Soil and water sampling
Soil cores and soil pore water were sampled at two sites within each of the two RWs (WetUp and WetDown). In transect A, WetUp A was 13 m downslope from the wetland-field interface and WetDown A was 40 m downslope from the interface. In transect B, distances from the wetland-field interface were 9 m and 52 m for WetUp B and WetDown B, respectively. The aim of this sampling design was to investigate the variability of soil P content and water table level, and their effect on soil MRP concentrations, along the flowpaths between the upslope and downslope side of the RWs (Figure 2.1).
Soil cores in the 0 – 15 cm and 25 – 40 cm horizons were collected with a 75 mm diameter sampler at each of the four sites in April 2014. WetDown B was only sampled in the surface due to water saturated soils below 15 cm. Soil pore water was collected with zero-tension lysimeters placed in triplicates (spaced ca. 1 m apart) at 10 – 15 cm and 50 – 55 cm depths, i.e. in the same soil horizons as the soil cores. The lysimeters were designed to collect free soil solution while maintaining in-situ anoxic conditions (Figure 2.2). After an equilibration period of three months, the 24 lysimeters were sampled weekly from October 2013 to January 2014 and every two weeks from February 2014 to May 2014. From June to October, soil moisture was too low to collect soil solution. Grab samples of stream water were collected at the same frequency downstream of the monitored RWs, and daily at the outlet of the 5 km2 watershed.
Because we were interested in subsurface transfer, we focused analyses on baseflow concentrations by discarding grab samples collected during storms (surface runoff may contribute to storm flow). We considered storms as events with > 10% discharge rise and > 20 l s-1 discharge (Dupas et al., 2015b). All samples were filtered (< 0.45 µm cellulose acetate filter) within 6 h after collection and kept refrigerated until analysis within 3 days.
Soil and water chemical analyses
Soil samples were air-dried, sieved to < 2 mm and analyzed for particle size fractions (NF X 31-107), organic matter/nitrogen/carbon contents (NF ISO 13878, NF ISO 10694), pH in water (1 :5 v :v water extraction NF ISO 10390) , extractable P (Dyer method, i.e. 1 :5 w/v extraction with citric acid 20 g l-1 NF X 31-160), total P (ICP-AES after total solubilization with hydrofluoric and perchloric acid NF X 31-147), Al and Fe (ICPAES after extraction with ammonium oxalate and oxalic acid, according to Tamm 1922) (Table 2.1). Equilibrium P concentration (EPCo) and maximum sorption capacity (Qmax) were estimated from 6-point batch isotherms (0, 0.1, 0.5, 50, 100, 200 mg P l-1; 1:25 w:v) in 0.01M CaCl2 according to Graetz and Nair (2000). One drop of chloroform was added to inhibit microbial activity. After 24 h equilibration at 20±2°C, samples were centrifuged (3000 rpm; 10 min), filtered (< 0.45 µm) and analyzed for MRP. Qmax was determined by fitting a Langmuir equation (Van der Zee and Bolt, 2001) to the last three points (50, 100, 200 mg P l-1): Q = (c * K * Qmax) / (1 + K * c).
where c is the concentration of P in the equilibrium solution (mg l-1), Q is the total amount of P sorbed (mg mg-1) and K is an a-nity parameter (l mg-1). EPCo represents the solution P concentration at which no net sorption or desorption of P would occur between soil and solution (Stutter and Lumsdon, 2008). EPCo was determined by fitting a linear equation to the first three points (0, 0.1, 0.5 mg P l-1). We consider EPCo as a reference MRP concentration in the soil solution, which we can compare to the actual MRP concentration of soil solution collected in-situ. Qmax served to calculate “Degree of P Saturation” (DPS), defined here as the ratio of Extractable P to Qmax. DPS is an index of P accumulation in the soil, either via direct application of fertilizers or enrichment via erosion (Schoumans and Chardon, 2015).
For each water sample collected in lysimeters or in the stream, MRP was determined colorimetrically by reaction with ammonium molybdate (ISO 15681). Because filtrates < 0.45 µm can contain colloidal forms of molybdate reactive phosphorus, we chose to use the term MRP rather than soluble reactive phosphorus (Haygarth and Sharpley, 2000). Precision of MRP measurement was ±4 µg l-1. Fe2+ was analyzed using the 1.10 phenantroline colorimetric method, according to AFNOR NF T90-017, with a precision of 5%. Nitrate concentration was measured by ionic chromatography (DIONEX DX 100), with a precision of 2.5%.
Soil P content and water table depths in riparian wetlands
Soil total P content at WetUp A and B was 13% and 33% higher than that at WetDown A and B, respectively (Figure 2.3). A probable explanation for the higher soil P levels on the upslope side of the RWs is that P delivery from the adjacent fields has enriched RWs in P (Ockenden et al., 2014). Previous studies in the Kervidy-Naizin watershed have indeed evidenced that its loamy soils are vulnerable to erosion, which leads to spatial redistribution of soil and nutrients across the landscape (Le Bissonnais et al., 2002). Longterm monitoring in a larger number of sites would be necessary to quantify the importance of P accumulation in RWs at the watershed scale. Such a monitoring is however beyond the scope of this study, which focuses on the role of groundwater dynamics in P remobilization in RWs.
