Search
2023 Volume 2
Article Contents
ARTICLE   Open Access    

Greenhouse gas emissions from managed freshwater wetlands under intensified aquaculture

More Information
  • Wetlands are important niches for aquatic organisms and biodiversity as well as for organic carbon sequestration. However, wetlands are frequently managed for aquaculture production, which jeopardized their ability to act as a natural carbon sink. In this experiment, greenhouse gases (GHG) including CO2, CH4 and N2O fluxes at the water-air interface, were monitored on a daily and monthly basis in four aquaculture wetlands: a natural wetland (NW), an enclosure wetland for intensive aquaculture (EWIA), a constructed wetland made from reclaimed paddy land for intensive aquaculture (CWIA), and a constructed wetland made from reclaimed paddy land for extensive aquaculture (CWEA). As a result, annual GHG fluxes for CWEA, CWIA, EWIA and NW wetlands averaged 0.81, 1.06, 2.43 and 1.12 kg N2O-N ha−1, 23.83, 457.08, 1,360.27 and 38.29 kg CH4-C ha−1, 1,321.32, 1,877.04, 2,246.79 and 1,305.48 kg CO2-C ha−1, respectively. Also, the highest yearly global warming potential (GWP) for EWIA, which was 60.02 t CO2-eq ha−1, was roughly ten times greater than that of NW, which was 6.04 t CO2-eq ha−1. Furthermore, the GHG emission intensity per unit of production ranged from 1.21 (CWEA) −5.30 (EWIA) kg CO2-eq kg−1, with EWIA having the highest N2O and CH4 emission factors per unit of aquaculture output at 0.34 g N2O kg−1 and 0.19 kg CH4 kg−1. The study also revealed that feed quantity and water parameters positively linked with GHG emissions, suggesting that improving human management and water environmental quality could achieve large mitigation for both natural and artificial aquaculture wetlands. Given the high heterogeneity of GHGs emissions in aquaculture wetlands, future research should conduct long-term continuous automatic monitoring of aquaculture wetlands to understand the patterns of GHGs emissions and their responses and mechanisms to human management.
  • 加载中
  • Allen D, Dalal RC, Rennenberg H, Schmidt S. 2011. Seasonal variation in nitrous oxide and methane emissions from subtropical estuary and coastal mangrove sediments, Australia. Plant Biology 13(1):126−33

    doi: 10.1111/j.1438-8677.2010.00331.x

    CrossRef    Google Scholar

    Altor AE, Mitsch WJ. 2006. Methane flux from created riparian marshes: relationship to intermittent versus continuous inundation and emergent macrophytes. Ecological Engineering 28:224−34

    doi: 10.1016/j.ecoleng.2006.06.006

    CrossRef    Google Scholar

    Aufdenkampe AK, Mayorga E, Raymond PA, Melack JM, Doney SC, et al. 2011. Riverine coupling of biogeochemical cycles between land, oceans, and atmosphere. Frontiers in Ecology and the Environment 9:53−60

    doi: 10.1890/100014

    CrossRef    Google Scholar

    Barnes J, Owens NJN. 1999. Denitrification and nitrous oxide concentrations in the Humber Estuary, UK, and adjacent coastal zones. Marine Pollution Bulletin 37:247−60

    doi: 10.1016/S0025-326X(99)00079-X

    CrossRef    Google Scholar

    Bastviken D, Tranvik LJ, Downing JA, Crill PM, Enrich-Prast A. 2011. Freshwater methane emissions offset the continental carbon sink. Science 331:50

    doi: 10.1126/science.1196808

    CrossRef    Google Scholar

    Bastviken D, Cole JJ, Pace ML, van de Bogert MC. 2008. Fates of methane from different lake habitats: connecting whole-lake budgets and CH4 emissions. Journal of Geophysical Research: Biogeosciences 113(G2):G02024

    doi: 10.1029/2007JG000608

    CrossRef    Google Scholar

    Beaulieu JJ, Tank JL, Hamilton SK, Wollheim WM, Hall RO Jr, et al. 2011. Nitrous oxide emission from denitrification in stream and river networks. Proceedings of the National Academy of Sciences of the United States of America 108(1):214−19

    doi: 10.1073/pnas.1011464108

    CrossRef    Google Scholar

    Bhattacharyya P, Sinhababu DP, Roy KS, Dash PK, Sahu PK, et al. 2013. Effect of fish species on methane and nitrous oxide emission in relation to soil C, N pools and enzymatic activities in rainfed shallow lowland rice-fish farming system. Agriculture Ecosystems & Environment 176:53−62

    doi: 10.1016/j.agee.2013.05.015

    CrossRef    Google Scholar

    Chen H, Zhu Q, Peng C, Wu N, Wang Y, et al. 2013. Methane emissions from rice paddies, natural wetlands, lakes in China: synthesis new estimate. Global Change Biology 19:19−32

    doi: 10.1111/gcb.12034

    CrossRef    Google Scholar

    Konnerup D, Betancourt-Portela JM, Villamil C, Parra JP. 2014. Nitrous oxide and methane emissions from the restored mangrove ecosystem of the Ciénaga Grande de Santa Marta, Colombia. Estuarine, Coastal and Shelf Science 140:43−51

    doi: 10.1016/j.ecss.2014.01.006

    CrossRef    Google Scholar

    Fang X, Wang C, Zhang T, Zheng F, Zhao J, et al. 2022. Ebullitive CH4 flux and its mitigation potential by aeration in freshwater aquaculture: Measurements and global data synthesis. Agriculture, Ecosystems & Environment 335:108016

    doi: 10.1016/j.agee.2022.108016

    CrossRef    Google Scholar

    Fourqurean JW, Duarte CM, Kennedy H, Marbà N, Holmer M, et al. 2012. Seagrass ecosystems as a globally significant carbon stock. Nature Geoscience 5(7):505−9

    doi: 10.1038/ngeo1477

    CrossRef    Google Scholar

    Friborg T, Christensen TR, Hansen BU, Nordstroem C, Soegaard H. 2000. Trace gas exchange in a high-arctic valley: 2. Landscape CH4 fluxes measured and modeled using eddycorrelation data. Global Biogeochemical Cycles 14(3):715−23

    doi: 10.1029/1999GB001136

    CrossRef    Google Scholar

    Gao J, Zheng X, Wang R, Liao T, Zou J. 2014. Preliminary comparison of the static floating chamber and the diffusion model methods for measuring water−atmosphere exchanges of methane and nitrous oxide from inland water bodies. Climatic and Environmental Research 19(3):290−302

    doi: 10.3878/j.issn.1006-9585.2012.12173

    CrossRef    Google Scholar

    Gorham E. 1991. Northern peatlands; role in the carbon cycle and probable responses to climatic warming. Ecological Applications 1:182−95

    doi: 10.2307/1941811

    CrossRef    Google Scholar

    Han Y, Zheng Y, Wu R, Yin J, Xu J et al. 2013. Greenhouse gases emission characteristics of Nanjing typical waters in spring. China Environmental Science 33(8):1360−71

