Foregone carbon sequestration loss dominates greenhouse gas budget in mariculture associated with coastal wetland conversion

doi:10.21203/rs.3.rs-3670767/v1

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Foregone carbon sequestration loss dominates greenhouse gas budget in mariculture associated with coastal wetland conversion

https://doi.org/10.21203/rs.3.rs-3670767/v1

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Coastal wetlands represent a disproportionately large global carbon (C) sink. However, they are threatened by the ever-expanding aquaculture and being lost at critical rates. Conversion of coastal wetlands to aquaculture systems has been predicted to result in significant C losses, yet accurate assessments of greenhouse gas (GHG) budgets associated with this major perturbation are rarely available. Here we show that the conversion of Spartina alterniflora saltmarsh to mariculture ponds in China induced a dramatic shift from net atmospheric GHG sink (‒13.8 Mg CO₂eq ha^–1 yr^–1) to net GHG sources (2.16 Mg CO₂eq ha^–1 yr^–1), creating a full GHG debt of 15.9 Mg CO₂eq ha^–1 yr^–1. The loss of foregone GHG mitigation capacity of saltmarsh makes the largest contribution (86.4%), while only 15.6% of the total debt arises from direct CO₂, CH₄, and N₂O fluxes in the mariculture ponds. But considering the main drivers of GHG emissions from animal protein production, mariculture has much lower GHG-cost than inland freshwater aquaculture and terrestrial beef, small ruminants, and pork production on a kg CO₂eq kg^‒1 protein basis. The low-C mariculture could be further realized by avoiding devastation of vegetated coastal wetlands and minimizing CH₄ emission by carrying out in high-salinity waters.

Scientific community and society/Agriculture

Earth and environmental sciences/Environmental social sciences/Environmental impact

Coastal wetlands

aquaculture

land-use change

net ecosystem exchange

methane

nitrous oxide

global warming potential

Coastal wetlands are one of the most productive ecosystems and represent an important component of global carbon (C) sinks¹. They can actively sequester carbon dioxide (CO₂) by trapping and burying autochthonous and allochthonous C into sediments and hold up 10–100s times greater rates of C storage per unit area than terrestrial forests². Coastal wetlands only occupy 0.07–0.22% of the Earth’s surface but bury 0.08–0.22 Pg C yr^–1, being equivalent to ~ 10% of the entire net residual land C sink^3,4. Nevertheless, coastal wetlands emit much lower methane (CH₄) relative to freshwater wetlands, and hence have long been recognized as more valuable C sinks for climate change mitigation⁵. However, coastal wetlands are among the most threatened by and vulnerable ecosystems to human activities such as agriculture and aquaculture development, coastal construction, and anticipated climate change and sea level rise^6–8. Currently, the area of global coastal wetlands decreases at an annual rate of 0.7–7% (ref. ²). Recent estimates have indicated that 29‒50% of global coastal wetlands were lost or degraded during the last four decades⁹.

Among all the anthropogenic and natural factors, conversion to aquaculture ponds is the major single cause of coastal wetlands degradation and decline, especially in the Asia-Pacific region^10,11. With the ever-rising demand for animal-based protein and the dwindling of wild fish stocks, global aquaculture has increased rapidly since the late 1980s (ref. ¹²). Nowadays, the contribution of aquaculture to food fish consumption has risen from 19% in 1990 to 52% in 2018 and is expected to increase to 62% by 2030 (ref. ^12,13). The ever-expanding aquaculture accounts for 52% of the 3.6 million ha global mangrove loss since 1980 (ref. ⁹). In China, more than 30% of the newly developed mariculture ponds during 1984‒2016 were converted from coastal saltmarshes¹¹. Upon the steadily increasing demand for fish protein, it is therefore likely that more natural coastal wetlands would be converted to aquaculture ponds in the future^8,10.

The high C densities per unit area of coastal wetlands, coupled with quick losing rates, imply that loss of global coastal wetlands may generate substantial C emissions. Pendleton et al.¹⁴ estimated that total C emissions from conversion and degradation of coastal wetlands were to be 0.15‒1.02 Pg CO₂eq yr^‒1 globally. Kauffman et al.¹⁵ found that the potential C losses from conversion of coastal wetlands (e.g., mangroves) to shrimp ponds in the Dominican Republic were up to 1,067–3,003 Mg CO₂eq ha^‒1. Consistently, Murdiyarso et al.¹⁰ reported that the conversion of mangrove to shrimp ponds in Indonesia resulted in C emission of 0.07‒0.21 Pg CO₂eq yr^‒1 due to the losses of aboveground and underground C pools. Moreover, aquaculture ponds often receive large amounts of fish feeds and fertilizers to improve production, while only a small portion of these supplied nutrients (11–36% for nitrogen) is converted into fish biomass¹⁶. The residual organic matter and nutrients could lead to high emissions of methane (CH₄) and nitrous oxide (N₂O) from aquaculture ponds^17,18. As CH₄ and N₂O are more powerful than CO₂ at trapping heat in the atmosphere, a comprehensive scaling of the net emissions from all the three major greenhouse gases (GHG) is necessary to evaluate how aquaculture expansion contributes to global climate change. Particularly, conversion of natural saltmarsh could not only result in the losses of soil and biomass C but also demolish one of their most important ecosystem services, e.g., high rates of CO₂ capturing, whereas this foregone GHG mitigation capacity losses and its contribution to the full GHG budget of mariculture has never been scaled in previous studies.

Recently, the updated Intergovernmental Panel on Climate Change (IPCC) reports provided the methodological guidance on GHG emissions from aquaculture systems^19,20. However, the lack of field direct measurement data increased the uncertainties of the IPCC default emission factors and budget estimates for GHG emissions from aquaculture. For instance, the IPCC Tier 1 CH₄ emission factor for saline agriculture and aquaculture ponds has been explored by only 3 published reports from tropical regions²⁰. As for the N₂O emission from aquaculture, the IPCC default emission factor and current estimates are also limited by modeling approaches using mathematic reckoning or known as indirect N₂O data²¹. Large variations in N₂O emission factors from aquaculture systems have been found due to the large variability in environmental conditions and yield performance of culturing species^21,22. Previous studies reported that N₂O production in extensively managed aquaculture ponds was limited by the high dissolved organic C (DOC) concentrations and anaerobic conditions, and the yield-scaled emission factor for N₂O in these ponds was 7.85-time lower than the IPCC default value^17,23. Therefore, field direct measurements are needed for validating current estimates, specific emission factors development for diverse aquaculture systems, and adoption of effective mitigation strategies.

