Minimizing the opportunity costs of carbon mitigation on farmland via restoration of salinized soils

doi:10.21203/rs.3.rs-2750754/v1

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Minimizing the opportunity costs of carbon mitigation on farmland via restoration of salinized soils

https://doi.org/10.21203/rs.3.rs-2750754/v1

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Mass conversion of native vegetation to agricultural land-use triggered secondary salinity, a hydrological imbalance, which has damaged more than 1.75 million ha of farmland in south-western Australia. Various types of reforestation have been proposed and tested to restore the hydrological balance, however the economic returns from these cannot compete with existing farm practice and land-holders thus have a reluctance to adopt. An alternative approach has been to reforest abandoned saline areas with salinity and/or water-logging tolerant trees to avoid displacement of farming activities and to minimize opportunity costs.

We propose an improved evaluation methodology for saline site reforestation and in this paper test its feasibility. Three eucalypts species (Eucalyptus salubris, E. camaldulensis and E. sargentii) were planted adjacent to a salt scald in Wickepin, Western Australia, and their survival and growth were monitored under increasing soil salinity ranging from moderate (ECe: 400-800 mS m^-1), high (800-1,600 mS m^-1) or extreme (>1,600 mS m^-1) with shallow (< 1 m depth) saline ground water (EC: 2,000-3,000 mS m^-1). Survival and growth of E. sargentii and E. salubris in the saline discharge areas were comparable to those in the arable area, and reforestation by these species can thus avoid land competition and opportunity costs.

bio-diversity restoration

co-benefits

carbon abatement

eucalypts

mitigation of salinity

opportunity costs compensation

The extensive dryland farming area of southern Australia has been damaged by secondary salinity (Harper et al. 2021), with estimates that 1.75+/-0.34 million ha has been affected in Western Australia, alone (Cacetta et al. 2022). Secondary salinity has been triggered by the extensive deforestation of native vegetation for agriculture which has resulted in a hydrological imbalance through rising and increased discharge of saline ground water (Turner and Ward, 2002). Ground water recharge under native forest, was normally around 0 mm year^− 1 with a maximum of 5 mm year^− 1 (Dunin, 2002), but after deforestation such recharge increased markedly (George et al., 1997). This has resulted in an increase in pressure in groundwater systems contained within the deep regolith, and the dissolution of stored salts. These salt stores have accumulated over long periods in the regolith and are estimated to be from 50 to 5000 Mg ha^− 1 (McFarlane and Williamson, 2002). The discharge of the saline groundwater in lower parts of the landscape results in a decrease in crop yields, or in extreme cases remnant vegetation death (Turner and Ward, 2002) and the abandonment of farmland.

Various measures to manage secondary salinity have been proposed and investigated, with reforestation across the farmed landscape a major methodology (Lefroy and Stirzaker, 1999; Harper et al. 2017) in an attempt to restore the landscape’s hydrological balance. Such displacement has a lower impact on the surrounding environment than civil engineering approaches such as drainage. Drainage has been shown to be ineffective in this region due to both the volumes of water that need to be removed and the cost of establishment (Horn, 2017). Various estimates of the amount of farmland that would need to be displaced by reforestation to resolve the water imbalance have been reported in Western Australia. For example, 24% (Lefroy and Stirzaker, 1999), 16% (White et al., 2002), 22% (Dunin, 2002), and at between 16 and 80% (George et al., 2001).

To treat secondary salinity, therefore, there are various difficulties to be overcome (Lefroy and Stirzaker, 1999; Flugge and Abadi, 2006). Among these difficulties, the largest is considered to be the opportunity costs associated with the displacement from farmland by reforestation, e.g. farmers no longer earn income from crop production or pasture grazing from the displaced farmland. When analysed, the business-as-usual scenario brought much larger profit to land-holders over more than half century compared with reforestation scenarios, and this resulted in a low incentive for land-holders to adopt such measures (Flugge and Abadi, 2006). To solve this challenge, the development of approaches that minimize such opportunity costs and thus increase the incentives to land-holders is necessary.

