and provide habitat for fish and birds. While global impact is limited, this intervention can be an important multi-benefit tool for building climate resilience and removing carbon in some countries and coastal regions. Restoration is already widely implemented. In some restorations, such as those that reestablish tidal exchange in previously impounded ecosystems, increases in salinity can reduce methane and nitrous oxide production relative to pre-restoration conditions.
The climate impact of salt marsh restoration is constrained by its limited adoption ceiling, variable but potentially high costs, vulnerability to future loss, and potentially low effectiveness. Adoption is limited by where marshes can actually be restored, such as on low-elevation coastal lands that are not heavily developed, and where they can be maintained into the future with climate change stressors, such as sea-level rise. If salt marshes are not restored with consideration of projected sea level rise, loss or conversion to mud flats or open water habitats in the future is possible, which would result in the loss of carbon benefits. Restored salt marshes can also emit potent GHGs such as methane and nitrous oxide as low oxygen conditions and ecosystem function are reestablished, which can offset some of the climate benefits of restoration. As a result, costs vary widely by site, and can exceed US$500/t CO₂‑eq (~US$1,000–7,000/ha), depending on site-specific effectiveness rates. Additionally, few data are available for understanding long-term, multi-decadal changes in carbon accumulation rates in restored sites, and some regions remain underrepresented globally.
Burden, A., Garbutt, A., & Evans, C. D. (2019). Effect of restoration on saltmarsh carbon accumulation in Eastern England. Biology Letters, 15(1). Link to source: https://doi.org/10.1098/rsbl.2018.0773
Convention on Wetlands. (2025). Global Wetland Outlook 2025: Valuing, conserving, restoring and financing wetlands (Scientific and Technical Review Panel report). Secretariat of the Convention on Wetlands. Link to source: https://www.ramsar.org/launch-global-wetland-outlook-2025
Danovaro, R., Aronson, J., Bianchelli, S., Boström, C., Chen, W., Cimino, R., Corinaldesi, C., Cortina-Segarra, J., D’Ambrosio, P., Gambi, C. and Garrabou, J. (2025). Assessing the success of marine ecosystem restoration using meta-analysis. Nature Communications, 16(1), 3062. Link to source: https://doi.org/10.1038/s41467-025-57254-2
Holmquist, J. R., Eagle, M., Molinari, R. L., Nick, S. K., Stachowicz, L. C., & Kroeger, K. D. (2023). Mapping methane reduction potential of tidal wetland restoration in the United States. Communications Earth & Environment, 4(1), 353. Link to source: https://doi.org/10.1038/s43247-023-00988-y
Mason, V. G., Burden, A., Epstein, G., Jupe, L. L., Wood, K. A., & Skov, M. W. (2024). Navigating research challenges to estimate blue carbon benefits from saltmarsh restoration. Global Change Biology, 30(10), 1–3. Link to source: https://doi.org/10.1111/gcb.17526
Pétillon, J., McKinley, E., Alexander, M., Adams, J.B., Angelini, C., Balke, T., Griffin, J.N., Bouma, T., Hacker, S., He, Q. and Hensel, M.J. (2023). Top ten priorities for global saltmarsh restoration, conservation and ecosystem service research. Science of the Total Environment, 898, 165544. Link to source: https://doi.org/10.1016/j.scitotenv.2023.165544
Reilly, A. V., Merrill, N. H., Mulvaney, K. K., Colarusso, P., & Burman, E. (2024). Fantastic wetlands and why to monitor them: Demonstrating the social and financial benefit potential of methane abatement through salt marsh restoration. PLOS Climate, 3(7), e0000317. Link to source: https://doi.org/10.1371/journal.pclm.0000317
Rolando, J., Hodges, M., Garcia, K., Krueger, G., Williams, N., Carr Jr, J., Robinson, J., George, A., Morris, J. and Kostka, J., (2023). Restoration and resilience to sea level rise of a salt marsh affected by dieback events. Ecosphere, 14(4), e4467. Link to source: https://doi.org/10.1002/ecs2.4467
Rowland, P. I., Wartman, M., Bursic, J., & Carnell, P. (2024). Restored and created tidal marshes recover ecosystem services over time. Environmental and Sustainability Indicators, 24, Article 100539. Link to source: https://doi.org/10.1016/j.indic.2024.100539
Taillardat, P., Thompson, B. S., Garneau, M., Trottier, K., & Friess, D. A. (2020). Climate change mitigation potential of wetlands and the cost-effectiveness of their restoration. Interface focus, 10(5). Link to source: https://doi.org/10.1098/rsfs.2019.0129
Williamson, P., Schlegel, R. W., Gattuso, J. P., Andrews, J. E., & Jickells, T. D. (2024). Climate benefits of saltmarsh restoration greatly overstated by Mason et al. (2023). Global Change Biology, 30(10). Link to source: https://doi.org/10.1111/gcb.17525
WWF UK. (2025, June 11). Vanishing saltmarshes threaten climate progress – but recovery is within reach, says new global report [Press release]. WWF UK. Link to source: https://www.wwf.org.uk/our-reports/state-worlds-saltmarshes
Christina Richardson, Ph.D.
Christina Swanson, Ph.D.
Paul West, Ph.D.
Peatland restoration is the process of returning peat-forming vegetation and natural hydrologic conditions to degraded peatlands. Restoration of peatlands stops GHG emissions from peat degradation and removes carbon through sequestration in biomass and peat.
This solution focuses on non-coastal peatlands that have been drained, cleared of vegetation, harvested for peat, or otherwise degraded. Coastal peatlands are addressed in the Restore Mangrove Ecosystems and Restore Salt Marsh Ecosystems solutions.
Peatlands are diverse ecosystems characterized by waterlogged, carbon-rich soils consisting of partially decomposed dead plant material (Figure 1). Because decomposition occurs very slowly under low oxygen, waterlogged conditions, large amounts of partially decomposed plant material accumulated over millennia in peatlands. These carbon-rich ecosystems occupy only 3–4% of land area (Xu et al., 2018; United Nations Environment Programme [UNEP], 2022), but store an estimated 600 Gt carbon (~2,200 Gt CO₂‑eq), roughly twice as much carbon as is stored in forest biomass globally (UNEP, 2022; Yu et al., 2010; Pan et al., 2024). Approximately 12% of global peatlands (~57 Mha) have been drained or otherwise degraded for agriculture, forestry, peat extraction, or other uses. Restoration of degraded peatlands through revegetation and rewetting (returning to a waterlogged state; Evans et al., 2021) reduces ongoing emissions and can restore the ability of the ecosystem to sequester carbon.
Figure 1. These photos show the diversity of peatlands that occur in different places, including a fen peatland and meadow complex in California (top left), a peat swamp in Indonesia (top right), a peat fen and forest in Canada (bottom left), and a peat bog in New Hampshire (bottom right).
