modelling the effects of bioturbation on the re-distribution of 137cs in an undisturbed grassland...

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European Journal of Soil Science, February 2010, 61, 24–34 doi: 10.1111/j.1365-2389.2009.01209.x Modelling the effects of bioturbation on the re-distribution of 137 Cs in an undisturbed grassland soil N. J. Jarvis a , A. Taylor b , M. Larsbo a , A. Etana a & K. Ros´ en a a Department of Soil and Environment, SLU, Box 7014, 750 07 Uppsala, Sweden, and b Department of Ecology, SLU, Box 7044, 750 07 Uppsala, Sweden Summary Under favourable conditions, soil ingestion by earthworm populations can be equivalent to approximately 5–10% of the topsoil mass per year. This suggests that for contaminants that are strongly bound to soil, earthworm ‘bioturbation’ may be a more important transport mechanism than water-borne advection dispersion. It is therefore quite surprising that few modelling studies to date have explicitly considered the effects of biological processes on contaminant transport in soil. In this study, we present a general model that incorporates the effects of both ‘local’ and ‘non-local’ biological mixing into the framework of the standard physical (advective-dispersive) transport model. The model is tested against measurements of the redistribution of caesium-137 ( 137 Cs) derived from the Chernobyl accident, in a grassland soil during 21 years after fallout. Three model parameters related to biological transport were calibrated within ranges defined by measured data and literature information on earthworm biomasses and feeding rates. Other parameters such as decay half-life and sorption constant were set to known or measured values. A physical advective-dispersive transport model based on measured sorption strongly underestimated the downward displacement of 137 Cs. A dye-tracing experiment suggested the occurrence of physical non-equilibrium transport in soil macropores, but this was inadequate to explain the extent of the deep penetration of 137 Cs observed at the site. A simple bio-diffusion model representing ‘local’ mixing worked reasonably well, but failed to reproduce the deep penetration of Cs as well as a dilution observed close to the soil surface. A comprehensive model including physical advective- dispersive transport, and both ‘local’ and ‘non-local’ mixing caused by the activities of both endogeic and anecic earthworms, gave an excellent match to the measured depth profiles of 137 Cs, with predictions mostly lying within confidence intervals for the means of measured data and model efficiencies exceeding 0.9 on all sampling occasions but the first. Introduction Under favourable conditions, soil ingestion rates of earthworm populations can be equivalent to approximately 5–10% of the topsoil mass per year (Lee, 1985; Curry & Schmidt, 2007). Many contaminants of ecological interest are very strongly sorbed to soil constituents, so that water-borne advective-dispersive transport through soil is extremely slow for such compounds that include persistent organic pollutants, trace metals and radionuclides. It therefore seems likely that biological transport in the solid phase induced by earthworm feeding and egestion (bioturbation) could alter significantly the distribution of these soil contaminants, which would in turn affect transfer to other Correspondence: N. J. Jarvis. E-mail: [email protected] Received 17 April 2009, revised version accepted 6 October 2009 environmental compartments of ecotoxicological interest such as plant uptake. M¨ uller-Lemans & van Dorp (1996) and Bunnenberg & Taeschner (2000) reviewed the literature on earthworm feeding habits and data on soil ingestion rates and made some simple calculations that demonstrated that soil turnover by soil fauna, and especially by earthworms, should be a very relevant transport mechanism for strongly sorbed contaminants such as radionuclides in comparison with physical transport mechanisms. Experimental evidence suggests that this may indeed be the case. McCabe et al. (1991) showed that downward movement of bomb fallout derived caesium-137 ( 137 Cs) into a loamy sand soil with small earthworm biomass was limited, whereas 137 Cs was much more evenly mixed and incorporated in the upper 5–8 cm of three finer-textured loamy soils, which had three to eight times larger earthworm biomasses. Similarly, relatively homogeneous depth- distributions of PCBs and PAHs (Cousins et al., 1999a; Armitage et al., 2006) and 137 Cs (Tyler et al., 2001) in uncultivated soils © 2009 The Authors 24 Journal compilation © 2009 British Society of Soil Science

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Page 1: Modelling the effects of bioturbation on the re-distribution of 137Cs in an undisturbed grassland soil

European Journal of Soil Science, February 2010, 61, 24–34 doi: 10.1111/j.1365-2389.2009.01209.x

Modelling the effects of bioturbation on there-distribution of 137Cs in an undisturbed grassland soil

N . J . J a r v i sa , A . T a y l o rb , M . L a r s b oa , A . E t a n aa & K . R o s e na

aDepartment of Soil and Environment, SLU, Box 7014, 750 07 Uppsala, Sweden, and bDepartment of Ecology, SLU, Box 7044, 750 07Uppsala, Sweden