Linking MRP concentration in riparian wetlands and in the stream
In-situ observations of soil solution MRP concentrations can be directly linked to stream concentrations. We used stream MRP concentration measured downstream of the two RWs and at the watershed outlet. This combined soil and stream observation could provide a complete view of the solubilization-mobilization-transport continuum from RWs soil to surface water (Haygarth et al., 2005; Haygarth et al., 2012; Mellander et al., 2012a).
The water table rise in December – January was accompanied by a sudden increase in discharge (Figure 2.6). Stream MRP concentrations also increased downstream of both RW A and B and at the watershed outlet. They were higher downstream of RW A than RW B, similar to differences observed at the same time in the soil solution (Figures 2.4 and 2.5). Stream MRP concentrations then decreased, probably as a result of exhaustion of the previously mobilized P pool in wetland soils. They increased again in February – March downstream of RW A and in April – May downstream of RW B. Again, the timing of these MRP peaks and their relative amplitude between RW A and B was consistent with observations of P release caused by reductive dissolution of Fe (hydr)oxides in wetland soils. Hence, variation in stream MRP concentrations could be related to the two successive P-release mechanisms identified in RWs. Despite this synchronization of peaks between wetland soils and the stream, differences in the relative amplitude of the peaks suggest that retention processes occurred: for example, the second peak was not as large as the first one in the stream downstream of RW A, whereas the amplitude of P release within the soil was similar during the first and second peak. This suggests that some of the MRP released in wetland soils via Fe (hydr)oxides reductive dissolution was not transferred to the stream, probably because of Fe2+ re-oxidation on its way to the stream and re-adsorption of P on the newly formed Fe (hydr)oxides (Baken et al., 2015; van der Grift et al., 2014).
Soil and water sampling
In April 2015, a 75-mm-diameter auger was used to collect soil samples in RWA and RWB at depths comparable to those of the lysimeters (i.e. 0-15 cm (“shallow”) and 30-40 cm (“deep”)) to determine P speciation in soils from these RWs. In RWA, soil samples were taken at WetUp-A and WetDown-A; in RWB, only the shallow soil (0-15 cm) at WetUp-B could be sampled due to the extreme moisture conditions in this RW at the time. Soil water samples were collected weekly to biweekly from October 2013 to June 2016. On each sampling date, samples from the triplicate lysimeters from each given depth and location were mixed to increase the amount of soil water available for chemical analysis.
Stream water samples were collected on the same dates as the soil water sampling in the stream reaches adjacent to the two RWs, and daily at the catchment outlet. Because this study focused on the release of P in RW soils and the subsequent transfer of this P to the stream through groundwater flow, samples collected during storms were removed from data analysis to avoid the potential contribution of overland flow to the stream’s dissolved P budget (included when calculating annual TDP flux). Note that the storm periods are also important concerning P release from RWs, due to the high hydraulic gradient between RWs and streams which leads to strong exchange of water and solutes. Storm events were identified when discharge exceeded 20 l s-1 and when increases in discharge over a period of 10 min were greater than 10% (Dupas et al., 2015b). All water samples were passed through <0.45 µm cellulose acetate filters within 6 h after sampling, then kept in a refrigerator in the dark at 4°C until analysis within 3 days.
Soil and water chemical analysis
Soil samples were transferred to the laboratory in plastic bags inside a cool box, then homogenized by wet sieving to <5 mm. Soil P speciation was determined using the sequential fractionation method developed by Hedley (1982) and modified by Ivanoff et al. (1998). This method is based on sequential P extraction with different reagents (H2O, 0.5 M NaHCO3 at pH 8.5, 1 M HCl, and 0.1 M NaOH). P was extracted by shaking soil suspensions (1 g of dry weight soil in 30 ml solution) for 16-21 h. After centrifugation at 10 000 RPM for 10 min, the supernatant was removed and analyzed for TDP and MRDP. P extracted by H2O and NaHCO3 is considered labile P, while HCl-and NaOH- extracted P is defined as moderately labile P. Residual (i.e. non-labile) P, which may contain a mixture of organic and inorganic P forms, was determined as the total P measured after calcination at 550°C and acid digestion of the residue after the final extraction step. Extraction efficiency was checked by comparing the sum of P extracted using the sequential extraction method with the total P content measured after calcination and acid digestion of the bulk soil samples. All sequential experiments were performed in triplicate.
For all water samples, MRDP was determined colorimetrically via reaction with ammonium molybdate (Murphy and Rile filtrates (natural water) or to the centrifuged solutions (soil extracts). The same method was used for TDP, but after digestion of the filtrates or centrifuged solution in acidic potassium persulfate. The precision of MRDP and TDP measurements was ±4 µg l-1 and ±13 µg l-1, respectively. The difference between TDP and MRDP was assumed to be organic P for soil extracts and MUDP for natural water samples. Development of reducing conditions in soil water was monitored by measuring Fe2+ concentrations using the 1.10 phenanthroline colorimetric method (AFNOR NF T90-017, 1997), with a precision of 5%. Fe2+ reactants were mixed with samples IN SITU to avoid the influence of oxidation on Fe2+ concentration during sample transport. The potential influence of Fe oxyhydroxide precipitation on MRDP concentrations during sample transport was checked by comparing MRDP concentrations in sample aliquots directly filtrated and acidified on-site to MRDP concentrations in sample aliquots filtrated in the laboratory. The results were found to be comparable with average differences below 12%, suggesting that potential Fe2+ oxidation had little influence on MRDP concentrations in the <0.45 µm filtrates (Figure S1).