Photo credits: Catie and Jim Bishop | U.S. Department of Agriculture; Rhett A. Butler; Garth Lenz; Linnea Hanson | U.S. Department of Agriculture
Draining and/or clearing peatlands increases CO₂ emissions, nitrous oxide emissions, and losses of dissolved carbon through waterways while also reducing carbon sequestration and methane emissions. CO₂ from peat decomposition is the largest source of GHG emissions from degraded peatlands (Intergovernmental Panel on Climate Change [IPCC] Task Force on National Greenhouse Gas Inventories, 2014; UNEP, 2022). When accumulated organic matter in the peat layer is exposed to air, it begins decomposing more rapidly, emitting CO₂ (Figure 2). Removal of overlying vegetation produces additional GHG emissions while also slowing or stopping carbon uptake and continued peat formation. Whereas emissions from vegetation removal occur rapidly following disturbance, peat decomposition and associated emissions can continue for centuries depending on environmental conditions and peat thickness (Leifield & Menichetti, 2018). Peatland restoration halts these emissions and, eventually, can restore the ecosystem's ability to take up carbon (Doelman et al., 2023; Humpenöder et al., 2020; Mander et al., 2024; Mander et al., 2025; Strack et al., 2022).
Peatland disturbance and restoration also have complex impacts on nitrous oxide emissions, methane emissions, and carbon loss through waterways (Figure 2; IPCC Task Force on National Greenhouse Gas Inventories, 2014; UNEP, 2022). Intact peatlands are a methane source because methane-producing microbes thrive under waterlogged conditions, so restoring peatlands typically increases methane emissions (Evans et al., 2021; Günther et al., 2020; Huang et al., 2021; Mander et al., 2024). However, restoration typically reduces nitrous oxide emissions from oxidation of the peat layer.
Figure 2. Intact peatlands (left) are a net greenhouse gas sink, sequestering carbon in peat through photosynthesis but also emitting methane due to waterlogged soils. Drained peatlands (right) are a GHG source, producing emissions from peat decomposition and drainage canals. Modified from the International Union for the Conservation of Nature (IUCN) UK Peatland Programme (2024).
Source: IUCN UK Peatland Programme. (2024, July 10). New briefing addresses the peatlands and methane debate.
Peatlands can be restored to native vegetation or used for paludiculture, the practice of growing crops or trees in saturated conditions (Tan et al., 2021; Temmink et al., 2026). We assume that effective restoration involves both rewetting and re-establishment of perennial vegetation, and in this assessment we did not differentiate between carbon dynamics on peatlands restored to native vegetation and those on peatlands used for paludiculture. Because peatland emissions dynamics, total peatland areas, and the degraded peatland area all vary geographically, we evaluated tropical, subtropical, temperate, and boreal regions separately in this analysis. Most peatlands occur in boreal regions (~310 Mha, or 64% of the total), with another 25% (~121 Mha) in tropical regions (UNEP, 2022; Protect Peatlands). These regions also contain the largest areas of degraded peatlands that can be restored (26 Mha and 19 Mha, respectively; UNEP, 2022).
Halting the ongoing drainage and degradation of peatlands through effective protection, which is covered in Protect Peatlands, is an essential complement to restoration (Austin et al., 2025).
Andersen, R., Farrell, C., Graf, M., Muller, F., Calvar, E., Frankard, P., Caporn, S., & Anderson, P. (2017). An overview of the progress and challenges of peatland restoration in Western Europe. Restoration Ecology, 25(2), 271–282. Link to source: https://doi.org/10.1111/rec.12415
Austin, K. G., Elsen, P. R., Coronado, E. N. H., DeGemmis, A., Gallego-Sala, A. V., Harris, L., Kretser, H. E., Melton, J. R., Murdiyarso, D., Sasmito, S. D., Swails, E., Wijaya, A., Winton, R. S., & Zarin, D. (2025). Mismatch between global importance of peatlands and the extent of their protection. Conservation Letters, 18(1), Article e13080. Link to source: https://doi.org/10.1111/conl.13080
Bonn, A., Reed, M. S., Evans, C. D., Joosten, H., Bain, C., Farmer, J., Emmer, I., Couwenberg, J., Moxey, A., Artz, R., Tanneberger, F., von Unger, M., Smyth, M.-A., & Birnie, D. (2014). Investing in nature: Developing ecosystem service markets for peatland restoration. Ecosystem Services, 9, 54–65. Link to source: https://doi.org/10.1016/j.ecoser.2014.06.011
Breznikar, A., Pönisch, D. L., Lorenz, M., Jurasinski, G., Rehder, G., & Voss, M. (2024). Rewetting effects on nitrogen cycling and nutrient export from coastal peatlands to the Baltic Sea. Biogeochemistry, 167(7), 967–987. Link to source: https://doi.org/10.1007/s10533-024-01149-9
Burney, J. A., Davis, S. J., & Lobell, D. B. (2010). Greenhouse gas mitigation by agricultural intensification. Proceedings of the National Academy of Sciences, 107(26), 12052–12057. Link to source: https://doi.org/10.1073/pnas.0914216107
Busch, J., Bukoski, J. J., Cook-Patton, S. C., Griscom, B., Kaczan, D., Potts, M. D., Yi, Y., & Vincent, J. R. (2024). Cost-effectiveness of natural forest regeneration and plantations for climate mitigation. Nature Climate Change, 14(9), 996–1002. Link to source: https://doi.org/10.1038/s41558-024-02068-1
Dietrich, O., & Behrendt, A. (2022). Wet grassland sites with shallow groundwater conditions: effects on local meteorological characteristics. Water, 14(21), Article 21. Link to source: https://doi.org/10.3390/w14213560
Doelman, J. C., Verhagen, W., Stehfest, E., & Vuuren, D. P. van. (2023). The role of peatland degradation, protection and restoration for climate change mitigation in the SSP scenarios. Environmental Research: Climate, 2(3), Article 035002. Link to source: https://doi.org/10.1088/2752-5295/acd5f4
Evans, C. D., Peacock, M., Baird, A. J., Artz, R. R. E., Burden, A., Callaghan, N., Chapman, P. J., Cooper, H. M., Coyle, M., Craig, E., Cumming, A., Dixon, S., Gauci, V., Grayson, R. P., Helfter, C., Heppell, C. M., Holden, J., Jones, D. L., Kaduk, J., … Morrison, R. (2021). Overriding water table control on managed peatland greenhouse gas emissions. Nature, 593(7860), 548–552. Link to source: https://doi.org/10.1038/s41586-021-03523-1
Global Peatlands Initiative. (2025). Info brief: The peatland breakthrough science-based framework for global peatland targets & guiding principles. Link to source: https://globalpeatlands.org/resource-library/info-brief-peatland-breakthrough-science-based-framework-global-peatland-targets