Summary

Under favourable conditions, soil ingestion by earthworm populations can be equivalent to approximately5–10% of the topsoil mass per year. This suggests that for contaminants that are strongly bound to soil,earthworm ‘bioturbation’ may be a more important transport mechanism than water-borne advection dispersion.It is therefore quite surprising that few modelling studies to date have explicitly considered the effectsof biological processes on contaminant transport in soil. In this study, we present a general model thatincorporates the effects of both ‘local’ and ‘non-local’ biological mixing into the framework of the standardphysical (advective-dispersive) transport model. The model is tested against measurements of the redistributionof caesium-137 (137Cs) derived from the Chernobyl accident, in a grassland soil during 21 years after fallout.Three model parameters related to biological transport were calibrated within ranges defined by measured dataand literature information on earthworm biomasses and feeding rates. Other parameters such as decay half-lifeand sorption constant were set to known or measured values. A physical advective-dispersive transport modelbased on measured sorption strongly underestimated the downward displacement of 137Cs. A dye-tracingexperiment suggested the occurrence of physical non-equilibrium transport in soil macropores, but this wasinadequate to explain the extent of the deep penetration of 137Cs observed at the site. A simple bio-diffusionmodel representing ‘local’ mixing worked reasonably well, but failed to reproduce the deep penetration of Csas well as a dilution observed close to the soil surface. A comprehensive model including physical advective-dispersive transport, and both ‘local’ and ‘non-local’ mixing caused by the activities of both endogeic andanecic earthworms, gave an excellent match to the measured depth profiles of 137Cs, with predictions mostlylying within confidence intervals for the means of measured data and model efficiencies exceeding 0.9 on allsampling occasions but the first.

Introduction

Under favourable conditions, soil ingestion rates of earthwormpopulations can be equivalent to approximately 5–10% of thetopsoil mass per year (Lee, 1985; Curry & Schmidt, 2007).Many contaminants of ecological interest are very strongly sorbedto soil constituents, so that water-borne advective-dispersivetransport through soil is extremely slow for such compoundsthat include persistent organic pollutants, trace metals andradionuclides. It therefore seems likely that biological transportin the solid phase induced by earthworm feeding and egestion(bioturbation) could alter significantly the distribution of thesesoil contaminants, which would in turn affect transfer to other

Correspondence: N. J. Jarvis. E-mail: [email protected]

Received 17 April 2009, revised version accepted 6 October 2009

environmental compartments of ecotoxicological interest such asplant uptake. Muller-Lemans & van Dorp (1996) and Bunnenberg& Taeschner (2000) reviewed the literature on earthworm feedinghabits and data on soil ingestion rates and made some simplecalculations that demonstrated that soil turnover by soil fauna,and especially by earthworms, should be a very relevant transportmechanism for strongly sorbed contaminants such as radionuclidesin comparison with physical transport mechanisms. Experimentalevidence suggests that this may indeed be the case. McCabeet al. (1991) showed that downward movement of bomb falloutderived caesium-137 (137Cs) into a loamy sand soil with smallearthworm biomass was limited, whereas 137Cs was much moreevenly mixed and incorporated in the upper 5–8 cm of threefiner-textured loamy soils, which had three to eight times largerearthworm biomasses. Similarly, relatively homogeneous depth-distributions of PCBs and PAHs (Cousins et al., 1999a; Armitageet al., 2006) and 137Cs (Tyler et al., 2001) in uncultivated soils

© 2009 The Authors24 Journal compilation © 2009 British Society of Soil Science

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Bioturbation and 137Cs distribution 25

have been attributed to the effects of thorough mixing byearthworms. Despite its significance, the redistribution of stronglysorbing contaminants in soil is often modelled without explicitconsideration of the effects of bioturbation (for example Smith& Elder, 1999; Bossew & Kirchner, 2004). One exception is thestudy by Bunzl (2002), who developed a random walk modelof bioturbation combined with physical transport mechanismsin a forest soil. Liming had increased the abundance of theepigeic earthworm L. rubellus L. by nearly 20 times comparedwith a control plot, which markedly changed the distribution ofChernobyl-derived 134Cs in the surface organic soil horizons.The model was able to mimic the different distributions of134Cs with depth in the limed and control plots. Bioturbationhas also been considered as a key transport mechanism inmodels of PCB re-distribution in soil (Cousins et al., 1999b;McLachlan et al., 2002), the evolution of soil carbon profiles(Elzein & Balesdent, 1995) and pedogenesis (Salvador-Blaneset al., 2007).

In contrast to the limited number of studies in soil science,modelling the effects of bioturbation on contaminant transportin marine and freshwater sediments is a long-established fieldof study (see the review by Boudreau, 1999). The effects ofbiological processes on solute transport in sediments are manyand varied, depending on community species composition, theecological niches occupied by these species and their behaviour.Various modelling approaches have been developed that aresuitable to describe these different activities (Meysman et al.,2003). For example, a diffusion equation can be employedto predict solute transport if sediment re-working is randomand takes place at a sufficiently small scale (‘local’ mixing;Boudreau, 1986a). However, these assumptions are violated whenthe locations of ingestion and egestion are widely separatedvertically. For example, ‘head-down’ feeders ingest sedimentat depth and egest their faeces at the sediment-water interface(‘conveyor-belt’ feeding, Boudreau, 1986b), which gives rise toa upward advective flux of particles and sorbed solutes. Forcases when egestion occurs within the sediment, Boudreau &Imboden (1987) developed a model of ‘non-local’ mixing inwhich bioturbation is given as a vertically-distributed integralsource-sink term.