In previous studies, compensation of opportunity costs by utilizing carbon mitigation investment was proposed. Approaches included reforestation as a secondary salinity measure as a co-benefit of carbon sequestration projects (Harper et al., 2007) or woody biomass obtained from reforestation as a bio-energy source (Sochacki et al., 2007; Harper et al., 2010, 2014). However, the reduction of opportunity costs is strongly dependent on the value of carbon credits, bio-energy (Flugge and Abadi, 2006), the value (e.g. rent) of the land being treated, and thus realization of such measures fully depends on external factors. The externalities associated with salinization of landscapes, such as the loss of biodiversity or damage to built public infrastructure also need to be taken into account.

To reduce opportunity costs to nearly zero without depending on external factors, reforestation in abandoned farmland damaged by secondary salinity (e.g., marginal area of salt scalds and saline seeps), has been investigated in south-western Australia. From these research trials, available woody species with saline tolerance and relatively high growth were screened, and appropriate planting density (trees ha^− 1) were investigated for several species (Archibald et al., 2006; Biddiscombe et al., 1985, 1989; Zohar et al., 2010; Harper et al., 2012; Sochacki et al., 2012). These studies suggested that reforestation of salinized lands, both as a treatment of secondary salinity measure and to mitigate carbon emissions would reduce the opportunity costs substantially and be highly feasible. Most of the previous studies, however, did not compare biomass production between such saline reforestation and farmland displacement, and also did not consider the risk of failure. In addition, some studies did not evaluate biomass or carbon sequestration yields but only height or crown volume of trees. The water use of trees is directly related to their growth rate in water-limited environments even under saline conditions (Hatton et al., 1998; Morris and Collopy, 1999; Benyon et al., 2001), and thus evaluation of biomass productivity is essential.

To make improved secondary salinity treatment measures that resolve water imbalance, mitigate carbon and also reduce opportunity costs to nearly zero we investigated the following:

1: Selection of woody species, with high salinity and water-logging tolerance, with comparable survival and growth in non-saline areas.

2: Exclusion of woody species which cannot survive with shallow ground water and/or highly saline conditions, so as to minimize the risk of failure of reforestation.

3: Evaluation of the biomass productivity of the selected woody species.

2.1. Research area

The research area is on the Tremayne farm in the Shire of Wickepin located approximately 200 km south-east from Perth, the capital of Western Australia. This area has a typical semi-arid warm Mediterranean climate with a seasonal drought (October to May) and winter rain (June to September) (Environment Australia, 2000). Mean, maximum and minimum values of annual rainfall for the past 30 years (1986–2015) were 375 mm (S.E. = 12.2 mm), 501 mm (2011) and 236 mm (2010), respectively (Australian Bureau of Meteorology data (weather station No. 10654; http://www.bom.gov.au/climate/data/stations/). The mean annual rainfall during the observation period (2006–2015) was similar (366 mm, S.E. = 26.2 mm), and the maximum and minimum values were the same.

The topography of this area is gently undulating with a landscape of low relief formed on the Archaean rocks of the Yilgarn Craton, these being deeply weathered in the Tertiary, with subsequent erosion and deposition (Environment Australia, 2000). Salt accumulated in the deep weathering profiles and had been subsequently been mobilized following the disruption of the hydrological balance caused by deforestation for agriculture.

The land-use of the research area is the typical dryland-farming in south-western Australia with rotation cropping of wheat or canola and pasture grazing.

2.2. Reforestation sites

A salt scald of about 0.8 km² (117º40’03.02”E, 32º44’03.67”S) previously reported by Sochacki et al. (2012) was selected. This area of the farm was formerly arable but has been salinized since farm development. Its surrounding area (more than 1 km²) had been abandoned as a result of secondary salinity. In 2006, within this abandoned marginal area, two experimental sites of about 0.2 ha each were established in the upstream (117º40’22.14”E, 32º44’13.77”S) and downstream (117º39’59.13”E, 32º43’15.53”S) areas, which were named “Upper salt scald site” and “Lower salt scald site”, respectively. As a control, one experimental site of about 0.15 ha was established in an upper slope of the adjoining arable area (117º40’47.85”E, 32º43’59.20”S), which was named “Arable area site”. This site is 27 m higher in elevation compared to the sites in the valley floor. By comparing tree survival and biomass between the salinized and arable sites appropriate tree species can be selected.

