Current status and emerging perspectives of coastal blue carbon ecosystems

Productivity and carbon balance in blue carbon ecosystems

Ecosystem respiration (R_E = sum of all respiration within an ecosystem) is greatest in mangroves, followed by seagrasses and tidal marshes [Table 1]. Global R_E (= total R_E × global area of each ecosystem) is greatest for seagrasses [Table 1] whereas mangroves account for most gross primary production (GPP or P_GPP), but seagrass meadows contribute the most to global GPP. All three ecosystems are net autotrophic (ratio of P_GPP/R_E > 1), producing more plant C_ORG than is lost via respiration (R_E). In contrast, the coastal ocean is in overall carbon balance with a mean P_GPP/R_E ratio of 0.98 [Table 1].

Net ecosystem production (NEP = ecosystem GPP - R_E) differs among the three ecosystems, with salt marshes having the lowest NEP and mangrove and seagrasses the highest NEP. However, all three ecosystems contribute disproportionately to oceanic C_ORG storage.

Blue carbon stocks, sequestration rates and CO₂ emissions

Unsurprisingly, such highly autotrophic ecosystems are carbon-rich habitats. Estimates of area-specific and global C_ORG stocks, C_ORG burial rates and CO₂ emissions due to habitat conversion [Table 2] show large differences between habitats, with seagrasses encompassing the most global area, C_ORG sequestration, the highest loss rates and subsequent CO₂ emissions. Mangroves have the greatest mean C_ORG stocks and lowest mean C_ORG sequestration rates, whereas salt marshes occupy the least global area, and have the lowest global mean C_ORG stocks and lowest global sequestration rates. Mangrove habitat losses are lowest among the coastal habitats, but all three have similar rates of CO₂ emissions [Table 2]. Most of these habitat differences reflect their global areas, but some differences do reflect inherent differences in rates of various biogeochemical processes^[1,4,5].

There is considerable uncertainty in these mean and global carbon values, given the inherent differences in various methods used and the masking of large and certain spatial and temporal trends in the data. Therefore, caution must be applied in assessing comparisons within and between ecosystems. For instance, wet tropical mangroves on average have larger C_ORG stocks and greater rates of greenhouse gas emissions than those measured in arid-zone mangroves^[1].

Trajectories for blue carbon

Nations can directly offset their carbon emissions, restore and preserve sequestered C_ORG and facilitate other ecosystem services (e.g., coastal protection, wood, fisheries) by protecting, planting, and rehabilitating blue carbon ecosystems. Having information on sediment sequestration rates and increases in plant C_ORG biomass with the growth and maturity of restored ecosystems is important to more accurately measure C_ORG increases via restoration and rehabilitation. These management tools assist in climate change mitigation via removal and burial of C_ORG in sediments.

Re-establishment of mangroves, seagrasses, and salt marshes results in a progressive increase in biomass C_ORG from early pioneering phases to maturity. A detailed example of such increases in a restored mangrove stand was found in Avicennia marina forests in southern Australia^[10]. Using dating techniques and gas measurements to estimate carbon increases in restored 13-35 year old forests had a positive impact on C_ORG additionality and C_ORG sequestration rate. The older forests (17- and 35-year old) had twice the total C_ORG standing amounts and sediment sequestration rates than the youngest stand [Figure 2]. Although sediment C_ORG stocks increased with forest age, aboveground stocks were greatest in the 17-year old forest. Older mangroves released 25% of CH₄ emissions and twice the CO₂ flux compared to young stands, reflecting the fact that forest age had a significant impact on sediment C_ORG dynamics. This study and others^[11-13] show that restored mangroves store increased amounts of C_ORG as they age, but this sink stabilizes once forests reach maturity. Blue carbon dynamics appears to be more complex in naturally regenerated (NR) mangroves^[14].

Figure 2. Increases in C_ORG stocks since establishment of mangrove forests aged 13-, 17- and 35-year old. Sediment C_ORG stocks (sediment C_ORG + belowground C_ORG stock) are brown bars. Values indicate mean ± 1 SE with different letters indicating significant differences in aboveground C_ORG stock (lowercase), sediment C_ORG stock (uppercase), and C_ORG accretion rate (italics). Reproduced with permission from Carnell et al.^[10].

In naturally expanding mangroves in Malaysia, standing amounts of ecosystem C_ORG in intact and different aged, naturally regenerated forests were similar despite a wide age difference (3-, 6-, 12-, 18-, 25- year). Sediment C_ORG stocks were unchanged, driving the constancy of total ecosystem C_ORG stocks over many years. Younger stands had greater C_ORG sequestration rates than the more mature stands. As the mangroves expand seaward, high suspended sediment loads upstream support higher bulk sediment density and C_ORG accumulation rates, suggesting that naturally regenerated mangroves composed mainly of young forests are significant C_ORG sequestrators, suggesting that nature-based solutions can be sought through avoidance of replanting^[14].

Like mangroves, standing stocks of blue carbon and burial rates in restored salt marshes often show positive trajectories of establishment and growth, but no clear linear increase in sediment C_ORG storage and accumulation^[15-18]. In restored salt marshes in England^[15], C_ORG accumulation rates over a 150-year restoration period were initially rapid during the first 20-year, then slowly declined at a steady rate thereafter. In the century following restoration, a total accumulation of 74 t C_ORG ha^-1 was estimated, with comparisons to natural marsh C_ORG content, inferring that about 100 year is needed for restored marsh C_ORG to approximate that in natural marshes. A comparison of C_ORG accumulation rates in restored versus old salt marshes in the United States^[16] found that restored salt marshes can rapidly accumulate sediment C_ORG, although both old and restored habitats can have nearly identical C_ORG accumulation rates. In a naturally restored chronosequence of Chinese marshes^[17], sediment C_ORG storage increased rapidly for the first 8-year after which it slowed down to the point that storage at the oldest restoration marsh (27-year) was 45% lower than in an adjacent natural marsh. Blue carbon may accumulate early in the chronosequence, but the patterns are complex and restoration of sediment C_ORG to the level of natural marshes may take decades. Across a 32-year chronosequence in Louisiana marshes^[18], there was no significant difference in C_ORG accumulation rates among marshes with increasing age; all marshes were influenced by site-specific environmental conditions such as plant stem density and sediment type, with sites with high stem densities facilitated high long-term C_ORG accumulation^[18].