Two main ecological groups of earthworms may inducebioturbation in soil: deep-burrowing anecic earthworm speciesconstruct permanent vertically-oriented burrows, feed mainlyon relatively undecomposed organic matter at the surface andegest their faeces both at the surface and in the linings oftheir burrows. In contrast, endogeic earthworm species feedthroughout the topsoil (usually at least to approximately 20-cmdepth), making temporary burrows in more or less randomdirections, as they eat their way through the soil. It thereforeseems reasonable to expect that a generally applicable model ofearthworm bioturbation in soil should account for both ‘local’ and‘non-local’ mixing.

The fallout resulting from the Chernobyl nuclear reactoraccident in 1986 has provided opportunities for studying the

processes affecting the re-distribution of radionuclides in soil,especially 137Cs (Riesen et al., 1999; Rosen et al., 1999). There-distribution of 137Cs has been measured in a silty clay grasslandsoil at Skogsvallen (Rosen et al., 1999; Persson, 2008) onsix occasions during the 21 years after the initial fallout fromChernobyl. In the present paper, a model of ‘local’ and ‘non-local’ biological transport mechanisms is coupled to the standardphysical advection-dispersion equation and tested against the long-term measurements of 137Cs re-distribution in the Skogsvallensoil. We also carried out a dye-staining experiment to investigatewhether preferential water flow in soil macropores such asearthworm channels or cracks could have contributed to thedownward migration of 137Cs observed in this well-structured soil(e.g. Bundt et al., 2000).

Theory

The model is derived by introducing additional terms forbio-diffusive and advective transport in the solid phase andan integral source-sink term, Tf , to account for ‘non-local’bioturbation into the classical advection-dispersion equation fortransport in soil water:

∂A

∂t= ∂

∂z

(Dθ

∂c

∂z− qc

)+ ∂

∂z

(Db

∂Sγ

∂z− qbSγ

)+ Tf − μA,

(1)

where A, c and S are the total caesium activity and activitiesin solution and sorbed phases, respectively (Bq m−3 soil, Bqm−3 water and Bq kg−1 soil), t is time (s), z is depth (m),μ is the first-order radioactive decay constant (s−1), θ is thevolumetric water content (m3 m−3), q is the soil water flowrate (m s−1), γ is the bulk density (kg m−3), qb is the rateof downward soil displacement (m s−1), Db is the bio-diffusioncoefficient (m2 s−1) and D is the physical dispersion-diffusioncoefficient (m2 s−1) given by ξDw + λq/θ where Dw is thediffusion coefficient in water (m2 s−1), ξ is a tortuosity factorand λ is the dispersivity.

The transfer function, Tf , describes the probability of egestionas a function of the vertical separation distance from the pointof ingestion. Little is known about the appropriate shape of thisfunction, but it should be symmetrical about the point of ingestionif mass is to be conserved, that is the soil fluxes to and froma given soil depth must be equal if the bulk density profileis at equilibrium. A Gaussian distribution has been proposed(Boudreau & Imboden, 1987), but to keep the model simple, weadopt a uniform transfer function, assuming that the probabilityof egestion is constant and independent of distance from the pointof ingestion within a depth L (m) from the soil surface. With thisassumption, we have

Tf = Inl

⎡⎣

⎛⎝1 − fsurf

L

L∫0

S dz

⎞⎠ − S

⎤⎦ , (2)

© 2009 The AuthorsJournal compilation © 2009 British Society of Soil Science, European Journal of Soil Science, 61, 24–34

Page 3: Modelling the effects of bioturbation on the re-distribution of 137Cs in an undisturbed grassland soil

26 N. J. Jarvis et al.

where Inl is the soil ingestion rate for non-local feeders(kg m−3 s−1), which is assumed to be constant within a depthL from the soil surface (m) and fsurf is the fraction of soil massingested by non-local feeders that is egested at the soil surface (–).To conserve mass, the soil displacement rate, qb, must vary withdepth according to:

qb(z) = fsurf

γ

L∫z

Inl dz. (3)

As Inl is assumed to be constant, the solution of the integral inEquation (3) shows that qb(z) decreases linearly with depth in thebioturbated zone:

qb(z) = fsurf Inl(L−z)

γ; z ≤ L (4)

qb(z) = 0; z > L.

Steady-state water flow is assumed (that is q is constant),because for strongly sorbing compounds, short-term and sea-sonal fluctuations in water flow rates are known to have lit-tle influence on solute transport predictions with the advection-dispersion equation (Destouni, 1991; Skaggs et al., 2007).Sorption is assumed to be instantaneous and to follow a linearisotherm:

S = Kdc, (5)

so that:

c = A

θ + γKd

, (6)

where Kd is the sorption constant (m3 kg−1). An increaseof sorption with time has been reported for caesium, butin mineral soils equilibrium appears to be reached at leastwithin a few weeks or months (Absalom et al., 1995). There-fore, the assumption of equilibrium sorption seems reasonablegiven the slow and long-term nature of the transport process,especially if Kd values are estimated from longer-term des-orption experiments rather than 24-hour sorption experiments(e.g. Sanchez et al., 2002). In principle, sorption should fol-low a non-linear isotherm, but a linear function may be a goodapproximation across a limited range of concentrations. Thelinear approximation will tend to under-estimate sorption andtherefore over-estimate the rate of downward displacement bywater.

The model equations were solved with an explicit finite differ-ence method correcting for numerical dispersion. The numericalsolution of the advection-dispersion equation was verified againstan analytical solution. The mass balance was checked and foundto be always accurate to within 0.3%.

