Introduction

Increasing anthropogenic terrestrial nitrogen (N) inputs and sewage pollution is regarded as one of the major cause for N enrichment in coastal wetlands. Coastal wetlands, estuaries and adjoining mangrove fringes which are in close proximity to human settlements can act as important sinks of terrestrial N where it can be stored and cycled. It is reported that 1% of reactive nitrogen entering estuaries are released as nitrous oxide (N2O) emissions1. N2O is a major contributor to stratospheric ozone depletion, has a lifetime of 118–131 years in the atmosphere2 and has about 300 times greater global warming potential than carbon dioxide (CO2)3. N2O is produced as an intermediate product during the denitrification process (heterotrophic NO3 reduction) and also as a by-product during the nitrification process (autotrophic oxidation of NH4 to NO3) and it is consumed during complete denitrification of N2O via dissimilatory reduction to N24. Therefore, the ratio of production and consumption of N2O determines whether these coastal wetlands are sinks or sources of N2O. However, the production and consumption of N2O in a system is determined by several drivers such as temperature, pH, dissolved oxygen (DO), carbon (C) and hydrologic parameters in the environment4.

Mangroves increase the ecosystem value of coastal wetlands through their significant role in nutrient cycling, moderation of extreme weather events, supply of food and raw materials, and provision of breeding, spawning and nursery habitat for a variety of invertebrates and commercial fish species. However, the effectiveness of mangroves to meet these services is affected by the degree of anthropogenic disturbance. Recent studies have shown that tropical mangrove wetlands are under high risk from anthropogenic stresses particularly due to the extensive use of terrestrial N fertilizers causing dissolved inorganic nitrogen (DIN) enrichment5,6. Major mangrove ecosystems around the globe have been assessed for their contribution to atmospheric N2O budgets. However, despite our knowledge about the contribution of mangrove wetlands in the global N2O budget, uncertainties still exist due to lack of information about the impact of fragmented mangrove wetlands which are numerous in the tropical regions and particularly along the Indian coastline. As tropical areas receive disproportionately more rainfall than temperate areas, this likely leads to greater terrestrial N into coastal wetlands from nearby urban and agricultural areas. The high seasonal variability in rainfall distribution patterns in tropical areas may also drive large temporally variability in coastal N2O dynamics.

Along the Indian coast (7,516.6 km long7), mangrove cover an area of 497,500 ha8. The India state Kerala in southern India, was once rich in mangrove habitats, but the area of intact mangrove has drastically declined due to anthropogenic pressures such as urbanisation, burgeoning power projects and rapid industrialisation. It is estimated that the Ernakulum district in Kerala has 943 mangrove stands of which 880 (~ 93%) of them have an area less than 1 hectare. The contribution of these smaller mangrove systems is often ignored and their role in global nutrient cycling has yet to be adequately determined. As most of these mangrove stands are generally observed near the vicinity of human settlements, they can be critically influenced by the N loading there. Recent global studies suggest that aquatic environments influenced by intensive agricultural practices and industrial activities can act as source of N2O to the atmosphere9,10,11 and this predicted rise in terrestrial N concentrations can potentially triple the global N2O emissions from coastal wetlands1. However, Indian mangrove stands have been mostly ignored from the N input and output studies due to their relatively smaller area and footprint. The few studies that have taken place in Indian mangroves suggest that they are highly influenced by anthropogenic nutrient inputs and are significant sources of atmospheric N2O12,13,14. In light of the increasing anthropogenic N inputs, it is important to monitor and assess the potential of fragmented/minor mangrove stands for their contribution of N2O to the atmosphere.

In order to understand the contribution of fragmented/minor mangrove stands to atmospheric N2O emissions, this study aimed to assess the Mangalavanam Coastal Wetland (MCW) a fragmented minor mangrove stand adjacent to the city of Kochi, India over a one year period incorporating a range of distinct seasons. The N2O flux rates were calculated using three different k models15,16,17 to reduce the uncertainty in calculated N2O flux rates. The study also tested the relation between N substrates and N2O fluxes with temperature, pH, DO, and C loading of the wetland to determine the drivers of N2O dynamics.