Plant invasion

Warming temperatures and subsequent habitat contraction and expansion are impacting C_ORG sequestration and standing stocks in mangroves and marshes. A pattern has emerged of increased concentrations of mangrove sediment and tree carbon as mangrove trees invade salt marsh habitats in warm temperate regions. In created Spartina marshes in Florida, most marshes have transitioned to mangrove habitat^[19]. Mature habitats had greater C_ORG stocks than middle-aged transitional sites or young salt marshes; belowground C_ORG stocks were similar along the ecotone, indicating that aboveground C_ORG stocks (mangrove and marsh) were responsible for the difference in total ecosystem stocks. Over the salt marsh-mangrove ecotone, C_ORG storage correlated with ecosystem age, indicating that these transitional habitats can potentially become significant sites of C_ORG storage.

Restoration is an important climate mitigation strategy for seagrass meadows^[20-24]. Across the Gulf of Mexico^[20], seagrass restoration sites increased in sediment C_ORG stocks with an increase in restoration age from 13 (young meadow) to 22 (15-16-year-old meadows) Mg C_ORG ha^-1 with a significant increase to 64 Mg C_ORG ha^-1 (30-35-year-old meadows). C_ORG content between natural and restored sediments was similar, suggesting the potential for seagrass restoration to offset anthropogenic losses. In Virginia, USA^[21], restored Zostera marina meadows of different ages (0-, 4-, and 10-year) showed higher sediment C_ORG content and accumulation rates in the 10-year meadow relative to the younger beds with restored seagrasses accumulating C_ORG at a similar rate to those in natural seagrass beds. Revegetated Posidonia australis meadows in Western Australia^[22] restored seagrass C_ORG standing stocks [Figure 3A] and sequestration capacity [Figure 3B] 18-year after planting, like that in natural seagrass meadows (26 g C_ORG m^-2 year^-1). Although seagrass restoration facilitated autochthonous and allochthonous C_ORG sedimentation and burial, sediment profiles indicated that seagrass losses triggered the erosion of historic C_ORG deposits. In carbonate seagrass meadows, CO₂ emissions driven by calcification can surpass C_ORG burial^[23]. Eddy covariance measurements in Floridan seagrass meadows indicated that net consumption of alkalinity by calcification explained most of the measured CO₂ emissions, exceeding anaerobic alkalinity production and C_ORG burial. In seagrass meadows containing carbonate-rich sediments, C_ORG sequestration may, therefore, be overestimated if CO₂ emissions from calcification are not considered.

Figure 3. Blue carbon stock (A) and burial rate (B) changes across revegetated plots of the seagrass Posidonia australis at Oyster Harbour, Western Australia. The fitted linear regression line in (A) is significant (P < 0.01). The horizontal and darkened lines in (B) are the mean and 95% confidence limits, respectively, of the carbon burial rate. Reproduced with permission from Marbá et al.^[22].

Macroalgal beds, such as kelp forests, may produce large standing stocks of blue carbon^[24], but such habitats do not sequester C_ORG but export C_ORG to adjacent coastal sediments^[1,25,26]. Measurements of carbon flux within a Japanese Sargassum bed^[26] found that 6%-35% of net community production was transported offshore as DOC.

Key international policy and carbon market frameworks

Restoration, management and conservation of blue carbon habitats have led to the development of climate mitigation, finance and policy frameworks that can be complex^[1,30,32,33]. For instance, sufficient financing remains a problem for starting and sustaining blue carbon projects.

Only net autotrophic ecosystems that sequester carbon meet the carbon market criteria of additionality and permanence

A genuine blue carbon ecosystem produces more net ecosystem C_ORG than it loses as ecosystem respiration (P:R ratio > 1). That is, it must be a net producer of blue carbon to allow sufficient C_ORG sequestration and storage. Coral reefs are not blue carbon systems because they are in overall metabolic balance (P:R ratio ≈ 1). Tidal flats, which store large amounts of C_ORG, are most often net heterotrophic (P:R ratio < 1). Tidal salt marshes mangroves and seagrass beds qualify, being net autotrophic and C_ORG storage habitats [Table 1]. Arguments have been raised suggesting that tidal flats, seaweeds and macroalgal beds constitute blue carbon ecosystems. Negative arguments have been raised^[34,35], including critical data gaps such as rates of CO₂ emissions and, in the case of macroalgae, the proportion of algal C_ORG exported to adjacent ecosystems, such as adjacent beaches and the deep ocean. Other workers, however, have concluded that macroalgae constitute a blue carbon ecosystem^[36-40]. For example, macroalgae contributed more C_ORG to seagrass sediments in Malaysia^[40] than seagrasses themselves , with mangrove C_ORG contributing the bulk of the material. Off Singapore^[41], Ulva and Sargassum assemblages accounted for 650 Mg C_ORG in biomass, greater than aboveground seagrass C_ORG but lower than mangrove C_ORG with a potential sequestration rate comparable to tropical grasslands and temperate forests. Similarly, arguments have been raised that although tidal flats are not necessarily net autotrophic, they often sequester large amounts of C_ORG. A global meta-analysis^[42] calculated that tidal flats accumulate C_ORG at a mean rate of 129.8 g C m^-2 year^-1 , slightly less than rates in acknowledged blue carbon habitats [Table 2]. However, as measured in Chinese tidal flats^[43], C_ORG sequestration rates vary widely from 35 to 361 g C m^-2 year^-1. Tidal flat sediments to a depth of 1 meter contain an average of 86.3 Mg C ha^-1 which is less than in blue carbon environments^[40]. At this stage, insufficient evidence exists to support the notion that macroalgal and tidal flat habitats are blue carbon ecosystems, especially since in the former case most macroalgal C_ORG tends to be sequestered in other habitats^[1,34,35]. In the latter case, tidal flats are declining globally due to extensive coastal degradation due to human developments, reduced riverine sediment export, sinking of river deltas, increased erosion and sea-level rise^[44].