Goib, B. K., Fitriani, N., Wicaksono, S., & Chitra, J. (2018). Restoring peat, improving welfare, and empowering women: Can we have it all? Link to source: https://wri-indonesia.org/en/insights/restoring-peat-improving-welfare-and-empowering-women-can-we-have-it-all
Granath, G., Moore, P. A., Lukenbach, M. C., & Waddington, J. M. (2016). Mitigating wildfire carbon loss in managed northern peatlands through restoration. Scientific Reports, 6(1), Article 28498. Link to source: https://doi.org/10.1038/srep28498
Günther, A., Barthelmes, A., Huth, V., Joosten, H., Jurasinski, G., Koebsch, F., & Couwenberg, J. (2020). Prompt rewetting of drained peatlands reduces climate warming despite methane emissions. Nature Communications, 11(1), Article 1644. Link to source: https://doi.org/10.1038/s41467-020-15499-z
Harrison, M. E., Ottay, J. B., D’Arcy, L. J., Cheyne, S. M., Anggodo, Belcher, C., Cole, L., Dohong, A., Ermiasi, Y., Feldpausch, T., Gallego-Sala, A., Gunawan, A., Höing, A., Husson, S. J., Kulu, I. P., Soebagio, S. M., Mang, S., Mercado, L., Morrogh-Bernard, H. C., … van Veen, F. J. F. (2020). Tropical forest and peatland conservation in Indonesia: Challenges and directions. People and Nature, 2(1), 4–28. Link to source: https://doi.org/10.1002/pan3.10060
Hein, L., Spadaro, J. V., Ostro, B., Hammer, M., Sumarga, E., Salmayenti, R., Boer, R., Tata, H., Atmoko, D., & Castañeda, J.-P. (2022). The health impacts of Indonesian peatland fires. Environmental Health, 21(1), Article 62. Link to source: https://doi.org/10.1186/s12940-022-00872-w
Helbig, M., Waddington, J. M., Alekseychik, P., Amiro, B., Aurela, M., Barr, A. G., Black, T. A., Carey, S. K., Chen, J., Chi, J., Desai, A. R., Dunn, A., Euskirchen, E. S., Flanagan, L. B., Friborg, T., Garneau, M., Grelle, A., Harder, S., Heliasz, M., … Schulze, C. (2020). The biophysical climate mitigation potential of boreal peatlands during the growing season. Environmental Research Letters, 15(10), Article 104004. Link to source: https://doi.org/10.1088/1748-9326/abab34
Heuts, T. S., Van Der Scheer, O. E., Harpenslager, S. F., Van Diggelen, J. M. H., Smolders, A. J. P., & Fritz, C. (2026). Phosphorus saturation ratio predicts phosphorus mobilization in peat and mineral soils after rewetting. Ecological Engineering, 230, Article 108019. Link to source: https://doi.org/10.1016/j.ecoleng.2026.108019
Hoffmann, C. C., Petersen, R. J., Genieys, A., Ovesen, N. B., Kjeldgaard, A., Egemose, S., & Zak, D. (2026). Reduced phosphorus loss in a pump‐drained wetland after harvesting and rewetting. Journal of Environmental Quality, 55(2), Article e70163. Link to source: https://doi.org/10.1002/jeq2.70163
Huang, Y., Ciais, P., Luo, Y., Zhu, D., Wang, Y., Qiu, C., Goll, D. S., Guenet, B., Makowski, D., De Graaf, I., Leifeld, J., Kwon, M. J., Hu, J., & Qu, L. (2021). Tradeoff of CO2 and CH4 emissions from global peatlands under water-table drawdown. Nature Climate Change, 11(7), 618–622. Link to source: https://doi.org/10.1038/s41558-021-01059-w
Humpenöder, F., Karstens, K., Lotze-Campen, H., Leifeld, J., Menichetti, L., Barthelmes, A., & Popp, A. (2020). Peatland protection and restoration are key for climate change mitigation. Environmental Research Letters, 15(10), Article 104093. https://doi.org/10.1088/1748-9326/abae2a
IPCC Task Force on National Greenhouse Gas Inventories. (2014). 2013 supplement to the 2006 IPCC guidelines for national greenhouse gas inventories: Wetlands (T. Hiraishi, T. Krug, K. Tanabe, N. Srivastava, J. Baasansuren, M. Fukuda, & T. G. Troxler, Eds.). Intergovernmental Panel on Climate Change. Link to source: https://www.ipcc.ch/site/assets/uploads/2018/03/Wetlands_Supplement_Entire_Report.pdf
IUCN. Peatlands and Climate Change (IUCN Issues Briefs). (2021). Link to source: https://iucn.org/sites/default/files/2022-04/iucn_issues_brief_peatlands_and_climate_change_final_nov21.pdf
Jong, H. N. (2023). Indonesia's peatland restoration claims in question as fires flare up. Mongabay. Link to source: https://news.mongabay.com/2023/10/indonesias-peatland-restoration-claims-in-question-as-fires-flare-up/
Kiely, L., Spracklen, D. V., Arnold, S. R., Papargyropoulou, E., Conibear, L., Wiedinmyer, C., Knote, C., & Adrianto, H. A. (2021). Assessing costs of Indonesian fires and the benefits of restoring peatland. Nature Communications, 12(1), Article 7044. Link to source: https://doi.org/10.1038/s41467-021-27353-x
Kreyling, J., Tanneberger, F., Jansen, F., van der Linden, S., Aggenbach, C., Blüml, V., Couwenberg, J., Emsens, W.-J., Joosten, H., Klimkowska, A., Kotowski, W., Kozub, L., Lennartz, B., Liczner, Y., Liu, H., Michaelis, D., Oehmke, C., Parakenings, K., Pleyl, E., … Jurasinski, G. (2021). Rewetting does not return drained fen peatlands to their old selves. Nature Communications, 12(1), Article 5693. Link to source: https://doi.org/10.1038/s41467-021-25619-y
Leifeld, J., & Menichetti, L. (2018). The underappreciated potential of peatlands in global climate change mitigation strategies. Nature Communications, 9(1), Article 1071. Link to source: https://doi.org/10.1038/s41467-018-03406-6
Leifeld, J., Wüst-Galley, C., & Page, S. (2019). Intact and managed peatland soils as a source and sink of GHGs from 1850 to 2100. Nature Climate Change, 9(12), 945–947. Link to source: https://doi.org/10.1038/s41558-019-0615-5
Lestari, N. S., Rochmayanto, Y., Salminah, M., Novita, N., Asyhari, A., Gangga, A., Ritonga, R., Yeo, S., & Albar, I. (2024). Opportunities and risk management of peat restoration in Indonesia: Lessons learned from peat restoration actors. Restoration Ecology, 32(1), Article e14054. Link to source: https://doi.org/10.1111/rec.14054
Lloyd, I. L., Thomas, V., Ofoegbu, C., Bradley, A. V., Bullard, P., D’Acunha, B., Delaney, B., Driver, H., Evans, C. D., Faulkner, K. J., Fonvielle, J. A., Francksen, R. M., Friday, L. E., Hose, G., Kaduk, J., Manning, F. R., Morrison, R., Novo, P., Page, S. E., … Balzter, H. (2023). State of Knowledge on UK Agricultural Peatlands for Food Production and the Net Zero Transition. Sustainability, 15(23), Article 16347. Link to source: https://doi.org/10.3390/su152316347
Lobell, D. B., & Villoria, N. B. (2023). Reduced benefits of climate-smart agricultural policies from land-use spillovers. Nature Sustainability, 6, 941–948. Link to source: https://doi.org/10.1038/s41893-023-01112-w
Loisel, J., & Gallego-Sala, A. (2022). Ecological resilience of restored peatlands to climate change. Communications Earth & Environment, 3(1), Article 208. Link to source: https://doi.org/10.1038/s43247-022-00547-x
