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The long-term Kervidy-Naizin Observatory: an ideal site for unraveling dissolved and colloidal phosphorus losses in agricultural catchments
A prerequisite for the implementation of mitigation strategies to reduce DP losses in the agricultural landscape is to identify sources within catchments, along with the mobilization mechanisms, transfer pathways, and their controlling factors. Excessive P (and nitrogen) in water bodies has been a regional challenge for water quality in Brittany, Western France, this region, representing a hotspot of intensive agriculture in France and Europe, and even in the world (Figure 1.8). Based on this situation, the Kervidy-Naizin catchment is monitored as a long-term field observatory (ORE AgrHys, https://www6.inra.fr/ore_agrhys) since 1993 to understand and model how the spatiotemporal variations in pedo-geology, hydroclimatology, landscape and agricultural activities control water and solute transport in this typical headwater catchment.
Figure 1.8. Maps of soil P surplus (a) and eutrophication risk (b) at the scale of Europe showing the unique, hotspot position of Brittany.
Because green algae blooms usually occur in coastal areas in Brittany, the research emphasis on the Kervidy-Naizin catchment was mainly nitrogen and dissolved organic carbon (DOC) until recently. However, 80% of the drinking water supply in Brittany comes from surface water reservoirs, which are frequently subjected to cyanobacteria blooms for which P is considered to be the limiting nutrient. The situation is similar for freshwater bathing sites in this region, which are regularly closed due to the proliferation of cyanobacteria in summer. For those reasons, the first systematic P study in this region was initiated in this catchment with the thesis of Dupas (2015) from 2013-2015, aiming at unraveling P transport processes and pathways at the catchment scale. Characterized by intensive agriculture (90% of Kervidy-Naizin land is dedicated to agriculture), very high animal densities (>35 000 pigs/5 km2), and high level of soil P content (up to 2.4 g/kg of total P) (Matos-Moreira et al., 2017), that study is rare even on the global scale. Before Dupas’s work, a 6-years’ of high-frequency monitoring of DP concentration at the outlet of the Kervidy-Naizin catchment indicated that DP is clearly a predominant part of P losses from this catchment during flood events (Figure 1.6). Other advantages of this catchment are that: i) it has been subject to ca. 20 years of numerous pedological, hydrological and geochemical studies (Aubert et al., 2013; Crave and Gascuel-Odoux, 1997; Curmi et al., 1998; Dia et al., 2000; Dupas et al., 2015 a, b, c; Durand and Torres, 1996; Humbert et al., 2015; Lambert et al., 2011; 2013; Mérot et al., 1995; Molénat et al., 2002, 2008; Morel et al., 2009), so that soil characteristics, groundwater and stream water dynamics, soil biogeochemical processes, and nitrate and DOC transfer pathways are well constrained, providing a unique framework for implementation of the P transfer study, ii) this site benefits from numerous water sampling equipment thanks to other projects on nitrogen and DOC undergoing at the same time with the present thesis, such as zero tension lysimeters, gauging station, automatic samplers at the catchment outlet for continuous stream discharge and stream water composition monitoring (as detailed in Dupas 2015; see Figure 1.9). These preliminary studies make this catchment a unique site to conduct a combined field and laboratory assessment of the release mechanisms of DP and controlling factors.
Figure 1.9. Sketch showing the main characteristics and equipment of the long-term observatory catchment of Kervidy-Naizin, Brittany, France.
General objectives and organization of the thesis
This thesis is based on the general context of understanding and reducing DP transfer in agricultural fields at the small catchment scale. This thesis estimated the DP release dynamics in soils in riparian wetlands (RWs) during baseflow periods, with the aim of identifying the mechanisms of DP (and colloidal P) release and their controlling factors. Field monitoring campaigns and laboratory simulation experiments are jointly conducted.