    Google Scholar

    Hashimoto S, Gojo K, Hikota S, Sendai N, Otsuki A. 1999. Nitrous oxide emissions from coastal waters in Tokyo Bay. Marine Environmental Research 47(3):213−23

    doi: 10.1016/S0141-1136(98)00118-4

    CrossRef    Google Scholar

    Healy RW, Striegl RG, Russell TF, Hutchinson GL, Livingston GP. 1996. Numerical evaluation of static-chamber measurements of soil—atmosphere gas exchange: identification of physical processes. Soil Science Society of America Journal 60:740−47

    doi: 10.2136/sssaj1996.03615995006000030009x

    CrossRef    Google Scholar

    Hou C, Song C, Li Y, Wang J, Song Y, et al. 2013. Effects of water table changes on soil CO2, CH4 and N2O fluxes during the growing season in freshwater marsh of Northeast China. Environmental Earth Sciences 69:1963−71

    Google Scholar

    Hirota M, Tang Y, Hu Q, Hirata S, Kato T, et al. 2004. Methane emissions from different vegetation zones in a Qinghai-Tibetan Plateau wetland. Soil Biology Biochemistry 36:737−48

    doi: 10.1016/j.soilbio.2003.12.009

    CrossRef    Google Scholar

    Hu Z, Lee JW, Chandran K, Kim S, Sharma K, et al. 2014. Influence of carbohydrate addition on nitrogen transformations and greenhouse gas emissions of intensive aquaculture system. Science of the Total Environment 470−471:193−200

    doi: 10.1016/j.scitotenv.2013.09.050

    CrossRef    Google Scholar

    Hu Z, Lee JW, Chandran K, Kim S, Sharma K, et al. 2013. Nitrogen transformations in intensive aquaculture system and its implication to climate change through nitrous oxide emission. Bioresource Technology 130:314−20

    doi: 10.1016/j.biortech.2012.12.033

    CrossRef    Google Scholar

    Hu Z, Lee JW, Chandran K, Kim S, Khanal SK. 2012. Nitrous oxide (N2O) emission from aquaculture: a review. Environmental Science & Technology 46(12):6470−80

    doi: 10.1021/es300110x

    CrossRef    Google Scholar

    Hu Z, Wu S, Ji C, Zou J, Zhou Q, et al. 2016. A comparison of methane emissions following rice paddies conversion to crab-fish farming wetlands in southeast China. Environmental Science and Pollution Research 23(2):1505−15

    doi: 10.1007/s11356-015-5383-9

    CrossRef    Google Scholar

    Intergovernmental Panel on Climate Change (IPCC). 2006. 2006 IPCC Guidelines for National Greenhouse Gas Inventories. Methodology Report. Institute for Global Environmental Strategies (IGES), Japan. www.ipcc.ch/report/2006-ipcc-guidelines-for-national-greenhouse-gas-inventories/

    Google Scholar

    Intergovernmental Panel on Climate Change (IPCC). 2013. Supplement to the 2006 IPCC guidelines for national greenhouse gas inventories: Wetlands. Methodology Report. Switzerland. Georgia: Batumi, 2014. www.ipcc.ch/site/assets/uploads/2018/03/Wetlands_Supplement_Entire_Report.pdf

    Google Scholar

    Jackowicz-Korczyński M, Christensen TR, Bäckstrand K, Crill P, Friborg T, et al. 2010. Annual cycle of methane emission from a subarctic peatland. Journal of Geophysical Research: Biogeosciences 115(G2):02009−10

    doi: 10.1029/2008JG000913

    CrossRef    Google Scholar

    Jenkinson DS, Adams DE, Wild A. 1991. Model estimates of CO2 emissions from soil in response to global warming. Nature 351:304−6

    doi: 10.1038/351304a0

    CrossRef    Google Scholar

    Kayranli B, Scholz M, Mustafa A, Hedmark Å. 2010. Carbon storage and fluxes within freshwater wetlands: a critical review. Wetlands 30:111−24

    doi: 10.1007/s13157-009-0003-4

    CrossRef    Google Scholar

    Liu S, Hu Z, Wu S, Li S, Li Z, et al. 2016. Methane and nitrous oxide emissions reduced following conversion of rice paddies to inland crab-fish aquaculture in southeast China. Environmental Science & Technology 50(2):633−42

    doi: 10.1021/acs.est.5b04343

    CrossRef    Google Scholar

    Lu R. 2000. Soil and agro-chemistry analytical methods. Beijing: China Agricultural Science and Technology Press.

    Google Scholar

    Mitsch WJ, Bernal B, Nahlik AM, Mander Ü, Zhang L, et al. 2013. Wetlands, carbon, and climate change. Landscape Ecology 28:583−97

    doi: 10.1007/s10980-012-9758-8

    CrossRef    Google Scholar

    Nahlik AM, Mitsch WJ. 2011. Methane emissions from tropical freshwater wetlands located in different climatic zones of Costa Rica. Global Change Biology 17(3):1321−34

    doi: 10.1111/j.1365-2486.2010.02190.x

    CrossRef    Google Scholar

    Pendleton L, Donato DC, Murray BC, Crooks S, Jenkins WA, et al. 2012. Estimating global 'Blue Carbon' emissions from conversion and degradation of vegetated coastal ecosystems. PLoS One 7:e43542

    doi: 10.1371/journal.pone.0043542

    CrossRef    Google Scholar

    Petrescu AMR, Lohila A, Tuovinen JP, Baldocchi DD, Desai AR, et al. 2015. The uncertain climate footprint of wetlands under human pressure. Proceedings of the National Academy of Sciences of the United States of America 112(15):4594−99

    doi: 10.1073/pnas.1416267112

    CrossRef    Google Scholar

    Sayre R. 2010. Microalgae: the potential for carbon capture. Bioscience, 60(9):722−27

    doi: 10.1525/bio.2010.60.9.9

    CrossRef    Google Scholar

    Seitzinger SP, Kroeze C. 1998. Global distribution of nitrous oxide production and N inputs in freshwater and coastal marine ecosystems. Global Biogeochemical Cycles 12(1):93−113

    doi: 10.1029/97GB03657

    CrossRef    Google Scholar

    Seitzinger SP, Kroeze C, Styles RV. 2000. Global distribution of N2O emissions from aquatic systems: natural emissions and anthropogenic effects. Chemosphere-Global Change Science 2:267−279

    doi: 10.1016/S1465-9972(00)00015-5

    CrossRef    Google Scholar

    Panneer Selvam B, Natchimuthu S, Arunachalam L, Bastviken D. 2014. Methane and carbon dioxide emissions from inland waters in India − implications for large scale greenhouse gas balances. Global Change Biology 20(11):3397−407

    doi: 10.1111/gcb.12575

    CrossRef    Google Scholar

    Smith LK, Lewis WMJ, Chanton JP, Cronin G, Hamilton S. 2000. Methane emissions from the Orinoco River floodplain, Venezuela. Biogeochemistry 51:113−40