China’s aquaculture industry has experienced rapid development during the past four decades. Nowadays, China alone contributes 58% of global mariculture volume and roughly half of global mariculture area^12,24. The ever-expanding mariculture has caused substantial loss of coastal wetlands in China and worldwide^9,11. There is clearly an emerging interest and urgent need for greater appreciation of the GHG emissions incurred during the severe land-use change. Here we quantified the annual fluxes of carbon dioxide (CO₂), methane (CH₄), and nitrous oxide (N₂O) from a natural coastal saltmarsh and three adjacent saltmarsh-converted mariculture ponds within 1 to 5 years, to access the full GHG debt of this typical mariculture mode in China (Supplementary Fig. 1). We also scaled the relative magnitude of each component of GHG from various aquaculture systems (e.g., inland freshwater aquaculture, and brackish shrimp farming and polyhaline polyculture in coastal regions) and compared emissions from aquaculture to those from livestock production. Findings here will help illuminate the role that aquaculture has in the environmental cost of global food-production systems and explore the potential mitigation strategies for aquaculture.

2.1 Impacts on CO₂ fluxes

Saltmarshes are among the most productive ecosystems in the world with the capacity to store large amounts of C per unit area, referred as “Blue Carbon” ². Our high-resolution eddy covariance (EC) measurements of CO₂ fluxes revealed that the S. alterniflora marsh was a net sink of atmospheric CO₂. Daily CO₂ fluxes in the marsh exhibited distinct seasonal patterns, with peak CO₂ fluxes, e.g., gross primary productivity (GPP), ecosystem respiration (Re), and net ecosystem CO₂ exchange (NEE), occurring in summer when vegetation has grown vigorously (Fig. 1a). From late-September, CO₂ fluxes gradually declined since both GPP and ER were inhibited by cool temperature. Because of the long green period of S. alterniflora, the saltmarsh switched from net CO₂ source to net CO₂ sink in early April and did not switch back to net CO₂ source until December. Over an annual cycle, 65.2 Mg CO₂ ha^‒1 (1,778 g C m^‒2) was accumulated as GPP and 47.4 Mg CO₂ ha^‒1 (1,293 g C m^‒2) was released as Re in this saltmarsh, resulting in a net uptake of 17.8 Mg CO₂ ha^‒1 (485 g C m^‒2) annually (Fig. 2a).

In contrast to S. alterniflora marsh, NEE in the mariculture ponds fluctuated between ‒34.1 and 215.3 mg CO₂ m^‒2 h^‒1 and no apparent seasonal variations were observed, except for the pulse of CO₂ emission in the initial drainage period (Fig. 1b). On an annual basis, mariculture ponds were net atmospheric CO₂ sources ranging from 0.91 to 1.76 Mg CO₂ ha^‒1, and the highest CO₂ emission was observed in the 1-year old mairculture pond (MP1) (Fig. 2a). The NEE in our ponds compared reasonably with observations from other mariculture systems which also receives feed inputs. Zhang et al.²⁵ reported an annual CO₂ emission of 4.69 Mg CO₂ ha^‒1 in a nearby crab-shrimp-clam polyculture system. Recent seasonal measurements from subtropical estuarine shrimp ponds also demonstrated culturing-period CO₂ emissions ranging from 0.29 to 0.55 Mg CO₂ ha^‒1 (ref. ²⁶). Aquaculture waters have been documented autotrophic and act as a sink of atmospheric CO₂, because primary production of phytoplankton is generally higher than the water-column respiration^26,27. However, phytoplankton is easily decomposed compared to plant biomass, resulting in minimal retention of primary production. Yang et al.²⁶ reported a net water-column CO₂ fixation of 2.02 Mg CO₂ ha^‒1 in a subtropical shrimp pond, which was ~ 1 order of magnitude lower than the CO₂ uptake rate in S. alterniflora marsh. Therefore, the shift of CO₂ sink to source following saltmarsh conversion can be primarily due to the loss of CO₂ sequestrating capacity of primary producer. Specifically, prevalence of CO₂ efflux in the ponds suggested that water-column CO₂ uptake was largely counterbalanced and surpassed by other components of respiration, e.g., soil respiration²⁶. It has been estimated that only < 20% of the feed C can be converted to biomass in aquaculture systems, the residuals are buried and then transformed to CO₂ and CH₄ by microbes^17,28. However, we found that NEE was significantly higher in MP1 than in the 4-year (MP2) and 5-year (MP3) ponds (Fig. 2a). As C inputs from feeds and fertilizers were similar among ponds, the difference in NEE indicated that the decomposition of residual fish feed and feces can only be partly responsible for the CO₂ emission. Decomposition of SOC, on the other hand, might account more for the CO₂ efflux. This can be supported by the significant lower SOC contents in the older ponds relative to MP1 (Table 1). Losses of SOC have been widely reported in wetlands subjecting to reclamation and conversion to other land uses^29,30. Arifanti et al.³¹ found that the SOC mineralization rate over 16 years following mangrove conversion to shrimp ponds was up to 24.5 Mg C ha^‒1 yr^‒1, accounting for 75.0% of total C losses. When coastal wetlands are converted into aquaculture ponds, the top layers of soil are excavated and exposed to oxygen which enhances the susceptibility of SOC to mineralization. Further, operation activities such as aeration, drainage and harvest could also destabilize the SOC, resulting in large amounts of CO₂ emission during the initial culturing years³². Thereafter, subsequent losses were much slower probably due to the increased proportion of recalcitrant organic C fractions^33,34, as exhibited in the relatively lower NEE in MP2 and MP3.

