ACTA UNIVERSITATIS AGRICULTURAE SUECIAE
AGRARIA 164
Quantifying Macropore Flow Effects on Nitrate and Pesticide Leaching in a Structured Clay Soil
Field experiments and modelling with the MACRO and SOILN models
Martin Larsson
SWEDISH UNIVERSITY OF AGRICULTURAL SCIENCES Quantifying Macropore Flow Effects on Nitrate and Pesticide Leaching in a Structured Clay Soil
Field experiments and modelling with the MACRO and SOILN models
Martin Larsson Department of Soil Sciences Uppsala
Doctoral thesis Swedish University of Agricultural Sciences Uppsala 1999 Acta Universitatis Agriculturae Sueciae Agraria 164
ISSN 1401-6249 ISBN 91-576-5489-1 © 1999 Martin Larsson, Uppsala Tryck: SLU Service/Repro, Uppsala 1999 Abstract
Larsson, M. 1999. Quantifying macropore flow effects on nitrate and pesticide leaching in a structured clay soil. Field experiments and modelling with the MACRO and SOILN models. Doctoral thesis. ISSN 1401-6249, ISBN 91-576-5489-1
Rapid non-equilibrium flow of water in soil macropores is one significant, yet poorly understood, process controlling solute transport in soils. To protect surface waters and groundwater from pollution of agro-chemicals, it is important to improve our understanding of the effects of macropore flow on solute leaching. In this study, the effect of macropore flow was quantified for nitrate and pesticides on a structured clay soil in south-west Sweden, using two models, the dual-porosity, dual-permeability model MACRO and the nitrogen turnover model SOILN. This was done by first calibrating the MACRO model against extensive field measurements of tile drain flow, soil moisture - contents, and Br tracer (drainage concentrations and amounts in the soil). Model simulations with MACRO and the coupled models MACRO-SOILN were then performed with and without macropore flow and compared with measurements of solute content in the soil profile and solute concentrations in drain discharge of NO3-N and a weakly- sorbed herbicide, bentazone. Ten-year simulations with 60 hypothetical compounds were also conducted to evaluate the effect of macropore flow on pesticides with widely different sorption and degradation properties.
The results showed that macropore flow reduced total leaching of bentazone to tile drains by as much as 50 % for the one-year experimental period. This is because much of the bentazone was stored in the micropores, moving at a ‘reduced’ velocity, ‘protected’ against bypass of water in the macropores. However, the scenario simulations indicated that macropore flow considerably increases leaching for less mobile pesticides, since this is the dominant leaching mechanism. The largest increase, up to five orders of magnitude, was found for moderately to strongly sorbed pesticides with half-lives <10 days. For most pesticides, macropore flow also significantly reduced the influence of compound properties on leaching. For nitrate, the simulated effect of macropore flow was a reduction in leaching by 28 % from June 1990 to July 1995, but due to the influence of climatic conditions, large variations were found between years (from 3 to 45 %).
Without taking macropore flow into consideration, the model could not depict the observed leaching pattern of bromide and bentazone. It was also impossible to match all the measured components of the nitrogen mass balance without accounting for macropore flow. Even though there are many uncertainties in the modelling approach, it is concluded that macropore flow models could be useful in the evaluation of alternative soil and crop management practices to minimize adverse impacts of non-point source pollution on water quality.
Key words: leaching, pesticide, nitrate, field-scale, macropore, model, MACRO, SOILN.
Author’s address: Martin Larsson, Department of Soil Sciences, P.O. Box 7072, SLU, SE-75007 Uppsala, Sweden. E-mail: [email protected] Preface
This thesis is based on the following papers, which are referred to in the text by their Roman numerals.
I. Larsson, M.H. & Jarvis, N.J. 1999. Evaluation of a dual-porosity model to predict field-scale solute transport in a macroporous soil. Journal of Hydrology, 215:153-171.
II. Larsson, M.H. & Jarvis, N.J. 2000. Quantifying interactions between compound properties and macropore flow effects on pesticide leaching. Pest Management Science, 56:133-141.
III. Larsson, M.H. & Jarvis, N.J. 1999. A dual-porosity model to quantify macropore flow effects on nitrate leaching. Journal of Environmental Quality, 28:1298-1307. Contents
Introduction, 7
Literature review, 7 Macropores, macropore flow and transport, 7 Experimental evidence of macropore flow and transport, 8 Extent of macropore flow, 9 Factors influencing macropore flow effects on leaching, 10 Quantifying macropore flow effects, 13 Field evaluation of macropore flow models, 13
Aims and objectives, 15
Materials and methods, 15 The models, 15 Description of the field experiments, 19 Modelling strategy and parameterization, 21
Results and discussion, 21 Effects of macropore flow on hydrology, 21 Effects of macropore flow on bromide leaching, 22 Effects of macropore flow on pesticide leaching, 22 Effects of macropore flow on nitrate leaching, 23 Uncertainties and limitations in the modelling approach, 24
Conclusions and recommendations, 27
References, 29
Acknowledgements, 34 Introduction
Macropore flow is the process by which infiltrating water rapidly moves downwards in structural pore spaces such as shrinkage cracks, worm channels, and root holes and thereby bypasses a part of the soil profile. Macropore flow may dramatically influence the transport of solutes (Thomas & Phillips, 1979) since the buffering capacity of the chemically and biologically reactive topsoil is bypassed. Consequently, the risk of leaching of surface-applied agrochemicals (e.g. pesticides and nitrogen) to surface waters and groundwater may increase. It is therefore of great importance to fully understand the effects of macropore flow on leaching to protect valuable water resources against non-point source pollution by agrochemicals. Only when transport through macropores can be accurately quantified and predicted, can the most efficient countermeasures be taken to minimize leaching from structured soils.
Many models have been developed in recent years to quantify and predict non- point source leaching of nitrate and/or pesticides to surface waters and groundwater, and some of these models now include a description of macropore flow (Jarvis, 1998). However, few macropore flow models have been thoroughly tested under natural field-scale conditions. The purpose of this thesis is to (i.) evaluate one widely-used macropore flow model MACRO (Jarvis, 1994; Jarvis & Larsson, 1998), and (ii.) quantify the effects of macropore flow on leaching of pesticides and nitrate by applying the model to measurements made in two comprehensive field-scale experiments in a structured clay soil in south-west Sweden.
Literature review
The aim of this brief literature review is to (i.) discuss definitions and concepts related to macropores and macropore flow (ii.) highlight some important factors governing solute leaching in macroporous soil, and (iii.) discuss methods to quantify macropore flow effects on solute leaching.