Study area

The MCW is a semi-closed fragmented mangrove stand in the Ernakulam district along the south-west coast of India at 9° 59′ 23.83″ N and 76° 16′ 26.74″ E (Fig. 1). It receives tidal input through a narrow channel/feeder canal from the adjacent nutrient rich Cochin Estuarine System (CES), where eutrophication is of great concern18. This tidal wetland has a wet and maritime tropical climate with three distinct seasons: pre-monsoon (February to May), southwest monsoon (June to September, from here on called monsoon) and northeast monsoon (October to January) or post-monsoon. The wetland is relatively shallow (< 0.5 m) and is largely inundated during high tide flooding. The dominant mangrove species present in the MCW are Rhizophora mucronata, Avicennia officinalis and Acanthus ilicifolius. Mangrove associates such as Acrostichum aureum, Derris trifoliate and Morinda citrifolia are also common. Kochi is the most densely inhabited city in Kerala with an urban population of more than of 2.1 million in an area of 440 km2. The MCW receives nutrient input from the urban drainage systems of the adjacent human settlements. Due to government restrictions on sampling in the area, only two sites could be sampled, a site heavily influenced by upstream estuarine inputs (Site 1) and a relatively pristine site (Site 2).

Figure 1
figure 1

Map of Mangalavanam Coastal Wetland, Kerala, India depicting the study stations [source: Google earth pro V 7.3. (November 12, 2019). Kochi, Kerala. 9° 59′ 23.83″ N, 76° 16′ 26.74″ E, Eye alt 7.6 km. https://earth.google.com/web/search/Mangalavanam+Bird+Sanctuary,+High+Court+Road,+Ayyappankavu,+Kochi,+Kerala,+India/@9.9893564,76.2747626,4.86064735a,2322.96914379d,35y,240.09539944h,45t,0r/data=Cr0BGpIBEosBCiUweDNiMDgwZDU5Mjk0ZWQxZDU6MHhlMjcwYmIzMGIwZGY3OGYwGTUtDOyM-iNAIX5K37WVEVNAKlBNYW5nYWxhdmFuYW0gQmlyZCBTYW5jdHVhcnksIEhpZ2ggQ291cnQgUm9hZCwgQXl5YXBwYW5rYXZ1LCBLb2NoaSwgS2VyYWxhLCBJbmRpYRgCIAEiJgokCf0tW8UnikNAEZyxjdwYtjXAGXdPXswSvVxAISlSbeKAkktA].

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Results

Physico-chemical attributes

The area received 3469 mm of rainfall over the study period, with 68.5% falling during the monsoon season (2376 mm) (Fig. 2, Table 1). The mean wind speed of the study area was 2.4 ± 0.4 m s−1, with monthly averages ranging from 2.0 m s−1 in April to 3.2 m s−1 in June. The wetland depth was generally shallow, with a mean depth of 0.3 ± 0.8 m and ranged from 0.1 m in February to 0.5 m in June. The mean annual water and sediment temperature was 27.0 ± 0.9 °C and 27.4 ± 0.8 °C, respectively. The mean mixo-mesohaline to mixo-polyhaline salinity was 12.3 ± 7.2 ‰, with significant differences between seasons (f = 107.20, ρ ˂ 0.001). The mean DO concentrations in the water column was 112.2 ± 45.6 µmol L−1 and decreased from the post-monsoon season (156.1 ± 16.0 µmol L−1) to the pre-monsoon season (109.0 ± 30.1 µmol L−1), and to the monsoon season (71.6 ± 39.8 µmol L−1). The annual mean DO saturation was 229.2 ± 11.2% and AOU was 117.0 ± 45.7 µmol L−1 with both varying significantly between seasons (f = 32.00, ρ ˂ 0.001 and f = 16.36, ρ ˂ 0.001, respectively). The pH of the water column over the year ranged from 6.5 to 7.6 with a mean of 6.8 ± 0.3, while in the sediments it ranged from 6.1 to 6.7. There was a significant difference in pH in the water column between seasons (f = 9.07, ρ = 0.002) and in sediment pH between stations (f = 4.72, ρ = 0.043).