Materials and methods

Site details

The model described above was tested against field measurementsof 137Cs distribution in soil made at Skogsvallen (17◦11′E,60◦10′N, 50 km north-west of Uppsala, Sweden) on six samplingoccasions during the 21 years after the Chernobyl fallout inSweden on 28 April 1986 (see Figure 1). The site, which islocated in a flat valley bottom, has been under permanent, un-grazed grass since the 1960s. The mean annual precipitation is566 mm and the mean annual temperature is 5.3◦C (1961–1990;Alexandersson et al., 1991). The silty clay soil (Table 1) wasformed in quaternary lacustrine deposits laid down when the areawas a part of the Baltic sea basin and is classified as a DystricCambisol (FAO). The A horizon has a fine crumb structure, andthe B horizon has a moderate to strong medium blocky structure.Roots, mainly from various grass species, are present throughoutthe sampled depth. In the B horizon, roots grow preferentiallyalong well-defined aggregate surfaces.

Figure 1 The deposition of 137Cs over Sweden after the Chernobylaccident, based on measurements made by aerial surveys in May toOctober 1986 by the Swedish Geological Company (SGAB, Uppsala) andthe location of the study site at Skogsvallen.

© 2009 The AuthorsJournal compilation © 2009 British Society of Soil Science, European Journal of Soil Science, 61, 24–34

Page 4: Modelling the effects of bioturbation on the re-distribution of 137Cs in an undisturbed grassland soil

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Experimental

Field measurements of 137Cs in soil. On each of 6 samplingoccasions, 15 samples were collected within 3 circles, each with aradius of 5 m. The circles were placed along a transect, with 20 mbetween the centre points of the circles (Figure 2). The soil wassampled down to 15 cm on the first sampling occasion in 1987,to 25 cm in 1992 and 1994 (Rosen et al., 1999), and to 60 cmon three subsequent occasions in 2000, 2004 and 2007. However,we only present data to 25-cm depth, because very little 137Csactivity was detected below this (Persson, 2008). The uppermost10 cm were sampled with a cylinder corer with a diameter of57 mm. The cores were placed in plastic bags during transportto the laboratory, where they were cut into 1-cm slices. Soil at10- to 60-cm depth was extracted with a core sampler, 22 mm indiameter. The core was sliced into 2.5-cm layers at the site.

Samples taken from the same circle were combined to formthree batched samples for each layer. The samples were dried at30◦C for at least a week. Coarse material was removed on a 2-mmsieve, before a representative amount of soil was taken out foranalysis. All samples were placed in plastic vials prior to analysis.The samples were then analysed with a high-purity germaniumdetector housed in a laboratory with low-background gammaemissions. A germanium detector measures gamma emission andis able to separate gamma emissions originating from differentelements using the specific energy emitted by each element. Anestimated uncertainty in the measurement of less than 10% wasconsidered to be acceptable. If the uncertainty was larger, emissiondetection was continued for 24 hours and then terminated.

The dose received at the site from Chernobyl fallout, Ao

(Bq m−2), was estimated from the measurements of total Cscontents in soil, At (Bq m−2), made on the six sampling occasions,and the known decay half-life DT50 of 137Cs (= 30.1 years):

Ao = At

exp(−t

ln(2)DT50

) , (7)

where t is the elapsed time after fallout in years. The averageAo value was estimated as 92.7 kBq m−2, with a coefficient ofvariation of only 12%, which is presumably related to spatialvariation in the initial deposition, subsequent lateral re-distributionprocesses in the soil (for example lateral flow) and sampling andmeasurement errors. For the purpose of comparing with modelpredictions, the measured Cs contents on each sampling occasionwere corrected to an initial dose of 92.7 kBq m−2. It can be notedthat137Cs originating from global fallout from weapons testinghas been estimated at 2.8 kBq m−2 in the Northern Hemisphere(UNSCEAR, 1982).

Earthworms. The numbers and biomass of anecic and endogeicearthworms at Skogsvallen were measured in May and October2008 by the soil-coring method. At four locations two replicatesoil samples were taken with a 20 × 20 cm frame to a depthof 20 cm. All soil was transferred to the laboratory, stored in

© 2009 The AuthorsJournal compilation © 2009 British Society of Soil Science, European Journal of Soil Science, 61, 24–34

Page 5: Modelling the effects of bioturbation on the re-distribution of 137Cs in an undisturbed grassland soil

28 N. J. Jarvis et al.

Figure 2 Field sampling scheme.

a climate chamber (8◦C) and sorted by hand the next day. Thetopsoil (0–5 cm, where most plant roots were found) was alsoextracted in Tullgren funnels to check for smaller individuals thatcan be missed by hand sorting (Lee, 1985). Earthworms wereweighed individually to obtain the fresh weight (without emptyingthe gut content), the body length was determined, and individualswere identified to species using the taxonomic key by Schwert(1990). The animals were then killed by freezing at −20◦C, driedfor 24 hours at 105◦C and weighed again to determine individualbody dry weight as well as biomass.