Groundwater control of biogeochemical processes causing phosphorus release from riparian wetlands
Because of the high sorption affinity of phosphorus (P) for the soil solid phase, mitigation options to reduce diffuse P transfer usually focus on trapping particulate P delivered via surface flow paths. Therefore, placing riparian buffers between croplands and watercourses has been promoted worldwide, sometimes in wetland areas. To investigate the risk of P-accumulating riparian wetlands (RWs) releasing dissolved P into streams, we monitored molybdate-reactive P (MRP) in the soil pore water of two RWs in an agricultural watershed. Two main mechanisms released MRP under the control of groundwater dynamics. First, soil rewetting after the dry summer period was associated with the presence of a pool of mobile P, limited in size. Its mobilization started under water saturated conditions caused by a rise in groundwater. Second, anoxic conditions at the end of winter caused reductive dissolution of Fe (hydr)oxides along with a release of MRP. Comparison of sites revealed that the first MRP release occurred only in RWs with P-enriched soils, whereas the second was observed even in RWs with low soil P status. Seasonal variations in stream MRP concentrations were similar to concentrations in RW soils. Hence, RWs can act as a key component of the P transfer continuum in agricultural landscapes by converting particulate P from croplands into MRP transferred to streams.
Introduction
In agricultural landscapes, riparian wetlands (RWs) are highly reactive biogeochemical interfaces located between croplands and watercourses (Vidon et al., 2010).
Due to their topographic position in valley bottoms, fuxes of sediments, nutrients and pesticides converge in these zones. Therefore, establishment of vegetated buffers has been promoted in riparian areas worldwide to reduce pollutant transfer to streams. Riparian wetlands have proved eective in sediment retention (Ockenden et al., 2014), denitrification (Anderson et al., 2014; Oehler et al., 2007) and microbial degradation of pesticides (Maillard and Imfeld, 2014). Their role in the phosphorus transfer continuum is less clear. Generally, RWs help decrease P delivery to watercourses by trapping particulate P and, to a lesser extent, sorbing dissolved P forms (Dorioz et al., 2006; Hoffmann et al., 2009).
However, RWs may also act as P sources for surface waters. This is due to the periodic water table fluctuations that affect these zones, which lead to a succession of dry periods and water-saturated periods (Obour et al., 2011; Song et al., 2007). Such unstable hydraulic conditions can increase P release from soil microbial biomass (Blackwell et al., 2010). Several laboratory studies have shown that soil rewetting after a dry period could lead to osmotic shock, causing microbial cell lysis and subsequent release of microbial P (e.g. Turner and Haygarth, 2001). According to Blackwell et al. (2009), up to ca. 70% of soil microbial biomass can be killed by osmotic shock caused by rewetting, with large variability in the amount of P released, depending on its recycling rate. Additionally, the periodic water table rises that affect RWs can modify the redox status of Fe (hydr)oxides, i.e. an important P sorbing compound in acidic soils (Li et al., 2012 ; Surridge et al., 2012). Anoxic conditions resulting from a high water table and low flow velocity can cause reductive dissolution of Fe (hydr)oxides in RW soils (Jeanneau et al., 2014 ; Knorr, 2013 ; Lambert et al., 2013 ; Li et al., 2012), which may cause release of previously adsorbed P (Carlyle and Hill, 2001 ; Hoffmann et al., 2009).
The presence of permanent vegetation in RWs can make biological P cycling in soils more intense than that in croplands as a result of more diversified plant and microbial communities (Roberts et al., 2012; Stutter et al., 2009). Permanent vegetation also results in high organic matter levels in RW soils, i.e. an important source of colloids (Haygarth et al., 2006), and provides favorable conditions for macropore formation. Hence, subsurface transfer of mobile P forms via preferential flow can be increased (Gachter et al., 1998; Haygarth et al., 1997). Although several laboratory experiments have highlighted mechanisms potentially involved in P solubilization and mobilization in RW soils (references above), a thorough literature search showed a lack of evidence regarding their contribution to diffuse P transfer in field conditions or the role of water table dynamics. In this study, we used zero-tension lysimeters to monitor molybdate-reactive P (MRP) concentrations in the free soil solution of two RWs in an intensively farmed watershed. The research questions addressed are: i) Does soil rewetting after a dry period and Fe (hydr)oxides reduction release MRP in RWs, and how are these production mechanisms linked to water table dynamics? ; ii) Can we relate MRP concentrations in streams with P-release mechanisms in RW soils? ; and, iii) Which soil factors control the spatial variability of P-release mechanisms in RWs?