Mander, Ü., Espenberg, M., Melling, L., & Kull, A. (2024). Peatland restoration pathways to mitigate greenhouse gas emissions and retain peat carbon. Biogeochemistry, 167(4), 523–543. Link to source: https://doi.org/10.1007/s10533-023-01103-1
Mander, Ü., Öpik, M., & Espenberg, M. (2025). Global peatland greenhouse gas dynamics: State of the art, processes, and perspectives. New Phytologist, 246(1), 94–102. Link to source: https://doi.org/10.1111/nph.20436
Marlier, M. E., Liu, T., Yu, K., Buonocore, J. J., Koplitz, S. N., DeFries, R. S., Mickley, L. J., Jacob, D. J., Schwartz, J., Wardhana, B. S., & Myers, S. S. (2019). Fires, smoke exposure, and public health: An integrative framework to maximize health benefits from peatland restoration. GeoHealth, 3(7), 178–189. Link to source: https://doi.org/10.1029/2019GH000191
Martens, H. R., Laage, K., Eickmanns, M., Drexler, A., Heinsohn, V., Wegner, N., Muster, C., Diekmann, M., Seeber, E., Kreyling, J., Michalik, P., & Tanneberger, F. (2023). Paludiculture can support biodiversity conservation in rewetted fen peatlands. Scientific Reports, 13(1), 18091. Link to source: https://doi.org/10.1038/s41598-023-44481-0
Melton, J. R., Chan, E., Millard, K., Fortier, M., Winton, R. S., Martín-López, J. M., Cadillo-Quiroz, H., Kidd, D., & Verchot, L. V. (2022). A map of global peatland extent created using machine learning (Peat-ML). Geoscientific Model Development, 15(12), 4709–4738. Link to source: https://doi.org/10.5194/gmd-15-4709-2022
Merten, J., Nielsen, J. Ø., Rosyani, & Faust, H. (2021). Climate change mitigation on tropical peatlands: A triple burden for smallholder farmers in Indonesia. Global Environmental Change, 71, Article 102388. Link to source: https://doi.org/10.1016/j.gloenvcha.2021.102388
Minasny, B., Adetsu, D. V., Aitkenhead, M., Artz, R. R. E., Baggaley, N., Barthelmes, A., Beucher, A., Caron, J., Conchedda, G., Connolly, J., Deragon, R., Evans, C., Fadnes, K., Fiantis, D., Gagkas, Z., Gilet, L., Gimona, A., Glatzel, S., Greve, M. H., … Zak, D. (2024). Mapping and monitoring peatland conditions from global to field scale. Biogeochemistry, 167(4), 383–425. Link to source: https://doi.org/10.1007/s10533-023-01084-1
Minayeva, T. Yu., Bragg, O. M., & Sirin, A. A. (2017). Towards ecosystem-based restoration of peatland biodiversity. Mires and Peat, 19, 1–36. Link to source: https://doi.org/10.19189/MaP.2013.OMB.150
Murdiyarso, D., Lilleskov, E., & Kolka, R. (2019). Tropical peatlands under siege: The need for evidence-based policies and strategies. Mitigation and Adaptation Strategies for Global Change, 24(4), 493–505. Link to source: https://doi.org/10.1007/s11027-019-9844-1
Oestmann, J., Dettmann, U., Düvel, D., & Tiemeyer, B. (2022). Experimental warming increased greenhouse gas emissions of a near-natural peatland and Sphagnum farming sites. Plant and Soil, 480(1), 85–104. Link to source: https://doi.org/10.1007/s11104-022-05561-8
Olson, D. M., Dinerstein, E., Wikramanayake, E. D., Burgess, N. D., Powell, G. V. N., Underwood, E. C., D’amico, J. A., Itoua, I., Strand, H. E., Morrison, J. C., Loucks, C. J., Allnutt, T. F., Ricketts, T. H., Kura, Y., Lamoreux, J. F., Wettengel, W. W., Hedao, P., & Kassem, K. R. (2001). Terrestrial ecoregions of the world: A new map of life on Earth: A new global map of terrestrial ecoregions provides an innovative tool for conserving biodiversity. BioScience, 51(11), 933–938. Link to source: https://doi.org/10.1641/0006-3568(2001)051[0933:TEOTWA]2.0.CO;2
Pan, Y., Birdsey, R. A., Phillips, O. L., Houghton, R. A., Fang, J., Kauppi, P. E., Keith, H., Kurz, W. A., Ito, A., Lewis, S. L., Nabuurs, G.-J., Shvidenko, A., Hashimoto, S., Lerink, B., Schepaschenko, D., Castanho, A., & Murdiyarso, D. (2024). The enduring world forest carbon sink. Nature, 631(8021), 563–569. Link to source: https://doi.org/10.1038/s41586-024-07602-x
Peacock, M., Audet, J., Bastviken, D., Futter, M. N., Gauci, V., Grinham, A., Harrison, J. A., Kent, M. S., Kosten, S., Lovelock, C. E., Veraart, A. J., & Evans, C. D. (2021). Global importance of methane emissions from drainage ditches and canals. Environmental Research Letters, 16(4), Article 044010. Link to source: https://doi.org/10.1088/1748-9326/abeb36
Posa, M. R. C., Wijedasa, L. S., & Corlett, R. T. (2011). Biodiversity and conservation of tropical peat swamp forests. BioScience, 61(1), 49–57. Link to source: https://doi.org/10.1525/bio.2011.61.1.10
Renou-Wilson, F., Moser, G., Fallon, D., Farrell, C. A., Müller, C., & Wilson, D. (2019). Rewetting degraded peatlands for climate and biodiversity benefits: Results from two raised bogs. Ecological Engineering, 127, 547–560. Link to source: https://doi.org/10.1016/j.ecoleng.2018.02.014
Ritson, J. P., Bell, M., Brazier, R. E., Grand-Clement, E., Graham, N. J. D., Freeman, C., Smith, D., Templeton, M. R., & Clark, J. M. (2016). Managing peatland vegetation for drinking water treatment. Scientific Reports, 6(1), Article 36751. Link to source: https://doi.org/10.1038/srep36751
Robinson, N., Drever, C. R., Gibbs, D. A., Lister, K., Esquivel-Muelbert, A., Heinrich, V., Ciais, P., Silva-Junior, C. H. L., Liu, Z., Pugh, T. A. M., Saatchi, S., Xu, Y., & Cook-Patton, S. C. (2025). Protect young secondary forests for optimum carbon removal. Nature Climate Change, 15, 793–800. Link to source: https://doi.org/10.1038/s41558-025-02355-5
Salmayenti, R., Holden, J., Baird, A. J., Rigby, R., & Spracklen, D. V. (2026). Tropical peatland restoration reduces fire occurrence. Geophysical Research Letters, 53(11), Article e2025GL121564. Link to source: https://doi.org/10.1029/2025GL121564
Schulz, C., Martín Brañas, M., Núñez Pérez, C., Del Aguila Villacorta, M., Laurie, N., Lawson, I. T., & Roucoux, K. H. (2019). Uses, cultural significance, and management of peatlands in the Peruvian Amazon: Implications for conservation. Biological Conservation, 235, 189–198. Link to source: https://doi.org/10.1016/j.biocon.2019.04.005
Silverthorn, T., Aberg, D., Baker, F., Bell, C., Evans, C. D., Gurd, C., Holman, I., Howell, G., McEwen, A., Morrison, R., Oakley, J., Southon, F., & Peacock, M. (2026). Land‐use effects on surface water quality in temperate lowland peatlands. Global Change Biology, 32(7), Article e70988. Link to source: https://doi.org/10.1111/gcb.70988
Spitzer, K., & Danks, H. V. (2006). Insect biodiversity of boreal peat bogs. Annual Review of Entomology, 51, 137–161. Link to source: https://doi.org/10.1146/annurev.ento.51.110104.151036
Strack, M., Davidson, S. J., Hirano, T., & Dunn, C. (2022). The Potential of Peatlands as Nature-Based Climate Solutions. Current Climate Change Reports, 8(3), 71–82. Link to source: https://doi.org/10.1007/s40641-022-00183-9