The ground water level of the Upper & Lower salt scald sites was shallow (0–1 m deep), but frequently rose to the ground surface after sustained rainfall. In addition, the ground water level observed in 2007 and 2016 at each site had not changed. Thus, the soil moisture content at 1 m depth at each site was saturated, or moistened through capillary rise throughout the year. Ground water at the Upper & Lower salt scald sites was extremely saline with EC values measured during March to September in 2007 of 2,760-2,960 mS m^− 1 and 1,930-2,920 mS m^− 1, respectively.

On the other hand, in the Arable area site the ground water table was estimated at about 5 m depth, which was estimated from the terrain and ground water table of monitoring bore about 300 m distant. Soon after establishment of the experiment, the soil moisture content at 1 m and 2 m depth was 10%, and ground water never reached the ground surface.

Soil salinity of each site was measured in 2006 (soon after establishment of the trees) and 2015 or 2017 from collected soil samples. Soil samples were collected at 10 cm intervals from the surface to 50 cm deep by a corer. Samples were taken from 5 points in the Arable site (center of site and center of four sides). The Upper & Lower salt scald sites both had a slight slope (less than 1 degrees), and samples were taken in a center line along this slope. Soil samples were collected from 6 points in the Upper salt scald site and 9 points in the Lower salt scald site. Collected samples were, air dried, and values of EC_1:5 (mS m^− 1) were measured according to the 1:5 (soil : water suspension) method (Rayment and Higginson, 1992). Measured values of EC_1:5 were transformed into ECe (mS m^− 1) using the following equation from George and Wren (1985):

ECe = (364/SP) × EC_1:5

where SP denotes saturation percentage of the soil. The observed porosity of the soil samples varied from 30–40% and thus 35% was used as a SP value in this study.

2.3. Reforestation experimentation

Tree seedlings were in June 2006 planted in Arable area site (0.15 ha), Upper salt scald site (0.21 ha) and Lower salt scald site (0.22 ha). Planted tree seedlings consisted of three tree species: Eucalyptus salubris F. Muell. (gimlet), E. camaldulensis Dehnh. (river gum) and E. sargentii Maiden (salt river gum). E. salubris is widespread across dry and semi-arid areas of south-western Australia (Gosper et al., 2013) and a native species of Avon Wheatbelt bio-region (FloraBase: https://florabase.dpaw.wa.gov.au/). Thus, E. salubris was regarded as a control species in this experimentation. E. camaldulensis is regarded as a primary candidate for remediating saline discharge areas in landscapes of low relief where shallow ground water has risen close to the soil surface under semi-arid conditions (Bush et al., 2013). In addition, with reports of its high saline and water-logging tolerance (Marcar et al., 1995) and reforestation success under highly saline conditions (more than 500 mS m^− 1 of EC_1:5 value) (Morris and Collopy, 1999), this species were selected although it was not a native species in this bio-region (FloraBase). E. sargentii was a native species in this bio-region (FloraBase), with a higher drought- and salt-tolerance than E. camaldulensis (Marcar et al., 1995) and reports of reforestation success in saline seep areas (Biddiscombe et al., 1985, 1989).

Effects of provenance and/or seedlot of elite trees of Eucalyptus species on seedling growth and survival under various conditions are well known and reported (e.g. Marcar et al., 2002; Zohar et al., 2010; Bush et al., 2013), however because of seedling availability from the local nursery, selected lines were not used in this experimentation.

In each reforestation site, four types of planting treatment were applied: direct planting, ripping + mounding, hole drilling and pipe setting. At right angle to the contours at each site, paralleled line blocks with a 2 m interval were set, and the four planting treatments were allocated to random line blocks. Each treatment was replicated 3 times. A set of three neighboring plants along a line block formed a species replicate with 2 m planting interval, and each replicate was allocated randomly within each line block. Among the four planting treatments, two (hole drilling and pipe setting) were excluded from further investigation because they failed.

Seedlings that died within one year after planting were replaced by new seedlings, which would be used for making allometric equations. Such replaced seedlings were excluded from further investigation (i.e., not in use for calculations of survival ratio and biomass).

