Research articleAssessing passive rehabilitation for carbon gains in rain-filled agricultural wetlands
Graphical abstract
Introduction
Facing global climate change, there is immense pressure on governments and land managers to develop and implement techniques to mitigate the impact of greenhouse gas (GHG) emissions. Biosequestration defines nature's ability to capture and store carbon dioxide (CO2), helping to offset anthropogenic emissions and balance the global carbon cycle (Fenstermacher et al., 2016; Lal, 2007; Pachauri, 2004; Pachauri and Meyer, 2014). Increasing the carbon sequestration capacity of earth's ecosystems is an essential part of the mitigation pathway to avoid global tipping points, such as a 2 °C air temperature increase (IPCC, 2018; Lal, 2003). As such, land management that optimizes carbon sequestration is crucial and should be underpinned by high quality scientific evidence pertaining to specific ecosystems (Zedler, 2003).
Wetlands are significant within the global carbon cycle, for their high capacity for removingatmospheric CO2 and storing it long-term in organic soils (Armentato and Menges, 1986; Fennessy et al., 2008). Freshwater wetlands store an estimated 33% of the global terrestrial carbon pool at a rate 30–40 times higher than forests (Bernal and Mitsch, 2011; Watkins et al., 2017). This sequestration process is made effective by high rates of primary productivity in wetland vegetation and the reduction of biomass decomposition by anaerobic soils during flooding (Anderson et al., 2016; Batson et al., 2015; Fenstermacher et al., 2016). However, these anaerobic conditions simultaneously make freshwater wetlands naturally high emitters of methane (CH4), which contributes to 40–60% of natural global CH4 emissions (Abdalla et al., 2016; Anderson et al., 2016; Bernal and Mitsch, 2011), or 20–30% of overall CH4 emissions (i.e. natural and industrial; (Neubauer, 2014). In the short term, CH4 production from freshwater wetlands can be high enough to negate the benefits of carbon sequestration (Batson et al., 2015; Mitsch et al., 2013; Nahlik and Fennessy, 2016).
When undisturbed, wetland soil carbon can be stored for over thousands of years (Bridgham et al., 1995). However, human alteration of wetlands through drainage, river regulation, and other hydrological disturbance, results in stored carbon being released back into the atmosphere as CO2 (Armentato and Menges, 1986; Bridgham et al., 2006; Fenstermacher et al., 2016). Depending on the degree of disturbance, natural GHG flux cycles can shift wetlands from functioning as carbon sinks to carbon sources (Mitra et al., 2005). The management of wetlands, both to protect undisturbed wetlands and to rehabilitate degraded wetlands, is therefore a key component for reducing global carbon emissions (Batson et al., 2015; Bortolotti et al., 2016).
Several studies have demonstrated increased carbon stocks and sediment accumulation in wetlands following restoration (Euliss et al., 2006; Knox et al., 2015; Mitsch and Hernandez, 2013; Waddington et al., 2010), and have subsequently generated interest for the inclusion of wetland rehabilitation in carbon offsetting programs (Anderson et al., 2016). However, the development of such programs is difficult with current data, which have significant knowledge gaps owing to high natural variation in rehabilitation response between wetlands. Effective wetland rehabilitation requires specific knowledge concerning wetland type, land use history, biogeographic/climatic region, sequestration capacity, and rates of recovery (Carnell et al., 2018; Fennessy et al., 2008). In addition, efforts are restricted by a shortage of data that monitors long-term (i.e. multi-decadal) rehabilitation outcomes, which may not allow sufficient time for the recovery of soil processes that impact carbon cycling (Ballantine and Schneider, 2009; Turner et al., 2001).