Figure 2
figure 2

Variation in environmental parameters observed in the Mangalavanam Coastal Wetland, India during the study.

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Table 1 Variation in selected environmental parameters observed in the Mangalavanam Coastal Wetland, India during the study.
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Carbon and nitrogen measurements

The total nitrogen (TN) concentration in the water column ranged from 172.1 to 877.1 µmol L−1, with an annual mean of 399.4 ± 244.8 µmol L−1 and in the sediments it ranged from 158.6 to 301.4 µmol g−1 (Table 1). The mean TN in the water column was highest during the monsoon season and the lowest during post-monsoon season, while in sediments the maximum concentration was observed during the post-monsoon season and the minimum during the pre-monsoon season. Dissolved nitrogen (DN) contributed 66 to 88% of the observed TN with an annual mean of 322.1 ± 182.0 µmol L−1 and varied significantly on temporal basis (f = 9.16, ρ = 0.002). Dissolved inorganic nitrogen constituted 12–41% of the DN concentrations with an annual mean of 60.7 ± 18.1 µmol L−1, while DON was 58–88% of the DN concentration with an annual mean of 248.6 ± 168.9 µmol L−1. The DIN and DON concentrations in the water column varied significantly on seasonal basis (f = 17.15, ρ ˂ 0.001 and f = 8.26, ρ = 0.003, respectively) with higher DIN and DON in the monsoon season. NO3 was the major constituent of DIN, with a mean of 62 ± 8% of DIN. NO3 was followed by NH4 and NO2, constituting about 36 ± 8% and 2 ± 1%, of the DIN respectively. There was a significant difference in water column NH4 (f = 8.89, ρ = 0.002), NO2 (f = 15.47, ρ < 0.001) and NO3 (f = 11.89, ρ = 0.001) seasonally. Particulate nitrogen (PN) in the water column ranged from 20.1 µmol L−1 to 232.9 µmol L−1 during the study period and varied significantly between seasons (f = 8.76, ρ = 0.002). The total carbon (TC) concentration in the water column ranged from 2473.3 to 7391.7 µmol L−1 (average 3916.3 ± 1453.0 µmol L−1) with total organic carbon (TOC) contributing 33 ± 10% of the observed TC concentrations (range from 484.2 to 3792.5 µmol L−1) which varied significantly on seasonal basis (f = 4.63, ρ = 0.02). In the sediment, TOC concentrations ranged between 2877.5 µmol g−1 to 9911.7 µmol g−1 over the study period. In the water column, the mean TOC concentration was highest during the monsoon season and lowest during the post-monsoon season, while in sediments, the maximum concentration was observed during the post-monsoon season (6230.3 ± 2292.4 µmol g−1) and the minimum concentration was observed during the pre-monsoon season (4000.8 ± 1323.8 µmol g−1). The total inorganic carbon was the major constituent of TC in the water column, with an annual mean of 2556.4 ± 771.6 µmol L−1, ranged from 1578.3 µmol L−1 in July to 4605.0 µmol L−1 in November and varied significantly between seasons (f = 7.57, ρ = 0.003).

Nitrous oxide dynamics

The dissolved N2O concentration in the water column of the MCW ranged from 50.3 nM in November to 131.5 nM in June (Fig. 3, Table 1). The mean percentage saturation of N2O was 1088 ± 252% and ranged from 696 to 1794%, with relatively little seasonal variations. Using Wanninkhof15 k values, N2O fluxes ranged from 17.5 to 117.9 μmol m−2 day−1, with an annual mean of 39.1 ± 22.6 μmol m−2 day−1. For comparisons sake, the flux calculations of Raymond and Cole17, N2O fluxes ranged from 47.7 to 207.9 μmol m−2 day−1, with an annual mean of 88.6 ± 33.6 μmol m−2 day−1 and Wanninkhof and McGillis16 N2O fluxes ranged from 3.4 to 34.3 μmol m−2 day−1, with an annual mean of 9.26 ± 7.27 μmol m−2 day−1. Using Wanninkhof15 and Wanninkhof and McGillis16, the N2O flux varied significantly on seasonal basis (f = 6.53, ρ = 0.007 and f = 7.55, ρ = 0.004, respectively), while no significant variation was observed with the N2O flux that was calculated using Raymond and Cole17 calculated fluxes (f = 3.43, ρ = 0.055). The N2O flux calculated using the three models was highest during the monsoon season and lowest during the post-monsoon season.