Macropores, macropore flow and transport
Luxmoore (1981) proposed a division of the soil pore system into three size classes: micro-, meso-, and macropores, based on fixed values of soil water pressure head (or equivalent pore size). Such an approach may be useful for some pragmatic purposes, such as in numerical modelling of non-equilibrium flow and transport processes, but the use of well-defined static terms representing fixed
7 physical limits may be unnecessarily restrictive. Beven (1981), Bouma (1981), and Skopp (1981) all argued that the definition of macropores should instead be based on measurements of transport characteristics or the hydraulic conductivity function rather than fixed physical limits. Based on these ideas, Othmer et al. (1991), Durner (1992), Wilson (1992), and Jarvis (1999) all demonstrated the existence of dual or multiple pore systems in soil by curve-fitting bimodal or multi-modal hydraulic functions to measurements of water retention and/or soil hydraulic conductivity. Many dual- and multi-domain flow and transport models have also been developed in recent years based on these concepts (Jarvis, 1998).
Skopp (1981) emphasised the transport processes occurring in the soil and introduced a functional definition of the soil pore system on this basis. He divided the pore space into macropores and matrix pores, where macropores are pores that provide “preferential paths of flow so that mixing and transfer between such pores and remaining pores is limited”, while matrix pores “transmit water and solute at a rate slow enough to result in extensive mixing.. ..of molecules between different pores”. As pointed out by Wilson et al. (1998) and Jarvis (1998), it is important in this respect to distinguish between macropore flow and macropore transport. Macropore flow can be defined in terms of non-equilibrium in hydraulic pressure between pore regions, while macropore transport can be described as non- equilibrium in solute concentrations between pore regions. Consequently, equilibrium water flow (e.g. in saturated soil) can cause non-equilibrium transport.
Experimental evidence of macropore flow and transport
The first experimental observations of macropore flow were made more than 100 years ago by Schumacher (1864) and Lawes et al. (1882). Nevertheless, as a whole, the research community did not recognize macropores as an important pathway for water flow and solute transport until recently. Only in recent years has the occurrence of macropore flow been studied more systematically, due to concerns arising from increasing concentrations of nitrate in groundwater and the detection of pesticides in surface waters and groundwater.
Even if some cases of pesticide detections in surface waters and groundwater soon after application can be attributed to isolated spills or other point sources, many findings are a result of diffuse application on arable land according to ‘Good Agricultural Practice’. A strong indication of macropore flow is when surface- applied solutes are detected in tile-drain discharge water soon after application and after only small amounts of precipitation (i.e. a precipitation amount that is much smaller than the storage capacity of the soil above the drainage depth). In this way, macropore transport of pesticides and nitrate has been demonstrated in several field experiments (e.g. Kladivko et al., 1991; Harris et al., 1994; Traub- Eberhard et al., 1994; Brown et al., 1995; Johnson et al., 1995). Undisturbed
8 lysimeters are also commonly used to study solute leaching under field-like conditions. Early solute breakthrough in lysimeter drainage water due to macropore flow has been reported for many soils and under various conditions (e.g. Kissel et al., 1974; Tyler & Thomas, 1977; Bergström & Jarvis, 1993; Baker & Timmons, 1994).
Another technique used to demonstrate the presence of macropores and macropore flow is the application of a dye solution to stain the flow paths in the soil during infiltration (Bouma & Dekker, 1978). By excavating the soil, the flow paths can then be detected by visual inspection. In a dye-tracing experiment at fourteen different field sites in Switzerland with a wide variety of soils, Flury et al. (1994) found bypass flow in most of these soils, and suggested that the principal cause was flow in macropores. In eight of the soils, dye was found below 80 cm depth after application of 40 mm of solution. Bergström & Shiromhammadi (1998) applied 35 mm of a dye solution to two lysimeters with Lanna clay soil at a rate of 15 mm h-1. Excavation of horizontal cross-sections showed that ca. 10 % of the area was stained at 20 cm depth and ca. 4 % at 60 cm depth. However, dye-tracing experiments are often conducted under ponded conditions or with high application rates and may therefore exaggerate the importance of macropore flow. For example, examining the rain intensities measured at Lanna between October 1994 and December 1995, it was found that there were only three occasions when more than 10 mm of rain fell during a 2 hour period, while the maximum amount recorded in any 2 hour period during these 15 months was not more than 18 mm.
Indirect evidence of macropore flow can also be obtained by model interpretation of data from field-scale leaching experiments. In some studies where models have been used which do not account for macropore flow, the mismatch between simulated and measured solute leaching has been attributed to macropore flow (e.g. Bergström & Jarvis, 1991; Jabro et al., 1994; Lengnick & Fox, 1994).
Extent of macropore flow
Macropore flow is of most importance in soils with well-developed structure, but may also occur along roots or through earthworm burrows in weakly-structured or structureless coarse-textured soils. Under normal field conditions without excessive irrigation or application amounts, pesticide leaching due to macropore flow has been indicated in soils with clay contents as small as 12 % (Kladivko et al., 1991). To get an idea of the extent of soils where macropore flow is likely to constitute an important transport pathway, Table 1 shows, for several countries, the proportion of agricultural soils with a clay content larger than 20 %. In the listed countries, about half of the arable land has clay contents of more than 20 %, which indicates that macropore flow is certainly not an exceptional or unusual phenomenon.
9 Table 1. The proportion of agricultural land with a clay content larger than 20%
Country Percent agricultural area with a clay content > 20 %
England and Wales a 53 % Finland b 50 % France c 44 % Germany d 52 % Sweden e 52 % a from the SEISMIC database, Hollis et al. (1993). b Puustinen (1994). c INRA http://viviane.roazhon.inra.fr/snas/france/argile/histmed.gif. d Not including vineyards/horticulture (CORINE land cover data set, 1:1000000 Soil map of Germany). e Eriksson et al. (1999).
Factors influencing macropore flow effects on leaching
Type of solute Although macropore flow influences the pattern of leaching of conservative - - surface-applied tracers such as Cl and Br , including the maximum tracer concentration found in groundwater or surface water (e.g. Kissel et al., 1974; Bronswijk et al., 1995), the long-term total leaching loss must be unaffected since sooner or later all the solute must be washed out from the soil regardless of the transport mechanism. In contrast, leaching of soil-indigenous conservative solutes (e.g. soil salts in dry climates where the evaporation is high and the net downward movement of water is small) may be significantly reduced due to macropore flow (e.g. Miller et al., 1965; Thomas & Phillips, 1979). This is because the rain and/or irrigation water has a smaller solute concentration than the resident soil water, and the mixing between the water infiltrating in the macropores and the resident water in the micropores is restricted, leaving the solute behind in the micropores (Tyler & Thomas, 1977).