Constraints on carbon accounting and investment are the large ecological and economic uncertainties; investor interest indicates a huge demand, but supply of blue carbon credits remains small

All ecosystems have large natural uncertainties that limit carbon accounting and investment that rely on knowing precise details to attract investors^[1]. Risk must be reduced to maximize the probability of return-on- investment; the likelihood of attractive returns needs to be demonstrated^[45]. Constraints include: (1) measurement uncertainty and significant variability in blue carbon storage, burial and greenhouse gas emissions; (2) logistical problems; (3) technical infrastructure and capacity; (4) governance issues; (5) existence (or not) of proper regulatory frameworks; (6) habitat permanence; (7) viability of property tenure; (8) long timeframe of blue carbon ecosystem maturity; and (9) possible need for different policy mechanisms for a specified finance type (e.g., public or private funding).

The potential need for carbon credits from investors and companies may be as much as $10 billion, but market-related challenges currently limit project implementation and the sale of resulting credits. Problems include blue carbon verification and costs, the small scale of blue carbon projects, and dual credit counting between national and commercial institutions^[46]. Other financial instruments such as ecosystem service insurance and bonds may be viable alternatives to more conventional forms of funding. A detailed analysis of the role of economics in blue carbon ecosystem marketing^[47] concluded that several actions are needed to view coastal blue carbon habitats as a natural climate solution: (1) improving legal and policy arrangements to ensure equitable sharing of benefits; (2) improving stewardship by incorporating local community knowledge and values; (3) clarifying property rights; (4) improving financial approaches and accounting tools to incorporate co-benefits; (5) developing technological solutions for low-cost measurement of blue carbon sequestration; and (6) clarifying information gaps of blue carbon cycling. Additional improvements to current practices include better stakeholder engagement, targeted use of public and philanthropic funding to subsidize pilot projects, reduce financial risks through collaterals, and promote low-risk but high-profit projects^[46,47].

Salt marshes, seagrasses and mangroves contribute to climate change mitigation at national, regional and global scales

Early studies^[3,31] justified their interest in blue carbon by asserting, without any empirical data, that blue carbon ecosystems must sequester significant amounts of C_ORG to the extent that they play a globally significant role in mitigating greenhouse gas emissions. Global areas of salt marshes and mangroves are small^[48] and their contributions are disproportionate, but recent calculations show that they mitigate only 0.4% (including seagrasses) of global CO₂ emissions related to fossil fuel use^[49]. Another estimate for mangroves indicates that they mitigate only 0.2% of total fossil fuel emissions. However, mangroves contribute significantly (18%) to emissions from the tropical ocean^[6] and both mangroves and salt marshes account for ≈70% of air-sea CO₂ exchange from global wetlands, and ≈30% of DIC export and ≈15% of DOC and POC export to the global coastal ocean^[6]. Nevertheless, revised estimates of global areas and C_ORG standing stocks of seagrasses [Tables 1 and 2] suggest that they are likely to be the main contributors to climate change mitigation. Global sequestration rates of seagrasses are ≈3 and 15 times, respectively, greater than salt marshes and mangroves [Table 2]. However, due to high habitat losses, seagrasses currently emit nearly double the amounts of CO₂ from salt marshes and mangroves [Table 2]. Salt marshes, mangroves and seagrasses account for 1.0%, 2.4% and 4.5% of global NEP of the global ocean, respectively^[42].

Climate change impacts on blue carbon dynamics are complex and highly variable over space and time

Climate change-related disturbances usually result in highly variable and significant changes in blue carbon stocks, sequestration rates, and greenhouse gas fluxes. Increasing sea temperatures, increasing frequency and intensity of heat waves, and extreme weather events lead to positive, negative or complex changes in blue carbon dynamics. A long-term record of sediment C_ORG stocks in a restored Zostera marina meadow^[50], depending on the position in the meadow, showed that a single heat wave led to significant losses of sediment C_ORG, with patterns of C_ORG losses and re-accumulation lagging seagrass recovery times within the central meadow. However, in the outer meadow, there was a net increase of 60% of sediment C_ORG throughout the following 3-year. Overall, areas of higher seagrass recovery maintained nearly twice as much C_ORG compared to areas of lower recovery. Under conditions of extreme weather events that result in mangrove dieback, degraded mangroves lose significant amounts of blue carbon, often as DIC outwelled to the adjacent coastal ocean^[51]. Degraded mangroves in the Dutch Caribbean, for example, lost 1.5 Mg CO₂ ha^-1 year^-1 from a 10% loss of sediment C_ORG compared with intact mangroves. Modeling exercises^[52,53] suggest complex effects of climate change on blue carbon dynamics will depend on future rates of sea-level rise and rates of temperature and precipitation increases and patterns. One model^[52] found a rise in global C_ORG storage by coastal tidal wetlands by 2100 based on projected emissions scenarios and current climate factors, with a rapid change driven by sea-level rise in salt marshes and higher temperature changes and changes in precipitation patterns in mangrove forests. Modeling the effects of two IPCC climate change scenarios ("business-as-usual" and "high-emissions") and using model estimates of past, present, and future deforestation rates of mangroves^[53], another model found that global C_ORG stocks will increase by ~7% regardless of scenario; this gain will be greater than losses from mangrove deforestation by 2100. However, results will vary regionally, with some areas seeing a drop in sediment sequestration rates. Thus, both empirical data and models indicate the complex effects of climate change on these ecosystems.