    doi: 10.1023/A:1006443429909

    CrossRef    Google Scholar

    Stief P, Poulsen M, Nielsen LP, Brix H, Schramm A. 2009. Nitrous oxide emission by aquatic macrofauna. Proceedings of the National Academy of Sciences of the United States of America 106(11):4296−300

    doi: 10.1073/pnas.0808228106

    CrossRef    Google Scholar

    Tan L, Ge Z, Zhou X, Li S, Li X, et al. 2020. Conversion of coastal wetlands, riparian wetlands, and peatlands increases greenhouse gas emissions: a global meta-analysis. Global Change Biology 26(3):1638−53

    doi: 10.1111/gcb.14933

    CrossRef    Google Scholar

    Tsuneda S, Mikami M, Kimochi Y, Hirata A. 2005. Effect of salinity on nitrous oxide emission in the biological nitrogen removal process for industrial wastewater. Journal of Hazardous Materials 119(1−3):93−98

    doi: 10.1016/j.jhazmat.2004.10.025

    CrossRef    Google Scholar

    Wang DQ, Chen ZL, Xu SY. 2009. Methane emission from Yangtze estuarine wetland, China. Journal of Geophysical Research: Biogeosciences 114:G02011

    doi: 10.1029/2008JG000857

    CrossRef    Google Scholar

    Yang P, Lai DYF, Yang H, Lin Y, Tong C, et al. 2022a. Large increase in CH4 emission following conversion of coastal marsh to aquaculture ponds caused by changing gas transport pathways. Water Research 222:118882

    doi: 10.1016/j.watres.2022.118882

    CrossRef    Google Scholar

    Yang P, Tang KW, Tong C, Lai DYF, Zhang L, et al. 2022b. Conversion of coastal wetland to aquaculture ponds decreased N2O emission: Evidence from a multi-year field study. Water Research 227:119326

    doi: 10.1016/j.watres.2022.119326

    CrossRef    Google Scholar

    Yang P, Tang KW, Yang H, Tong C, Zhang L, et al. 2023. Contrasting effects of aeration on methane (CH4) and nitrous oxide (N2O) emissions from subtropical aquaculture ponds and implications for global warming mitigation. Journal of Hydrology 617:128876

    doi: 10.1016/j.jhydrol.2022.128876

    CrossRef    Google Scholar

    Yang J, Liu J, Hu X, Li X, Wang Y, et al. 2013. Effect of water table level on CO2, CH4 and N2O emissions in a freshwater marsh of Northeast China. Soil Biology and Biochemistry 61:52−60

    doi: 10.1016/j.soilbio.2013.02.009

    CrossRef    Google Scholar

    Yang P, He Q, Huang J, Tong C. 2015. Fluxes of greenhouse gases at two different aquaculture ponds in the coastal zone of southeastern China. Atmospheric Environment 115:269−77

    doi: 10.1016/j.atmosenv.2015.05.067

    CrossRef    Google Scholar

    Zhang Y, Tang KW, Yang P, Yang H, Tong C, et al. 2022. Assessing carbon greenhouse gas emissions from aquaculture in China based on aquaculture system types, species, environmental conditions and management practices. Agriculture, Ecosystems & Environment 338:108110

    doi: 10.1016/j.agee.2022.108110

    CrossRef    Google Scholar

    Zhang Y, Li C, Trettin CC, Li H, Sun G. 2002. An integrated model of soil, hydrology, and vegetation for carbon dynamics in wetland ecosystems. Global Biogeochemical Cycles 16(4):1061

    doi: 10.1029/2001GB001838

    CrossRef    Google Scholar

    Zheng J, Pan G, Cheng K, Zhang X. 2014. A discussion on quantification of greenhouse gas emissions from wetlands based on 2013 supplement to the 2006 IPCC guidelines for national greenhouse gas inventories: Wetlands. Advances in Earth Science 29(10):1120−25

    doi: 10.11867/j.issn.1001-8166.2014.10.1120

    CrossRef    Google Scholar

    Zou J. 2005. A study on greenhouse gases (CO2, CH4 and N2O) emissions from rice-winter wheat rotations in southeast China. China: Nanjing Agricultural University Press

    Google Scholar

  • Cite this article

    Yue Q, Cheng K, Sheng J, Wang L, Ji C, et al. 2023. Greenhouse gas emissions from managed freshwater wetlands under intensified aquaculture. Soil Science and Environment 2:3 doi: 10.48130/SSE-2023-0003
    Yue Q, Cheng K, Sheng J, Wang L, Ji C, et al. 2023. Greenhouse gas emissions from managed freshwater wetlands under intensified aquaculture. Soil Science and Environment 2:3 doi: 10.48130/SSE-2023-0003

Figures(4)  /  Tables(6)

Article Metrics

Article views(3198) PDF downloads(610)

ARTICLE   Open Access    

Greenhouse gas emissions from managed freshwater wetlands under intensified aquaculture

Soil Science and Environment  2 Article number: 3  (2023)  |  Cite this article

Abstract: Wetlands are important niches for aquatic organisms and biodiversity as well as for organic carbon sequestration. However, wetlands are frequently managed for aquaculture production, which jeopardized their ability to act as a natural carbon sink. In this experiment, greenhouse gases (GHG) including CO2, CH4 and N2O fluxes at the water-air interface, were monitored on a daily and monthly basis in four aquaculture wetlands: a natural wetland (NW), an enclosure wetland for intensive aquaculture (EWIA), a constructed wetland made from reclaimed paddy land for intensive aquaculture (CWIA), and a constructed wetland made from reclaimed paddy land for extensive aquaculture (CWEA). As a result, annual GHG fluxes for CWEA, CWIA, EWIA and NW wetlands averaged 0.81, 1.06, 2.43 and 1.12 kg N2O-N ha−1, 23.83, 457.08, 1,360.27 and 38.29 kg CH4-C ha−1, 1,321.32, 1,877.04, 2,246.79 and 1,305.48 kg CO2-C ha−1, respectively. Also, the highest yearly global warming potential (GWP) for EWIA, which was 60.02 t CO2-eq ha−1, was roughly ten times greater than that of NW, which was 6.04 t CO2-eq ha−1. Furthermore, the GHG emission intensity per unit of production ranged from 1.21 (CWEA) −5.30 (EWIA) kg CO2-eq kg−1, with EWIA having the highest N2O and CH4 emission factors per unit of aquaculture output at 0.34 g N2O kg−1 and 0.19 kg CH4 kg−1. The study also revealed that feed quantity and water parameters positively linked with GHG emissions, suggesting that improving human management and water environmental quality could achieve large mitigation for both natural and artificial aquaculture wetlands. Given the high heterogeneity of GHGs emissions in aquaculture wetlands, future research should conduct long-term continuous automatic monitoring of aquaculture wetlands to understand the patterns of GHGs emissions and their responses and mechanisms to human management.