Figure 3
figure 3

Variation in N2O dynamics observed in the Mangalavanam Coastal Wetland, India during the study.

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Discussion

This study found that the MCW fragmented mangrove stand is a C and N rich environment which favours substantial N2O production and its release into the atmosphere. The N pool and the atmospheric flux of N2O in the MCW (39.1 ± 22.6 μmol m−2 day−1) was found to be higher than that reported in the adjacent CES (11.4 ± 6.9 μmol m−2 day−1) where there are significant anthropogenic N inputs19. In the Bhitarkanika mangrove on the east coast of India12, high N2O emissions were observed (5–113.38 μmol m−2 day−1) driven by high anthropogenic nutrient loading. In other Indian mangroves–Muthupet mangrove in southern India, comparatively low N2O emissions (9.3–18.18 μmol m−2 day−1) were reported even though the mangrove is subjected to a range of anthropogenic inputs including aquaculture (shrimp-farming effluent) and seasonal agricultural run-off13. Our observations were also in the range of previous studies from the large eutrophic Pearl River Estuary in China (37 ± 15 μmol m−2 day−1) which receives extensive urban, industrial and agricultural run-off10.

In a range of pristine mangroves in Australia20, the uptake of N2O has been reported with an uptake rate of 1.52 ± 0.17 μmol m−2 day−1. However, the relative pristine mangrove site selected in our study (Site 2) was supersaturated with dissolved N2O (1122.8 ± 266.9%) and acted as atmospheric source of N2O (41.0 ± 26.1 μmol m−2 day−1). A recent study on N2O dynamics from four southern hemisphere subtropical estuaries11 (an urbanised estuary, mixed, urban and agricultural estuary and pristine estuary) reported that the N2O saturation ranged from 77.7 to 381.5% with highest saturations in the pristine estuary and lowest in the agricultural estuary, which was much lower than the range seen in this study (696.7–1794.3%). The study also reported that estuaries in the two urbanised catchments had the highest mean N2O flux rates during summer (18.8 ± 11.6 μmol m−2 day−1 and 18.7 ± 12.8 μmol m−2 day−1, respectively) with the pristine estuary having a summer flux rate of 5.1 ± 6.9 μmol m−2 day−111. Other studies from estuaries in urbanised catchments have reported low mean N2O flux rates, with 4.0 μmol m−2 day−1 reported in the Yarra River in Melbourne21, and 7.3 μmol m−2 day−1 in the Brisbane River22. The high flux rates observed here are likely a result of high rainfall in the tropics driving higher freshwater inputs, terrestrial runoff and the enhanced anthropogenic N enrichment compared to other estuaries cited above. The pristine site (Site 2) also received tidal inputs from the nutrient enriched CES which would lead to higher N2O concentrations in the water column compared to Site 1 that was heavily influenced by upstream estuarine inputs. Also, the water column of Site 1 drains out completely into the CES during low tides; however the water column depth of Site 2 was less affected during low tides which likely reduced the export of N to adjacent waters.