Strack, M., Keith, A. M., & Xu, B. (2014). Growing season carbon dioxide and methane exchange at a restored peatland on the Western Boreal Plain. Ecological Engineering, 64, 231–239. Link to source: https://doi.org/10.1016/j.ecoleng.2013.12.013
Strack, M., & Zuback, Y. C. A. (2013). Annual carbon balance of a peatland 10 yr following restoration. Biogeosciences, 10(5), 2885–2896. Link to source: https://doi.org/10.5194/bg-10-2885-2013
Tan, Z. D., Lupascu, M., & Wijedasa, L. S. (2021). Paludiculture as a sustainable land use alternative for tropical peatlands: A review. Science of The Total Environment, 753, Article 142111. Link to source: https://doi.org/10.1016/j.scitotenv.2020.142111
Temmink, R. J. M., Lång, K., Vroom, R. J. E., Leifeld, J., Fritz, C., Zeug, W., Thrän, D., Kleinspehn, C., Gaudig, G., Neubert, J., Kreyling, J., Rhymes, J. M., Evans, C. D., Kotowski, W., Nordt, A., & Tanneberger, F. (2026). Agriculture on wet peatlands: The sustainability potential of paludiculture. Agricultural Systems, 231, Article 104561. Link to source: https://doi.org/10.1016/j.agsy.2025.104561
Terzano, D., Attorre, F., Parish, F., Moss, P., Bresciani, F., Cooke, R., & Dargusch, P. (2022). Community-led peatland restoration in Southeast Asia: 5Rs approach. Restoration Ecology, 30(8), Article e13642. Link to source: https://doi.org/10.1111/rec.13642
The Gecko Project. (2023). Success of Indonesian peatland restoration in doubt as fire season sets in. Link to source: https://thegeckoproject.org/articles/success-of-indonesian-peatland-restoration-in-doubt-as-fire-season-sets-in/
Thornton, S. A., Setiana, E., Yoyo, K., Dudin, Yulintine, Harrison, M. E., Page, S. E., & Upton, C. (2020). Towards biocultural approaches to peatland conservation: The case for fish and livelihoods in Indonesia. Environmental Science & Policy, 114, 341–351. Link to source: https://doi.org/10.1016/j.envsci.2020.08.018
Toumbourou, T. D., Lestari, S., Yuwati, T. W., Treby, S., Winarno, B., Rachmanadi, D., Idrus, N. I., Sakuntaladewi, N., Budiningsih, K., Grover, S. P. P., & Rawluk, A. (2024). Principles for equitable and resilient tropical peatland restoration in Central Kalimantan, Indonesia. Restoration Ecology, 32(7), Article e14221. Link to source: https://doi.org/10.1111/rec.14221
Tubiello, F. N., Conchedda, G., Casse, L., Hao, P., De Santis, G., & Chen, Z. (2023). A new cropland area database by country circa 2020. Earth System Science Data, 15(11), 4997–5015. Link to source: https://doi.org/10.5194/essd-15-4997-2023
Uda, S. K., Hein, L., & Atmoko, D. (2019). Assessing the health impacts of peatland fires: A case study for Central Kalimantan, Indonesia. Environmental Science and Pollution Research, 26(30), 31315–31327. Link to source: https://doi.org/10.1007/s11356-019-06264-x
Uda, S. K., Hein, L., & Adventa, A. (2020). Towards better use of Indonesian peatlands with paludiculture and low-drainage food crops. Wetlands Ecology and Management, 28(3), 509–526. Link to source: https://doi.org/10.1007/s11273-020-09728-x
UNEP. (2022). Global peatlands assessment: The state of the world’s peatlands: Evidence for action toward the conservation, restoration, and sustainable management of peatlands. Link to source: https://www.unep.org/resources/global-peatlands-assessment-2022
van der Laan, A., van Dijk, J., Rebel, K. T., & Wassen, M. J. (2024). Rewet without regret? Nutrient dynamics in fen peat exposed to different rewetting degrees. Biogeochemistry, 167(5), 705–721. Link to source: https://doi.org/10.1007/s10533-024-01139-x
Vanselow-Algan, M., Schmidt, S. R., Greven, M., Fiencke, C., Kutzbach, L., & Pfeiffer, E.-M. (2015). High methane emissions dominated annual greenhouse gas balances 30 years after bog rewetting. Biogeosciences, 12(14), 4361–4371. Link to source: https://doi.org/10.5194/bg-12-4361-2015
Ward, C., Stringer, L. C., Warren-Thomas, E., Agus, F., Crowson, M., Hamer, K., Hariyadi, B., Kartika, W. D., Lucey, J., McClean, C., Nurida, N. L., Petorelli, N., Pratiwi, E., Saad, A., Andriyani, R., Ariani, T., Sriwahyuni, H., & Hill, J. K. (2020). Smallholder perceptions of land restoration activities: Rewetting tropical peatland oil palm areas in Sumatra, Indonesia. Regional Environmental Change, 21(1), Article 1. Link to source: https://doi.org/10.1007/s10113-020-01737-z
Wetlands International. (2025). The Peatland Breakthrough: Science-Based Framework for Global Peatland Targets and Guiding Principles. Link to source: https://www.wetlands.org/the-peatland-breakthrough/
Wilson, D., Farrell, C. A., Fallon, D., Moser, G., Müller, C., & Renou-Wilson, F. (2016). Multiyear greenhouse gas balances at a rewetted temperate peatland. Global Change Biology, 22(12), 4080–4095. Link to source: https://doi.org/10.1111/gcb.13325
Worrall, F., Howden, N. J. K., Burt, T. P., Rico-Ramirez, M. A., & Kohler, T. (2022). Local climate impacts from ongoing restoration of a peatland. Hydrological Processes, 36(3), Article e14496. Link to source: https://doi.org/10.1002/hyp.14496
Xu, J., Morris, P. J., Liu, J., & Holden, J. (2018a). Hotspots of peatland-derived potable water use identified by global analysis. Nature Sustainability, 1(5), 246–253. Link to source: https://doi.org/10.1038/s41893-018-0064-6
Xu, J., Morris, P. J., Liu, J., & Holden, J. (2018b). PEATMAP: Refining estimates of global peatland distribution based on a meta-analysis. CATENA, 160, 134–140. Link to source: https://doi.org/10.1016/j.catena.2017.09.010
Yu, Z., Loisel, J., Brosseau, D. P., Beilman, D. W., & Hunt, S. J. (2010). Global peatland dynamics since the Last Glacial Maximum. Geophysical Research Letters, 37(13). Link to source: https://doi.org/10.1029/2010GL043584
Yuwati, T. W., & Pratiwi, D. (2022). Paludiculture: Peatland utilization for food security. IOP Conference Series: Earth and Environmental Science, 1107(1), Article 012075. Link to source: https://doi.org/10.1088/1755-1315/1107/1/012075
Zeng, Y., Sarira, T. V., Carrasco, L. R., Chong, K. Y., Friess, D. A., Lee, J. S. H., Taillardat, P., Worthington, T. A., Zhang, Y., & Koh, L. P. (2020). Economic and social constraints on reforestation for climate mitigation in Southeast Asia. Nature Climate Change, 10(9), 842–844. Link to source: https://doi.org/10.1038/s41558-020-0856-3
Ziegler, R., Wichtmann, W., Abel, S., Kemp, R., Simard, M., & Joosten, H. (2021). Wet peatland utilisation for climate protection – An international survey of paludiculture innovation. Cleaner Engineering and Technology, 5, Article 100305. Link to source: https://doi.org/10.1016/j.clet.2021.100305
Avery W. Driscoll, Ph.D.