    • Wetlands are critical terrestrial ecosystems that contain a huge global carbon (C) store (Gorham, 1991; Fourqurean et al., 2012; Mitsch et al., 2013) and may exert a significant impact on climate change (Zhang et al., 2002). Peat-accumulating wetlands currently hold approximately 390−455 Pg of terrestrial C, or nearly one-third of the world’s soil C stock (Jenkinson et al., 1991). C capture and sequestration within wetlands is a complicated process as it involves both aerobic and anaerobic processes (Kayranli et al., 2010). Additionally, a variety of substrates could potentially be used to capture and sequester CO2, such as microorganism, algae (Sayre, 2010). In contrast, wetlands also have a major part to play in C cycling, with high GHG emission, estimated at 0.65 Pg C per year in the form of CH4 (Bastviken et al., 2011), totalling up to 2.1 Pg C per year as CO2 (Aufdenkampe et al., 2011). Therefore, maintaining wetlands as a large carbon sequestration site is an essential way to balance the whole terrestrial ecosystem.

      Wetland reclamation under human factors would generate many additional anthropogenic GHG emissions and removals, of which aquaculture could be a typical artificially managed freshwater wetland system (Pendleton et al., 2012; Tan et al., 2020; Yang et al., 2022b). Yang et al. (2022a) reported a large increase in CH4 emission following conversion of coastal marsh to aquaculture ponds. Zhang et al. (2022) reported that aquaculture-emissions in China offset about 7% of terrestrial carbon burial, and revealed that the growth and intensification of the aquaculture sector raised climate concerns. Seitzinger (2000) showed that one third of the global anthropogenic N2O emissions were from aquatic ecosystems, and Hu et al. (2012) reported that the global N2O emission from aquaculture would be up to 383 Gg comprising 5.72% of anthropogenic N2O emission by 2030 from 93 Gg in 2009 if the aquaculture industry kept its current annual growth rate of roughly 7.1%. Wetlands Supplement (IPCC, 2013) for aquaculture emission adopted the literature from Hu et al. (2012 & 2013), and mentioned the need for in-situ experiments on the measurement of GHGs emissions from aquaculture systems primarily for developing emission factors of different wetland systems (Hu et al., 2013). Furthermore, some of the emission factors of Wetlands Supplement Guidelines were based on changes in soil C stocks or lab experiments, rather than on the limited field measurements available in the literature (Zheng et al., 2014). Given the limited available data to estimate GHGs emissions from constructed wetland system (IPCC, 2013), more scientific data from in-situ GHGs emissions measurement could support aquatic wetland emission estimation and management.

      In addition, the reclamation of paddy fields for aquaculture was also an important mode of aquatic wetland system in China, which caused more GHGs emissions. Liu et al. (2016) demonstrated that CH4 and N2O emissions reduced following conversion of rice paddies to inland crab-fish aquaculture. Particularly, N2O and CH4 emission from aquaculture had become violent and vital because a vast input of feeding forage enriched substrates for biological activities in water and sediment (Hu et al., 2013); and Fang et al. (2022) confirmed that ebullition dominated the pathways of CH4 emissions from inland aquaculture, which could be mainly influenced by water temperature and dissolved oxygen (Yang et al., 2023). Hence, the impact of natural and constructed wetlands for intensive aquatic culture on GHGs emissions should vary greatly due to differences in environment factors (water and sediment quality) and humanistic management practices (feed types, stocking density). Given this, the main purposes of this research were to (i) quantify the diurnal CHGs (including CO2, CH4 and N2O) fluxes and seasonal variations in different types of managed wetlands; (ii) ascertain the effect of human-based management and environmental factors on CHGs emissions; and (iii) determine the influence of wetland utilization change on CHGs emissions.

    • There was an abundance of aquaculture resources in Anhui province in China, with total water resources of 68 billion m3, 580,000 ha of inland water aquaculture, and an annual output of 2.3 Tg of aquatic products. Chizhou city was known for its aquatic products, with an annual output of 111 Gg, accounting for 11% of water resource in Anhui province. There were various types of aquaculture land with lakes accounting for 63% of the aquaculture area in inland waters, ponds for 20%, and paddy fields for 15%. Thus, Chizhou city (30°41' N, 117°34' E) was selected for this case study to measure GHGs emissions from different managed wetlands with a static floating gas-sampling chamber. The study area is close to the Yangtze River, and has a subtropical monsoon climate. Chizhou city has an annual mean temperature of 16.5 °C, an annual mean precipitation of 1,800 mm, a flat topography, dense network of rivers, and a suitable climate.

    • Four wetlands, including natural wetland (NW), enclosure wetland for intensive aquaculture (EWIA), constructed wetland for intensive aquaculture (CWIA) and constructed wetland for extensive aquaculture (CWEA), were selected for this study and they differ in terms of production and management. Of the four wetlands, one was natural wetland, while the other three were constructed. The information of the four wetlands are shown in Table 1.

      Table 1.  Basic information of the four experimental sites.

      SitesOriginAreaWater tableLocationFeedForage contentYield
      (hm2)mLongitudeLatitudeTotal C
      (g kg−1)
      Total N
      (g kg−1)
      C/N*t ha−1 year−1
      NWA natural lake1,0003−930.67°117.52°NANANANA/
      EWIAWetland shifted to aquatic production532−430.64°117.46°Fed on forage from March to October127.12 ± 7.4177.85 ± 2.031.6:111.32
      CWEAPaddy converted constructed wetland with2130.73°117.58°Fed on wheat grain or grass or fertilizer irregularly5.00
      CWIAPaddy converted constructed wetlands with intensive aquaculture21−230.66°117.50°Fed on forage from April to October151.31 ± 2.8072.39 ± 1.292.1:17.50
      * C/N represented the carbon-nitrogen ratio.
    • The static floating gas-sampling chamber was used for the experiments at four sites (Gao et al., 2014). The chambers were made of PVC and covered by a layer of foam and aluminum foil in order to insulating and reducing heat transmission and reflect light. Every chamber (30 cm inner diameter ×18 cm net height above the surface of the water after being dipped 5 cm into the water) had two holes on top, one for a sampling pipe with a three-way airtight valve and the other for a thermometer to monitor the temperature inside the chamber. All connections were keeping "air tight".

    • For monitoring the diurnal dynamics of GHGs under different wetland management, the campaigns were conducted for 3 d at each site during the months of September to December 2014. Each site contained three replicate points, each of which was a tripod constructed in advance from three bamboo sticks. On each sampling day, five gas samples were taken every 2 h from 8:00 AM to 4:00 PM. For each chamber run, gas samples were collected at 0, 10, 20 and 30 min following the chamber installation (with a wire attached to a tripod to hang over the water). A syringe was used to extract a 40 mL gas sample, which was then injected into a vacuum gas sampling battle (27 mL). At the end of each gas sample collection, the three-way airtight valve should be closed and the water, air and inside-chamber temperatures were noted. After four gas samples, floating chambers were raised and transported to the following sampling location. After collection, all samples were delivered to the lab to be measured over the course of two days.