Export of porewater dissolved carbon and alkalinity via subsurface groundwater discharge is a major pathway of blue carbon loss, but alkalinity and DIC discharge can result in localized buffering of coastal acidification

Two preliminary mangrove carbon budgets^[54,55] show that ≈50% of mangrove carbon fixation was unaccounted for (known as "the missing carbon"), representing a large global carbon flux. This gap prompted extensive research, which ultimately revealed that large amounts of porewater DIC and DOC are exported from mangroves via subsurface discharge. This finding has resulted in a revision of the blue carbon budget^[56], in which the missing carbon is a large export of dissolved carbon. Tidal pumping^[57] is a major driver of mangrove porewater exchange and associated dissolved CO₂ export; global porewater-derived CO₂ export from mangroves (83 Tg C year^-1) was considerably greater than earlier estimates of flow via water-air exchange (34.1 Tg C year^-1) and C burial (18.4-34.4 Tg C a^-1), suggesting that the escape of porewater CO₂ is an important pathway of CO₂ flux to adjacent coastal waters^[57]. Work in Chinese mangroves^[58] indicates that about half of DIC produced in mangrove sediments is derived from both sulfate reduction and terrestrial groundwater discharge. Salt marshes are also an important reservoir of dissolved carbon in the coastal waters^[59-62], with most DIC generated in summer when higher temperatures stimulate anaerobic respiration^[59-62]. In some salt marshes, groundwater export of dissolved carbon exceeds sediment C_ORG burial rates^[62], reflecting the large amounts of dissolved carbon released to adjacent waters. Seagrass meadows laterally transport DOC with their rapid metabolism causing overlying waters to contain low CO₂ concentrations and enhancing atmospheric CO₂ uptake^[63,64]. Carbon fluxes are more complex in seagrasses inhabiting carbonate environments^[64], with variable rates of CO₂ flux in relation to seagrass composition. Downward fluxes were found only over seagrasses, whereas the presence of calcifying algae resulted in CO₂ release._. DIC export from a Posidonia meadow^[65] exceeded above-meadow CO₂ outgassing, with DIC uncoupled from alkalinity fluxes. The slight release of alkalinity suggested that ≈90% of exported DIC was eventually lost offshore as CO₂ was released to the atmosphere; combined flows indicated a carbon loss exceeding C_ORG sequestration.

Microbial decomposition of benthic organic matter via fermentation, sulfate, iron and manganese reduction, and methanogenesis results in several metabolic by-products (e.g., DIC and alkalinity) that are then transported via subsurface groundwater to adjacent coastal waters^[56,66]. In mangroves inhabiting carbonate deposits in the Red Sea^[66], total alkalinity emission from carbonate dissolution supported a CO₂ uptake of 345 g C m^-2 year^-1, 23-fold the C_ORG sequestration rate; total alkalinity emission was large (403 mmol m^-2 d^-1). In the Everglades^[67], most DIC was exported as total alkalinity (97 mmol m^-2 d^-1 of a total dissolved carbon export of 142 mol m^-2 d^-1), with both values several times greater than C_ORG burial rates. Large alkalinity export was also measured across an inshore to offshore mangrove continuum on the Amazon coast^[68], with export rates of DIC (20 mmol m^-2 d^-1) and alkalinity (15.2 mmol m^-2 d^-1) equating to nearly 80% of C burial rates. Globally, mangroves export 17 T_MOL year^-1 of alkalinity^[56]. Seagrass meadows and salt marshes produce and export alkalinity but at lower rates than mangroves^[68,69]. Alkalinity is produced via sulfate reduction coupled with pyrite formation, with greater pyrite stocks in mangroves than in salt marshes and seagrasses^[69]. Global alkalinity release from mangroves, apparently linked to the formation of pyrite, equals ~24% of the global mangrove C_ORG sequestration rate. Macrotidal salt marshes such as those in China^[70] export large amounts of DIC (202 mmol m^-2 d^-1) and alkalinity (78 mmol m^-2 d^-1) with a ratio which implies alkalinity production from anaerobic mineralization in sediment. Further, lateral exports were equivalent to C_ORG burial rates.

DIC and alkalinity export results in an observable, localized increase in pH in adjacent coastal waters, partially counteracting coastal acidification. The magnitude and impact of this effect will be site-specific, depending on residence times of tidal waters and on other sources or sinks of DIC and alkalinity. For mangroves^[71,72], a pH change is likely to be greater in greater areas of mangrove and likely to be more significant where coral reefs also occur. For seagrasses^[73], several factors influence their ability to buffer ocean pH, including photosynthetic activity^[73], mode of DIC uptake^[74], extent of seagrass loss^[75] and seasonality^[76]. For instance, in 7 Zostera marina meadows spanning the US west coast^[76], amelioration by seagrasses showed an improvement compared to non-vegetated areas, with an overall increase of 65% over 6 years, accompanied by a mean pH increase of 0.07. Sustained elevations of pH > 0.1 occurred for up to 21 days, even at night, with maximum pH increases in warmer months during maximum growth season. The buffering influence of seagrasses on ocean pH can also positively impact their epibionts and neighboring calcifying algae^[77,78]. Salt marshes similarly exhibit variations in pH, reflecting oscillations in DIC and alkalinity export^[79], complex mixing of estuarine water^[80], and climate^[81]. Macroalgae in the form of kelp beds^[82] and seaweed farms^[83] have the capacity to locally ameliorate ocean acidification via photosynthetic drawdown of CO₂ and may serve as refugia for carbonate-rich organisms, such as sea urchins and foraminifera^[84]. Furthermore, coastal acidification stimulates macroalgal germination^[85], growth^[86] and photosynthetic^[87] rates.

Mineralization of ancient wetland deposits contributes to DIC and DOC groundwater discharge to adjacent coastal waters