Table 1

Soil properties in the *Spartina alterniflora* saltmarsh and mariculture ponds.
Systems	Culture time (year)	Area (ha)	SOC (g C kg^–1)	TN (g N kg^–1)	DOC (mg C kg^–1)	NH₄⁺ (mg N kg^–1)	NO₃^– (mg N kg^–1)	WD (cm)	DO (mg L^–1)	Salinity (ppt)
Saltmarsh	–	–	15.8 ± 0.2a	1.04 ± 0.04b	135 ± 9a	2.31 ± 0.18b	0.14 ± 0.03b	0.12 ± 0.47b	–	28.2 ± 0.1b
MP1	1	3.87	13.5 ± 0.1b	0.82 ± 0.03a	47.0 ± 3.8b	24.4 ± 1.3a	1.41 ± 0.21a	88.2 ± 4.0a	6.83 ± 0.25a	30.8 ± 0.5a
MP2	4	4.83	11.9 ± 0.2c	0.78 ± 0.01a	36.3 ± 3.0c	26.8 ± 1.9a	1.35 ± 0.23a	88.9 ± 5.6a	8.12 ± 0.33b	29.0 ± 0.4b
MP3	5	4.47	11.1 ± 0.2d	0.77 ± 0.01a	35.0 ± 3.2c	28.2 ± 2.0a	1.23 ± 0.21a	82.4 ± 5.3a	8.89 ± 0.31a	28.6 ± 0.4b
F-value	–	–	104.14	30.17	55.09	115.75	18.40	598.30	12.07	19.34
P-value	–	–	< 0.001	< 0.001	< 0.001	< 0.001	< 0.001	< 0.001	< 0.001	< 0.001
MP, mariculture pond; SOC, soil organic carbon; TN, total nitrogen; DOC, dissolved organic carbon; NH₄⁺, ammonium; NO₃^–, nitrate; WD, water depth; DO, dissolved oxygen. Different letters denote significant differences (P < 0.05) between groups based on ANOVA followed by Tukey’s HSD test. Values are means ± 1 standard errors during the entire year but not for SOC and TN.

It is noteworthy that the activities of animals, bivalve molluscs in particular, can also accelerate SOC decomposition in mariculture systems. Declines of quantity and quality of SOC have been documented in the no-feeding shellfish farms in a subtropical estuary (Ao River estuary, China)³⁵. Bioturbation and bioirrigation associated with the burrowing, feeding, and other activities of molluscs can mix, flush and suspend the sediments, resulting in oxygenation of sediments and elevated accessibility of organic matters to microorganisms, which favors the sediment respiration and decomposition of SOC (ref. ³⁶).

2.2 Impacts on CH₄ fluxes

Significant differences were also observed in the magnitude and seasonal pattern of CH₄ fluxes between S. alterniflora marsh and mariculture ponds (Fig. 1c). In the saltmarsh, daily CH₄ fluxes gradually increased from the start of the growth season in April, peaked on August, and then decreased to near zero during the wintertime. In the ponds, CH₄ fluxes were low at the beginning of the culturing period, then steadily increased and reached the first peak in July-August. From mid-August, CH₄ fluxes declined and were maintained low until draining. Drainage triggered the second peaks of CH₄ emission and the highest CH₄ flux was measured in MP1 (Fig. 1d). Annual CH₄ emission in the S. alterniflora marsh was 144 kg CH₄ ha^–1, 83.8% of which occurred in growth season. Converting saltmarsh to mariculture ponds dramatically decreased CH₄ emission by 80.3‒96.0%, with annual CH₄ emission in the ponds ranging from 5.77 to 28.3 kg CH₄ ha^–1. Similar to NEE, CH₄ emission exhibited a decreasing trend with culturing time, which was 3.91 times greater in MP1 than in MP3 (Fig. 2b). CH₄ emission during culturing period in the ponds was 3.93–19.2 kg CH₄ ha^–1, contributing 62.2–68.1% annual budget. While CH₄ emission in drainage period mainly occurred in the initial 10 days, contributing more than half of the emission (Figs. 1c and 2b).

Poffenbarger et al.³⁷ proposed a salinity threshold of > 18 ppt for coastal wetlands that CH₄ emission could be neglected, which has recently been adopted by IPCC (ref. ²⁰). However, our results suggested that vegetated coastal saltmarsh could still emit appreciable CH₄ albeit the high salinity (28.4 ppt). In a previous study, we found that 67.3% of CH₄ produced in S. alterniflora marsh originated from ‘noncompetitive’ trimethylamine³⁸, as methanogens are often outcompeted for acetate or H₂ by sulfate reduction in marine environments. These methylamines can be produced during the decomposition of the cytoplasmic osmolytes (e.g., glycine betaine, choline, and proline) of marine plants³⁹, supporting substantial CH₄ production in S. alterniflora marsh. We had expected that converting S. alterniflora marsh to mariculture ponds would increase CH₄ emission, as anthropogenic labile organic C inputs and continuous inundation may favor CH₄ production in the ponds. In contrast to our expectations, this conversion significantly reduced CH₄ emission (Fig. 2b), which could be related to the decreased total C inputs and substrate availability. In S. alterniflora marsh, annual gross C assimilation was 17.8 Mg C ha^‒1, which was 18 times higher than the C inputs from feeds and fertilizers in mariculture ponds (0.73 Mg C ha^‒1, Supplementary Table 1). Here, C inputs from primary production of phytoplankton and periphyton in the ponds were not quantified. However, previous study reported an annual net autotrophic C fixation of 182‒218 g C m^‒1 in an analogous crab-shrimp-clam polyculture system adjacent to our sites⁴⁰. Therefore, we consider that when primary productivity in the ponds was counted, C inputs for mariculture ponds were still much lower than that in S. alterniflora saltmarsh, which eventually decreased the supply of methanogenic substrates. We have speculated that, compared with plant residues, organic compounds in the feeds (e.g., starch and protein) can be more easily decomposed and become as substrates, especially “noncompetitive” methylamines^17,23. Indeed, CH₄ emission factor of applied/inputted C (EF_C) in test ponds (1.43%) was 2.35 times that in S. alterniflora marsh (0.61%), indicating that the inputted C could be more efficiently converted to CH₄ in mariculture ponds. However, the considerably lower CH₄ emissions in the mariculture ponds suggested that the benefit of higher C decomposability was completely offset by the loss of primary production.

More important, conversion of S. alterniflora saltmarsh to mariculture ponds significantly reduced SOC by 14.6‒29.7%, and especially DOC by 65.2‒74.1% (Table 1). Although the importance of recent photosynthates as source of CH₄ has been widely reported, SOC still accounted for a large fraction of CH₄ production in wetlands^41,42. For example, 40‒96% of emitted CH₄ from paddy fields originated from SOC (ref. ^42,43). Consistently, Kankaala and Bergström⁴⁴ found that the SOC-derived CH₄ contributed more than 90% to the total CH₄ emission from a shallow boreal lake. Particularly, CH₄ emission, SOC and DOC concentration in the ponds decreased with culturing age, and MP1 had the higher CH₄ emission and organic C contents relative to MP2 and MP3. Hence, SOC decomposition partly, at least, supported CH₄ production, since the application rates of feed and fertilizer were identical. Alongside with decreased substrate availability, changes in hydrological condition may also contribute to the reduction of CH₄ emission. Compared with the intermittently flooded saltmarsh, the ponds were permanently inundated during the culturing period (Supplementary Fig. 2). Generally, permanent inundation can create ideal anaerobic condition favoring CH₄ production⁴⁵. However, the standing water can also act as a diffusion barrier which may suppress CH₄ emission by increasing CH₄ oxidation during transport⁴⁶. Here, the maximal CH₄ fluxes in the ponds were observed in the early drainage period (Fig. 2d). Given the low temperature of this stage (early winter) and aerobic/semi-anaerobic condition, we consider that the emitted CH₄ was not currently produced but was a release of trapped CH₄ produced in culturing period. Inevitably, prolonged transport process increased the chance of CH₄ being oxidized. Nykänen et al.⁴⁷ found that up to 97.5% of CH₄ produced in a shallow small lake was oxidized before reaching the atmosphere. In contrast, CH₄ transport could proceed quickly in saltmarsh via aerenchyma tissue of S. alterniflora, which may also help bypass oxidation in soil zones⁴⁸.