Materials and Methods
Study sites
The monitored RWs were located in Kervidy-Naizin, a 5 km2 agricultural watershed belonging to the Agrhys environmental research observatory (http://www6.inra.fr/ ore_agrhys_eng) in Brittany, France (Aubert et al., 2013; Aubert et al., 2014). The Kervidy-Naizin watershed is drained by a stream of second Strahler order. Climate is temperate oceanic, with a mean annual (2007 – 2013) temperature of 10.6°C and annual rainfall of 867 mm. Lithology consists of impervious Brioverian schists capped by up to 30 m of unconsolidated weathered materials, in which a shallow aquifer develops. The schist contains mainly quartz, muscovite and chlorite and, to a lesser extent, K-feldspar and plagioclase (Pauwels et al., 1998). The weathered schist is not likely to be a source of MRP in the watershed as mean MRP concentration measured below the soil depth (1m) was 7 µg l-1 (unpublished results), which is lower than the long term mean baseflow concentration in the stream (18 µg l-1, Dupas et al., 2015b). Soils are silty loams, classiffed as Luvisols. They are well-drained in the upland domain and hydromorphic in valley bottoms, where RWs develop. Agricultural activities are dominated by arable crops (cereals, maize) and animal production (pigs, dairy cows). We confined investigations to two RWs at the footslope of two transects equipped with piezometers (Molenat et al., 2008) (Figure 2.1). Transect A ranged in elevation from 110 to 120 m a.s.l. with a mean slope of 3.8% (max=6%). The RW in transect A was 51 m wide, and its vegetation consisted of unfertilized herbaceous species (Dactylis glomerata and Agrostis canina). Winter barley (Hordeum vulgare) was grown in the adjacent field during the study period (October 2013 – May 2014). Soil P content (0 – 15 cm) in this field was 315 mg P kg-1 extractable P (Dyer method NF X 31-160) and 1283 mg P kg-1 total P (NF X 31-147). Pig slurry was applied on this field on April 9, 2014 (61 kg P ha-1). Transect B ranged in elevation from 104 to 109 m a.s.l. with a mean slope of 2.8% (max=4.5%). The RW in transect B was 64 m wide, and its vegetation consisted of shrubs and trees (Populus negro, Salix caprea, Betula alba). The adjacent field was left fallow during the study period, but maize (Zea mays) residues remained on the soil surface. Soil P content (0 – 15 cm) in this field was 157 mg P kg-1 extractable P (Dyer method NF X 31-160) and 1244 mg P kg-1 total P (NF X 31-147). Pig slurry was applied on this field on May 2, 2014 (52 kg P ha-1). Both RWs were managed as unfertilized riparian buffers for 20 years without exportation of biomass. Prior to conversion into a buffer-zone, RW A received up to 60 kg P ha-1 yr-1 as pig slurry and mineral fertilizer, whereas RW B received P input from grazing cattle (unquantified inputs).
Figure 2.1. a) map of the Kervidy-Naizin catchment ; b) location of the sampling sites (red dots) and variability in water table level along the two transects. Vertical lines represent piezometers (depth 3 – 8 m; screening 1.5 – 4 m).
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, 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%.
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
2.1 Introduction
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
2.4 Conclusion
2.5 Supplementary materials
2.6 Conclusion of chapter
Chapter 3 Release of dissolved phosphorus from riparian wetlands: Evidence for complex interactions among hydroclimate variability, topography and soil properties
3.1 Introduction
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 Results
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 Discussion
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
3.5 Conclusions
3.6 Supplementary materials
3.7 Conclusion of chapter
Chapter 4 Drying/rewetting cycles stimulate release of colloidal-bound phosphorus in riparian soils
4.1 Introduction
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 Results
4.3.1 P and DOC concentrations in RF samples
4.3.2 UF leachate results
4.4 Discussion
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
4.5 Conclusions
4.6 Supplementary materials
4.7 Conclusion of chapter
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
5.1 Introduction
5.2 Materials and methods
5.2.1 Sampling sites and soil preparation
5.2.2 Experimental setup
5.2.3 Analyses
5.3 Results
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 Discussion
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
5.5 Conclusions
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 Perspectives
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