    • The seasonal dynamic monitoring lasted 13 months, from June 2015 to June 2016 (Fig. 1) showed the temperature and precipitation during the experiments. The four sites of CWIA, CWEA, EWIA and NW had five, five, six and six replicate sampling points, respectively. On each sampling day, gas samples were taken between 9:00 to 12:00 AM. The same procedure as outlined above was applied to each chamber run. The seasonal monitoring frequency was once a month. Nevertheless, due to freezing temperatures, no data on GHG fluxes were available in January 2016.

      Figure 1. 

      Temporal variation of environmental temperature and precipitation at the experimental sites.

    • The gas samples were analyzed by a gas chromatograph (Agilent 7890A GC, USA), and its measurement accuracy (reproducibility), which was the average relative standard deviation, was less than 1%. The detector used for CH4 and CO2 analysis was the FID with detection sensitivity of 0.5 and 1 ppm for CO2 and CH4, and detection limit ≤ 5.1 × 10−12 g CO2 s−1 and 1.9 × 10−12 g CH4 s−1, the column temperature was 80 °C, and the detector temperature was 200 °C; N2 as the carrier gas with the flow rate of 40 mL min−1; H2 as the fuel with the flow rate of 35 mL min−1; air as auxiliary fuel with the flow rate of 350 mL min−1. The detector for N2O was the ECD with detection sensitivity of 0.5 ppm for N2O and detection limit ≤ 4.4 × 10−15 g mL−1, the column temperature was 65 °C, and the detector temperature was 320 °C, argon methane gas as the carrier gas with a flow rate of 30 mL min−1. Gases flux rates were calculated (Healy et al., 1996; Altor & Mitsch, 2006) according to the following equation:

      $ {F}_{G}=kV/A\times (dC/dT)\times \rho \times \left[\frac{{T}_{0}}{\left({T}_{0}+{T}_{C}\right)}\right]\times \frac{P}{{P}_{0}}\times 60 $ (1)

      Where, $ {F}_{G} $ represented the flux rate (mg m−2 h−1 for CO2 and NH4; μg m−2 h−1 for N2O); $ V $ was the chamber volume (m3); $ A $ was the sample surface area of the chamber (m2); $ dC/dT $ donated the ratio of GHGs concentration change over time in a static chamber (ppm min−1);$ \rho $ was the gas density (g m3); $ {T}_{C} $ was the temperature (°C) of the chamber at the time of each sample; $ {T}_{0} $ and $ {P}_{0} $ were the standard temperature (273 K) and atmospheric pressure (assume 1 atm), respectively; $ k $ was the gas dimensional conversion constant; 60 was conversion efficiency of per minute to per hour emission. As N2O was involved, the equation should be multiplied by 103 for converting mg to μg. CO2, CH4 and N2O were accounted with global warming potentials (GWP) of 1, 28 and 265, respectively (IPCC, 2013).

      In general, the flux rate by different time intervals of 4 gas sample concentration by linear regression analysis (Zou et al., 2005). The daily and seasonal cumulative GHGs emissions could be obtained by average value and observation interval times.

    • At the beginning of the season's monitoring in June 2015, sediment samples at 0−15 cm depth were collected for four points by soil sample collectors and kept in plastic bags. From June 2015 to February 2016, water samples under the water's surface at a depth of 30 cm were taken monthly during the seasonal monitoring, and stored in 250 mL plastic bottles. Accordingly, no samples were collected in January 2016 due to the freezing temperatures. All the samples were stored in an ice-filled cooler box, and were delivered to the lab being stored in a 4 °C refrigerator, then indicator analysis was completed within 2 d.

      In the laboratory, the soil samples were air-dried, and finely ground to pass a 2 mm sieve prior to analysis. A portion of soil was ground to pass a 0.15 mm sieve for physical and chemical analysis of pH, soil organic matter (SOM), ammonium-N (${\text{NH}^+_4} $-N), nitrate-N (${\text{NO}^-_3} $-N), and total nitrogen (TN) using the protocol described by Lu (2000). For the water samples, properties analysis of pH, ${\text{NH}^+_4} $-N, ${\text{NO}^-_3} $-N, nitrite-N (${\text{NO}^-_2} $-N), TN, dissolved organic carbon (DOC) were carried out after filtration by 0.45 μm filter membrane using the method reported by Lu (2000).

    • Data processing was performed using Microsoft Office Excel 2013 and Origin 2016, and all statistical analyses were conducted using one-way ANOVA of SPSS and the least significant difference test were used to check the differences among the GHGs emissions of different aquaculture wetlands types.

    • The diurnal variations of GHGs fluxes for four different aquaculture wetlands varied greatly at both time and spatial scales, of which were measured for three days with different measurement dates and temperature. As shown in Fig. 2, there was no consistency in the daily variation dynamics of N2O-N, CH4-C and CO2-C emission fluxes among the four sites. For the three human-managed wetland systems of CWIA, CWEA and EWIA, the GWP values were mainly determined by the CH4 fluxes based on the similar trends. Particularly when the temperature dropped below 10 °C, the CH4 emissions and GWP values at the CWIA site significantly decreased (Fig. 2a). However, this was not the case at the NW site, where the GHG fluxes remained relatively stable despite the temperature dropping below 10 °C, and the GWP values largely following the same dynamics as CO2-C fluxes (Fig. 2d).

      Figure 2. 

      Diurnal variation of CO2, CH4 and N2O fluxes for (a) CWIA, (b) CWEA, (c) EWIA, (d) NW. D1, D2 and D3 for CWIA were October 16, November 3 and December 9 with daily mean temperature ranging from 21.3 to 9.0 °C; D1, D2 and D3 for CWEA were September 4, September 7, and September 22 with temperatures ranging from 29.7 to 23.1 °C; D1, D2 and D3 for EWIA were September 21, September 22, and October 14 with temperatures ranging from 24.9 to 19.6 °C; D1, D2 and D3 for NW were November 4, November 20, and December 11 with temperatures ranging from 17.6 to 8.4 °C. GWP, the global warming potential.

    • As shown in Fig. 3, the emissions of N2O-N, CH4-C and CO2-C exhibited pronounced seasonal variations, with the highest emission rates occurring in the summer months. Firstly, the N2O-N fluxes ranged from 1.90 to 41.74 μg m−2 h−1, 3.03 to 20.59 μg m−2 h−1, 4.16 to 51.33 μg m−2 h−1 and 3.25 to 24.82 μg m−2 h−1 for CWIA, CWEA, EWIA and NW, respectively (Fig. 3a). Secondly, the CH4-C fluxes from EWIA and CWIA were higher compared to CWEA and NW during the periods of June to September 2016 and May to June 2017. However, all sites showed low emission values ranging from 0.03 to 3.76 mg m−2 h−1 during October 2016 to March 2017 (Fig. 3b). The CO2-C fluxes varied across the sites, with values ranging from −9.71 to 67.24 mg m−2 h−1, −8.30 to 35.20 mg m−2 h−1, 4.14 to 55.08 mg m−2 h−1 and −7.05 to 29.40 mg m−2 h−1 for CWIA, CWEA, EWIA and NW, respectively (Fig. 3c). In summary, the GHGs fluxes were consistently positive, except for the negatives values observed in April 2017, indicating that the four aquaculture systems were net emitters of GHGs.