Ruthie Burrows, Ph.D.
James Gerber, Ph.D.
Daniel Jasper
Alex Sweeney
Christina Richardson, Ph.D.
Christina Swanson, Ph.D.
Paul C. West, Ph.D.
Using the IPCC’s Tier 1 methodology (IPCC Task Force on National Greenhouse Gas Inventories, 2014), we estimated the net effectiveness of peatland restoration at 19.0–56.0 t CO₂‑eq /ha/yr (100-yr basis), depending on the climate zone (Table 1a–d).
The primary pathway by which peatland restoration reduces emissions is by slowing or stopping CO₂ emissions associated with loss of peat through oxidation. Because peat oxidation occurs more rapidly in warmer climates, the effectiveness of restoration is higher in tropical and subtropical regions than in temperate and boreal regions (IPCC Task Force on National Greenhouse Gas Inventories, 2014). We estimated that drained peatlands emit 15.1–49.7 t CO₂‑eq /ha/yr, while restored peatlands can range from carbon sinks (–1.6 t CO₂‑eq /ha/yr on average in the boreal) to small sources (0.5 t CO₂‑eq /ha/yr on average in temperate areas). However, site-level variability is high, with a recent review reporting values ranging from a sink of –42 t CO₂‑eq /ha/yr (in a restored Canadian sphagnum bog) to a source of 8 t CO₂‑eq /ha/yr (at a restored German sphagnum paludiculture site) (Mander et al., 2024; Oestmann et al., 2022; Strack et al., 2014).
Peatland restoration influences GHG emissions from several other pathways as well:
Table 1. Effectiveness at reducing emissions and, in some cases, removing carbon.
Unit: t CO₂‑eq /ha/yr, 100-yr basis
| Estimate | 19.0 |
Unit: t CO₂‑eq /ha/yr, 100-yr basis
| Estimate | 29.7 |
Unit: t CO₂‑eq /ha/yr, 100-yr basis
| Estimate | 47.8 |
Unit: t CO₂‑eq /ha/yr, 100-yr basis
| Estimate | 56.0 |
The costs of peatland restoration are highly uncertain and variable. We estimated that the median cost of peatland restoration is approximately US$55/t CO₂‑eq (2023 US$), or US$1,884/ha restored (Table 2). These estimates are based on data from only three countries. In Scotland, peatland restoration costs are submitted as part of the Peatland Action Programme, which provides grants for peat restoration (Glenk et al., 2022). Based on data from 300 sites, they report a median cost of US$1,884/ha (equal to US$55/t CO₂‑eq , 100-year basis). In the United States, the U.S. Department of Agriculture developed models of restoration costs based on contracts with agricultural landowners for wetland restoration through the Wetlands Reserve Program, with costs ranging from US$170/ha to US$6,100/ha depending on the region. All other cost data used in our calculations are from Indonesia, with estimates ranging from US$237/ha to more than US$25,000/ha (Hansson & Dargusch 2018; Tan et al., 2022). Peatland restoration costs include the costs of restoring the natural hydrologic condition and of revegetation, and depend largely on the size of canals that must be dammed or filled (Hansson & Dargusch 2018). These estimates do not include forgone revenues from economic uses of drained peatlands, nor do they include potential new revenues from paludiculture, tourism, or ecosystem service payments.
Table 2. Cost per unit of climate impact.
Unit: 2023 US$/t CO₂‑eq , 100-yr basis
| Median | 55 |
We define a learning curve as falling costs with increased adoption. There is no evidence that the costs of peatland rewetting and revegetation will fall with increasing adoption. Therefore, there is no learning curve for this solution.
Speed of action refers to how quickly a climate solution physically affects the atmosphere after it is deployed. This is different from speed of deployment, which is the pace at which solutions are adopted.
At Project Drawdown, we define the speed of action for each climate solution as emergency brake, gradual, or delayed. Restore Peatlands is a GRADUAL climate solution, because the primary mode of action is a reduction in emissions that would otherwise occur on an annual basis over a long time. It has a steady, linear impact on the atmosphere.
GHG emissions dynamics of rewetted peatlands remain uncertain because they can be highly variable depending on site level factors, such as land use history, type of restoration interventions, and environmental conditions (Mander et al., 2024).
Effective peatland restoration faces many barriers, including cost of restoration, disruption to economic use, willingness of local community and landowners to support restoration, and technical challenges, particularly on severely subsided sites (Andersen et al., 2017; Harrison et al., 2020; Lestari et al., 2024; Zeng et al., 2020).
Peatland restoration, like other nature-based climate solutions, faces concerns related to the permanence of emissions reductions and carbon sequestration (Loisel & Gallego-Sala, 2022; Turetsky et al., 2015). Restored peatlands are vulnerable to drainage or clearing, which can be mitigated by coupling restoration with long-term protections. Restored peatlands are also vulnerable to natural disturbances, such as fires and droughts, though they tend to be more resilient than drained peatlands (Granath et al., 2016).
Data on the current adoption of peatland restoration are very sparse, noncentralized, and largely not validated. Therefore, we do not provide a global estimate of the currently restored area. We found two countries for which there are national-scale estimates of restored area. The United Kingdom’s Peatland Programme reports that restoration activities were undertaken on 254,254 ha of peatland between the 1990s and 2024 (IUCN UK Peatland Programme, 2024). In Indonesia, the Ministry of Environment and Forestry reported 3.7 Mha of peatland restoration in a 2022 report based on estimates of the area with a water table depth of 40 cm or less as of December, 2021. However, only 6,000 of these ha were reported to have undergone some form of vegetation restoration, and water table depth is a limited proxy for peatland restoration status because it fluctuates substantially over time due to natural variability in water supplies. For instance, the area meeting the water table threshold had fallen to just 0.5 Mha by June 2022 (Jong, 2023; The Gecko Project, 2022). Based on published estimates, it seems likely that the peatland area that has been successfully restored is in the range of 0.5–5Mha.