Significant uncertainties exist in global N2O budget due in part to the differences in methods used in calculating flux rates. Two meta‐analyses published in the past 20 years highlight that the lack of direct measurements of k significantly hampers the ability to parameterize air‐water gas exchange in estuaries and that care should be given when choosing values of k, with respect to location‐dependent controls on gas exchange17,23. Here we offer a comparison of the different flux rates calculated using three different and commonly used k value models (Table 1) and the relationship of various measured parameters to N2O flux rates calculated using three models (Table 2). The parameterization of Raymond and Cole17 is widely used for rivers and streams, where the turbulence reaching the air‐water interface is chiefly generated by friction with the bottom. The parameterization of Wanninkhof and McGillis16 is best used for open water bodies but can underestimate N2O flux rates where wind speeds are low such as sheltered mangrove systems. The k value model of Wanninkhof15 is also commonly used and uses the Schmidt number of the water related to viscosity and with its exponent reflecting the surface layer’s rate of turbulent renewal. As water column depth, wind speed, and current velocity act as main drivers of these k models and influences turbulence at the air–water interface, in shallow flowing ecosystems24,25, such as in mangroves, we felt the k value model of Wanninkhof15 best fits for the study.

Table 2 Relationship of N2O flux calculated using different models to other environmental parameters observed in the Mangalavanam Coastal Wetland, India during the study.
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The study observed a relationship between N2O dynamics and the availability of nitrogen species (Table 2). Concentrations of TN, DN, DON, NO3, NH4 and NO2 in the study area were high and at the high end of the range of concentrations seen in other highly polluted estuaries26. MCW is located in the heart of the densely populated Kochi city and has several sewage inlets from surrounding human settlements. This significantly increases the availability of N in both sediments and the water column leading to higher N2O production in the study area. Recent studies have shown that sewage influxes increase N2O production19,22. The release of nitrogenous compounds through the excreta of water-birds can also cause N enrichment in mangroves26 and the abundant water bird population of MCW would also lead to higher nitrogen loads in the study sites.

The MCW also receives significant N inputs through the narrow channel from CES that receives substantial fresh water discharge (approx. 1.2 × 1010 m3 year−1) from the six major rivers in the area (Pamba, Manimala, Achancoil, Meenchil, Periyar and Muvattupuzha Rivers) during the monsoon season27,28 and also receives inputs of approximately 260 m3 of domestic and 104 × 103 m3 of industrial waste discharges daily, most of which are unprocessed29,30. The CES (and subsequently the MCW) receives high inputs of N from the surrounding agricultural areas which are flushed into receiving waters (both from surface runoff and groundwater inputs). There were significant relationships between rainfall distribution pattern and NH4 (r = 0.415, ρ = 0.044), NO2 (r = 0.489, ρ = 0.015), NO3 (r = 0.755, ρ = 0.000) concentrations and N2O flux calculated using Wanninkhof15 (Table 2). Similar observations of increased N loads and associated N2O fluxes due to rainfall have been reported in the eutrophic Taihu Lake in China31. The high resident time of the water column in the study lead to greater accumulation of N in the water column and longer processing times of N before being exported which likely influence N2O production and flux rates in the study area.

High N and organic carbon concentrations have been shown to accelerate the nitrification–denitrification processes31,32,33 and lead to higher N2O production although C and N concentrations did not show a clear relationship to dissolved N2O concentration in the MCW. Non-significant relationships between dissolved N2O concentration and DIN have been observed in other recent studies34,35,36. However, during this study significant relationships were observed between TC, TIC, TN, DN, DIN, NO3, NH4, DON, PN and N2O flux calculated using Wanninkhof15 (Table 2). This indicates a consistent relationship existed between N, C inputs and N2O outputs. Significant relationships between DIN inputs and N2O fluxes have also been shown in the nearby CES in India19, rivers37 and estuaries globally38. Positive relationships were also observed between N2O and NH4 and NO3 + NO2 during the summer dry season, while during winter, N2O saturation was strongly correlated to NO3 + NO2 but not with NH411.