Macropore flow should normally increase leaching of pesticides, since they are degradable compounds that are surface-applied and foreign to the soil. Indeed, it is sometimes speculated that the effect of macropore flow on leaching may completely overshadow the influence of the properties of the particular compound (Klein, 1994; Barbash & Resek, 1996). The results of some field studies with compounds of different sorption characteristics do suggest a reduced influence of
10 pesticide properties due to macropore flow (Kladivko et al., 1991; Traub- Eberhard et al., 1994; Brown et al., 1995). In these cited studies, pesticides with widely different sorption properties were found at the same time in the first major drainage event after application. However, the rank-order in total leaching followed the rank order in sorption, indicating that an influence of the compound properties still exists. In a limited desktop study with the MACRO model, (Jarvis, 1995) suggested that macropore flow would increase total leaching most for otherwise ‘non-leachable’ compounds, while ‘leachable’ compounds would be less affected by macropore transport.
The effects of macropore flow on nitrate leaching are complex and not well understood since nitrate is both produced within the soil by biological transformation processes and is also normally applied at the surface as manure or mineral fertilizer (White, 1985). In experiments with surface application of mineral nitrogen, dramatic increases in nitrate concentrations in lysimeter outflow and tile drain discharge water have been found as a result of macropore flow (e.g. Kissel et al., 1974; Baker & Timmons, 1994; Johnson et al., 1995). In a 4-year field study on a heavy clay soil with surface-applied mineral fertilizer (150 kg N ha-1), Booltink (1995) also suggested that macropore flow increased total nitrate leaching to groundwater. In contrast, Thomas & Phillips (1979) proposed that macropore flow could reduce total nitrate leaching. A decrease in nitrate leaching due to macropore flow was also indicated by Lengnick & Fox (1994) who found that the NCSWAP model, which does not account for macropore flow, overestimated transport by more than 37 % in a 3-year field experiment with surface-applied mineral fertilizer (200 kg N ha-1). This overestimation was attributed to the inability of the model to account for bypass flow of infiltrating rain water of low nitrate concentration. Studies of macropore flow effects on nitrate transport have often focused on short periods after fertilizer application (e.g. Priebe & Blackmer, 1989). It is therefore possible that demonstrated increases in nitrate leaching due to macropore flow are exaggerated since most of the annual leaching loss normally occurs in the winter, when macropore flow may decrease leaching of nitrogen mineralized from organic matter.
Application time There are two important aspects concerning the effects of application timing on leaching due to macropore flow, (i.) the precipitation amount and pattern shortly after application, and (ii.) the soil moisture content at the time of application. The initiation of macropore flow and transport may be particularly important close to the soil surface as a consequence of infiltration of rain or irrigation water (Beven & Germann, 1982). When the rain or irrigation intensity exceeds the infiltration capacity of the micropores, the water at the soil surface will be routed into the macropores. This process may involve small-scale ‘surface runoff’ with extensive mixing of the infiltrating water and the water stored in the uppermost soil layer (Steenhuis & Walter, 1980; Ahuja, 1986). If surface-applied, the concentration of the solute will be at its maximum at the soil surface directly after application,
11 implying a potential for macropore transport of significant amounts of the solute. This can be illustrated by the results from a field study on a loamy soil, where the highest atrazine concentration in groundwater (40 µg l-1) was found only 6 days after application, after 48 mm of rainfall during a two day period starting 12 hours after application (Isensee et al., 1990). In the preceding year, there were only small amounts of precipitation after application, and the atrazine concentration in groundwater was always smaller than 1 µg l-1 (Isensee et al., 1990). Thus, once surface-applied solutes are transported into the micropores below the ‘mixing layer’ at the surface, they are effectively ‘protected’ against leaching of water bypassing in the macropores, since the volume of macropores is generally small in comparison to the volume of micropores, and lateral diffusion rates are slow. This ‘protective’ effect of the micropores was also demonstrated in a study with small lysimeters irrigated soon after solute application (Shipitalo et al., 1990). Leaching was considerably reduced by applying 5 mm of water during a one hour period, two days prior to irrigating with 30 mm of water at an intensity of 60 mm h-1. Compared to a control treatment, the pre-wetting treatment reduced transport - of Br by 7 times and atrazine by 2 times (Shipitalo et al., 1990). In a field study on a heavy clay soil, Brown et al. (1995) concluded that even though the total losses of pesticides of widely different properties were largely determined by the total amount of precipitation in any year, the first rainfall event after application was disproportionally important. They also found that the effect of the first rain event was most important for rapidly-degrading pesticides.
The effect of soil moisture content at the time of application on leaching through macropores is not fully understood. In a field study with fourteen different soils, Flury et al. (1994) could not find any pronounced effect of initial water content on bypass flow using a weakly-sorbed dye. In a lysimeter study with surface application of two pesticides and two conservative tracers, Shipitalo & Edwards (1996) found higher leachate concentrations of the pesticides after the first storm following application when the compounds were applied to dry soil. However, total leaching was not significantly different since the amount of percolating water was considerably higher from the initially wetter soil. In two years with very different soil moisture contents at the time of pesticide application, Brown et al. (1995) detected pesticides in both years in drain discharge water soon after application due to bypass flow in macropores. However, drier soil conditions at the time of application resulted in somewhat larger pesticide concentrations. Larger total leaching losses from dry soil conditions at the time of application have also been demonstrated (e.g. White et al., 1986). However, some experimental results point in the opposite direction (Coles & Trudgill, 1985; Seyfried & Rao, 1987). One reason for these differences could be the difference in availability of the compound in the ‘mixing layer’ after application. In wet soil, the solute may move by diffusion below the mixing depth or into small pores at the surface (Hance, 1976) reducing the exposure to macropore transport. In dry soil, the solute would remain at the soil surface to a greater extent and therefore be more exposed to macropore transport. Another factor that may increase the
12 potential for macropore transport in dry soil is a decrease in lateral mixing between micropores and macropores (Shipitalo & Edwards, 1996). Aggregate size and stability (Coles & Trudgill, 1985), non-equilibrium sorption, water solubility, formulation type, and application rate (Hance & Embling, 1979) are other factors which may explain the influence of soil moisture content at the time of application on leaching.
Quantifying macropore flow effects
Field and lysimeter experiments are useful in demonstrating the occurrence of macropore flow, and can also serve to illustrate the effects of different management practices on macropore flow and transport. However, on their own, they do not enable quantification of macropore flow effects on leaching. To achieve this aim, field experiments must be interpreted using models. Even though quite a few models of different types now account for macropore flow and transport (see review by Jarvis, 1998), only a few might be considered suitable for field-scale applications under normal agro-environmental conditions. Some models that have been successfully used to describe macropore flow effects on agrochemical leaching from lysimeters or field experiments are the mechanistic models MACRO (Jarvis, 1994), RZWQM (Ahuja et al., 1995), and TRANSMIT (Hutson & Wagenet, 1995), and the functional models SLIM (Addiscott & Whitmore, 1991) and PLM (Nicholls & Hall, 1995). The soil porosity in these models is divided into two or more separate but interacting pore regions characterized by their own flow rate and solute concentration.