Table of contents :
Chapter 1- General Introduction
1.1 The global story of phosphorus
1.1.1 Why do we study phosphorus?
1.1.2 The global challenge of P resources
1.1.3 The P dilemma: too much vs. too little
1.2 P losses to waters: point vs. diffuse sources
1.3 Knowledge and research trends on diffuse P loss processes
1.3.1 The need for an integrated catchment approach
1.3.2 Some key concepts about P mobilization processes in agricultural catchments
1.3.3 Are dissolved and colloidal P species major components of diffuse P losses in agricultural landscapes?
1.3.4 Riparian buffer zones as potential sources of dissolved and colloidal P agricultural catchments
1.4 The long-term Kervidy-Naizin observatory: an ideal site for unravelling
dissolved and colloidal phosphorus losses in agricultural catchments
1.5 General objectives and organization of the thesis
Chapter 2 Groundwater control of biogeochemical processes causing phosphorus release from riparian wetlands
2.2 Materials and Methods
2.2.1 Study sites
2.2.2 Soil and water sampling
2.2.3 Soil and water chemical analyses
2.3 Results and discussion
2.3.1 Soil P content and water table depths in riparian wetlands
2.3.2 Groundwater level controls P release in riparian wetlands
2.3.3 Linking MRP concentration in riparian wetlands and in the stream
2.5 Supplementary materials
2.6 Conclusion of chapter 2
Chapter 3 Release of dissolved phosphorus from riparian wetlands: Evidence for complex interactions among hydroclimate variability, topography and soil properties
3.2 Materials and Methods
3.2.1 Research site
3.2.2 Soil and water sampling
3.2.3 Soil and water chemical analysis
3.3.1 Hedley P fractionation
3.3.2 Rainfall, discharge and water-table variations
3.3.3 Soil water chemistry
3.3.4 Stream water chemistry
3.4.1 Influence of soil P content and soil P speciation
3.4.2 Key influence of interannual hydroclimatic variability on P release dynamics
3.4.3 Topography as the potential ultimate driver of dissolved P release in RW soils
3.4.4 Delivery and retention of mobilized P
3.6 Supplementary materials
3.7 Conclusion of chapter 3
Chapter 4 Drying/rewetting cycles stimulate release of colloidal-bound phosphorus in riparian soils
4.2 Materials and Methods
4.2.1 Soil properties and preparation
4.2.2 Experimental setup and conduct of DRW experiments
4.2.3 Leachate treatments
4.2.4 Chemical analysis
4.2.5 UF data treatment
4.2.6 Statistical analysis
4.3.1 P and DOC concentrations in RF samples
4.3.2 UF leachate results
4.4.1 Soil rewetting stimulates release of colloidal P
4.4.2 Co-existence of physically- and biologically-driven P release during rewetting
4.4.3 Influence of soil properties
4.4.4 Linking sources and production mechanisms of P forms released during soil rewetting
4.4.5 Environmental and ecological implications
4.6 Supplementary materials
4.7 Conclusion of chapter 4
Chapter 5 Release of dissolved phosphorus upon reduction of wetland soils: a laboratory study of the respective roles of soil Fe-oxyhydroxides dissolution, pH changes, sediment inputs and soil phosphorus speciation
5.2 Materials and methods
5.2.1 Sampling sites and soil preparation
5.2.2 Experimental setup
5.3.1 Soil/sediment composition
5.3.2 Anaerobic incubations of RW soils
5.3.3 Aerobic incubations of RW soils
5.3.4 Anaerobic incubations of sediment with and without RW soil addition
5.4.1 Controls of soil properties on concentration and speciation of released DP
5.4.2 Assessing the respective roles of reductive dissolution of Fe-oxyhydroxides and pH rise
5.4.3 Influence of sediment deposition in RWs on DP release under anoxic conditions
5.6 Supplementary materials
5.7 Conclusion of chapter 5 (laboratory simulation of reduction processes) .
Chapter 6 General conclusions
6.1 Recall of thesis objectives
6.2 Summary of conclusions
6.2.1 Constraints from field monitoring on the mechanisms and factors causing DP releases in riparian wetlands
6.2.2 Highlighting how drying-wetting cycles stimulate the release of colloidal P in wetland using column leaching experiments
6.2.3 Constraints on the processes releasing P under anoxic conditions .
6.3 Possible implications for management
6.4.1 Nature, source and significance of organic P fraction
6.4.2 Towards a better characterization of colloid composition and colloid properties regarding P transfer in soils
6.4.3 Test the generality of the conceptual model developed from the Kervidy- Naizin catchment
Chapter 7 General references