      Figure 3. 

      Seasonal GHG emission dynamics of (a) N2O, (b) CH4 and (c) CO2 fluxes for different wetlands. The management practices of drainage and flooding occurred only to the sites of CWIA, CWEA and EWIA.

      Regarding cumulative GHG emissions, EWIA had the highest cumulative N2O, CH4 and CO2 emissions with values of 2.43 kg N2O-N ha−1, 1,360.27 kg CH4-C ha−1, and 2,246.79 kg CH4-C ha−1 (Table 2). Additionally, the GHGs emission intensities per unit product at the yield-scaled ranged from 1.21 to 5.30 kg CO2-eq kg−1, and the highest N2O and CH4 emission factors of aquaculture production were both from EWIA, with the values of 0.34 g N2O kg−1 and 0.19 g CH4 kg−1, respectively (Table 2). In Table 3, the GWP of EWIA, with a value of 60.02 t CO2-eq ha−1 (the highest), was about 10 times higher than that of NW being 6.04 t CO2-eq ha−1. For intensive aquaculture wetlands, the biggest GWP contributions came from CH4 emission, with proportions of 85% and 70% for EWIA and CWIA, respectively. In contrast, for NW and CWEA, respectively, CO2 emissions made up the greatest quantities, at 72% and 80%.

      Table 2.  Cumulative CO2, CH4 and N2O fluxes and yield-scaled GHG emission intensity for different wetland systems.

      SitesCumulative emissionGHG emission intensity per kg product
      N2O-N (kg ha−1)CH4-C (kg ha−1)CO2-C (kg ha−1)N2O (g kg−1)CH4 (kg kg−1)Total (kg CO2-eq kg−1)
      CWIA1.06 ± 0.20457.08 ± 161.201,877.04 ± 527.950.220.103.25
      CWEA0.81 ± 0.1223.83 ± 7.801321.32 ± 91.110.250.011.21
      EWIA2.43 ± 0.201,360.27 ± 1552.982,246.79 ± 488.230.340.195.30
      NW1.12 ± 0.4238.29 ± 11.991,305.48 ± 216.63///
      /: represented no data. The values were presented as the mean ± standard deviation.

      Table 3.  Global warming potentials and their contributions for different wetland systems.

      SitesGWPContributions (%)
      t CO2-eq ha−1 y−1N2OCH4CO2
      CWIA24.382%70%28%
      CWEA6.046%14%80%
      EWIA60.022%85%14%
      NW6.677%21%72%
    • Table 4 shows the analysis of environmental indicators at four different sites. For water quality, the pH values were slightly alkaline with an average of 7.48, ranging from 7.39 to 7.57. The concentrations of ${\text{NH}^+_4}{\text {-N}} $ ranged between 0.25 to 0.66 mg L−1, and there was no significant difference among the four sites. The concentrations of ${\text{NO}^-_3} $ were higher at, the EWIA site with a value of 1.36 mg L−1 compared to the other three aquatic-wetland systems. For TN concentrations, NW had a lower value than CWIA and EWIA. The DOC content varied between 20.01 mg L−1 at NW and 29.92 mg L−1 at EWIA.

      Table 4.  Water and sediment qualities at the four experimental sites.

      SitespH$ {\text{NH}^+_4}{\text {-N}}$ (mg L−1)${\text{NO}^-_3}{\text {-N}} $ (mg L−1)${\text{NO}^-_2}{\text {-N}} $ (mg L−1)TN (g L−1)DOC (mg L−1)
      Water qualityCWIA7.52 ± 0.14a0.60 ± 0.35a0.57 ± 0.26b0.48 ± 0.51a1.93 ± 0.69ab24.99 ± 8.50ab
      CWEA7.39 ± 0.48a0.46 ± 0.33a0.36 ± 0.12b0.10 ± 0.16ab1.30 ± 0.52bc21.83 ± 7.67ab
      EWIA7.50 ± 0.28a0.66 ± 0.61a1.36 ± 0.56a0.36 ± 0.49ab2.07 ± 0.88a29.93 ± 8.92a
      NW7.57 ± 0.29a0.25 ± 0.20a0.28 ± 0.08b0.03 ± 0.03b0.66 ± 0.23c20.01 ± 7.71b
      SitespH$ {\text{NH}^+_4}{\text {-N}}\;{\rm{ (mg\; kg^{−1})}} $$ {\text{NO}^-_3}{\text {-N}}\;{\rm{ (mg\; kg^{−1})}} $SOM (g kg−1)TN (g kg−1)C/N
      Sediment qualityCWIA6.81 ± 0.13a36.67 ± 4.32b273.70 ± 185.08a39.69 ± 2.75ab2.95 ± 0.18ab7.81 ± 0.51ab
      CWEA6.22 ± 0.18b33.21 ± 6.39b54.57 ± 20.31a39.66 ± 4.16ab2.68 ± 0.19b8.58 ± 0.42a
      EWIA6.59 ± 0.29a203.22 ± 16.08a251.03 ± 263.67a44.00 ± 1.15a3.34 ± 0.06a7.46 ± 0.19b
      NW6.58 ± 0.14a27.86 ± 5.07b21.23 ± 5.10a35.78 ± 4.09b2.91 ± 0.47ab7.00 ± 0.73b
      TN, DOC, SOM and C/N represented total nitrogen, dissolved organic carbon, soil organic matter and carbon : nitrogen ratio, respectively.

      The sediment properties at all sites were weakly acidic, with pH values ranging from 6.0 to 6.9. Significantly, the pH value of CWEA was lower than that of the other three sites. The highest TN content was found at EWIA with a value of 3.34 g kg−1. The SOM contents showed a significant difference between NW and EWIA. The ${\text{NH}^+_4}{\text {-N}} $ concentrations was significantly higher at EWIA with a value of 203.22 mg kg−1 compared to other three systems. The concentrations of ${\text{NO}^-_3}{\text {-N}} $ varied between 21.23 to 273.70 mg kg−1. The C/N ratio was between 7.00 to 8.58.

    • As shown in Table 5, there were strong positive correlations between ${\text{NH}^+_4}{\text {-N}} $ concentrations in water and the fluxes of CH4 (R2 = 0.60, P < 0.01), N2O (R2 = 0.37, P < 0.05), and CO2 (R2 = 0.70, P < 0.01). Additionally, ${\text{NO}^-_3}{\text {-N}} $ concentrations in water were also strongly and positively correlated with CH4 (R2 = 0.80, P < 0.01), N2O (R2 = 0.60, P < 0.01), and CO2 fluxes (R2 = 0.54, P < 0.01). In summary, the form and concentration of nitrogen in water had a significant impact on GHG emissions.