An estimate of the egestion rate of earthworms at the soilsurface was obtained by measuring faeces production on fourreplicate observation plots each 0.25 m2 in size in September2008. The plots were located 1 m from the spots where thesoil samples had been removed. For each plot, a surface areaof 50 × 50 cm was carefully cleared of vegetation, litter and oldearthworm casts, to facilitate the detection of new casts. After18 days, all casts produced after clearance of the plots werecollected and dried for 24 hours at 105◦C to obtain the dryweight. Cast production was estimated as dry matter per unit areaand time. Because the coring method tends to underestimate thenumbers of deep-burrowing anecic earthworm species (Lawrence& Bowers, 2002), mustard extraction (Gunn, 1992) was alsoapplied in October. Following Hale (2007), 40 g of groundmustard seeds (commercial hot mustard powder) was mixed with4 litres of water and allowed to rest for 3 hours. Immediatelyprior to sampling, the expellant was once again mixed thoroughlyand was poured slowly and evenly across each of the fourobservation plots that had previously been used to estimate soilegestion rates. The expellant was applied twice with an intervalof 15 minutes, applying 4 litres on each occasion. Each plot wasobserved for a period of 30 minutes after the first application of theexpellant. All emerging earthworms were collected with forcepsand processed in the same way as the animals captured with thecoring method.

Dye-tracing. A dye-tracing experiment was carried out accordingto the procedure described by Hagedorn & Bundt (2002).A solution of the weakly sorbed dye-tracer Brilliant Blue FCFwas prepared at a concentration of 4 g litre−1. On 6 October2008, 12 litres of the dye solution was applied to four 0.6 × 0.6m replicate plots using a watering can. About 1 hour after applyingthe dye, a vertical soil profile was prepared 15-cm into thedyed area. The soil was uniformly dyed to a depth of about10 cm. Below this depth the dye was concentrated in well-defined preferential flow paths. Soil samples were, therefore,taken from stained and unstained soil at 10–20-cm depth. Newvertical profiles were subsequently prepared and sampled untilapproximately 100 g of both stained and unstained soil had beencollected from each replicate plot. 137Cs activity was measuredon these samples as described above. Each vertical profile wasphotographed.

Sorption. Soil at the field water content was sampled at depthsof 0–5 and 5–20 cm at Skogsvallen in June 2004 and kept insealed plastic bags. In the laboratory, the soil was passed throughan 8-mm sieve and spiked with 50.5 kBq 134Cs kg−1, equivalentto roughly two and a half times the fallout of 137Cs received atthe site. 134Cs was used because its shorter half-life meant thatthe background concentration in the soil was negligible. The soilwas mixed uniformly into plastic pots at approximately 1.5 kgper pot, and was left to equilibrate for periods of up to 10weeks. Four replicate 10-g soil samples were taken after 1 dayand 1, 4, 9 and 10 weeks and 134Cs activities measured afterextraction by ‘end-over-end’ shaking for 24 hours in 20 ml of0.01 m Ca(NO3)2. After centrifugation and filtration, the 134Csactivity in the solution was measured by gamma emission in thegermanium detector as described earlier and Kd was subsequentlycalculated. Extraction with 0.01 m Ca(NO3)2 is not a standardmethod for studies of Cs sorption, but was chosen because itreflects reasonably well the soil solution composition in the field(Holm et al., 1998).

© 2009 The AuthorsJournal compilation © 2009 British Society of Soil Science, European Journal of Soil Science, 61, 24–34

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Bioturbation and 137Cs distribution 29

Table 2 Model parameter values (values within parentheses for Db , Inl

and fsurf indicate the parameter ranges used in the calibration procedure)

Parameter Value

Dispersivity, λ 5, 10 cmDiffusion coefficient in

water, Dw

2 × 10−9 m2 s−1

Tortuosity factor, ξ 0.5Water content, θ 0.4 m3 m−3

Darcy flow rate, q 0.166, 0.266 m year−1

Bulk density, γ See Table 1Sorption constant, Kd 7.3/20.1 m3 kg−1a

Bio-diffusioncoefficient, Db

3 × 10−5 m2 year−1 (0–6 × 10−5 m2 year−1)

Ingestion rate, Inl 6 kg m−3 year−1 (0–10 kg m−3 year−1)

Depth ofbioturbation, L

50 cm

Fraction egested atsurface, fsurf

(non-local feeders)

0.1 (0–1)

First-order rateconstant, μ

0.023 year−1

aAt 0–5 and 5–20 cm, respectively.

Results and discussion

Model parameterization and modelling strategy

Table 2 shows the parameter values used in the simulations.Some parameters controlling physical transport (Dw , γ , Kd ) anddecay (μ) were set to either known or measured values. Otherswere set to fixed values (ξ , θ ) because preliminary simulationsshowed that they were very insensitive within physically rea-sonable ranges. The extreme insensitivity of transport to watercontent was also indicated by substitution of appropriate valuesfor θ and γ.Kd in Equation (6). A ‘best guess’ for the waterflow rate, q, was arrived at by subtracting an assumed evapo-transpiration of 400 mm year−1 from the known annual rainfall.Likewise, the dispersivity was set to a typical value of 5 cm(Vanderborght & Vereecken, 2007). For comparative purposes,a second simulation was run with extreme values for both dis-persivity (10 cm) and Darcy water flow rate (0.266 m year−1,assuming that evapotranspiration was only 300 mm year−1, seeTable 2). The sorption constant Kd was set to values based on themeasurements in bulk soil samples collected from Skogsvallen,backed up by literature estimates. Figure 3 shows that sorptionKd values measured on soil at 0- to 5-cm depth increased dur-ing the first week, but were thereafter stable, with no significantdifference between sampling dates, and a mean value of 7.3 m3

kg−1 from 1 to 10 weeks after application (see Table 2). Unfortu-nately, no reliable data were obtained from the experiments on soilat 5–20 cm depth, because sorption was so strong that estimatesof the activity in solution (and therefore the Kd value) were toouncertain. We therefore relied on literature data to estimate thesorption constant at 5–20-cm depth. Using a similar methodol-ogy to our experiment, Sanchez et al. (2002) measured desorption