2.4. Evaluation of tree survival and biomass

In September 2007, surviving trees were counted, and in September 2015, surviving trees were counted and their tree height and trunk diameter at 0.3 m height were measured. Percentage survival was calculated based on this tree counts.

For biomass calculation, allometric equations of each tree species were developed from destructive sampling of aerial parts cut from replaced trees or trees from other reforestation sites inside the catchment. Form of allometric equations was a power function with tree parameters D²H or H as independent variables and dry-biomass (W_s+b or W_l) as dependent variables (Table 1). D, H, W_s+b and W_l denotes diameter [m] at 0.3 m high, tree height [m], stem and branch biomass [kg] and leaf biomass [kg], respectively. From the 2015 tree measurements and allometric equations in Table 1, mean tree biomass [kg tree^− 1] was calculated. Equations of their independent variables as D²H (CAM1, CAM2, SAL1, SAL2, SAR1 and SAR2) were used for biomass calculation of large trees, and those as H (CAM3, CAM4, SAL3, SAL4, SAR3 and SAR4) were used for small trees.

By comparing significant difference of survival ratio and mean tree biomass among three experimental sites within each tree species, and when such data in Upper & Lower salt scald sites had no significant difference with those in Arable area site, such tree species meets objective of this study.

Tukey’s multiple comparison test of population ratio (α = 0.05) was applied to evaluate tree survival. This can test significant differences of survival values and then rank them.

For evaluation of biomass difference, firstly ANOVA (α = 0.05) was adopted and only after significant difference of mean tree biomass was found, Tukey’s multiple comparison test of population mean (α = 0.05) was used. If homoscedasticity of grouped data was denied by Bartlett test, Steel-Dwass multiple comparison method (nonparametric rank test) was applied instead of this Tukey’s test.

Details are described in section 3.2, but there was a case that only a single tree survived in the Lower salt scald site. In such case, biomass difference of that species in Arable area site and Upper salt scald site were tested by Student’s one side t-test (α = 0.05).

At last, among selected tree species and planting methods in Upper & Lower salt scald sites, each survival and mean tree biomass was compared by using above mentioned statistical analyses.

3.1 Site salinity

Figure 1 shows measured ECe values in each site, and changes of ECe from 2006 to 2015 or 2017. During the period of observation in this study, the Arable area site was free from salinity damage, on the other hand, salinity increased at both the Upper & Lower salt scald sites. At the start of the experiment, ECe values of both sites were from 400 to 800 mS m^− 1 categorized as moderate soil salinity (Marcar and Crawford, 2004). But after only about 10 years, the salinity of the Upper & Lower salt scald sites became severe and was categorized as high (800-1,600 mS m^− 1) and extreme (more than 1,600 mS m^− 1) soil salinity (Marcar and Crawford. 2004) in most measurement depth.

3.2. Selection of feasible tree species

Figure 2 shows mean tree biomass and survival ratio observed in September 2015 (9.25 years after planting) at the three reforestation sites. Figure 2A shows the results of the direct planting method and Fig. 2B the results of the ripping + mounding planting method.

There was no significant difference on survival ratio and mean tree biomass among Arable area site and Upper & Lower salt scald sites for E. salubris, planted as control, and E. sargentii, planted as promising species. On the other hand, E. camaldulensis, planted as primary candidate for secondary salinity measure, had significantly a lower survival ratio in the Lower salt scald site and lower mean tree biomass in the Upper & Lower salt scald sites. Both planting treatments (direct planting and ripping + mounding) had similar results. As a result, E. salubris and E. sargentii met the objective of this study.

3.3. Comparison of survival and biomass between selected tree species

Figure 3 compares the survival ratio and mean tree biomass for E. salubris and E. sargentii and planting methods in the Upper & Lower salt scald sites. At both sites, mean tree biomass of E. sargentii was significantly larger than that of E. salubris, but the survival ratio had no significant difference. In addition, there was no significant difference between planting methodologies.

Table 2 summarizes observed results of this study consisting of survival ratio of 2007 and 2015, mean tree biomass of 2015, stand biomass in 2015 [Mg ha^− 1], stand biomass increment during 9.25 years [Mg ha^− 1 year^− 1], initially planted tree number and initial stand density [trees ha^− 1].