The large rates of dissolved carbon and alkalinity export from blue carbon ecosystems imply correspondingly rapid rates of microbial C_ORG mineralization in sediments, the source of these solutes. It was once assumed that the bulk of bacterial activity in wetland sediments occurs mostly in surface sediments (< 1-20 cm), mainly in association with recent deposition and subsequent decomposition of fresh organic material or surface microalgae and in proximity to roots and rhizomes, animal tubes and burrows^[55,56]. However, large C_ORG stocks exist several meters into the sediment and, in at least one location, to a mangrove sediment depth of 6 meters^[88]. Such deep deposits have been dated to 1,000-5,000 ¹⁴C BP (before present)^[56] and are thus derived from extinct habitats that are not directly linked to contemporary wetland habitats. Studies have also found that old porewater can be exported and that bioturbation can stimulate the decomposition of ancient, buried carbon^[89,90]. ²²⁶Ra measurements in mangrove sediments indicate that construction of new crab burrows exposes old deposits, enhancing mineralization of old carbon via deep sediment oxygenation by crab bioturbation^[89]. Measurements of ∆¹⁴C in exported DIC from subtropical Australian mangroves indicate the release of old DIC that is isotopically depleted, derived from at least a century-old sediment deposit^[90]. Century-old C_ORG buried in deep deposits is still in the process of being decomposed and transported via porewater and/or submarine groundwater discharge. This scenario is reminiscent of results in a coastal marsh^[91] where sediment ∆¹⁴C profiles indicated that sequestration rates at the decadal scale were much higher than rates at the millennial scale. This finding indicates either that old buried carbon was still being mineralized or that there was a recent increase in plant productivity. Bacterial metabolism in mangrove sediments has been measured to a depth of 1 meter^[56], showing no clear trends in bacterial metabolism with increasing depth, suggesting that mineralization of old buried carbon may persist to even greater depths. More research is needed to clarify the role of ancient wetland deposits on blue carbon cycling and wetland biogeochemistry.

Blue carbon ecosystems require allochthonous organic carbon inputs to achieve carbon mass balance

A recent carbon mass balance for mangrove ecosystems^[56] found that rates of autochthonous C_ORG incorporation (root, wood, leaf litter production) into forest sediments were not sufficient to balance microbial decomposition rates. One possibility to redress this imbalance is to include mineralization from deep ancient deposits (Perspective 6) as an ecosystem carbon input. Another possibility is that mangroves are reliant on allochthonous inputs to balance ecosystem carbon flow. In fact, about 58% of δ¹³C data of mangrove sediment^[92] is low (< -25‰), suggesting mangrove litter, but relatively high values indicate large inputs of ¹³C-enriched carbon sources, such as various algae and river-derived POC rather than mangrove-derived plant matter. About 60% of known mangrove study sites^[56] show that sediment C_ORG is derived mostly from allochthonous sources, including seaweed and algal-derived detritus from seagrasses and coral reefs, higher terrestrial land plants, dead bacterial biomass, and phytoplankton and POC derived from adjacent rivers, estuaries, coastal and open ocean sources. However, the proportion of mangrove litter, roots and wood detritus to the sediment C_ORG pool increases with stand age. In many seagrass meadows, non-seagrass detritus accounts for ~50% of sediment C_ORG, likely originating from allochthonous organic material that is transported and trapped by seagrass sediments. The percentage of allochthonous inputs to seagrasses is partly a function of distance from the source^[93]. For example, salt marshes at San Quintin (an arid Eastern Pacific lagoon) receive 40% of their C_ORG from external inputs, whereas in seagrass meadows open to lagoonal currents, the C_ORG source is largely allochthonous^[94].

Restored ecosystems often reveal temporal differences in blue carbon sources. In restored and pristine salt marshes in Washington, USA^[95], autochthonous marsh C3 plants contributed 73%-89% to unaltered marshes, whereas restored marshes received a wide array of mostly allochthonous C_ORG, which varied in relative contribution by distance from the source and abundance. All three coastal ecosystems thus receive and are greatly reliant on allochthonous sources of C_ORG to maintain their metabolic requirements and to balance their carbon inputs and outputs.

Habitat connectivity enhances blue carbon uptake and release

The export of allochthonous matter from one ecosystem to another illustrates the importance of connectivity between habitats, enhancing the flow of blue carbon. Mangrove-derived seston exported to adjacent seagrasses in Gazi Bay, Kenya^[96] resulted in increased CO₂ concentrations and increased C_ORG content in the water column; flow reverses via tides transport high-quality seagrass carbon to the mangrove forest. The extent of input from one habitat to another partly depends on distance. On the north Sulawesi coast, Indonesia^[97], seagrasses adjacent to mangroves have significantly higher sediment C_ORG concentrations than in non-adjacent meadows; in the adjacent meadows, mangrove C_ORG contributes 34%-83% to sediment C. Several mechanisms exist through which one blue carbon habitat may influence C_ORG concentration and sequestration in adjacent habitats, such as outwelling, sediment trapping, nutrient fluxes, wave dampening, trophic cascades, and tidal regimes^[98]. The co-metabolism effect may also result in enhanced carbon flow^[99]. In experiments where macroalgal detritus was mixed with seagrass detritus, high density macroalgal addition stimulated the decomposition of 20% more seagrass compared to other treatments. Coral reefs and seagrass meadows often co-occur and are connected by various ways, including enhancement of water column and sediment carbon. For instance, seagrasses benefited by close association with Colombian coral reefs^[100], with seagrasses within reef barriers having greater sediment C_ORG concentrations than seagrasses without a reef barrier. Seagrasses also benefit from close physical proximity to mangroves as found on the Andaman and Nicobar Islands, India^[101], and along a mangrove-seagrass-reef continuum fringing Iriomote Island, Japan^[102], where carbon transport from adjacent mangroves and rivers offsets uptake of CO₂ by seagrasses; the coral reefs take up DIC, DOC and alkalinity, fostered by suppression of high sedimentation by the ability of mangroves to trap sediment particles.

Bioturbation, grazing, and predation affect blue carbon dynamics

Biota can affect several aspects of blue carbon via a network of trophic interactions among plants, animals and microbes, exerting strong effects on biogeochemical dynamics in blue carbon habitats^[103]. For instance, a variety of predators influence blue carbon stocks and sequestration rates^[104], including the effects of reduced predation on C_ORG sequestration in salt marshes and mangroves and C_ORG concentrations in seagrasses. Declines in various predators (e.g., New England blue crabs, predatory mangrove fish in Australia) result in an increase in bioturbator numbers but a decline in the capacity for C_ORG sequestration^[104]; in Shark Bay, Western Australia, the presence of tiger sharks results in sea turtles and dugongs foraging in seagrass communities that are low in terms of risk from predation, but leads to these seagrasses having lower C_ORG stocks than areas associated with high predation risk. Arguably the greatest and most direct impacts by biota on blue carbon are produced by bioturbators^[105,106]. For example, salt marsh crabs produce burrows that enhance CO₂ release across the sediment-water interface due to tidally- driven hydraulic gradients^[105,106]. Grazing has a similar effect on seagrass meadows, resulting in a several-fold increase in CH₄ emissions; this increase can be attributed to the decline in loss of downward oxygen transport caused by the grazing decline in living seagrass biomass^[107]. An assessment of risks^[108] indicates that indirect, large-scale threats from ocean warming, heatwaves, and sea-level rise and direct small-scale threats from activities such as dredging and the building of jetties or a marina have a higher risk of increasing CO₂ emissions from seagrass than to recover their ability to sequester C_ORG.