The significant positive correlation between NO3 and NH4 concentrations to N2O flux calculated using Wanninkhof15 and strong negative correlation between DO and NH4 (r = − 0.558, ρ = 0.005), NO2 (r = − 0.586, ρ = 0.003) and NO3 (r = − 0.788, ρ ˂ 0.001) suggest high N2O flux rates produced through nitrification. However, lower DO during the monsoon months along with higher availability of NO3 in the presence of high organic carbon loading likely promotes denitrification increasing N2O production and its fluxes. Several studies suggest that estuaries receiving higher N inputs generally show oxygen depletion, which in turn triggers denitrification34,39,40. Increasing anthropogenic nitrogen inputs to the mangrove sediments along with periodic tidal flooding can generate anoxic conditions in the sediment; thereby further enhancing denitrification41. This process along with nitrification could significantly contribute to nutrient turnover and N2O production in mangroves. However, the magnitude at which a wetland can act as source of N2O cannot be explained solely based on N loads as the magnitude of N2O flux depends on both denitrification-nitrification rates (how efficiently N is cycled) and N2O reduction rates (how efficiently N2O is reduced to N2)31.

The highest dissolved N2O concentrations and saturations were observed during the pre-monsoon season when temperatures were highest, while the N2O flux rates were highest during monsoon season. However, there was no significant relationship between water column and sediment temperatures and N2O concentrations, saturations and flux rates. Recent studies suggest that higher temperatures during summer seasons (pre-monsoon) can enhance microbial activity favouring N2O production9, while the reduced freshwater discharge during summers can increase the resident time of the water column, leading to greater accumulation of N2O in water and reducing the N2O loss to the atmosphere due to the decrease in gas transfer velocity42.

The shallow depth profile of the MCW likely promoted easier diffusion of oxygen molecules into the water column and to the sediments, favouring nitrification. This can drive higher dissolved N2O concentrations and its saturation during the pre-monsoon than the monsoon and post-monsoon seasons. A recent report suggests that the shallow bathymetry of the CES favoured easier exchange of N2O that was produced in the estuarine sediment and water column to the atmosphere19. A positive correlation between water column depth and N2O flux was observed using all models (Table 2). The deepest water column depth was mainly observed during monsoon season when the C and N loads were highest; there was low DO saturation and wind speeds were highest. However the mean water column depth during the monsoon season remained shallow (less than 0.5 m) even though it had the highest range in depths (0.3–0.5 m). Combined with the higher wind speeds during the monsoon season, this resulted in higher N2O fluxes rates. The significant negative correlation between salinity and N2O flux calculated using Wanninkhof15 was likely driven by either nutrient rich seawater returned from the CES or the reduced solubility of gases in saline conditions.

During the study, DON was the major constituent of the N and signified that the N enrichment in the study area was mainly through organic matter inputs or oceanic inputs where DON is the most prevalent form of N. As the major source of organic carbon in the study area was likely from mangrove litter which is generally nitrogen deficient43,44, the positive correlation of TC to DON in the water column (r = 0.870, ρ < 0.001) suggests that the study area receives significant amount of organic inputs from allochthonous sources, particularly with high N concentrations. The significant positive correlation between TC and TIC in the water column and N2O flux indicates that the N2O production and higher fluxes occurs in N rich environments with regard to C availability. However, the study failed to explain the relationship of TOC to N2O production and fluxes. A recent study in the tropical estuaries of north-western Borneo45 also failed to explain the significant correlation between DOC and N2O concentrations. Although temperature and pH are important factors affecting N2O production, no clear relationship was found between these variables and dissolved N2O concentrations and fluxes. Several studies also suggest that the influence of water temperature on the denitrification rate may vary between regions and seasons31,46,47.

As the MCW is a fragmented mangrove, it’s atmospheric contribution of N2O calculated using Wanninkhof15 is quite small (ranging from 1.1 ± 0.2 × 106 μmol day−1 during post-monsoon season to 2.4 ± 1.2 × 106 μmol day−1 during monsoon season) and on its own may not be significant. However considering the numerous smaller mangrove stands that exist all over the tropics, fragmented mangroves may represent a significant source of N2O to the atmosphere. For example, the Ernakulam district in the state Kerala alone has about 933 minor mangrove stands, covering a total area of 206 hectares. This suggests that fragmented/minor mangrove stands may have an as yet unquantified but significant influence on atmospheric N2O budgets and atmospheric warming into the future.