Field evaluation of macropore flow models
Model validation If models are to be used as predictive management tools, validation is important to create confidence in the model in the user-community (Rykiel, 1996). Armstrong et al. (1996) proposed a stepwise procedure to validate the predictive ability (i.e. without calibration) of pesticide leaching models using field data on (i.) soil hydrology, (ii.) non-reactive solute movement, (iii.) pesticide fate in the soil, and finally (iv.) pesticide leaching. However, Addiscott et al. (1995) pointed out that some parameters can be difficult to measure, so that model calibration may be necessary, and consequently, the model cannot be easily validated in this strict sense. Jarvis (1998) proposed instead an alternative definition of validation, where a validated model is a model that has been ‘successfully applied’ and tested with minimum possible calibration at a number of sites with different environmental conditions. The condition ‘successfully applied’ will differ depending on the purpose of the model application (Parrish & Smith, 1990). For example, if a model is used to increase our understanding, the criterion for ‘successfully applied’ may differ significantly from a model application where it
13 is used for prediction (Caswell, 1976). Since all simulation models are wrong, but some are useful, Mankin et al. (1977) suggested that models should be evaluated on their usefulness rather than their validity. Rykiel (1996) added that the evaluation should include a judgement of the amount of knowledge available and the consequences of decisions based on the model results.
Data requirements for model calibration and validation Calibration is the process of adjusting model parameters repeatedly until a ‘best fit’ between measured and simulated data is obtained. This is done to estimate unknown parameter values and to minimize parameter errors, in order to reveal the extent of model error (Loague & Green, 1991). Since some parameters are interdependent, non-unique parameter sets may result (Beven, 1989), particularly if the calibration procedure is not highly constrained by measurements. Consequently, to minimize the uncertainty in calibration of parameters related to the water balance, both soil water contents and water discharges should be measured (Armstrong et al., 1996). A proper model evaluation requires also that outputs are compared to measured data of all major components of the solute mass balance, including state variables (e.g. solute contents in soil) and complementary data from flux measurements such as water samples from tile drainage or lysimeter outflow. This is especially critical when macropore transport occurs, since flux and resident concentrations will differ considerably and because preferential movement of small amounts of solute may not always be detected by soil core samples. This is because of the often large spatial variability of the solute contents in relation to the sampling intensity and the larger detection limits for soil extracts compared to water samples (Jarvis et al., 1995).
For a proper evaluation, it is important that the measurements are made at appropriate spatial and temporal resolutions. Since macropore flow can vary considerably over short distances, results from a few column or lysimeter experiments may not be representative of the real behaviour at the field scale. As has been noted previously, macropore flow is also highly transient and may strongly affect the temporal pattern of leaching, emphasizing the requirement in model evaluation for measurements made at appropriate time resolutions.
Assessing model performance The performance of simulation models is commonly evaluated by comparing model output with measured data either in qualitative graphical displays, or by using a range of different quantitative statistical measures (Loague & Green, 1991; Walker et al., 1996). Statistical measures may be especially useful when comparing different model concepts or different modelling strategies applied to the same dataset. However, since the amount and quality of measured data differs in different applications, it may be difficult to judge absolute model performance only in terms of statistical measures.
14 Aims and objectives
The overall aim of this thesis was to quantify the effects of macropore flow on leaching of nitrate and pesticides from a structured clay soil at Lanna, south-west Sweden. This aim has been achieved by (i.) measuring water flow and solute transport to tile drains at the field scale, (ii.) calibrating the dual-porosity/dual- permeability model MACRO (Jarvis & Larsson, 1998) against these extensive measurements, (iii.) evaluating the ability of the model to describe solute leaching, and (iv.) using the calibrated model to quantify macropore flow effects on leaching of pesticides and nitrate. To quantify the effects of macropore flow on nitrate leaching, it was also necessary to couple MACRO to an existing model of nitrogen turnover in soil, SOILN (Johnsson et al., 1987).
Materials and methods
The models
The models used in this study, MACRO (Version 4.1, Jarvis & Larsson, 1998) and SOILN (Johnsson et al., 1987; Paper III), are rather simple and user-friendly, yet sufficiently comprehensive, with a degree of detail (and mechanism) suitable for the purpose. The MACRO and SOILN models have been widely used and successfully tested for a range of different agro-environmental conditions (Hoffmann, 1999; Jarvis, 1998). Both models are physically-based, one- dimensional, for application in layered soils at the pedon/small plot scale. The soil profile is divided into homogeneous layers on the basis of physical, chemical, and biological characteristics. The MACRO model calculates a full water balance (e.g. precipitation and irrigation, water flows in micropores and macropores, losses to field drains, evapotranspiration and water uptake by roots), soil temperatures, and solute transport and transformations. The SOILN model deals with the major processes of importance for inputs, transformations, and outputs of nitrogen. MACRO and SOILN were coupled in this study, such that the driving data for SOILN (e.g. water storages and flows, water exchange between the flow domains, soil temperatures) are provided by the MACRO model. This combined modelling system enables simulation of macropore flow effects on nitrate leaching. Since both the models, and the coupling between them, are presented elsewhere (Johnsson et al., 1987; Jarvis & Larsson, 1998; Paper III), I will here give a brief overview and only describe in detail those parts that are most relevant for this thesis.
15 Fig. 1. Hydraulic functions in micro- and macropores as described by the MACRO model.
The MACRO model In the MACRO model, the soil porosity is divided into micropores and macropores. The two pore regions act as separate, but interacting, flow domains with their own conductivity, flow rate, and solute concentration. As illustrated in Fig. 1, the boundary between the micro- and macropores is defined by a characteristic soil water pressure ψb with a corresponding hydraulic conductivity, Kb and water content, θ b (Jarvis, 1994). In the micropores, the soil water pressure is calculated with the Brooks & Corey (1964) equation and the unsaturated hydraulic conductivity with Mualem’s model (Mualem, 1976). In the macropores, hydraulic conductivity is calculated with a simple power law equation (Jarvis, 1994). Vertical water flow in the micropores is calculated with Richards equation while gravity flow is assumed in the macropores. Lateral water flow from macropores to micropores, Sw is treated as a first-order process:
3Dwγ w S w = ()θ b −θ mi ()1 d 2 where Dw is an effective water diffusivity, γw is a scaling factor (set to 0.8) introduced to match the approximate and exact solutions to the diffusion problem (van Genuchten, 1985; Gerke & van Genuchten, 1993), θmi is the water content in
16 the micropores, and d is the aggregate half-width, reflecting a diffusion pathlength assuming a rectangular-slab geometry for the aggregates (van Genuchten & Dalton, 1986). Lateral water flow to tile drains and groundwater seepage to secondary drainage systems (e.g. perimeter field ditches) is calculated from saturated soil layers using seepage potential theory (Youngs, 1980).