2.3 Impacts on N₂O fluxes

In S. alterniflora saltmarsh, N₂O fluxes were at a low level and fluctuated within a narrow range between weak uptakes (‒24.1 µg N₂O m^‒2 h^‒1) and emissions (27.7 µg N₂O m^‒2 h^‒1), while uptake frequently occurred in summer and autumn (Fig. 1e). In mariculture ponds, basal manure application and the first supplemental fertilizing did not stimulate N₂O fluxes. After the second supplemental fertilizing, a peak of N₂O flux was observed, albeit the relatively lower rates. Similar to CH₄ fluxes, a burst of N₂O was observed immediately upon drainage (Fig. 1f). Annually, S. alterniflora marsh was a weak sink of atmospheric N₂O (‒0.04 kg N₂O ha^–1) and 68.3% of the N₂O uptakes occurred in growth season (Fig. 2c). However, converting saltmarsh to mariculture ponds produced a shift to N₂O source. Annual N₂O emission in the ponds ranged from 1.06 to 1.11 kg N₂O ha^–1. In spite of the shorter time and lower temperature, drainage period N₂O emission was higher than that in culturing period. Particularly, the initial 10 days of drainage period contributed 22.1‒31.4% annual N₂O budget (Fig. 2c).

Our annual N₂O budget in the S. alterniflora marsh was comparable to the reported atmospheric N₂O consumptions of 0.01‒0.14 kg N₂O ha^‒1 yr^‒1 in coastal saltmarshes which receive limited anthropogenic N loading^49,50. These data supported the view of productive and pristine saltmarshes as negligible sources or even small sinks of atmospheric N₂O (ref. ⁴⁹). However, converting S. alterniflora marsh to mariculture ponds fundamentally shifted the N₂O sink to a net source of N₂O. At least two explanations can be proposed. First, soil mineral N for nitrifiers and denitrifiers was considerably increased following the conversion (Table 1). Generally, N is the key limiting nutrient in pristine coastal wetlands. Here annual mean NO₃^‒ in the S. alterniflora marsh was only 0.14 mg N kg^‒1, a value that was lower than the threshold (5 mg N kg^‒1) for denitrification⁵¹ and was close to the threshold (~ 1 mg N kg^‒1) for denitrifiers using N₂O as the unique electron acceptor⁵². Hence, the lack of NO₃^‒ might be responsible for negative N₂O fluxes in S. alterniflora marsh. Conversely, mineral N concentrations in the ponds were a magnitude higher (Table 1). The mariculture ponds received higher N inputs from feeds and fertilizers while only a small part of the added N (~ 25%) was converted to fish biomass, leading to large amount of residual active N accumulated and facilitating N₂O production in the ponds^16,17. Second, Chen et al.⁵³ found that N₂O production via denitrification relies on NO₃^‒/DOC ratio in soils. The relatively lower DOC concentration and higher NO₃^‒/DOC ratio in ponds could increase the proportion of N₂O in denitrification products since high DOC content increases electron pressure, facilitating N₂O reduction and resulted in a lower ratio of N₂O/(N₂O + N₂) or even net N₂O consumption, especially in anaerobic environments⁵⁴. Senbayram et al.⁵⁵ pointed out that when soil NO₃^‒ concentrations fell below 20 mg kg^‒1, organic substrate addition would sharply reduce N₂O emission in anaerobic soils. In aquaculture systems, the lack of organic electron donors for N₂O reduction was also recognized as a major reason for N₂O emissions⁵⁶, whereas carbohydrate (starch) addition was noted to dramatically reduce N₂O emission by 83.4% in an intensive tilapia culturing system⁵⁷. Therefore, it is likely that the elevated NO₃^‒ and lowered DOC availability collectively increased N₂O emission via denitrification following the conversion of saltmarsh to mariculture ponds.

As well as with the change in magnitude, N₂O fluxes in mariculture ponds showed distinct temporal patterns from that in S. alterniflora marsh. Substantial N₂O fluxes were triggered by pond draining. This finding agreed well with previous studies that observed drainage episodes as the critical period for N₂O emission from aquaculture ponds^21,22,58. As discussed above, abundant soil NO₃^‒ usually generates a higher ratio of N₂O/(N₂ + N₂O) because NO₃^‒ is preferred over N₂O as an electron acceptor⁵⁹. The ratio also increases if some oxygen (O₂) is present, because the inhibition of N₂O reductase by O₂ is stronger than the other reductases involved in denitrification⁶⁰. On the other hand, the draining pond soil was still saturated initially and might cause optimum moisture condition (~ 80% water-filled pore space in soil) for a large proportion of N₂O produced during denitrification^58,59. Furthermore, with no standing water, N₂O could diffuse more easily and escape from consumption compared to the waterlogging condition. Hence, we consider these factors have collectively contributed to the N₂O emission pulse in the early stage of the drainage period.

2.4 Full GHG debt of mariculture ponds

Although aquaculture is often inferred as important sources of GHG to the atmosphere^61,62, only a few studies could comprehensively account for the GHG emissions associating with the conversion of natural coastal wetlands. Here, we incorporated the direct fluxes of CO₂, CH₄, and N₂O to evaluate their net global warming potential (nGWP) of each system. Annually, mariculture ponds exerted an overall positive net radiative forcing impact of 2.16 Mg CO₂eq ha^–1 over a 100-year time horizon, with CO₂, CH₄, and N₂O accounting for 68.9%, 18.0%, and 13.2% of the nGWP, respectively (Fig. 2d). Among the ponds, MP1 held up the highest nGWP (3.22 Mg CO₂eq ha^–1 yr^–1), primarily due to its higher NEE and CH₄ emissions relative to the older ponds. In contrast to mariculture ponds, S. alterniflora marsh showed a negative nGWP of ‒13.8 Mg CO₂eq ha^–1 yr^–1. This substantial negative value was mainly attributed to the high CO₂ uptake by plants, despite 22.6% of its climatic cooling impact was offset by the simultaneous CH₄ emission. Specifically, the negative nGWP of S. alterniflora marsh indicated that its conversion to mariculture ponds caused the loss of foregone GHG mitigation capacity by the saltmarsh⁶³, yielding a full GHG debt of 15.9 Mg CO₂eq ha^–1 yr^–1. This value was 7.36-fold higher than their direct GHG emissions from the mariculture ponds.