Through this feasibility study, E. sargentii and E. salubris met the objective of this study, because these species showed statistically the same survival ratio and mean tree biomass between agricultural area (Arable area site) and saline discharge area (Upper & Lower salt scald sites) as shown in Fig. 2. Considering biomass productivity, in other word land use efficiency, E. sargentii planted by direct planting or the ripping + mounding method in marginal area of salt scald is the recommended reforestation methodology. Climatic condition (rainfall) during observation period of this study (from 2006 to 2015) were similar to that during the recent 30 years, and thus observed data in this study were were considered representative values of this region.

The experimental location of Upper & Lower salt scald sites in this study had quite severe saline and water-logged condition, and such condition was considered one of the severest among previous studies reported in south-western Australia as described in the following three points.

The first point was the occurrence of a quite shallow ground water table at around 1 m deep throughout a year. This was not limited to the experimental sites but was more widespread across at least more than half of the area. Such shallow ground water table (1 m deep) was rare case. For example, reported ground water depth by Archibald et al. (2006) and Biddiscombe et al. (1985, 1989) were more than 2 m deep, that by Zohar et al. (2010) varied from 1 to 4 m deep, and that by Bush et al. (2013) was from 1.5 to 4.5 m deep. In addition, this ground water frequently rose up to the ground surface (nearly zero) after consecutive rainfall in winter (during Jun. to Sep.), which is the common situation in this region.

Second point was extremely high EC value of ground water of Upper & Lower salt scald sites ranging from 2,000 to 3,000 mS m^− 1. Reported EC values of ground water was at most 2,000 mS m^− 1 in Boundain which is close to our research area (Wickepin). In other cases (Archibald et al., 2006; Bush et al., 2013) salinity was around 1,000 mS m^− 1 or less.

Third point was increased soil salinity strength during observation period (within 10 years). Judging from soil salinity criteria reported by Marcar and Crawford (2004), soil salinity of Upper & Lower salt scald sites increased from moderate to high and from moderate to extreme, respectively. Even severe situation, ECe value per se was not the worst condition, because other studies (e.g., Archibald et al., 2006; Harper et al., 2012) reported higher soil ECe values, however increasing salinity strength was not reported.

Because of increasing soil salinity and shallow water table with highly saline ground water as above described, root zone of Upper & Lower salt scald sites was highly damaged by salinity and water-logging. This was considered as reason why all tree species decreased their survival during 2007 to 2015 (Table 2).

Selected tree species in this study (E. sargentii and E. salubris) which could survive and grow under increasing soil salinity and severe saline water-logging stress were considered to have low failure risk. Considering failure risk of reforestation especially for carbon abatement projects is recently regarded as important (Nolan et al., 2018)

E. sargentii showed results in line with expectations that there was no significant difference on survival ratio and mean tree biomass among Arable area site and Upper & Lower salt scald sites (Fig. 2). In previous studies, availability of E. sargentii for secondary salinity measure was controversial. Stolte et al. (1997) stated this tree species was vulnerable to regional rises in water table, and Zohar et al. (2010) reported its salinity tolerance was variable to seed source (tolerant or unknown, but not sensitive). On the contrary, Biddiscombe et al. (1989) and Marcar et al. (1995) stated that this tree species was suitable for planting in saline seep area. Marcar et al. (1995) stated that its survival decreased when soil ECe exceeds 2,000 mS m^− 1, but Harper et al. (2012) reported good biomass growth (about 4 Mg ha^− 1 year^− 1 of 26 years average) under extremely high soil salinity condition (over 4,000 mS m^− 1). In this study, E. sargentii could survive and grow under increasing soil salinity strength (Fig. 1) with quite shallow saline ground water (0–1 m), and its biomass growth reached about 3–4 Mg ha^− 1 year^− 1 (Table 2), which was comparable to the value reported by Harper et al. (2012). We thus considered E. sargentii was not vulnerable to shallow ground water table or extreme soil salinity.