      Table 5.  Correlation coefficients between GHGs emission and water parameters in all four wetlands.

      ItemspHDOC${\text{NH}^+_4}{\text {-N}} $${\text{NO}^-_3}{\text {-N}} $${\text{NO}^-_2}{\text {-N}} $TNWater temperature
      CorrelationCH40.110.090.60**0.80**−0.170.130.46*
      N2O0.310.200.37*0.60**−0.220.100.55**
      CO20.28−0.110.70**0.54**−0.230.110.65**
      P valueCH40.5740.6330.0020.0000.3990.5170.014
      N2O0.1090.2990.0490.0010.2490.6050.003
      CO20.1440.5760.0000.0030.2460.5750.000
      * Significant at P < 0.05; ** Significant at P < 0.01.

      Furthermore, upon exploring forage-fed impact of CWIA and EWIA (Fig. 4), it was found that there were positive correlations between forage feeding and the CH4 fluxes (R2 = 0.68), N2O fluxes (R2 = 0.50) and CO2 fluxes (R2 = 0.41) at the EWIA site; and there were positive correlations between forage feeding and the CH4 fluxes (R2 = 0.69), N2O fluxes (R2 = 0.70) and CO2 fluxes (R2 = 0.57) at the CWIA site.

      Figure 4. 

      Dependence of CO2, CH4 and N2O fluxes on feeding amount at the sites of (a) CWIA and (b) EWIA.

    • This study tracked GHGs emissions for four different aquatic-wetland systems, partially reflecting the variations in how human management practices affect wetland GHG emissions. In the current study, the average CO2, CH4 and N2O emissions from nature wetland (NW) were 51.09 mg m−2 h−1, 0.55 mg m−2 h−1and 19.50 μg m−2 h−1, respectively, which were similar to those values ranging from 15.40 to 1082.00 mg CO2 m−2 h−1, 0.04 to 175.94 mg CH4 m−2 h−1 and 13.3 to 747 μg N2O m−2 h−1 shown in Table 6. The high fluxes of CH4 and CO2 at the canal were mostly caused by domestic waste pollution (Han et al., 2013). The N2O fluxes from aquatic-wetlands converted by agricultural land in the study were 14.08 μg m−2 h−1 (CWEA) and 18.65 μg m−2 h−1 (CWIA), which were higher than the values of 10.74 and 11.8 μg m−2 h−1 for the aquatic wetlands systems of shrimp and mixed shrimp-fish pond (Yang et al., 2015), but lower than that of the crab-fish aquatic system with a value of 48.1 μg m−2 h−1 (Liu et al., 2016). The CH4 fluxes of 0.35 mg m−2 h−1 at the CWEA site was the lowest among these inland aquaculture systems possibly due to the absence of amount forage feed.

      Table 6.  Comparison of greenhouse gases fluxes from different wetland types with previous studies in China and other regions.

      LocationSourcesTime horizonWetland typesN2O fluxesCH4 fluxesCO2 fluxes
      μg m−2 h−1mg m−2 h−1mg m−2 h−1
      ChinaEast ChinaCurrent study13 monthsConstructed wetland for intensive aquaculture18.65 ± 20.486.92 ± 9.4177.23 ± 68.95
      Constructed wetland for extensive aquaculture14.08 ± 9.580.35 ± 0.4551.21 ± 40.97
      Enclosure wetland for intensive aquaculture41.89 ± 24.8320.60 ± 19.9090.34 ± 65.45
      Natural wetland19.50 ± 12.480.55 ± 0.4151.09 ± 33.48
      East ChinaLiu et al. (2016)12 monthsRice paddies110.50.71
      Crab-fish aquaculture48.10.37
      Southeastern ChinaYang et al. (2015)5 monthsShrimp pond10.7419.9520.78
      4 monthsShrimp-fish pond11.81.65−60.46
      Lab simulation of aquatic systemHu et al. (2014)2 monthsControl5,504 ± 1,2212,401 ± 343
      Treatment911 ± 4884,590 ± 546
      Overseas
      of Nature wetlands
      Yangtze estuarine, ChinaWang et al. (2009)12 monthsMarsh site2.06
      Bare tidal flat0.04
      East ChinaHan et al. (2013)3 monthsOpen river62.17 ± 2.1316.41 ± 3.06162.18 ± 10.55
      Canal111.74 ± 7.41175.94 ± 18.421,082.00 ± 90.53
      Reservoir50.99 ± 2.681.70 ± 0.1425.61 ± 4.08
      Northeast ChinaYang et al. (2013)3 monthsFreshwater marsh13.38.92394.5
      IndiaSelvam et al. (2014)24 h/1−8 hPonds11.93 ± 12.33123.02 ± 117.33
      Rivers4.13 ± 8.2736.85 ± 70.58
      Reservoirs2.13 ± 2.3315.40 ± 35.75
      ColombiaDennis et al. (2014)11 monthsRestored mangrove162 ± 483.47 ± 1.52
      97 ± 4611.78 ± 10.54
      747 ± 2996.12 ± 9.69
      Costa RicaNahlik et al. (2011)29 monthsForested marsh5.05
      Rainforest swamp33.39
      Alluvial marsh39.94
      AustraliaAllen et al. (2011)6 monthsMangrove30.000.30
      Orinoco River Floodplain, VenezuelaSmith et al. (2000)17 monthsOpen river1.33
      Flooded forest4.67
      Macrophyte mats1.33

      The estimates of global wetland emissions were still subject to severe data limitations and distributional biases, especially for the emission under human intervention (Selvam et al., 2014). The 2006 IPCC Guidelines, which was limited to peatlands that had been drained and managed for peat extraction, provided methods for estimating national anthropogenic emissions (IPCC, 2006). The 2013 Supplement to the 2006 IPCC Guidelines for National Greenhouse Gas Inventories was published to reduce uncertainties of wetland emissions estimates. It included information on inland organic soils, wetlands on mineral soils, coastal wetlands, as well as constructed wetlands for wastewater treatment (IPCC, 2013). However, the evaluation methods for artificially managed aquatic wetlands still needed to be improved. Hence, based on more reliable observational data, scientists can simulate GHGs emissions through modeling and analyze human impacts on the GHG balance of wetland ecosystems (Petrescu et al., 2015).