Calcium carbonate chemistry affects blue carbon budgets

The presence of calcium carbonate (CaCO₃), especially in tropical seagrass meadows, plays a crucial role in inorganic carbon cycling (Perspective 5) and cycling of blue carbon. CaCO₃ production emits CO₂, partially offsetting the role of carbonate-rich ecosystems as CO₂ sinks through C_ORG sequestration^[109]. A global analysis^[109] of inorganic carbon (C_INORG) burial rates indicates rates of 0.8 Tg C_INORG year^-1 and 15 to 62 Tg C_INORG year^-1 in mangroves and seagrasses, respectively. Burial of CaCO₃ may correspond to an offset of as much as ≈30% of seagrass CO₂ sequestration. Seagrasses living in CaCO₃-rich sediments may have CO₂ sequestration neutralized by CO₂ released via calcification^[110]. Thus, CO₂ emissions driven by calcification processes may exceed rates of C_ORG sequestration in carbonate seagrass habitats^[23]. In Florida seagrass meadows, net consumption of alkalinity by calcification accounts for nearly all CO₂ emissions^[23]. Therefore, net carbon sequestration in carbonate-rich environments may be overestimated if calcification is not included in chemical budgets. Some tropical seagrass meadows with calcareous algae may have higher seagrass productivity and C_ORG storage^[111]. Off Zanzibar, Tanzania, biomass production was higher in meadows containing both seagrass and calcareous algae than in meadows without algae; sediment C_ORG levels were also highest in seagrass meadows with calcareous algae.

Methane and nitrous oxide emissions reduce blue carbon potential

As both CH₄ and N₂O are significantly more potent greenhouse gases than CO₂, their emissions reduce the value of a given habitat as a carbon sink, and in salt marshes, may negate the value of stored C_ORG towards mitigating climate change^[112]. In coastal marshes in Georgia, USA^[113], CH₄ fluxes via groundwater average ≈3 g CH₄ m^-2 year^-1 as a greenhouse gas over a century timeframe, which is equal to rates from other salt marshes. Submarine groundwater discharge results in a large increase in CH₄ export that offsets about 30% of the cooling potential from C_ORG sequestration. In mangroves, CH₄ efflux rates similarly have the potential to partly offset C_ORG burial rates, possibly by as much as 20%^[114]. Both dead and living mangrove tree stems release CH₄ to the atmosphere, with rates 8-fold higher from standing dead stems than from living stems; upscaling to mangrove dieback in the Gulf of Carpentaria, Australia, dead tree emissions may account for as much as 26% of net ecosystem CH₄ flux^[115]. Processes regulating CH₄ and N₂O fluxes in mangroves, marshes, and seagrasses are complex, complicating efforts to derive blue carbon benefits^[116].

Black carbon reduces estimates of labile carbon in blue carbon budgets

"Black carbon", defined as carbon formed by incomplete combustion of fossil fuels, burning biomass, and other pollution, is highly recalcitrant and found in increasing abundance and proportion in soils and sediments, entering coastal waters via rivers and atmospheric deposition. Black carbon can account for as much as 9% to 25% of organic carbon content in subtropical-temperate deltas of the major rivers in China^[117], with an accumulation rate of 33 g m^-2 year^-1 in China’s coastal wetlands, suggesting they act as a hotspot for black carbon and a substantial contributor to blue carbon storage. Similar results have been measured in a temperate salt marsh in Tasmania, Australia, and in a tropical Malaysian seagrass meadow^[118]. Black carbon accounted for 38% and 22% of the sediment TOC fraction in the seagrass and marsh ecosystems, respectively, demonstrating the need to account for black carbon in future blue carbon assessments. The contribution of black carbon to total C_ORG standing stocks depends partly on environmental history of the ecosystem^[119]. For example, in a fire-affected region of Malaysia^[120], the black carbon to total C_ORG ratio was 43% in an adjacent seagrass meadow but 18% in a sandy, low organic content meadow; in moderately polluted mangroves, the ratio was 2% but higher (6%) in highly polluted mangroves. The black carbon to total C_ORG ratio for seagrasses also varies between urban and rural meadows and is twice as predicted, at 28% and 36%, respectively. The high fractions are likely the result of variability of black carbon atmospheric supply, an increase in seagrass litter loss close to exposed meadow edges, and sediment resuspension within the dispersed patchy meadow. Black carbon thus has an overall negative impact on the storage of labile blue carbon.

Habitat destruction results in blue carbon stock losses, but variable gas emissions

Rates of sediment C_ORG changes and greenhouse gas emissions vary partially depending on the type of disturbance to which an ecosystem is subjected. Clear-felling of mangroves, for instance, reduces sediment C_ORG stocks via erosion and subsequent transport^[1], but short-term increases in CO₂ release may occur^[121]. In Mexican mangroves^[122] subjected to hydrological perturbations, CH₄ emissions were highest in the wet season in a dead mangrove area, whereas CO₂ emissions were highest during the dry season with no significant differences among disturbance levels; N₂O release did not vary among seasons and degradation levels. In a mangrove area in Vietnam^[123] that was cleared 2-year prior to sampling, sediment losses amounted to ~10 Mg C_ORG ha^-1 year^-1 with mean sediment CO₂ fluxes of 24 mmol m^-2 d^-1, nearly 3 times lower than rates in planted 16-year old forests. Reforested and rehabilitated forests reveal different rates of greenhouse gas release depending on the state of the forest and the type of gas. In managed forests in the Ranong Biosphere Reserve, Thailand^[124], the highest CO₂ effluxes were measured in an undisturbed primary forest compared to reforested and naturally rehabilitated forests. However, CH₄ emissions were significantly higher from the natural rehabilitation forest, especially in low-intertidal areas with longer tidal flooding; N₂O emissions did not vary among mangrove forest types. In cyclone areas, mangroves can lose blue carbon mostly as enhanced export of litter and DIC^[125]. For salt marshes, various disturbances can affect blue carbon fluxes such as sediment metabolism. In high salt marshes in Virginia, USA^[126], impacted by sea-level rise, modeling indicates that increased inundation would reduce respiration rates to a greater extent than primary production, resulting in a moderate loss of blue carbon. Impoundment can also have significant impacts on blue carbon gas emissions in salt marshes. In Phragmites- invaded marshes in Massachusetts, USA, freshening to lower salinity led to a 50-fold increase in CH₄ emissions but with only moderate increases in CO₂ efflux^[127]. Although freshened, impounded marshes can be strong carbon sinks with CH₄ release partially offsetting C sink capacity.