E. salubris was also considered a feasible tree species for reforestation of saline sites for the same reasons as E. sargentii (Fig. 2). There are no explicit reports of salinity or water-logging tolerance of E. salubris so far. When reforestation of saline sites is required to be conducted not as monoculture planting but as environmental planting consisting of mixture of native tree species (Paul et al., 2013), E. salubris was considered important. Environmental reforestation of saline sites has an important potential co-benefit of bio-diversity restoration (George et al., 2012), which may in time attract additional incentive payments (Maraseni and Cockfield, 2015; Andres et al., 2022). However, mean tree biomass of E. salubris was significantly lower than that of E. sargentii (Fig. 3), and this suggests the need to seek other native tree species which have similar biomass productivity between saline and non-saline land.

The potential of other tree species for reforestation of saline sites has been considered elsewhere. This includes E. occidentalis Endl. (Flat-topped Yate) reported by Archibald et al. (2006), Biddiscombe et al. (1989) and Sochacki et al. (2012), E. astringens (Maiden) Maiden (Brown Mallet) reported by Biddiscombe et al. (1989)d spathulata Hook. (Swamp Mallet) and E. salicola Brooker (Salt Gum) reported by Zohar et al. (2010). However, the growth of these species except E. occidentalis were reported based on not biomass but tree sizes (e.g. height or volume). Therefore, they should be reevaluated based on biomass [kg tree^− 1]. Because the relationship between size and biomass varies species by species and planting density, the same volume does not mean the same biomass. For example, E. sargentii, E. salubris and E. camaldulensis in this study had exactly the same tree size (5.0 m tree height and 0.10 m diameter at 0.3 m high), but quite different biomass calculated based on allometric equations (Table 1) of 24.7 kg tree^− 1, 23.6 kg tree^− 1 and 12.7 kg tree^− 1, respectively.

Among these potential candidates, E. occidentalis was considered the most promising because of following two reasons. One was that the biomass growth of E. occidentalis (around 4 Mg ha^− 1 year^− 1) at the Wickepin site (Sochacki et al., 2012) was comparable to that of E. sargentii in this study. The other was that at another experimental site with highly saline (> 4,000 mS m^− 1) conditions at Dryandra had similar growth rates (around 4 Mg^− 1 ha^− 1 year^− 1) for E. sargentii and E. occidentalis at 26 years of age (Harper et al., 2012), which proved their long-term survival and growth.

For the tree planting methodology, survival ratio and mean tree biomass of two tree species (E. sargentii and E. salubris) planted by direct planting method was not significantly different to those planted by the ripping + mounding method (Fig. 3), and thus from a cost-effectiveness perspective, direct planting was beneficial due to the reduced initial cost.

From this study, we found two tree species (E. sargentii and E. salubris) which had the same survival ratio and mean tree biomass under highly saline and water-logged condition as those in a non-saline arable area, and thus provides an option for the treatment of secondary salinity. Moreover, these two tree species maintained their high survival ratio and grew well even under severe soil salinity conditions and quite shallow saline ground water. In addition, because more than one tree species native in this bio-region were found feasible, reforestation of saline sites will have an opportunity to obtain bio-diversity restoration co-benefit as environmental planting.

For further studies in future, the transpiration and water use efficiency of selected tree species will be investigated to directly quantify their ability of decreasing saline ground water tables. And additional potential candidate tree species will be tested to determine whether they have comparable biomass productivity between saline and non-saline areas.

Acknowledgement

This research work was supported by the following four research funds. JSPS KAKENHI Grant number 16H02764, the Endeavour Research Fellowship 2015 program (Department of Education and Training, Australian Government), the Mitsui & Co., Ltd. Environment Fund and the Global Environment Research Fund of The Ministry of Environment of Japan (GHG-SSCP-S2-1a). Authors remark great contributions of Mr. John Law and Mr. Ken Martin for their continuous support to our research from site establishment to subsequent observations.

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Table 1　Allometric equations of each tree species

Table 2　Summary of observed data

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Minimizing the opportunity costs of carbon mitigation on farmland via restoration of salinized soils

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Abstract

Figures

1. Introduction

2. Materials And Methods

2.1. Research area

2.2. Reforestation sites

2.3. Reforestation experimentation

2.4. Evaluation of tree survival and biomass

3. Results

3.1 Site salinity

3.2. Selection of feasible tree species

3.3. Comparison of survival and biomass between selected tree species

4. Discussion

5. Concluding Remarks

Declarations

References

Tables

Additional Declarations

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