2.5 Comparison with other animal protein production sectors

We compared the direct GHG emissions from our mariculture ponds with those from other aquaculture systems in China, e.g., freshwater crab ponds (< 0.5 ppt) and brackish shrimp ponds (4.2 ppt) (ref. ^17,26,64). We found that, in spite of the differences in C and N inputs and operation scenarios, nGWP and yield-scaled GWP (GHGI) in aquaculture systems decreased significantly along the gradient of salinity (Fig. 3a,b). Our estimates of nGWP and GHGI in the polyhaline mariculture ponds (28.4 ppt) were just 8.51‒13.7% and 1.38‒14.9% of those in freshwater and brackish ponds. This difference was mainly attributed to the minor CH₄ emission in our ponds, given N₂O and CO₂ emissions were much comparable among three systems. Salinity can strongly impact CH₄ production by regulating the availability of sulfate. In contrast to polyhaline water ponds where CH₄ production was inhibited by presence of sulfate, lower salinity and sulfate availability in freshwater and brackish aquaculture ponds allowed intense CH₄ production from H₂/CO₂ and acetate, whereas these “competitive” substrates can be easily decomposed from residues of feeds and submerged macrophytes^17,23. This result substantiates the role of CH₄ emission in GHG budgets of aquaculture systems and suggests that mariculture at high salinity could have substantial lower direct GHG-cost relative to freshwater and brackish aquaculture due to their virtual low CH₄ emissions.

We also compared the full GHG debt in aquaculture systems with the estimates from other animal protein production sectors on a kg CO₂eq kg^‒1 protein basis (GHGI-protein) (Fig. 3c). Similar to nGWP and GHGI, the direct GHGI-protein in our mariculture ponds (6.48 kg CO₂eq kg^‒1 protein) was significantly lower than that in freshwater crab (185 kg CO₂eq kg^‒1 protein) and brackish shrimp (14.7 kg CO₂eq kg^‒1 protein) ponds, but was higher than that in a no-feeding oyster farm (0.13 kg CO₂eq kg^‒1 protein) (ref. ⁶⁵). Notably, we found the GHGI-protein in aquaculture systems were much lower than those from terrestrial livestock production (39.5‒466 kg CO₂eq kg^‒1 protein), except for freshwater crab ponds. However, when the loss of foregone GHG mitigation capacity of pristine wetlands was counted, GHGI-protein in our mariculture ponds would increase to 46.7 CO₂eq kg^‒1 protein, which is higher than poultry but still lower than beef, small ruminants, or pork production. These results imply that mariculture can provide a low-GHG animal protein source relative to freshwater aquaculture and livestock production. Developing approaches to attenuate net emissions associated with the foregone GHG mitigation loss will be crucial for minimizing the climate impact of mariculture. Future aquaculture programs could preferentially engineer aquaculture systems by conversion of lowly productive sites (e.g., saline bare land, abandoned ponds, etc.) and carrying out it in high-salinity waters^30,31, as reduced foregone C sequestration losses and CH₄ emissions could be simultaneously achieved under such conditions.

Overall, our results show that the conversion of saltmarsh to mariculture ponds in coastal region of China creates a GHG debt of 15.9 Mg CO₂eq ha^–1 yr^–1. Within them, the loss of foregone GHG mitigation capacity of pristine wetlands makes the largest contributions (86.4%) to overall GHG debt, whereas the other 15.6% arises from the direct emission of CO₂, CH₄, and N₂O in the mariculture ponds. This finding contributes to growing recognition of the climatic roles of mariculture and highlights the importance of taking the foregone C sequestration losses associated with coast wetlands conversion into the full GHG accounting. Nonetheless, the GHG-cost of protein in mariculture was much lower than inland freshwater aquaculture and the major terrestrial livestock productions (such as beef, small ruminants, and pork), indicating mariculture can provide a low-GHG animal protein source. For initiatives aiming at reduction of GHG emissions, future aquaculture should primarily avoid conversion of the highly productive coastal wetlands, and further minimize CH₄ emissions by using high-salinity waters.

3.1. Site locations

The study site is located at the coast of Yellow Sea, Yancheng City, Jiangsu Province, China (33°23'N, 120°43'E). This region is characterized by a subtropical monsoon climate with the long-term (1981–2010) mean annual air temperature of 14.3°C and precipitation of 1067 mm. Tides in this region are typically semidiurnal with amplitude of 2–3 m and tidewater salinity of 30.0‒32.0‰. Native plant Suaeda salsa used to be the dominant species in this site until S. alterniflora was intentionally introduced in 1982. Since then, exotic S. alterniflora gradually replaced the native plants and established monoculture communities in the intertidal zone⁶⁶. Expansion of aquaculture in this region initiated in early 1980s and represented the primary causes of coastal wetlands loss in recent years. During 2000‒2017, Yancheng City lost 608 km² natural coastal wetlands and more than 35% of them were converted to aquaculture ponds⁶⁷.

3.2. Experimental design and field measurement

Field experiments were carried out throughout April 2016 to May 2017 in a S. alterniflora marsh and three adjacent saltmarsh-converted mariculture ponds (Supplementary Fig. 1). These ponds were converted from the S. alterniflora marsh in 2010, 2011, and 2014, respectively, which represented a conversion chronosequence of 1- (MP1), 4- (MP2), and 5-year (MP3) at beginning of the experiment. Generally, aquaculture period in this region usually starts in April and ends in November. In line with the local aquaculture practices, we adopted the similar aquaculture mode and operating scenario in the experimental ponds. Briefly, common prawn (Mercenaria mercenaria)-crab (Portunus trituberculatus)-clam (Exopalaemon carinicauda) polyculture was employed in these ponds. Crabs and prawn were fed with commercial feed pellets and trash fish until they were harvested, and fertilizers were also applied to stimulate phytoplankton and clam production (Supplementary Tables 1 and 2). Annually, total inputs of C and N were 725 kg C ha^–1 and 268 kg N ha^–1, respectively. The water regime was also in line with local practice. Prior to clams stocking, sea water was introduced into ponds from an adjacent tidal creek. In the end of the culturing period, these ponds were drained to facilitate harvesting. Crab and prawn were harvested in late September, followed by clam harvesting. Detailed management practices in these mariculture ponds are shown in Supplementary Table 2 and elaborated in Supplementary Text 1 (Methods-Extensive).