Peatland restoration data are very limited and rarely temporally resolved. The available data are insufficient to calculate an adoption trend for this solution.
We estimated the adoption ceiling for peatland restoration to be 56.9 Mha (Table 3). This is the area of peatland that has been drained or converted to other land uses (UNEP, 2022) and represents ~12% of the total global peatland area. Approximately 37.8 Mha of drained peatlands are currently used for agriculture (croplands or grazing), 18.1 Mha are used for forestry, and 1.0 Mha are used for peat extraction (UNEP, 2022). Although restoring peatlands will involve displacing some of these activities to other lands. Though drained peatlands can be particularly high-yielding (Lloyd et al., 2023), these are relatively small areas in the context of global land use. For example, agricultural peatlands (including both croplands and grazing lands) represent less than 3% of total global croplands, which were estimated to cover ~1,500 Mha circa 2020 (Tubiello et al., 2023). Indirect land use change emissions arising from relocation of croplands are accounted for in our effectiveness estimates.
Drained peatlands are concentrated in boreal (26.3 Mha) and tropical regions (18.5 Mha), with smaller areas in temperate (9.9 Mha) and subtropical (2.1 Mha) regions (UNEP, 2022). The proportion of peatlands that have been drained is largest in temperate regions (28%), followed by the tropics (15%), subtropics (14%), and then the boreal (8%).
These values reflect best estimates of drained peatland area from the Global Peatland Assessment (UNEP, 2022). However, there is substantial uncertainty in the distribution of intact and drained peatlands, and efforts to map global peatlands are ongoing (Melton et al., 2022; Minasny et al., 2024).
Table 3. Adoption ceiling.
Unit: ha available for restoration
| Estimate | 26,300,000 |
Unit: ha available for restoration
| Estimate | 9,930,000 |
Unit: ha available for restoration
| Estimate | 2,140,000 |
Unit: ha available for restoration
| Estimate | 18,500,000 |
Unit: ha available for restoration
| Estimate | 56,900,000 |
We assumed an achievable range of 25–50% of the adoption ceiling, equal to 14.2 – 28.4 Mha of peatland restoration (Table 4a–e).
The momentum behind peatland restoration has been growing rapidly over the past several decades, including major international efforts such as the U.N. Decade on Ecosystem Restoration (2021–2030), the Global Peatlands Initiative, and the Peatland Breakthrough (Global Peatlands Initiative, 2025; UNEP, 2022; Wetlands International, 2025). Our estimate of the high achievable level of peatland restoration roughly aligns with the 30 Mha target of the Peatland Breakthrough, a global initiative to guide peatland restoration through unified targets and restoration principles that was launched at COP30 (Global Peatlands Initiative, 2025). Given that the current area successfully restored is likely in the range of 0.5–5 Mha, both the low and high achievable estimates will require an ambitious expansion of peatland restoration.
Peatland restoration will very likely need to be directly incentivized through payments for ecosystem services such as emissions reductions, water quality, flood management, and/or biodiversity to achieve this level of adoption (Bonn et al., 2014). Paludiculture, or using restored peatlands for production of commodities that tolerate saturated conditions, may also boost restoration (Mander et al., 2024; Tan et al., 2021; Ziegler et al., 2021). In temperate and lower boreal regions, paludiculture is typically characterized by herbaceous vegetation, including cattails, reeds, canary grass, sedges, and alder, used for grazing or biomass and by cultivation of sphagnum for horticultural use (Temmink et al., 2026). Additional research is needed to develop suitable paludiculture systems for restored tropical peatlands (Uda et al., 2020), which are typically characterized by tree cover rather than herbaceous plants. Sustainable gathering and hunting of native peatland species is a form of paludiculture that has been practiced by indigenous communities for centuries, and tree crops such as sago and jelutong show some potential for economically viable paludiculture (Tan et al., 2021; Ziegler et al, 2021).
Table 4. Range of achievable adoption levels.
Unit: ha
| Current adoption | NA |
| Achievable – low | 6,560,000 |
| Achievable – high | 13,100,000 |
| Adoption ceiling | 26,300,000 |
Unit: ha
| Current adoption | NA |
| Achievable – low | 2,480,000 |
| Achievable – high | 4,960,000 |
| Adoption ceiling | 9,930,000 |
Unit: ha
| Current adoption | NA |
| Achievable – low | 535,000 |
| Achievable – high | 1,070,000 |
| Adoption ceiling | 2,140,000 |
Unit: ha
| Current adoption | NA |
| Achievable – low | 4,630,000 |
| Achievable – high | 9,260,000 |
| Adoption ceiling | 18,500,000 |
Unit: ha
| Current adoption | NA |
| Achievable – low | 14,200,000 |
| Achievable – high | 28,400,000 |
| Adoption ceiling | 56,900,000 |
We estimated that peatland restoration could reduce emissions by 0.48 Gt CO₂‑eq/yr at the low-achievable adoption scenario, 0.97 Gt CO₂‑eq/yr at the high-achievable adoption scenario, and 1.93 Gt CO₂‑eq/yr at the adoption ceiling (Table 5a–e). Our estimate of the climate impact at the adoption ceiling is approximately in line with literature estimates of annual emissions from drained peatlands (e.g., 1.98 Gt CO₂‑eq/yr from Leifield and Menichetti [2018]).
Although the total area available for restoration is largest in the boreal biome, more than half of the adoption ceiling climate impact occurs in the tropics due to larger per-hectare emissions reductions from peatland restoration in warmer climates. Restoration in boreal regions is estimated to provide 26% of the climate impact at the adoption ceiling, with an additional 15% and 5% of the climate impact available through temperate and subtropical peatland restoration, respectively.
Table 5. Climate impact at different levels of adoption.
Unit: Gt CO₂‑eq/yr, 100-year basis
| Current adoption | NA |
| Achievable – low | 0.125 |
| Achievable – high | 0.249 |
| Adoption ceiling | 0.498 |
Unit: Gt CO₂‑eq/yr, 100-year basis
| Current adoption | NA |
| Achievable – low | 0.074 |
| Achievable – high | 0.148 |
| Adoption ceiling | 0.295 |
Unit: Gt CO₂‑eq/yr, 100-year basis
| Current adoption | NA |
| Achievable – low | 0.026 |
| Achievable – high | 0.051 |
| Adoption ceiling | 0.102 |
Unit: Gt CO₂‑eq/yr, 100-year basis
| Current adoption | NA |
| Achievable – low | 0.259 |
| Achievable – high | 0.518 |
| Adoption ceiling | 1.037 |
Unit: Gt CO₂‑eq/yr, 100-year basis
| Current adoption | NA |
| Achievable – low | 0.483 |
| Achievable – high | 0.966 |
| Adoption ceiling | 1.932 |
Peatland restoration can help communities adapt to extreme weather, though restored peatlands may be more sensitive than undisturbed peatlands to extreme weather (Loisel & Gallego-Sala, 2022). Because peatlands regulate water flows, they can reduce the risk of droughts and floods (IUCN, 2021; Ritson et al., 2016). Evidence suggests that peatlands can cool the immediate environment, lowering daytime temperatures and reducing daily temperature extremes (Dietrich & Behrendt, 2022; Helbig et al., 2020; Worrall et al., 2022).