Vertical solute transport in the micropores is estimated using the convective- dispersive equation, while mass flow is assumed to dominate in macropores. Lateral solute exchange between micropores and macropores, Ue, is controlled both by diffusion and convection:
3Deθ mi U e = ()cma − cmi + S w c' ()2 d 2 where De is an effective diffusion coefficient, cma and cmi are the solute concentrations in the macropores and micropores respectively, and c' indicates either cma or cmi depending on the direction of the lateral water flow. The quantity of solute routed into the macropores at the soil surface is calculated assuming instantaneous local equilibrium and complete mixing of incoming net rainfall with the water stored in a shallow surface soil layer or ‘mixing depth’, zd (Steenhuis & Walter, 1980). The solute concentration c*ma in the water routed into the macropores is calculated as:
∗ Qd ()t−∆t + Rct cma = ()3 R + ()zd ()θ + ()()1− f γ kd where Qd(t-∆) is the solute amount stored in the mixing depth at the previous time step, R is the net rainfall with the solute concentration ct, θ is the water content in the top layer, f is the fraction of sorption sites in the macropores, γ is the bulk density and kd is the sorption distribution coefficient in the top layer.
Pesticide degradation is estimated using first-order kinetics with simple functions to account for the effects of soil moisture and temperature (Boesten & van der Linden, 1991). Sorption is calculated according to the Freundlich isotherm assuming an instantaneous equilibrium between the soil solid and liquid phases and with a distribution of sorption sites between micropores and macropores specified by the user.
As driving data, the MACRO model uses daily air temperature, wind speed, vapour pressure and solar radiation to calculate potential evapotranspiration with the Penman-Monteith equation, while precipitation can be given either as hourly or daily data. Partitioning of precipitation into snow and rain is controlled by the measured air temperatures, while a simple degree-day factor together with the air temperature controls the rate of snowmelt. Soil temperatures are calculated from
17 air temperatures and the physical properties of the soil using the heat conductivity equation.
The SOILN model + - In the SOILN model, nitrogen is divided into two inorganic pools (NH4 and NO3) and three organic pools: a litter pool comprising microbes and fresh organic material, a faeces pool containing added manure, and a humus pool with stabilised organic matter derived from litter and faeces decomposition. Inputs of N can be derived from added mineral and organic fertilizer, atmospheric deposition and from inflowing groundwater, whereas outputs include harvest, denitrification and leaching. Only NO3-N is considered to be transported with the moving water, while NH4-N is regarded as immobile. The rates of litter decomposition and humus mineralization are controlled by first-order coefficients. Transformations of organic N depend on the C-N ratios of litter and roots, and are also regulated by soil temperature and moisture response functions. Nitrogen uptake by plants is described by an empirical logistic curve and depends on the availability of inorganic N, as well as a parameter describing the potential maximum uptake, and two ‘shape’ parameters. At harvest, plant N is partitioned into harvested N and N in above- and below-ground residues.
Fig. 2. Schematic representation of the SOILN model structure, including flows in micro- and macropores, and between the two pore regions. Parts within the dashed line represent the uppermost layer of the soil profile. Subsurface layers have the same model structure as the surface layer but receive no direct input from fertilizer and deposition (adapted from Johnsson et al., 1987).
18 The coupling between MACRO and SOILN The MACRO model was coupled with the SOILN model to enable simulation of macropore flow effects on nitrate leaching (Fig. 2). Abiotic processes were simulated with MACRO and the simulation outputs (e.g. water flows and temperatures) used as driving data to SOILN. Descriptions of transport of NO3-N in the macropores and lateral exchange between the two pore regions (equation 2) were incorporated in SOILN. These processes were treated exactly as described in the MACRO model, while NO3-N movement in the micropores was calculated by SOILN assuming mass flow with numerical dispersion. The macropores are treated as a rapid transport pathway where transformations of N and root uptake are assumed to be negligible.
Description of the field experiments
The model simulations presented in this thesis are based on field studies at the Lanna experimental farm, situated on a flat plain in south-west Sweden, lat. 58º21'N, long. 13º08'E (Brink & Lindén, 1980; Bergström & Brink, 1986; Lindén et al., 1993). The site consists of seven experimental plots, each 0.4 ha in size (Fig. 3). In 1935, tile drains were installed at 1 m depth, running lengthways in the plots at 13.5 m spacing. Three tile lines in each plot discharge into a single observation well. A secondary drainage system surrounds the experimental field with the aim of preventing inflow from the surroundings. The experiment and simulations described in Paper I, and the scenario simulations described in Paper
123 45 67
**** ** *
plot border observation well tile drain * ground water pipe
Fig. 3. Layout of the field-plots and tile drain system at the Lanna experimental farm.
19 II are based on plot 6, which has been under no-till practice since 1988. A one- year experiment (October 1994 to November 1995) was conducted with - application and monitoring of the conservative tracer Br and the pesticide bentazone (Paper I). Extensive sampling was performed during this period to measure water and solute amounts to 90 cm depth in the soil (Paper I, Table 2) together with water discharge and solute leaching from tile drains. The model simulations of nitrate leaching and turnover presented in Paper III are based on a long-term field study carried out on plot 1, which is conventionally-tilled. The simulated results were compared with measurements made between 1 January 1988 and 30 June 1995 of water discharge and nitrate leaching from tile drains, mineral N in soil, and N removed with the harvest.
The Lanna soil has a clay content varying from ca. 46 % in the topsoil to ca. 67 % at 2 m depth (Bergström et al., 1994). Some selected soil physical properties are shown in Table 2.
Table 2. Soil physical properties (after Bergström et al., 1994)
Depth Particle size distribution Organic Structure interval carbon Clay Silt Sand contentb (< 2 µm) (2-60 µm) (> 60 µm)
__ cm ______% ______0-30 46.5 46.2 7.3 2.0 Strong coarse blocky 30-60 56.1 40.6 3.3 0.8 Strong fine to medium blocky 60-100 60.6 37.4 2.0 0.3 Strong coarse blocky 105-275 66.6a 30.5a 2.9a 0.2c n.i.d
a From Wiklert et al. (1983). b From Lindén et al. (1993). c Estimated. d No information.
The soil in the unsaturated zone is characterized by numerous cracks and biotic macropores, while below ca. 2.2 m depth the clay is normally water-saturated and structureless. At 11 m depth, the clay is underlain by a 0.2 m thick layer of sandy till on top of the bedrock (Brink & Lindén, 1980). During the period 1985 to 1995, the mean annual temperature at Lanna was 6°C (-3°C in January and 15°C in July), and the mean annual precipitation was 573 mm, with a minimum of 506 mm in 1989 and a maximum of 685 mm in 1985.