Figure 3 Sorption Kd at 0- to 5-cm depth in Skogsvallen soil (n = 4,error bars indicate 95% confidence intervals of the mean).

of caesium from 30 mineral soils from Belgium in 7-week-longpot experiments. Applying multiple linear regression to their data(R2 = 0.75, standard deviation of the estimate = 0.478, n = 30)gives Kd (l kg−1) as:

log Kd = −0.22 + 0.455pH + 1.172 log fclay − 0.969 log K,

(8)

where pH is measured in 0.01 m calcium chloride, fclay is thepercentage clay content (on a whole soil basis) and K is theexchangeable potassium content (cmolc kg−1) measured in 1 mNH4OAc (at pH 7). An estimated Kd value of 20.1 m3 kg−1 at5–20 cm depth is obtained (Table 2) by substituting appropriatemeasured values for pH, fclay and K at Skogsvallen (Table 1)into Equation (8). Furthermore, the Kd value estimated byEquation (8) at 0–5-cm depth (= 9.36 m3 kg−1) was similar toour measured value (7.3 m3 kg−1), which gives us confidencethat the Kd value estimated at 5–20 cm depth should also bereasonable. The larger Kd value predicted deeper in the soil isprimarily because of the smaller exchangeable potassium content(see Table 1 and Equation (8)).

We initially assumed that endogeic earthworms would beresponsible for ‘local’ bio-diffusion, while anecic earthwormswould cause ‘non-local’ transport within the soil. The depth ofsoil consumption/egestion by ‘non-local’ mixers, L (Table 2),was fixed at 50 cm, on the basis of field observations andthe assumption that high water tables in winter at this poorlydrained site should prevent deeper burrow formation. Preliminarytests also showed that simulation results were insensitive to L,within the range tested (30–80 cm). Initial estimates of twosensitive model parameters describing biological transport (Db

and Inl , Table 2) were estimated by combining measurements ofearthworm biomass at Skogsvallen (see Table 3) with literaturedata on feeding rates. The bio-diffusion coefficient, Db, can be

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30 N. J. Jarvis et al.

Table 3 Abundance and biomass of endogeic (Aporrectodea caliginosa

Savigny and A. rosea Savigny) and anecic (Lumbricus terrestris L.)earthworms at Skogsvallen in 2008 (n = 4) measured by soil coring at0–20-cm depth and mustard extraction, respectively

Numbers m−2Biomass / g fresh

weight m−2

Type Season MeanStandarddeviation Mean

Standarddeviation

Endogeic Spring 169.0 93.7 70.5 48.9Autumn 215.8 57.1 96.5 54.6

Anecic Autumn 7.0 6.8 8.5 11.0

defined as one half of the square of a ‘step length’ (the averagedistance particles are moved) divided by the elapsed time betweenmovements (the ‘rest period’, Wheatcroft et al., 1990). Theseterms can be related to biological parameters. The rest period� is given by (Wheatcroft et al., 1990):

� = γ

Il

= γ

flBl

, (9)

where Il , fl and Bl are the soil ingestion rate (kg soil m−3 s−1),feeding rate (kg soil kg−1 biomass s−1) and biomass (kg m−3) of‘local’ mixers, while the step length can be approximated by halfthe average body length of the animal, l, providing the mixing isisotropic. Thus, we have:

Db = l2Il

8γ. (10)

Db is often assumed to decrease with depth (e.g. Cousins et al.,1999b), because earthworm biomass, and thus Il , tends to decreasewith depth, while γ increases (Table 1). However, this may becompensated for by the increase in earthworm body length, l,with depth found at Skogsvallen (data not shown). We thereforeassumed a constant value for Db in the entire simulated profileto 22.5-cm depth, even though the change in soil structure at theinterface between A and B horizons in the soil profile suggests thatbioturbation may be most intense in the uppermost 10 cm (see sitedetails and Table 1). However, little 137Cs had penetrated below10 cm by the time of the last sampling in 2007, so assumptionsabout changes in Db below this depth are presently of littlepractical consequence.