Peatlands support the livelihoods of nearby communities, especially those in low- and middle-income countries. In the peatlands of the Amazon and Congo basins, fishing livelihoods depend on aquatic wildlife (Thornton et al., 2020). Peatlands in the Peruvian Amazon provide important goods for trade, such as palm fruit and timber and are used for hunting (Schulz et al., 2019). Peatlands can also support the livelihoods of women, allowing them to access more economic opportunities and contributing to gender equality. For example, in Indonesia women use raw materials from peatlands to create mats sold for use in significant events such as births, weddings, and burials (Goib et al., 2018). Selling commodities made from peatland materials increases household income and can increase the position of women to make decisions in a household (Goib et al., 2018).
Restoring peatlands may come with some trade-offs to livelihoods if it prevents other productive uses. Paludiculture may help local communities to maintain their livelihoods on restored peatlands (Temmink et al., 2026; Yuwati & Pratiwi, 2022). Case studies in Southeast Asia have found that community-led restoration projects that consider local context and livelihoods are more successful than conventional restoration projects (Terzano et al., 2022).
Peatlands mitigate exposure to air pollution and can save money from reduced health-care expenditures (Kiely et al., 2021).
When peatlands are drained, they are susceptible to fire, and rewetting reduces fire risk (Salmayenti et al., 2026). Peatland fires can contribute to air pollution because of the way these fires smolder (Uda et al., 2019). Smoke and pollutants, particularly PM2.5, from peatland fires can harm respiratory health and lead to premature mortality (Marlier et al., 2019). A study in Indonesia estimated peatland fires contribute to the premature mortality of about 33,100 adults and about 2,900 infants annually (Hein et al., 2022). Researchers have linked exposure to PM2.5 from peatland fires to increased hospitalizations, asthma, and lost workdays (Hein et al., 2022).
Peatlands support diverse flora and wildlife (UNEP, 2022; Minayeva et al., 2017; Posa et al., 2011), including many rare and threatened species (Posa et al., 2011). A study of Indonesian peat swamps found that the IUCN Red List classified approximately 45% of mammals and 33% of birds living in these ecosystems as threatened, vulnerable, or endangered (Posa et al., 2011). Peatlands also support a variety of insect species (Spitzer & Danks, 2006). Because of their sensitivity to environmental changes, some peatland insects can act as indicators of peatland health and play a role in conservation (Spitzer & Danks, 2006).
While restored peatlands are more biodiverse than degraded ecosystems, restored peatlands are often less biodiverse than intact peatlands (Renou-Wilson et al., 2019). For example, a large study of peatlands in Europe found that rewetting increased biodiversity of peatlands, but plant composition was different from natural, intact peatlands (Kreyling et al., 2021). Biodiversity benefits have also been identified in European paludiculture systems, though more research on the biodiversity impacts of paludiculture is needed (Martens et al., 2023).
Peatlands can filter water pollutants and improve water quality and are important sources of potable water (Minayeva et al., 2017). Xu et al. (2018a) estimated that peatlands store about 10% of freshwater globally, not including glacial water. Peatlands provide potable water for about 71.4 million people, primarily in the United Kingdom and Ireland (Xu et al., 2018a).
For a description of water quality benefits, please refer to the “water resources” subsection above.
Where peatland restoration involves removing agricultural or forestry land from economic use, it risks compromising the livelihoods of local communities (Lestari et al., 2024; Merten et al., 2021; Ward et al., 2020). Close collaboration with local communities can facilitate procedural equity and enable positive, durable outcomes for both the ecosystem and the community (e.g., Toumbourou et al., 2024). Additionally, restoring peatlands can trigger displacement of production activities onto other lands, resulting in land clearing elsewhere (i.e., leakage). We account for the emissions impacts of leakage in our effectiveness calculations. While restoration can reduce emissions associated with agriculture locally, these emissions can occur elsewhere if the production area shifts. Paludiculture can reduce the risk of negative economic impacts and leakage (Mander et al., 2024; Tan et al., 2021; Ziegler et al., 2021).
Though healthy peatlands provide water quality benefits, peatland rewetting can mobilize nutrients and other contaminants into surface waters (Breznikar et al., 2024; Heuts et al., 2026; Silverthorn et al., 2026; van der Laan). Studies have found increases in phosphorus, nitrogen, and heavy metal loads from peatland rewetting, though effects are mixed. Water quality monitoring and appropriate mitigation near restoration sites can help manage water quality risks (e.g. Hoffman et al., 2026).
Peatland restoration can improve the health and function of adjacent ecosystems that are being protected or restored.
These solutions are all suitable to implement on degraded peatland, and thus are in competition with restoration.
ha under restoration
CO₂ , N₂O
ha under restoration
CO₂ , N₂O
ha under restoration
CO₂ , N₂O
ha under restoration
CO₂ , N₂O
Peatland restoration can divert resources from other climate solutions, including protecting intact peatlands. Preventing peatland drainage is typically more cost effective and offers larger emissions benefits than does peatland restoration (see Protect Peatlands). Restoration should therefore complement, rather than compete with, efforts to reduce further disturbance of intact peatlands.
Degraded peatlands currently emit an estimated 1.3–1.9 Gt CO₂ ‑eq/yr (excluding fires), equal to ~2–4% of total global GHG emissions (Leifield and Menichetti, 2018; UNEP, 2022). Leifeld et al. (2019) projected that without protection or restoration, drained peatlands could produce enough emissions to consume 12–41% of the remaining emissions budget for keeping warming below 1.5–2.0 °C. Peatland drainage had produced a cumulative 80 Gt CO₂ ‑eq by 2015, equal to nearly 16 months’ worth of total global emissions (60.3 Gt CO₂‑eq/yr as of 2024).
Peatland rewetting and restoration has been widely shown to stop ongoing CO₂ emissions from oxidation of drained peat and reduce lateral transport of carbon through waterways (Günther et al., 2020; Mander et al., 2024; Wilson et al., 2016). The net emissions reduction in rewetted relative to drained conditions is clear and consistent, despite large variability in the magnitude of carbon uptake and methane and nitrous oxide emissions following rewetting. While many restored peatlands regain their capacity to act as net carbon sinks, some sites remain emissions sources for decades following rewetting (Strack & Zuback 2013; Vanselow-Algan et al., 2015; Wilson et al., 2016).
The results presented in this document summarize findings from 19 reviews, meta-analyses, and global databases reflecting current evidence for peatland restoration. The majority of the non-global literature is focused on Indonesia, the United Kingdom, and Northern Europe. We recognize this limited geographic scope creates bias, and hope this work inspires research and data sharing on this topic in underrepresented regions.
Join the 80,000+ subscribers discovering how to drive meaningful climate action around the world! Every other week, you'll get expert insights, cutting-edge research, and inspiring stories.