20 Modelling strategy and parameterization
The modelling strategy adopted was first to calibrate the MACRO model against measurements of soil water contents, drain discharge, bromide contents in soil, and concentrations in tile drainage in the 1-year experiment on plot 6 (Paper I). Simulations with bentazone (Paper I) were made without any further calibration and the simulation results were compared to measurements of bentazone contents in the soil and leaching to tile drains.
The model parameterization derived in Paper I was then used in a study to illustrate the interactions of pesticide compound properties and macropore flow effects on leaching (Paper II). Scenario simulations were performed for a 10-year period with climatic data from 1985 to 1994. Sixty hypothetical compounds of widely different sorption and degradation properties were simulated for two different pesticide application scenarios: a spring sown cereal with application in June, and winter wheat with application in October.
- The parameter set related to soil hydraulic properties obtained from the Br -tracer experiment (Paper I) was also used in the study of macropore flow effects on nitrate leaching performed on plot 1 (Paper III). However, the effective diffusion pathlength, d, and the macroporosity (i.e. the difference between θb and θs) was changed in the topsoil to account for the effects of soil tillage, since plot 1 was managed with conventional tillage while plot 6 was under no-till management (Paper III, Table 4).
To quantify the effects of macropore flow on solute leaching (Papers I, II, and III), simulations were performed both with macropore flow (WMF) and without (WOMF). To eliminate macropore flow in the simulations, the effective diffusion pathlength, d (equations 1 and 2) was reduced to 1 mm, which results in physical equilibrium between the two pore regions, while larger values of d (50 to 300 mm) reflected the influence of macropore flow in the structured clay soil at Lanna.
Results and discussion
Effects of macropore flow on hydrology
The drain discharges simulated considering macropore flow were generally close to those measured (Paper I, Fig. 3a), both with respect to magnitude and timing of flow events, while in the WOMF simulations, some peak flows were considerably delayed and others much smaller than the measurements indicated. Simulations considering macropore flow gave ca. 9 % more drain discharge than simulations
21 without macropore flow (Papers I and II) on plot 6 and ca. 7 % more on plot 1 (Paper III). This is because some of the infiltrating water is predicted to move rapidly to deeper soil layers where it is less susceptible to evaporation and root uptake.
Effects of macropore flow on bromide leaching
- The leaching of Br at Lanna (Paper I, Fig 6a) showed clear evidence of macropore transport, with high concentrations in the first major tile-drain discharge after application. This was followed by a rapid decrease in concentration and a delayed ‘dispersed’ breakthrough of the bulk of the mass of bromide to drain depth. Despite the high concentrations in the first drain discharge, only a limited portion (ca. 7 %) of the total leaching occurred as a result of this event (Paper I, Fig. 6b).
In the WMF simulation, a good match was obtained with the measured initial - breakthrough of Br in the drainflow, the accumulated leaching to tile drains, and - also the Br contents in the soil, while the pattern of short-term fluctuations in drainage water concentrations were poorly captured (Paper I). In the WOMF - simulation, it was impossible to depict the early initial breakthrough of Br in the tile drain discharge without seriously overestimating the concentrations in the subsequent months and also the total accumulated leaching. For the 1-year period - considered in Paper I, the WMF simulation gave ca. 40% less Br leaching to tile- drains compared to the WOMF simulation.
Effects of macropore flow on pesticide leaching
The pattern of leaching of the weakly-sorbed herbicide bentazone was similar to - that of Br (Paper I, Fig. 8a), with high concentrations found in the tile-drain discharge soon after application. For the experimental period, the simulated effect of macropore flow was to reduce total leaching of bentazone to tile drains by as much as 50 %. The reason for this is clearly illustrated by the bentazone transport predicted in the soil (Paper I, Fig. 7), where the WOMF simulation overestimates the downward transport of the bulk of the mass, while the estimated downward transport in the WMF simulation is considerably slower and satisfactorily matches the measurements. In the WMF simulation, the weakly-sorbed bentazone was readily transported downwards in the micropores soon after application, where it moved at a reduced convective velocity, since the infiltrating water routed into the macropores at the surface effectively bypasses the compound stored in the micropores. In the WOMF simulation, where equilibrium in solute concentration is maintained between all pores at any depth, the percolating water leached the bentazone more efficiently.
22 However, scenario simulations with 60 hypothetical pesticides (Paper II) indicated that macropore flow in the structured clay soil at Lanna will usually increase pesticide leaching, since macropore flow is the dominant leaching mechanism for less mobile compounds. The largest effects were found for 3 -1 moderately to strongly sorbed compounds (200 < Koc < 2000 cm g ) with half- lives less than 5 days (Paper II, Fig. 3b). A reduction in leaching due to 3 macropore flow was only found for pesticides with Koc-values less than ca. 30 cm g-1 and half-lives longer than ca. 80 days (Paper II, Fig. 3). These compounds generally leached more than 10 % of the applied amount and would most likely fail the groundwater (drinking water) criterion for EU registration of 0.1 µg litre-1 (Paper II).
The scenario simulations (Paper II) also indicated that macropore flow may significantly reduce the influence of pesticide properties on leaching. For example, a 100 times difference in simulated leaching between two compounds in the absence of macropore flow was reduced to a fourfold difference when macropore flow was considered. This suggests that reductions in dose may be a relatively more attractive means of decreasing leaching of pesticides from macroporous soils, since the difference in total leaching between compounds with diverse sorption and degradation properties is reduced.
A simple power law function could summarise the effects of macropore flow for those pesticides with a total leaching of more than 0.0001 % of the applied dose (simulated WOMF; Paper II, Fig. 4b):
(1−n) lm = ble ()4 where lm and le are the percent leached of the applied dose predicted with and without macropore flow respectively, b is a constant, and n is an index accounting for the effects of macropore flow for a certain soil/climate scenario. If there is no effect of macropore flow (n = 0), the properties of the solute will influence total leaching equally with and without macropore flow, while for the theoretical maximum effect (n = 1), total leaching will be the same for all compounds in the presence of macropore flow (i.e. pesticide properties will have no effect on leaching). For the soil and climate conditions at Lanna, the macropore flow index n was 0.71.
Effects of macropore flow on nitrate leaching
It is well-known that NO3-N leaching from sandy soils is generally larger than from clay soils, but it is not generally recognised that macropore flow may be one important reason for this difference (e.g. Simmelsgaard, 1998; Hoffmann & Johnsson, 1999). In contrast, it is often suggested that macropore flow will
23 increase the total leaching of NO3-N (e.g. Booltink, 1995). Heavy rain following fertilizer application in one year (135 mm in one month) did result in a simulated increase in leaching by 34 % due to macropore flow (Paper III). However, the general effect of macropore flow under the prevailing agro-environmental conditions at Lanna was clearly to reduce NO3-N leaching. From July 1990 to June 1995, the WMF simulation showed ca. 30 % smaller accumulated leaching from tile-drains compared to the WOMF simulation. However, the differences between individual years were large, with reductions in leaching varying from 3 to 45 %. The higher predicted NO3-N leaching in the WOMF simulation was maintained by higher mineralization and decomposition rates of organic nitrogen.