Table 3 shows the numbers and biomass of endogeic (mostlyAporrectodea caliginosa Savigny but also Aporrectodea roseaSavigny) and anecic (Lumbricus terrestris L.) earthworms mea-sured by soil coring at 0–20-cm depth and mustard extraction,respectively. These values are at the low end of the range previ-ously reported for earthworm abundance and biomass in humidtemperate grasslands (see Edwards & Bohlen, 1996), but aresimilar to those reported in colder Nordic climates (e.g. Bostrom& Lofs, 1996; Lagerlof et al., 2002; Persson et al., 2007). Nostatistically significant difference in abundance or biomass wasfound between spring and autumn samplings for the endogeic

earthworms. Thus, Bl was set to the overall mean value of 83.5 gfresh weight m−2 (= 0.42 kg m−3 in the 20 cm sampling depth).The mean body length, l, for endogeic earthworms at Skogsvallenwas 5.5 cm. Assuming a typical feeding rate, fl , of 1 g soil g−1

fresh weight d−1 (Curry & Schmidt, 2007) and a feeding seasonof 200 days, gives estimates of Il and Db of 84 kg m−3 year−1

(equivalent to a turnover rate of topsoil of c. 8% per year) and3.1 m2 year−1, respectively.

The ingestion of anecic earthworms (‘non-local’ feeders) wasestimated in a similar way. The measured biomass was 8.5 g freshweight m−2 (= 0.017 kg m−3 if the sampling depth is assumed tobe 0.5 m, Table 2). The low end of the range of the few reportedsoil ingestion rates for Lumbricus terrestris L. is of the order of0.5 to 0.7 g soil g−1 fresh weight day−1 (Curry & Schmidt, 2007).We estimated a value of 0.17 g soil g−1 fresh weight day−1, on thebasis of our own (unpublished) experiments on gut contents andgut transit times for L. terrestris. Assuming a value in the middleof this range (0.3 g soil g−1 fresh weight day−1) and, as before,a feeding season of 200 days, gives an initial estimate of Inl of1.0 kg m−3 year−1, equivalent to a turnover rate of approximately0.1% per year. The measured excretion rate at the soil surface was0.77 g m−2 day−1 or 0.154 kg m−2 year−1 for a 200-day season,corresponding to a value of fsurf of 0.31. The measured excretionrate at the soil surface is similar to previous measurements made atother pasture sites in Sweden (Persson et al., 2007). It is, however,small compared with data reported from temperate grasslands,which range from 1 to 10 kg m−2 year−1 (Edwards & Bohlen,1996). This may result from several factors, including the smallbiomass of anecic earthworms (Table 3), a short active seasonbecause of the cold climate and a small topsoil bulk density(Table 1), which may also decrease casting on the surface (Binet& Le Bayon, 1999).

These initial estimates of Db and Inl were used to define arange within which optimum parameter values were identified bycalibration (Table 2). fsurf was also included in this calibration,which was performed with SUFI, a forward, iterative, globalsearch procedure (Abbaspour et al., 1997). The relative contri-butions of physical (i.e. water-borne) and biological transportmechanisms were illustrated by a step-wise simulation approach.The full model was first fitted to the data. ‘Non-local’ biolog-ical transport was then excluded from the model by setting Inl

to zero. In this simulation, Db was re-calibrated. Finally, onlyphysical transport was simulated by also setting Db to zero.

Comparisons of model simulations and measurements

Figure 4 compares model simulations of the re-distributionof 137Cs in soil with the field measurements made on sixoccasions during the 20-year period (1987–2007). The agreementis generally very good, with predictions mostly lying withinconfidence intervals for the means of measured data and modelefficiencies (Loague & Green, 1991) exceeding 0.9 on all samplingoccasions but the first. The calibrated parameter value for Db (seeTable 2) was remarkably close to the value initially estimated from

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Bioturbation and 137Cs distribution 31

(a) (b)

(c) (d)

(e) (f)

Figure 4 Comparison of measurements(symbols are mean values, with 95% confi-dence intervals of the means shown as dottedlines) and model predictions (solid lines) of137Cs distribution in the soil at Skogsvallenon the six sampling occasions between 1987and 2007. EF is the model efficiency (Loague& Green, 1991), calculated on log10 values inorder to give more equal weight to the datafrom different depths.

measured earthworm biomass and literature data on soil ingestionrates. In contrast, the calibrated value of Inl (= 6 kg m−3 year−1,Table 2) was six times larger than the initial estimate. The best-fit to the data was obtained by assuming that only 10% of thesoil ingested by non-local mixers in the 0.5-m thick bioturbatedlayer was egested at the soil surface (fsurf = 0.1, Table 2). Takentogether, these calibrated parameter estimates would predict anegestion rate at the soil surface of 0.3 kg m−2 year−1, which canbe compared with a measured rate of 0.154 kg m−2 year−1. Thislevel of agreement is encouraging when the uncertainties in modelprocess descriptions and parameterization and also the limitedspatial and temporal resolution of the soil excretion measurements(18 days in autumn, four replicate 0.25-m2 plots) are considered.

The simpler model that excluded ‘non-local’ biological trans-port also matched the data reasonably well (Figure 5), assum-ing a slightly larger bio-diffusion coefficient (= 3.4 × 10−5 m2

year−1). This suggests that ‘local mixing’ attributed to endogeicearthworms is the dominant transport mechanism for 137Cs at

Skogsvallen. However, this simpler model failed to predict thereduced activity close to the soil surface (Figures 4 and 5), whichis presumably caused by surface egestion of soil (McCabe et al.,1991; VandenBygaart et al., 1998). Unlike the model accountingfor ‘non-local’ biological transport, it also fails to predict the deeppenetration of small amounts of 137Cs that was observed belowapproximately 10–12-cm depth. Despite the assumption of lin-ear sorption, the physical advection-dispersion model seriouslyunder-estimates the downward displacement of 137Cs in soil (seeFigure 5) even when extreme values are assumed for dispersivityand Darcy water flow rate.