Further, current knowledge on freshwater wetland restoration for carbon gains is overwhelmingly biased towards the northern hemisphere (Beringer et al., 2013; Boon et al., 1997; Page and Dalal, 2011). Australia has an estimated 24 million ha of wetlands, but following global trends (Davidson, 2014; Hu et al., 2017) has lost up to 89% of historic wetland extent, due to the regulation of major river basins, urban expansion, and agricultural development (Finlayson et al., 2013; Page and Dalal, 2011). In the state of New South Wales on the east coast of Australia, the vast majority of wetlands occur in areas with less than 500 mm annual rainfall (i.e. semi-arid climates). Owing to their temporary nature, rain-filled (or depressional) wetlands are often undervalued and are typically subjected to agricultural land-use practices, to the extent that 84% of wetlands in the Murray-Darling basin are currently grazed by livestock (Nairn and Kingsford, 2012). Despite covering a substantial area (~300,000 ha in NSW alone (Kingsford et al., 2003)), there is relatively little information on the carbon sequestration, carbon stocks, GHG emissions, and soil microbial communities of rain-filled wetlands in Australia. However, recent work on temperate seasonal wetlands in southeastern Australia show that they function as significant carbon sinks compared to adjacent terrestrial soils (Pearse et al., 2017). Additionally, wet-dry cycles in wetlands have been shown to influence the community structure of soil microbes, including those involved in GHG flux, i.e. methanogens, methanotrophs, and fermenters (He et al., 2015), and thus could influence the carbon sink capacity of seasonal wetlands.
Ecosystem rehabilitation can be either passive or active. Passive ecosystem rehabilitation removes the pressure(s) causing degradation, relying on natural, autogenic processes such as seed bank re-establishment and species recolonization, to restore ecological function (McIver and Starr, 2001; Meli et al., 2017). Conversely, active rehabilitation in wetlands requires physical intervention to restore the topography, hydrology, or biota of the system (Holl and Aide, 2011; McIver and Starr, 2001). Active rehabilitation provides better control over abiotic and biotic conditions, and may expedite recovery, however typically demands substantially greater financial resources than passive rehabilitation (Holl and Aide, 2011; Meli et al., 2017). Therefore, it is important to assess the effectiveness of passive rehabilitation before further resources are invested (McIver and Starr, 2001).
Given the prevalence of wetland drainage and disturbance in agricultural regions, we aimed to present a comprehensive study of carbon cycling in semi-arid rain-filled wetlands of southeastern Australia. We compared GHG emissions, soil microbial communities, soil carbon, and carbon stored in aboveground plant biomass, between passively rehabilitated and adjacent control (or ‘business-as-usual’, i.e. grazed and/or cropped) wetlands in the Riverina region of New South Wales. We compared carbon cycling characteristics across wetlands of two different dominant vegetation types, eucalypt and graminoid, and compared newer (2–5 years) and older (10–20 years) rehabilitation ages. Our aim was to assess passive rehabilitation for increasing carbon storage and discuss the potential inclusion of rain-filled wetland rehabilitation in carbon offsetting programs. We hypothesized that: (1) Soil carbon stocks and aboveground plant carbon stocks would increase with time following rehabilitation, while GHG emissions would decrease, (2) Wetlands would support distinct microbial communities according to their wet or dry phase, corresponding to variation in GHG emissions, and (3) Owing to higher overall biomass, eucalypt-dominated wetlands would store more carbon than graminoid-dominated wetlands.
Section snippets
Site selection & experimental design
All sites were located in the Riverina region of the Murray River catchment,New South Wales (NSW), Australia (Fig. 1; Table S1). The region is semi-arid, characterized by cool winters and hot summers (Eardley, 1999), with mean temperatures ranging from 2.8 °C in July to 32 °C in January (Bureau of Meteorology, 2018). Average annual rainfall is 544 mm, with the highest rainfall occurring from June to August (mean of 53.3–55.8 mm d-1; BOM). The region is part of the Murray-Darling Basin,
Soil carbon
Mean soil Corg was 3.54 ± 3.24% and ranged from 0.39 to 28.19% across all sites. Soil depth significantly influenced Corg percentage (Table S2). The top 0–5 cm of soil had 5.55 ± 0.66% mean Corg, which decreased to 3.13 ± 0.26% at 5–10 cm, then 1.88 ± 0.18% at 10–15 cm, followed by a slight increase to 1.9 ± 0.38% at 15–20 cm (Table S2). No significant difference in Corg % was found as a result of wetland condition (control, newly rehabilitated or older rehabilitated; p = 0.361; Fig. 2),
Discussion
Passive rehabilitation is a cost-effective method that removes the pressure(s) causing ecosystem degradation to allow natural recovery (Bullock et al., 2011; Holl and Aide, 2011; Meli et al., 2017). When effective, passive rehabilitation increases ecosystem service value that exceeds the cost of restoration, which is rarely achieved for active intervention, due to high restoration costs (Birch et al., 2010). We found no significant effect of passive rehabilitation on the soil carbon or
Conclusions
In conclusion, this study demonstrates the importance of developing an appropriate rehabilitation method for the type of wetland and the objective of rehabilitation. Rain-filled wetlands in semi-arid southeastern Australia likely need active intervention to restore their natural hydrology, and/or topsoil addition to expedite carbon sink recovery, and even then, may take decades or centuries before a benefit is detectable. Carbon offsetting opportunities may be better found in systems with
Funding & permissions
This work was supported by funding from Murray Local Land Services and Australian Research Council Linkage Grant (LP160100061).