3.3. Measurement of greenhouse gas fluxes

Eddy covariance (EC) method was used to quantify the ecosystem-scale CH₄ and CO₂ (net ecosystem CO₂ exchange, NEE) fluxes from S. alterniflora marsh. CO₂ and CH₄ fluxes were computed as the covariance between fluctuations in gas mixing ratio and vertical wind velocity over 30 min periods. We gap-filled the missing NEE using the marginal distribution sampling algorithm and further partitioned it into gross primary productivity (GPP) and ecosystem respiration (Re). GPP and Re were all presented with positive signs (NEE = Re ‒ GPP). For CH₄ fluxes, we used the random forest (RF) to fill the data gaps with meteorological variables. Details regarding the EC flux and ancillary measurements, data processing, gap filling are elaborated in Supplementary Text 1. As EC N₂O flux measurements are limited by the detection accuracy of the instrumentation, especially for natural ecosystems where N₂O fluxes are generally weak⁶⁸, here we measured N₂O flux in S. alterniflora marsh using the static closed chamber technique. Four gas sampling plots were randomly established in S. alterniflora marsh ~ 20 m away from the EC tower, and N₂O flux was measured using the transparent Plexiglass chambers.

The static closed chamber method was also used to measure CO₂, CH₄ and N₂O fluxes in mariculture ponds. Four subsites were randomly selected in each pond to minimize the spatial uncertainties of GHG emissions. GHG fluxes were also measured by the transparent chambers, and thus the measured CO₂ fluxes in mariculture ponds were recognized as NEE (ref. ⁵⁰). The chamber deployment, sampling procedures and configuration of the GC for measuring GHG fluxes are extensively included in Supplementary Text 1.

3.4. Annual GHG budgets and radioactive forcing calculation

Annual CO₂ and CH₄ budgets in S. alterniflora marsh were calculated by integrating the gap-filled and partitioned fluxes over the course of a full year. For N₂O emission in S. alterniflora saltmarsh and CO₂, CH₄, and N₂O emissions in mariculture ponds, the annual budgets were calculated by linear interpolation between measuring days. The net radiative forcing arising from annual CO₂ (NEE), CH₄, and N₂O budgets was calculated using the net global warming potential (nGWP, kg CO₂eq ha^‒1 year^‒1) methodology⁶⁹:

where CH₄, N₂O and CO₂ denote annual emissions of CH₄ (kg CH₄ ha^–1), N₂O (kg N₂O ha^–1) and NEE (kg CO₂ ha^–1), respectively.

Thereafter, the greenhouse gas intensity (GHGI, kg CO₂eq kg^–1 yield) and protein-based GHGI (GHGI-protein, kg CO₂eq kg^–1 protein) in mariculture ponds was calculated by dividing nGWP by yield biomass or protein¹⁷. Besides, the CH₄ emission factor of applied C (EF_C) and N₂O emission factor of applied N (EF_N) in the ponds was calculated by dividing annual CH₄ or N₂O emissions by total C or N inputs, respectively.

3.5. Environmental data

Micrometeorological instrumentation was deployed in the S. alterniflora saltmarsh to accompany EC measurements. Air temperature, precipitation, soil temperature at 10 cm depth, and water depth were synchronously measured. Salinity of the tide water was measured in a tidal creek about 60 m southeast of the tower. In mariculture ponds, water depth, water temperature (or soil temperature in drainage period), salinity, and dissolved O₂ of the pond water were measured simultaneously with gas sampling.

The 0‒10 cm soil samples were collected weekly from the saltmarsh and mariculture ponds for mineral N and DOC measuring. SOC and total N (TN) contents were determined by the wet-oxidation method and the Kjeldahl procedure, respectively. The procedures for auxiliary measurements were also detailed in Supplementary Text 1.

3.6. Statistical analysis

Statistical analyses were performed using R version 3.2.2 unless otherwise stated. Data normality was examined using the Kolmogorov-Smirnov test, and natural-log transformation was used if necessary. One-way analysis of variance (ANOVA) with a Tukey’s HSD test was used for comparison of soil properties cross sites and annual GHG emissions among ponds.

Data availability

All data needed to evaluate the conclusions in the paper are present in the paper and/or the Supplementary Materials. Source data for all CO₂, CH₄, and N₂O fluxes and environmental data are provided with this paper.