Conclusion

The net N2O flux from MCW suggest that it is a source of atmospheric N2O, however due to its small area of coverage its contribution to atmospheric N2O is minor. The high precipitation rates in the study, through terrestrial runoff and river water discharge and high N inputs, influenced N2O flux rates particularly during monsoon season. The study suggests that when taken as whole, fragmented/minor mangroves which are abundant in tropical regions, need to be critically assessed and protected from further anthropogenic loading. Further studies in a range of different geologic and hydrologic conditions will help to include this potentially significant ecosystem type in global N2O budgets. The study highlights that N2O flux rates were dependant on the availability of DIN as well as salinity, DO, estuarine depth, rainfall, wind speed and availability of carbon.

Materials and methods

Field work was carried out over a period of 12 months from June 2014 to May 2015 with observations made during the morning hours on a monthly basis. Rainfall data was obtained from the Indian Meteorological Department. The depth of the wetland was measured by lowering a graduated weighted rope until it touched the top of the sediments. Transparency was measured using a 20 cm diameter Secchi disc48. Triplicates of nutrient samples were collected from surface waters at each site. Temperature, salinity and pH were measured using an Eutech water quality analyser (CyberScan Series SCD 650). Water samples for DO were taken in 60 ml glass bottles fixed with 0.5 ml each of Winkler reagents and titrated against sodium thiosulphate using visual endpoint detection49. Apparent oxygen utilisation (AOU) was calculated as outlined by Garcia and Gordon50, using the measured DO concentration. Dissolved inorganic nitrogen (DIN, which includes NH4, NO2, and NO3) were analysed spectrophotometrically following standard procedures51.

Water samples for total organic carbon (TOC) and total inorganic carbon (TIC), total nitrogen (TN), dissolved nitrogen (DN) and dissolved N2O were collected in 120 ml glass bottles. The samples were fixed using saturated mercuric chloride (0.6 ml/120 ml) to arrest microbial activity52. An aliquot of each sample was filtered through a 25 mm GF/F filter and the filtrate was collected for the measurement of DN, while the remainder of the sample was analysed for TOC, TIC and TN by wet combustion method using a TOC elemental analyser (Multi N/C 2100 S Analytik jena). Particulate nitrogen (PN) and dissolved organic nitrogen (DON) concentrations in the water column were calculated by subtracting DN from TN and DIN from DN, respectively.

Sediment samples were obtained using a van-Veen grab sampler (area 0.04 m2) with a glass corer (3 cm diameter) inserted into each grab sample. As the surface sediment contained mangrove litter, samples were collected at a depth of 2 cm from the surface of the sediment and sieved (usually < 2 mm), dried at 60 °C for 24–72 h and ground to a fine powder. An aliquot of the dried sediment sample was acidified using 1 M hydrochloric acid to remove the inorganic carbon present in the sample53. The acidified samples were then washed with distilled water, dried and ground to powder. The samples were then analysed for TOC using dry combustion method (TOC elemental analyser Multi N/C 2100 S, Analytik jena). The other part of the sample that was not treated with acid was used to measure TN in the sediment using Pyro-cube IRMS.

Dissolved N2O was determined by the multiple phase equilibration technique54.

The N2O water to air exchange fluxes (f) were computed using:

$$f = {\text{ k}{\alpha}}_{{}} \left( {{\text{C}}_{{\text{w}}} - {\text{ C}}_{{\text{a}}} } \right),$$
(1)

where k is the gas transfer coefficient of N2O23,55, α is the solubility coefficient of N2O calculated using temperature and salinity56, Cw is the water column N2O partial pressure and Ca is the local measured atmospheric value for N2O (323 ppb as recently reported)57. To provide a comparison of different k values, the wind parameterization k models of Wanninkhof15, Wanninkhof and McGillis16 and Raymond and Cole17 were calculated. Wind speed data (10 m height) were obtained from ERA interim, European Centre for Medium Range Weather Forecast (ECMWF) with a data resolution of ~ 80 km from the study site.

SPSS v16 (Statistical Programme for Social Sciences v16) was used for Pearson’s correlation analyses and the two-way analysis of variance (ANOVA) and Origin Pro 8.5 used to plot the graphs.