Without considering macropore flow in the simulations at Lanna, it was impossible to match all the measured components of the measured N mass balance (Paper III). It would have been possible to obtain a considerably better match between measured and simulated NO3-N leaching to drains by increasing the potential denitrification or by reducing the humus mineralization and litter decomposition rates. However, this would have resulted in a serious underestimation of the simulated amounts of NO3-N in the soil, and in some years, it would also have been difficult to simulate sufficient N uptake in the plant.
Uncertainties and limitations in the modelling approach
There are many uncertainties in using simulation models such as MACRO and SOILN. Therefore, in the following paragraphs, some of the difficulties encountered in the applications described in this thesis are discussed. These uncertainties and potential errors can be divided into (i.) uncertainties and errors in the measurement of driving data or model parameter values, and (ii.) assumptions and simplifications in the model descriptions.
The MACRO model For the field-scale applications reported in this thesis, uncertainties with respect to the correct specification of poorly-defined boundary conditions were especially troublesome. For example, in order to match the measured drain discharge it was necessary to increase the measured precipitation amounts to account for presumed wind effects (Papers I and III) and also to adjust for inflowing groundwater to plot 6 (Paper I) and water flow to secondary drains from plot 1 (Paper III). Even though the accumulated amounts of tile drain discharge could be matched in this way, there is still considerable uncertainty concerning the validity of these adjustments, especially with respect to the temporal patterns assumed, since they are based on model interpretation and not direct measurement. However, the risk of parameterisation errors should be small for the 1-year study on plot 6 (Paper I), since both soil moisture contents and tile drain flows were monitored. The risk for parameterisation errors is considerably greater for the model application on plot 1
24 (Paper III), since no soil moisture measurements were available. Other parameters influencing the water balance (e.g. root distribution and water uptake by plants) may then be inadvertently set to incorrect values to compensate for errors in the water balance caused by an improper parameterisation of groundwater inflow and/or erroneous precipitation measurements.
The temporal resolution of driving data is another source of uncertainty. In Paper III, daily precipitation amounts were used as driving data. In MACRO, a constant rain intensity must then be specified for the whole simulation, while in reality, the rain intensity will vary in time both within and between rain events. This approximation may considerably influence the short-term leaching pattern, and might also have consequences for predictions of the total leaching, especially if the saturated hydraulic conductivity of the micropores, Kb (Fig. 1), lies within the range of normal rain intensities.
Another uncertainty is connected to the estimation of the ‘mixing-depth’, zd (equation 3). The thickness of this shallow surface soil layer will control the quantity of solute that will be dissolved by incoming rainwater and routed into the macropores. The ‘mixing depth’ will depend on the structure of the soil surface (e.g. surface roughness, crusting, aggregate stability), cover conditions, the kinetic energy of rainfall and the water content in the surface layer at the time of precipitation (Ahuja, 1986). The ‘mixing depth’ may therefore vary considerably over time due to, for example, loosening of the soil surface by tillage and subsequent settling by rainfall impact. One of the most important parameters influencing macropore transport is the effective diffusion pathlength, d (equations 1 and 2), governing the exchange of water and solute between macropores and micropores. This parameter should ideally equal the aggregate half-width. However, to satisfactorily match the - measured Br concentrations in tile drain discharge (Paper I), d had to be increased considerably compared to the initial values obtained by visual observation of the soil structure. Organic and clay coatings on the aggregate surfaces may reduce mass transfer rates between macropores and the intra- aggregate soil matrix (e.g. Thoma et al., 1992) resulting in large ‘effective’ d values. Another, perhaps more speculative, explanation is that the model assumption of instantaneously-attained uniform concentrations in the micropore region could result in an erroneous estimation of the lateral solute exchange resulting in large ‘effective’ d values.
The saturated hydraulic conductivity of the micropores, Kb is another important parameter influencing macropore transport since it sets the limit between flow through the micropores and bypass flow in macropores. This parameter can be roughly estimated by tension infiltrometer measurements (Jarvis & Messing, 1995). However, since larger pores are easily destroyed if the soil is disturbed, measurements with tension infiltrometers can only be performed with ease for the soil surface layer. Since the soil structure is not constant, but changes due to
25 tillage, raindrop impact, freezing/thawing and drying/wetting cycles (e.g. Benoit, 1973; Bronswijk & Evers-Vermeer, 1990; Messing & Jarvis, 1993), Kb may also vary considerably over time.
In the applications presented in this thesis, the MACRO model was only evaluated for non-sorbing and weakly-sorbed compounds. It is possible that if the model were tested for more strongly sorbed compounds, additional difficulties in the parameterization may have been encountered. For example, Beulke et al. (1998) found that the leaching simulated by MACRO was sensitive to the parameter partitioning the number of sorption sites between macropores and micropores. This parameter is difficult to estimate directly, and therefore requires calibration. Model assumptions concerning the nature of the sorption process would also be more critical for more strongly sorbed compounds. In MACRO, sorption is described as an instantaneous process described by the Freundlich isotherm. However, this assumption may not always be valid, in that sorption may be time-dependent, while different adsorption and desorption characteristics (hysteresis) have also been reported for many pesticides (see review by Koskinen & Harper, 1990).
The SOILN model Since not all the different pools (e.g. litter from roots, above-ground litter) and fluxes (e.g. denitrification, mineralization, litter decomposition) of nitrogen and carbon were directly measured in the application described in Paper III, parameter estimation by calibration was not satisfactorily constrained, leading to some uncertainties in the interpretation of the results. For example, the underestimation of NO3-N contents in the soil at 60 to 90 cm depth (Paper III, Fig. 3) was thought to be caused by the over-simplified model description of the flow and transport processes (Paper III). However, it is theoretically possible that it was at least partly caused by erroneous parameterization of the sizes of the N and C pools or nitrogen transformations in the soil-plant system.
To correctly estimate the macropore flow effect on transport of mineral fertilizer nitrogen (salt) soon after surface application, the parameter in SOILN governing the dissolution rate of the fertilizer may be particularly important. The dissolution rate constant will depend on fertilizer type but also on soil moisture conditions. This parameter might also require calibration between years, since the soil moisture conditions can vary considerably between application and the first rain event causing macropore flow.