Kirwan, M. L. & Megonigal, J. P. Tidal wetland stability in the face of human impacts and sea-level rise. Nature 504, 53‒60 (2013).
Mcleod, E. et al. A blueprint for blue carbon: toward an improved understanding of the role of vegetated coastal habitats in sequestering CO₂. Fron. Ecol. Environ. 9, 552‒560 (2011).
Duarte, C. M., Losada, I. J., Hendriks, I. E., Mazarrasa, I. & Marbà, N. The role of coastal plant communities for climate change mitigation and adaptation. Nat. Clim. Change 3, 961–968 (2013).
Spivak, A. C., Sanderman, J., Bowen, J. L., Canuel, E. A. & Hopkinson, C.S. Global-change controls on soil-carbon accumulation and loss in coastal vegetated ecosystems. Nat. Geosci. 12, 685‒692 (2019).
Livesley, S. J. & Andrusiak, S. M. Temperate mangrove and salt marsh sediments are a small methane and nitrous oxide source but important carbon store. Estuar. Coast. Shelf S. 97, 19‒27 (2012).
Lovelock, C. E. et al. The vulnerability of Indo-Pacific mangrove forests to sea-level rise. Nature 526, 559‒563 (2015).
Ma, Z. et al. Rethinking China’s new great wall. Science, 346, 912‒914 (2014).
Richards, D. R. & Friess, D. A. Rates and drivers of mangrove deforestation in Southeast Asia, 2000–2012. Proc. Natl. Acad. Sci. USA. 113, 344‒349 (2016).
Barbier, E. B. et al. The value of estuarine and coastal ecosystem services. Ecol. Monogr. 81, 169‒193 (2011).
Murdiyarso, D. et al. The potential of Indonesian mangrove forests for global climate change mitigation. Nat. Clim. Change 5, 1089‒1092 (2015).
Ren, C. et al. Rapid expansion of coastal aquaculture ponds in China from Landsat observations during 1984–2016. Int. J. Appl. Earth Obs. Geoinform. 82, 101902 (2019).
FAO. The State of World Fisheries and Aquaculture 2020. Sustainability in action. (FAO Fisheries and Aquaculture Department, 2020).
FAO. The State of World Fisheries and Aquaculture 2014. Opportunities and challenges. (FAO Fisheries and Aquaculture Department, 2014).
Pendleton, L. et al. Estimating global “blue carbon” emissions from conversion and degradation of vegetated coastal ecosystems. Plos One 7, e43542 (2012).
Kauffman, J. B. et al. The jumbo carbon footprint of a shrimp: carbon losses from mangrove deforestation. Fron. Ecol. Environ. 15, 183‒188 (2017).
Hargreaves, J. A. Nitrogen biogeochemistry of aquaculture ponds. Aquaculture 166, 181‒212 (1998).
Yuan, J. et al. Rapid growth in greenhouse gas emissions from the adoption of industrial-scale aquaculture. Nat. Clim. Change 9, 318–322 (2019).
Yang, P. et al. Effects of coastal marsh conversion to shrimp aquaculture ponds on CH₄ and N₂O emissions. Estuar. Coast. Shelf S. 199, 125‒131 (2017).
Hiraishi, T. (eds) 2013 Supplement to the 2006 IPCC Guidelines for National Greenhouse Gas Inventories: Wetlands (IPCC, 2014).
Buendia, E. C. et al. (eds) 2019 Refinement to the 2006 IPCC Guidelines for National Greenhouse Gas Inventories—Volume 4 Agriculture, Forestry and Other Land Use (IPCC, 2019).
Liu, S. et al. Methane and nitrous oxide emissions reduced following conversion of rice paddies to inland crab-fish aquaculture in Southeast China. Environ. Sci. Technol. 50, 633‒642 (2016).
Yang, P. et al. Spatial variations of N₂O fluxes across the water-air interface of mariculture ponds in a subtropical estuary in Southeast China. J. Geophys. Res.-Biogeo. 125, e2019JG005605 (2020).
Yuan, J. et al. Methane and nitrous oxide have separated production zones and distinct emission pathways in freshwater aquaculture ponds. Water Res. 190, 116739 (2021).
Verdegem, M. C. J. & Bosma, R. H. Water withdrawal for brackish and inland aquaculture, and options to produce more fish in ponds with present water use. Water Policy 11, 52‒68 (2009).
Zhang, D. et al. Carbon dioxide fluxes from two typical mariculture polyculture systems in coastal China. Aquaculture 521, 735041 (2020).
Yang, P. et al. Insights into the farming-season carbon budget of coastal earthen aquaculture ponds in southeastern China. Agr. Ecosyst. Environ. 335, 107995 (2022).
Almeida, R. M. et al. High primary production contrasts with intense carbon emission in a eutrophic tropical reservoir. Fron. Microbiol. 7, 717 (2016).
Boyd, C. E., Wood, C. W., Chaney, P. L. & Queiroz, J. F. Role of aquaculture pond sediments in sequestration of annual global carbon emissions. Environ. Poll. 158, 2537‒2540 (2010).
Arias-Ortiz, A. et al. Losses of soil organic carbon with deforestation in mangroves of Madagascar. Ecosystems 24, 1‒19 (2021).
Kauffman, J. B., Heider, C., Norfolk, J. & Payton, F. Carbon stocks of intact mangroves and carbon emissions arising from their conversion in the Dominican Republic. Ecol. Appl. 24, 518‒527 (2014).
Arifanti, V. B., Kauffman, J. B., Hadriyanto, D., Murdiyarso, D. & Diana, R. Carbon dynamics and land use carbon footprints in mangrove-converted aquaculture: The case of the Mahakam Delta, Indonesia. Forest Ecol. Manag. 432, 17‒29 (2019).
Kosten, S. et al. Better assessments of greenhouse gas emissions from global fish ponds needed to adequately evaluate aquaculture footprint. Sci. Total Environ. 748, 141247 (2020).
Barros, N. et al. Carbon emission from hydroelectric reservoirs linked to reservoir age and latitude. Nat. Geosci. 4, 593‒596 (2011).
Zhang, J., Song, C. & Wenyan, Y. Tillage effects on soil carbon fractions in the Sanjiang Plain, Northeast China. Soil Till. Res. 93, 102‒108 (2007).
Lei, J., Yang, L. & Zhu, Z. Testing the effects of coastal culture on particulate organic matter using absorption and fluorescence spectroscopy. J. Clean. Prod. 325, 129203 (2021).
Norkko, J. & Shumway, S. E. in Shellfish Aquaculture and the Environment (ed Shumway, S. E.) pp. 297–317 (Wiley-Blackwell, 2011).
Poffenbarger, H. J., Needelman, B. A. & Megonigal, J. P. Salinity influence on methane emissions from tidal marshes. Wetlands 31, 831‒842 (2011).
Yuan, J. et al. Spartina alterniflora invasion drastically increases methane production potential by shifting methanogenesis from hydrogenotrophic to methylotrophic pathway in a coastal marsh. J. Ecol. 107, 2436‒2450 (2019).
Wang, X. C. & Lee, C. Sources and distribution of aliphatic amines in salt marsh sediment. Org. Geochem. 22, 1005‒1021 (1994).
Zhang, K., Tian, X., Dong, S., Feng, J. & He, R. An experimental study on the budget of organic carbon in polyculture systems of swimming crab with white shrimp and short-necked clam. Aquaculture 451, 58‒64 (2016).
Ström, L., Ekberg, A., Mastepanov, M. & Christensen, T. R. The effect of vascular plants on carbon turnover and methane emissions from a tundra wetland. Global Change Biol. 9, 1185‒1192 (2003).
Tokida, T. et al. Falling atmospheric pressure as a trigger for methane ebullition from peatland. Global Biogeochem. Cy. 21, GB2003 (2007).
Minoda, T., Kimura, M. & Wada, E. Photosynthates as dominant source of CH₄ and CO₂ in soil water and CH₄ emitted to the atmosphere from paddy fields. J. Geophys. Res.-Atmos. 101, 21091‒21097 (1996).
Kankaala, P. & Bergström, I. Emission and oxidation of methane in Equisetum fluviatile stands growing on organic sediment and sand bottoms. Biogeochemistry 67, 21‒37 (2004).
Altor, A. E. & Mitsch, W. J. 2008. Methane and carbon dioxide dynamics in wetland mesocosms: effects of hydrology and soils. Ecol. Appl. 18, 1307‒1320 (2008).
Bastviken, D., Cole, J., Pace, M. & Tranvik, L. Methane emissions from lakes: Dependence of lake characteristics, two regional assessments, and a global estimate. Global Biogeochem. Cy. 18, GB4009 (2004).
Nykänen, H., Peura, S., Kankaala, P. & Jones, R. Recycling and fluxes of carbon gases in a stratified boreal lake following experimental carbon addition. Biogeosciences 11, 16447‒16495 (2014).
Ding, W., Zhang, Y. & Cai, Z. Impact of permanent inundation on methane emissions from a Spartina alterniflora coastal salt marsh. Atmos. Environ. 44, 3894‒3900 (2010).
Moseman-Valtierra, S. et al. Short-term nitrogen additions can shift a coastal wetland from a sink to a source of N₂O. Atmos. Environ. 45, 4390‒4397 (2011).
Yuan, J. et al. Exotic Spartina alterniflora invasion alters ecosystem-atmosphere exchange of CH₄ and N₂O and carbon sequestration in a coastal salt marsh in China. Global Change Biol. 21, 1567‒1580 (2015).
Dobbie, K. E. & Smith, K. A. Impact of different forms of N fertilizer on N₂O emissions from intensive grassland. Nutr. Cycl. Agroecosys. 67, 37–46 (2003).
Chapuis-Lardy, L., Wrage, N., Metay, A., Chotte, J.-L. & Bernoux, M. Soils, a sink for N₂O? A review. Global Change Biol. 13, 1‒17 (2007).
Chen, Z. et al. Increased N₂O emissions during soil drying after waterlogging and spring thaw in a record wet year. Soil Biol. Biochem. 101, 152‒164 (2016).
Miller, M. N. et al. Crop residue influence on denitrification, N₂O emissions and denitrifier community abundance in soil. Soil Biol. Biochem. 40, 2553‒2562 (2008).
Senbayram, M., Chen, R., Budai, A., Bakken, L. & Dittert, K. N₂O emission and the N₂O/(N₂O+N₂) product ratio of denitrification as controlled by available carbon substrates and nitrate concentrations. Agr. Ecosyst. Environ. 147, 4‒12 (2012).
Avnimelech, Y. Carbon/nitrogen ratio as a control element in aquaculture systems. Aquaculture 176, 227‒235 (1999).
Hu, Z. et al. Influence of carbohydrate addition on nitrogen transformations and greenhouse gas emissions of intensive aquaculture system. Sci. Total Environ. 470, 193‒200 (2014).
Ma, Y. et al. A comparison of methane and nitrous oxide emissions from inland mixed-fish and crab aquaculture ponds. Sci. Total Environ. 637‒638, 517‒523 (2018).
Murray, R. H., Erler, D. V. & Eyre, B. D. Nitrous oxide fluxes in estuarine environments: response to global change. Global Change Biol. 21, 3219‒3245 (2015).
Richardson, D., Felgate, H., Watmough, N., Thomson, A. & Baggs, E. Mitigating release of the potent greenhouse gas N₂O from the nitrogen cycle-could enzymic regulation hold the key? Trends in Biotechnol. 27, 388‒397 (2009).
Hu, Z., Lee, J.W., Chandran, K., Kim, S. & Khanal, S. K. Nitrous oxide (N₂O) emission from aquaculture: a review. Environ. Sci. Technol. 46, 6470‒6480 (2012).
Williams, J. & Crutzen, P. Nitrous oxide from aquaculture. Nat. Geosci. 3, 14 (2010).
Gelfand, I. et al. Carbon debt of Conservation Reserve Program (CRP) grasslands converted to bioenergy production. Proc. Natl. Acad. Sci. USA. 108, 13864‒13869 (2011).
Yang, P. et al. Assessing nutrient budgets and environmental impacts of coastal land-based aquaculture system in southeastern China. Agr. Ecosyst. Environ. 322, 107662 (2021).
Ray, N. E., Maguire, T. J., Al-Haj, A. N., Henning, M. C. & Fulweiler, R. W. Low greenhouse gas emissions from oyster aquaculture. Environ. Sci. Technol. 53, 9118‒9127 (2019).
Liu, C. et al. Habitat changes for breeding waterbirds in Yancheng National Nature Reserve, China: A remote sensing study. Wetlands 30, 879‒888 (2010).
Wang, G. et al. Research on driving forces of reclamation dynamics and land use change in coastal areas of Jiangsu Province. J. Hohai U. (Nat. Sci.), 48, 163‒170 (2020) (in Chinese).
Kroon, P. S. et al. Uncertainties in eddy covariance flux measurements assessed from CH₄ and N₂O observations. Agr. Forest Meteorol. 150, 806–816 (2010).
Stocker, T. F. et al. (eds) Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change (Cambridge University Press, Cambridge, 2013).