Benthic blue carbon stocks are increasing in polar and subpolar regions

Blue carbon stocks are increasing in polar seas partly due to prolonged phytoplankton blooms driven by losses of sea ice and collapsing ice shelf as a result of climate change; these blooms, in turn, fuel productivity of benthic infauna and, to a much lesser extent, epibenthos^[128]. Sea ice losses also make the West Antarctica continental shelf a central location for iceberg disturbance as ice scour limits C_ORG storage and resets succession of benthic biota, fertilizing the ocean with nutrients and generating phytoplankton blooms in which excess organic carbon eventually deposits onto the seabed where it is partly consumed and buried^[129]. Along Antarctic fjords, glacial retreat also results in blue carbon gains, leading to an increase in zoobenthic blue carbon of up to 4,536 t C_ORG year^-1[130]. At three fjords along the West Antarctic Peninsula, further pathways of C_ORG storage were measured along a temporal gradient since the last glacier cover to test the idea that seabed carbon standing stocks would be a function of glacial ice cover^[131]. The results were complex, with highly variable infauna numbers and carbon concentrations over time, like the epifauna. However, sediment C_ORG concentrations were much greater than total benthic faunal carbon. In shallower waters around the Antarctic Peninsula, infauna contained 60% less blue carbon (391 t C_ORG km^-2) than epibenthic communities (648 t C_ORG km^-2)^[132]. This carbon sink may double as climate-mediated ice loss continues. Increasing melt from glaciers not only results in blue carbon storage but shifts benthic communities from net autotrophs to heterotrophs^[133]. Clear waters during cold El Niño in spring 2015 facilitated high benthic microalgal productivity, shifting the benthic community to net autotrophy. However, increased glacial melt during the La Niña spring of 2016 resulted in water column turbidity, limiting benthic algal productivity, and turning the benthic faunal community from net autotrophic to net heterotrophic.

Data from the Arctic is extensive for carbon accumulation in infauna and sediments, whereas most Antarctic blue carbon data was confined until recently to epifauna. In Arctic fjords, pelagic and benthic carbon is abundant, with high rates of C_ORG deposition, sediment burial and faunal consumption^[134-136]. Both terrestrial and macroalgal C_ORG were important sources of sedimentary C_ORG in most fjords, many of which are hotspots of C_ORG sequestration. However, there are strong spatial gradients in sedimentation rate and primary productivity that in turn impact benthic C_ORG uptake and burial^[135,136]. Driven by warmer temperatures, glaciated fjords may become less efficient at burying C_ORG in deep sediments^[136]. One cause for the long-term persistence and high C_ORG sequestration rates in Arctic sediments may lie in the tight binding of C_ORG to reactive iron and manganese, resulting in preservation for thousands of years^[137]. In Arctic coastal wetlands in Norway^[138], carbon sequestration rates were highly variable over space and time (9-603 g C_ORG m^-2 year^-1), correlating with an increased growing season. Blue carbon storage is low mostly due to new marsh formation with a thin organic sediment layer. Stocks of C_ORG and the increasing frequency of shallow polar habitats may be greater than previously assumed, equating to negative feedback on the rapidly changing polar climate.

Historical ecological shifts occur among blue carbon ecosystems

Historical ecological shifts occur in coastal blue carbon ecosystems. Tidal mudflats can be colonized by mangrove propagules until successive changes result in a mangrove forest beneath which lies the remnants of the original tidal flat. These historical deposits can continue to participate in contemporary carbon cycles (Perspective 6). In four Australian wetlands, environmental histories were reconstructed with respect to blue carbon conditions to assess the impact of historical factors on modern C_ORG stocks and burial rates^[139]. High C_ORG stocks were linked to thick mangrove peat layers below the modern salt marsh deposits in both marine and fluvial sites where historic mangrove roots were found to be chemically stable, considering their advanced age. Cores from marine-influenced and fluvial-influenced salt marshes did not contain mangrove root C_ORG, but there was significant preservation of charcoal at the fluvial-influenced marsh. These results suggest that both classification of modern vegetation structure and geomorphic setting are insufficient to account for carbon stock variability and accumulation rates. Ecosystem changes are not uncommon and can still result in the preservation of old C_ORG. Stratigraphic evidence of extensive mangrove root networks beneath modern grasslands and salt flats in northern Australia suggests the existence of extensive, pre-historic mangroves approximately 5,500-7,000 years BP^[140]. Other studies such as in the Mekong River delta have similarly revealed the presence of ancient mangroves, salt marshes, and tidal flats below modern deltaic sediments^[141]. The significance of this ancient preserved carbon remains unknown, but may be of great biogeochemical and geological significance considering the sizes of many deltaic systems. Evidence of successional patterns in mangrove vegetation in locations such as the Mekong Delta^[142] may also be of great importance with regard to blue carbon, but similarly are undetermined. Molecular fingerprinting can now identify sources and preservation state of organic matter in blue carbon habitats^[143,144]. Moreover, habitat shifts affecting blue carbon stocks have occurred, and continue to occur, among coastal communities in Spencer Gulf, South Australia. The main source of sediment organic matter in these habitats is vascular plant material that is well-preserved due to enhanced lignocellulose preservation under anoxic conditions^[145]. There was a clear molecular change discovered at a salt marsh where encroachment of young mangroves is occurring, but with a sediment core from the marsh-mangrove location reflecting a complex history. However, most C_ORG sequestered in the upper sediment layers may have originated from invading mangroves. Further, a sediment core from a mangrove habitat shows a dramatic shift in C_ORG content between surficial and deeper sediment layers, coinciding with a major shift in δ¹³C. This change indicates that about a millennia ago, seagrasses inhabited this area, delimiting a change from a subtidal to a more intertidal environment, coinciding with a known sea-level decrease in this region^[144].