    • For aquatic wetland system, human factors such as C and N inputs, water level control, fish type and fish intensity could affect the intensity of GHGs emission (Yang et al., 2013; Zheng et al., 2014; Yang et al., 2023). Forage feeding rate directly affected ammonia concentrations in the aqueous phase, which then affected the fate or release rate of CH4 and N2O from the sediment-water interface to the atmosphere (Liu et al., 2016). Beaulieu et al. (2011) reported that approximately 1.00% of N input may be transformed to N2O for rivers and estuaries, and Seitzinger & Kroeze (1998) discovered that roughly 0.75 % of N for streams was released to the atmosphere as N2O. Hu et al. (2013) also revealed that about 1.3% of the nitrogen input was emitted as N2O gas in laboratory-scaled aquaculture system. While this study found that approximately 0.2%−0.3% of N input became N2O emissions. In addition, changes in water levels (such as drainage or irrigation) were common management practices in aquaculture, which could significantly affect CH4 emissions (Friborg et al., 2000; Jackowicz‐Korczyński et al., 2010).

      Environmental factors were closely related to the input of C and N, directly affecting GHGs emissions. Numerous research had demonstrated that the CH4 and N2O fluxes were associated with environmental factors, such as air and soil temperature, dissolved oxygen and DOC in water, and the availability of C and N substrate in sediment (Bhattacharyya et al., 2013; Hu et al., 2016). According to this study, both the ${\text{NH}^+_4}{\text {-N}} $ and ${\text{NO}^-_3}{\text {-N}} $ concentrations showed strong positive correlations with CH4, N2O and CO2 fluxes (Table 5). Liu et al. (2016) studied that CH4 and N2O emissions were negatively related to the water DO concentration. Tsuneda et al. (2005) found that N2O conversion increased when the salinity increased from 10 to 20 mg L−1 during single nitrification. Whereas Barnes et al. (1999) and Hashimoto et al. (1999) revealed a negative relationship between salinity and dissolved N2O concentration. Moreover, GHG emissions in aquaculture systems were linked to aquatic animal species due to variations in their digestion processes, such as filter- and deposit-feeding, as observed in studies by Stief et al. (2009) and Bhattacharyya et al. (2013). Therefore, managing the entire range of aquatic organisms in aquaculture can play a crucial role in mitigating GHG emissions.

      Among various forms of aquatic management, GHGs emission factors on the scale of area and yield varied significantly. This study found a positive correlation between cumulative GHG emissions intensity and yield in aquatic wetlands. Generally speaking, the production of CWEA, CWIA and EWIA increased sequentially, reaching 5.00, 7.50 and 11.32 t ha−1, respectively. The corresponding annual cumulative emissions also increased sequentially, reaching 0.81, 1.06 and 2.43 kg N2O-N ha−1, 23.83, 1,877.04 and 2,246.79 kg CH4-C ha−1, 1,321.32, 1,877.04 and 2,246.79 CO2-C kg ha−1. Petrescu et al. (2015) reported that management intensity strongly influenced the net climate footprint of wetlands. The highest N2O emission factor of aquaculture production from the intensive aquaculture type of EWIA was 0.34 g kg−1, which was still much lower than the values of 2.65 and 2.57 g kg−1 reported by Hu et al. (2012) and Liu et al. (2016) using lab simulation experiments. In terms of GWP, aquatic system with high forage feeding intensity had significantly higher annual GWP compared to the wetland systems with no or low forage feeding intensity. After the intensive aquaculture management of wetlands, the main GHG emissions of the system shifted from CO2 to CH4, which greatly increased the GWP. For instance, the largest contributions to GWP for the intensive aquaculture of EWIA and CWIA was CH4 emission, while for natural aquatic-wetland and exclusive aquaculture, the main contributor turned out to be CO2 emission (Table 3). Therefore, scientific management of aquaculture density, forage feeding intensity, and water level in aquatic wetland systems would play a crucial role for the trade-off between CH4 and CO2 emission, as well as GHG mitigation (Petrescu et al., 2015).

    • The GHGs emission fluxes found in this investigation were usually characterized by high standard deviations (Figs 2 and 3), which were in agreement with the results from Nahlik et al. (2011) and Dennis et al. (2014). The large variances indicated significant differences in emissions between parallel points at the same site, which might be caused by changes in water level, temperature, ebullition, and environmental disturbance (Liu et al., 2016; Petrescu et al., 2015; Fang et al., 2022; Yang et al., 2022b). For example, the sediment and water quality in the areas near the feeding machines were complicated, and the breeding density in this region was frequently high, both of which greatly affected the GHGs emission intensity (Chen et al., 2013; Hou et al., 2013).

      Besides, the sampling points were uniformly set at different locations of each site, and the process of sailing for sampling could also cause some interference to the water body, like creating some surface waves, thereby affecting the detection results and causing uncertainties to the results (Gao et al., 2014). More importantly, the bamboo poles used for fixation were more or less influenced by ships, causing disturbance to the sediment and generating bubbles, further affecting the fate or release rate of GHGs through the sediment-water interface system (Hirota et al., 2004; Bastviken et al., 2008). Therefore, future work should improve the equipment technology for monitoring GHGs emission in wetlands. Finally, diurnal dynamic measurements were carried out at the four sites for eight hours due to the safety issues, night time monitoring was excluded. This may consequently lead to the overestimation of the GHGs fluxes because of lower night time temperatures. So, it is necessary to monitor the emission characteristics throughout the whole day as much as possible.

    • For the four aquatic-wetlands, the extensive aquaculture system from reclaimed paddy land (CWEA) had the lowest GHG emission intensity with 0.81 kg N2O-N ha−1, 23.83 kg CH4-C ha−1 and 1,321.32 kg CO2-C ha−1, while the intensive aquaculture system from wetland enclosure (EWIA) had the highest cumulative GHG emission with the values of 2.43 kg N2O-N ha−1, 1360.27 kg CH4-C ha−1, 2,246.79 kg CO2-C ha−1, respectively. And the GHG emission intensity per unit of production ranged from 1.21 (CWEA)−5.30 (EWIA) kg CO2-eq kg−1. The current study supported a valuable reference for N2O and CH4 emission factors in constructed wetlands and demonstrated that the intensive aquaculture ultimately led to the high area-scaled and yield-scaled GHGs emission intensity. These measurement data served as a crucial support for the follow-up model simulation and GHGs emission prediction on a large scale.

      • This work was financially supported by Natural Science Foundation of China under a grant number 41907073 and 41877546.

      • The authors declare that they have no conflict of interest.

      • Copyright: © 2023 by the author(s). Published by Maximum Academic Press, Fayetteville, GA. This article is an open access article distributed under Creative Commons Attribution License (CC BY 4.0), visit https://creativecommons.org/licenses/by/4.0/.
    Figure (4)  Table (6) References (53)
  • About this article
    Cite this article
    Yue Q, Cheng K, Sheng J, Wang L, Ji C, et al. 2023. Greenhouse gas emissions from managed freshwater wetlands under intensified aquaculture. Soil Science and Environment 2:3 doi: 10.48130/SSE-2023-0003
    Yue Q, Cheng K, Sheng J, Wang L, Ji C, et al. 2023. Greenhouse gas emissions from managed freshwater wetlands under intensified aquaculture. Soil Science and Environment 2:3 doi: 10.48130/SSE-2023-0003

Catalog

    /

    DownLoad:  Full-Size Img  PowerPoint
    Return
    Return