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Foregone carbon sequestration loss dominates greenhouse gas budget in mariculture associated with coastal wetland conversion

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Version 1

Abstract

Figures

1. Introduction

2. Results and discussion

2.1 Impacts on CO₂ fluxes

2.2 Impacts on CH₄ fluxes

2.3 Impacts on N₂O fluxes

2.4 Full GHG debt of mariculture ponds

2.5 Comparison with other animal protein production sectors

3. Methods

3.1. Site locations

3.2. Experimental design and field measurement

3.3. Measurement of greenhouse gas fluxes

3.4. Annual GHG budgets and radioactive forcing calculation

3.5. Environmental data

3.6. Statistical analysis

Declarations

Data availability

References

Additional Declarations

Supplementary Files

Status:

Version 1

Foregone carbon sequestration loss dominates greenhouse gas budget in mariculture associated with coastal wetland conversion

Status:

Version 1

Abstract

Figures

1. Introduction

2. Results and discussion

2.1 Impacts on CO2 fluxes

2.2 Impacts on CH4 fluxes

2.3 Impacts on N2O fluxes

2.4 Full GHG debt of mariculture ponds

2.5 Comparison with other animal protein production sectors

3. Methods

3.1. Site locations

3.2. Experimental design and field measurement

3.3. Measurement of greenhouse gas fluxes

3.4. Annual GHG budgets and radioactive forcing calculation

3.5. Environmental data

3.6. Statistical analysis

Declarations

Data availability

References

Additional Declarations

Supplementary Files

Status:

Version 1

2.1 Impacts on CO₂ fluxes

2.2 Impacts on CH₄ fluxes

2.3 Impacts on N₂O fluxes