Our assumption that the feeding activity of endogeic earth-worms only produces ‘local’ mixing may be the reason for thelarger than expected calibrated estimate of Inl . It seems reason-able to suppose that large adults could also produce ‘non-local’mixing. However, it is also possible that a large estimate of Inl

may be compensating for the fact that the model does not accountfor preferential transport of 137Cs in soil macropores. The effects

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32 N. J. Jarvis et al.

D C BA

0

2

4

6

Dep

th (

cm)

8

10

12

0

137 Cs activity (kBq kg-1)

1 2 3 4 5 6

Figure 5 Comparison of measurements in 2004 (symbols) with modelpredictions: A (dotted line), physical transport only, with best-guessestimates of λ (5 cm) and q (0.166 m year−1); B (short dashedline, physical transport only, with extreme values of λ (10 cm) andq (0.266 m year−1); C (long dashed line), as A, but includes bio-diffusion;D (solid line), the complete model, i.e. as C, but also accounting for‘non-local’ biological transport. Confidence intervals for the mean of themeasurements (see Figure 4) are not shown in order to improve clarity.

of this physical non-equilibrium process on 137Cs distribution insoil would be similar to ‘non-local’ biological transport. The dye-tracing experiment was carried out to investigate this possibility.

Dye tracing

Figure 6 shows one typical profile excavated after the dyestaining experiment. The uppermost 8–10 cm of soil (the Ahhorizon, Table 1) was uniformly stained, indicating homogeneousinfiltration and a lack of marked preferential flow, presumably as aresult of the fine crumb structure produced by intense bioturbation.Preferential flow pathways were generated at the interface with themore coarsely aggregated B horizon (Figure 6, Table 1). Almostall of these stained flow pathways were inter-aggregate fissuresand cracks, although a few stained earthworm channels were alsoobserved.

137Cs activities were, on average, approximately three timeslarger in stained flow pathways than in the unstained matrix(Table 4). This suggests that the deep penetration of 137Cs inthis structured clay soil may, at least partly, result from preferen-tial transport, predominantly in inter-aggregate fissures. However,there was considerable variation between plots (Table 4) and thenumber of replicates was small. Thus although a t-test assumingequal variances suggested that 137Cs activities were significantlylarger (P = 0.049) in stained flow pathways than in unstained soilmatrix at 10–20 cm depth, a non-parametric test (Wilcoxon RankSum Test) suggested that they were not (P = 0.11). Furthermore,both parametric (P = 0.21) and non-parametric (Kruskal-Wallis,

Figure 6 Photograph of one vertical soil profile prepared after thedye-staining experiment.

Table 4 137Cs activity measured on stained and unstained soil at10–20-cm depth at Skogsvallen in 2008 and on bulk core samples atthe same depth in 2004 and 2007

137Cs activity / B q kg−1 dry weight

Dye tracing in 2008 Core sampling

Replicate Stained flow paths Unstained soil matrix 2004 2007

1 39 61 51 1192 60 20 37 363 113 5 31 444 118 17 – –Mean 83 26 39 66Standard

deviation39 24 10 46

P = 0.31) one-way analysis of variance tests showed that 137Csactivities measured on bulk soil core data in 2004 and 2007 at10- to 20-cm depth were not significantly different from thosemeasured in the unstained soil matrix in 2008 (Table 4), whichsuggests that the effect of preferential flow on the depth distribu-tion of Cs measured by core sampling in the soil profile has beenquite limited. The reason for this must be that soil with enhanced137Cs contents in contact with preferential flow pathways consti-tutes only a small fraction of the total soil mass extracted by thecores. Thus, although some preferential transport of caesium mayhave occurred, biological transport mechanisms seem dominantand must be invoked to explain the significant activities of 137Csmeasured in bulk soil at 10–20-cm depth.

Conclusions

A physical advective-dispersive transport model based onmeasured sorption at 0- to 5-cm depth strongly under-estimateddownward movement of 137Cs at Skogsvallen. Although a dye-tracing experiment suggested the occurrence of physical non-equilibrium transport in soil macropores, this was insufficient toexplain the extent of the deep penetration of 137Cs. It is concluded

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Bioturbation and 137Cs distribution 33

that biological transport mechanisms dominate the downwarddisplacement of 137Cs observed at the site. A simple bio-diffusionmodel representing ‘local’ mixing processes matched the observeddepth distributions reasonably well, but could not reproduce thedeep penetration of Cs, or a dilution observed close to the soil sur-face. These phenomena appear to result from ‘non-local’ mixing,resulting from the activities of large individuals of both endogeicand anecic earthworm species. A comprehensive model includ-ing physical (advective-dispersive) transport, and both ‘local’ and‘non-local’ bioturbation gave an excellent match to the measureddepth profiles of 137Cs.

Acknowledgements

We are grateful to the many students over the years who havehelped with field sampling and analysis and to Anna Stromqvistwho carried out the desorption experiments on Skogsvallen soil.The measurement programme at Skogsvallen was financed bySLU (Swedish University of Agricultural Sciences) and SSM(Swedish Radiation Safety Authority).

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