Data collection at site TO4 was carried out under Scientific Licence SL102069, issued by the New South Wales Office of Environment and Heritage (National Parks and Wildlife Service).
Credit author statement
Sarah Treby: Methodology, Investigation, Formal analysis, Writing – Original draft preparation, Visualisation, Project administration. Paul Carnell: Methodology, Investigation, Writing – review & editing, Supervision, Funding acquisition. Stacey Trevathan-Tackett: Methodology, Investigation, Writing – review & editing, Supervision. Giuditta Bonetti: Investigation, Formal Analysis, Writing – Original draft preparation, Visualisation. Peter Macreadie: Supervision, Funding acquisition, Writing –
Acknowledgements
The authors thank volunteers Alice Gavoille, Dionysia Evitaputri, James Mackey, and colleague Katy Limpert (Blue Carbon Lab) for help with field data collection. Further thanks go to Jerry Lai (Deakin University) for advice on statistical analyses. We are very grateful to the landholders that participated in this study for their kindness and cooperation. Thanks also to Susanne Watkins and Trish Bowen at Murray Local Land Services for assistance with funding acquisition, and to Sarah Ning at
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2021, Applied Soil EcologyCitation Excerpt :Floating chambers (0.03 m3 volume, 0.12 m2 surface area) were used to measure diffusive GHG emissions between the water-air interface (Frankignoulle, 1988). Floating chambers were made from plastic rectangular tubs, painted silver to reduce solar radiative heating of chamber air (Treby et al., 2020). Chambers were cautiously inserted into the water surface at about 2 cm depth to avoid surface boundary layer disturbance that can potentially affect measurement accuracy (Matthews et al., 2003).
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2021, Applied Soil EcologyCitation Excerpt :Regardless, we expected rehabilitation to decrease organic matter quality (high C:N) across vegetation types, due to disturbance removal and restoration of wetland functionality as carbon sink. While we did not measure plant biomass in this study, previous work at these sites (Treby et al., 2020) showed that rehabilitation through fencing can have a positive outcome on aboveground carbon biomass but without a corresponding benefit in belowground carbon stock (Treby et al., 2020). This variability might be related to the slow recovery of soil conditions after disturbance removal (Ballantine and Schneider, 2009; Moreno-Mateos et al., 2012) and to the potential positive benefit of grazing on carbon sequestration (Tanentzap and Coomes, 2012).
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2021, Water ResearchCitation Excerpt :In the last two decades, governments and institutions internationally have started to recommend either active or passive rehabilitation practices that remove pressures that cause wetland degradation such as grazing, agriculture, and water diversion (Ramsar, 2016; Tanentzap and Coomes, 2012; Wong, 2008). Rehabilitation practices that actively aim to restore the original wetland hydrology or do this through once-off managed events (also known as “environmental watering”) have been previously assessed as more effective compared to passive rehabilitation practices (i.e., grazing removal) for reducing carbon emissions and/or promoting biodiversity more rapidly (Bossio et al., 2006; Chandra et al., 2020; Limpert et al., 2020; Treby et al., 2020). However, although rehabilitation practices have had global positive recognition for biodiversity, further empirical evidence evaluating the microbial response to hydrological rehabilitation is needed to better understand the effects of short-term manipulation on the microbial communities involved in the carbon cycle.