26 Conclusions and recommendations
The results from this study show that, in the structured clay soil at Lanna, macropore flow significantly reduced leaching of a moderately persistent and weakly sorbed pesticide. This is because the bulk of the solute mass was effectively ‘protected’ against bypass of water in the macropores, being stored in micropore water moving at a ‘reduced’ convective transport velocity. However, simulations with 60 hypothetical pesticides of widely different sorption and degradation properties indicated that the dominating effect of macropore flow will be an increase in leaching, since for less mobile compounds, macropore flow is the most important leaching mechanism. The simulation results also indicated that macropore flow can reduce the influence of pesticide properties on leaching. This suggests that reductions in dose may be a relatively more effective means of reducing leaching of pesticides from structured soils, since the difference in total leaching between compounds with diverse sorption and degradation properties is reduced. Several new application techniques are available to reduce the dose, for example, target-sensing sprayer control systems (Giles et al., 1991) and precision band spraying (Giles & Slaughter, 1997). A reduced dose can also be achieved by substituting high-dose compounds with low-dose compounds. However, it should also be borne in mind that it is not only the leaching risk that is of interest when selecting pesticides for low adverse environmental impact. Ecotoxicological characteristics must also be taken into consideration.
For nitrate, the calculated effect of macropore flow at Lanna was to reduce leaching by ca. 30%, although large differences were found between individual years (from 3 to 45 %). However, it is quite possible that for contrasting agro- environmental conditions (e.g. type and quantity of fertilizer, crop system, tillage management system, and climate), the effect of macropore flow on nitrate leaching may well be different. For example, if most of the annual precipitation occurs soon after application, or if fertilizer is applied in the autumn, it is possible that macropore flow could result in overall increases in nitrate leaching.
By taking macropore flow into account, the MACRO model, and MACRO coupled with SOILN, could satisfactorily depict the measured solute - concentrations (Br , bentazone, and nitrate) in the tile drain discharge and the accumulated leaching (Papers I and III). However, measured short-term variations in flux concentrations were often poorly described by the model. It is suggested that the main reason for this discrepancy is the model assumption of
27 instantaneously-attained uniform concentrations in the micropore region at a given depth in the soil. Without taking macropore flow into consideration, it was - impossible to simulate the large concentrations of Br and bentazone in tile drain discharge soon after application and the subsequent rapid decrease. It was also impossible to match all the measured components of the nitrogen mass balance without accounting for macropore flow. There are still many uncertainties in modelling macropore flow effects on solute leaching, and simulation results should therefore be judged with caution and careful attention to the possible consequences of decisions based on these results. This is especially important if the model is to be used for extrapolation to domains where it has not been thoroughly evaluated (e.g. significantly different soils, climates, or substances). Nevertheless, despite such uncertainties, models must sometimes be used predictively, for example, in comparative simulations to answer ‘what if’ questions intended to evaluate and improve management practices to reduce leaching. It is therefore desirable that efforts are made to improve our knowledge concerning critical processes influencing non-equilibrium macropore flow and transport. The problems with model parameterisation encountered in the applications presented in this thesis suggest that further research in the following areas could result in significant model improvements: • The influence of soil-surface properties (texture, structure, roughness) on the ‘mixing depth’, the surface layer of soil controlling solute routing into the macropores. • Lateral solute exchange between macropores and micropores. The significance of the assumption concerning instantaneously-attained uniform concentrations in the micropore region and the importance of aggregate coatings and macropore linings. • The significance of temporal changes in important soil physical and hydraulic parameters (e.g. saturated hydraulic conductivity of the micropores, effective diffusion pathlength, macropore volume) caused by soil tillage and subsequent consolidation by raindrop impact.
The results of this research should be incorporated into improved versions of simulation models such as MACRO and SOILN. This would help in the evaluation of alternative soil and crop management practices to minimize adverse impacts of macropore flow on non-point source pollution. Some promising management strategies include: • avoiding applications during ‘high-risk’ periods. Given some flexibility in application timing, it may be possible to reduce losses, especially of quickly- degrading solutes, since the precipitation pattern soon after application is a crucial factor influencing total leaching. However, what constitutes a ‘high- risk’ period is not well understood and requires further study. In addition, it will probably not be easy to predict these ‘high-risk’ periods in rain-fed agriculture since weather forecasts are normally only reliable for a very limited time. However, this strategy would be possible in irrigated agriculture, where irrigation timings, amounts and intensities can be more easily managed
28 and controlled to avoid generation of macropore flow during sensitive periods.
• applying appropriate soil management methods (e.g. tillage practices and surface cover management) to reduce solute leaching through macropores, since the significance of macropore flow is strongly affected by soil structure, especially at and close to the surface. However, since macropore flow may reduce nitrate leaching and increase pesticide leaching, the total effect of this kind of measure must be judged carefully.
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33 Acknowledgements
First, I would like to express my sincere gratitude and appreciation to my main supervisor Prof. Nicholas Jarvis, who has been an excellent and extraordinarily helpful guide and mentor. Always encouraging, with a contagious enthusiasm and a genuine interest for science.
Thanks are also due to my other supervisors, Prof. Lars Bergström who has always been available for valuable discussions and revision of manuscripts, and Assoc. Prof. Holger Johnsson for helpful discussions regarding the (not so much any more) mysterious world of nitrogen modelling. Thanks are also due to Dr Henrik Eckersten for advice concerning the development of the MACRO-SOILN coupling.
Thanks to all the colleagues at the Division of Water Quality Management, especially Dr Jörg Brücher, Dr Markus Hoffmann, and Katarina Kyllmar for their help and friendship, Prof. Arne Gustafsson for his flexibility and help in all kind of practical and economic issues, Stefan Ekberg and co-workers for skillful help in the laboratory, and Dr Gunnar Torstensson for invaluable technical help and fruitful discussions.
I would also like to thank the staff at the Lanna experimental farm, especially Rolf Tunared for help in the field, and for providing me with climatic data and other useful information about the experiments.
Even though the results from my measurements at Winand Staring Centre-DLO, in Wageningen were in the end excluded from this thesis, thanks are due to all the people who helped me during that stay. I would specially like to thank Drs. Jûnt Halbertsma, Jannes Stolte, Henk Wösten, Coen Ritsema and Luis Dekker.
Special thanks are due to Dr Ettore Capri for making my stay at Universita Cattolica in Piacenza, Italy most memorable and for his extraordinary flexibility during the completion of this thesis.
Finally, I would like to thank my father Karl-Henry Larsson, who has been waiting patiently for me to finish my studies, so he can retire peacefully from the family farming business he has wanted to quit for many years now.
I also wish to acknowledge the financial support provided by the European Commission - Environment Research Program (Contract EV5V-CT94-0467), within the project ‘Analysis and improvement of existing models of field-scale solute transport through the vadose zone of differently textured soils, with special reference to preferential flow’. Additional funding was received from The Swedish Environmental Protection Agency, SNV.
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