Land-use and land-cover changes (LULCC) drive changes in emissions and blue carbon stocks

Blue carbon ecosystems are impacted by various types of natural and anthropogenic changes due to LULCC, the most common being unsustainable extraction of wood and other products, clearing and conversion for agriculture (pastures, plantations) and aquaculture (shrimp ponds, salt farms), impoundment and drainage for pest control, urban development (construction for housing and industry), constructed and restored wetlands, pollution, climate change, and natural and coastal erosion induced by land-use change. The level of impact depends on factors such as spatial and temporal change, and the frequency and magnitude of the disturbance. A meta-analysis determining the blue carbon response of coastal wetlands to LULCC^[145] concluded that conversion to various land-use changes resulted in decreased sediment C_ORG concentrations and stocks by an average of 18%, with the loss of sediment C_ORG causing likely CO₂ emissions in the range of 2 to 23 Mg CO_2eq ha^-1 year^-1; in contrast, conversion to artificial wetlands may result in a net uptake of about -3 Mg CO_2eq ha^-1 year^-1.

A model forecast of CO₂ emissions from blue carbon ecosystems post-disturbance^[146] indicates that more than 40-year post-disturbance, the total CO₂ emitted from blue carbon habitats was about 70%-80% of the initial C_ORG standing stock. Empirical evidence from blue carbon ecosystems along the Mexican coasts^[147-149] indicates historical changes in blue carbon stocks in sediments. Increasing salinization in the Gulf of Mexico^[147], as a result of sea-level rise, resulted in higher C_ORG stocks and sequestration in Halodule wrightii beds than in Thalassia testudinium meadows where only sequestration rates increased due to erosion caused by land-use change and sea-level rise. In Mexican mangroves, seagrasses, and salt marshes^[148], most stored sediment C_ORG (60%-86%) during the Anthropocene (circa 1950) and contemporary mangrove C_ORG stocks were up to 6 times higher than pre-1950, probably because of land-use changes. Increased C_ORG sequestration rates in seagrasses were primarily due to increases in sediment accumulation as a result of land erosion and changes in sea level. Changes in C_ORG concentrations in mangroves and salt marshes were likely due to productivity increases in the runoff of fertilizer and sewage. Salt marshes in the northeast Pacific coast of Mexico have low input rates of terrestrial matter and sediment accumulation rates than in adjacent seagrass meadows. However, throughout the past century, accumulation rates have steadily increased due to soil erosion from land-use change (agriculture, aquaculture)^[149]. In tropical seagrass meadow sediments off Zanzibar, Tanzania, mass accumulation rates similarly increased due to land-use change from 1998 until 2018, linked to shifts in C_ORG sources^[150]. However, different land-based activities resulted in different δ¹³C signatures of sediment C_ORG between two meadows, with one meadow indicating increased δ¹³C signatures after a sewage outlet was moved and the other meadow showing a decrease in δ¹³C signatures because of increased sediment runoff post-deforestation. Long-term data from salt marshes in the Yellow River delta, China, show that continuous LULCC transformation driven by natural and anthropogenic processes decreased carbon storage by 10.2% during 1970-2010^[151].

Globally, mangrove conversion to shrimp ponds remains the most common type of land-use change, with significant loss of sediment and vegetation carbon from land clearing. In the Dominican Republic^[152], where such conversion and subsequent abandonment of ponds is common, C_ORG stocks were lowest in abandoned shrimp ponds (11% of that of mangroves) and highest in intact mangroves. Estimated emissions from such conversions range from 2,244-3,799 Mg CO_2eq ha^-1. The conversion of mangroves to aquaculture ponds in the Dominican Republic resulted in total emissions of ≈4 million Mg CO_2eq or ~21% of pre-converted blue carbon. Shrimp ponds similarly led to massive losses of mangrove carbon in Brazil^[153] with 58%-82% loss of C_ORG stocks and estimated emissions of 1,390 Mg CO_2eq ha^-1 year^-1. In Thailand, where mangrove losses to aquaculture and agriculture have been historically severe^[154], high-resolution data sets were used to quantify mangrove C_ORG stock changes over the past 60-year. Net loss of mangrove carbon over the period was about 32.3 million Mg C_ORG, but during the same period, 2.6 million Mg C_ORG accumulated due to recovery gains in mangrove extent, mostly in areas of pristine forests; substantial increases in C_ORG stock (1.3 million Mg C) were measured as a result of the establishment of new mangrove on accreting tidal flats. Conversion to other land uses similarly resulted in large losses of mangrove C_ORG and CO₂ as such pasture conversion^[155] resulted in emissions of ~ 1,460 Mg CO_2eq ha^-1.

Several estimates^[156,157] have documented the net loss of mangrove carbon stocks globally from LULCC. These estimates reveal a net reduction in mangrove C_ORG biomass by 82%, while sediment C_ORG experienced a reduction of 54%, with relative losses depending on LULCC type, the time since LULCC occurred, and the geographical and climatic conditions of the respective locations. Losses of sediment C_ORG stocks are linked to decreased sediment C_ORG content and increased sediment bulk density over the upper 1 m depth, but there are no significant effects of LULCC on greenhouse gas emissions from mangrove sediments. The global mangrove C_ORG stock declined by 158.4 million t C_ORG between 1996 and 2016^[157], representing a 1.8% reduction compared to the stock present in 1996. The relatively small to moderate losses of C_ORG stocks over the last two decades can be partly attributed to efforts to conserve and restore mangroves.