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Effect of forest edges on deposition of radioactive aerosols
Atmospheric Environment 36 (2002) 5595–5606
Effect of forest edges on deposition of radioactive aerosols Z. Ould-Dadaa,*, D. Copplestoneb, M. Toalc, G. Shawd a
Food Standards Agency, Radiological Protection and Research Management Division, Room 715B, Aviation House, 125 Kingsway, London WC2B 6NH, UK b ERC, University of Liverpool, Vanguard Way, Birkenhead, Wirral CH41 9HX, UK c Centre for Ecology and Hydrology, Pollution and Ecotoxicology Section, Monks Wood, Abbots Ripton, Huntingdon, Cambridgeshire PE17 2LS, UK d Department of Environmental Science & Technology, Faculty of Life Sciences, Imperial College at Silwood Park, Ascot, Berkshire SL5 7PY, UK Received 20 March 2002; received in revised form 1 August 2002; accepted 23 August 2002
Abstract The possible enhancement of aerosol deposition at forest edges was investigated in a wind tunnel and in the field. The wind tunnel study was carried out using 0.82 mm mass median aerodynamic diameter uranium particles and a composite canopy of rye grass and spruce saplings. The field study was undertaken at a coniferous woodland near to BNFL Sellafield, Cumbria, UK. Two transects were set through the woodland to determine the influence of the forest edge on atmospheric deposition of radionuclides released under authorisation from the Sellafield site. Results from the wind tunnel study showed that the deposition flux of uranium particles decreased with distance downwind from the grass– tree edge towards the interior of the canopy. The deposition flux at the edge was maximal at about 4 107 mg of U cm2 s1. This was 3 times higher than that observed over grass where a constant flux of about 1.32 107 mg of U cm2 s1 occurred. Results from the field study showed a clear influence of the forest edge on the atmospheric deposition of 241Am and 137Cs. Activity depositions of around 4750 and 230 Bqm2 for 137Cs and 241Am, respectively, were measured in front of the woodland. Activity deposition inside the forest edge, however, rose to levels of between 20,200 and 50,900 Bq m2 and 1100 and 3200 Bq m2 for 137Cs and 241Am, respectively, depending upon the transect. Similar activity concentrations were measured in the pasture to the front and behind Lady Wood. Results from these studies corroborate those obtained from various studies on air pollutants including radionuclides. This underlines the importance of deposition at the edge of forests and its contribution to the overall canopy deposition. The edge effect is therefore an important factor that should be considered in the assessment of fallout impact, whether this is to be made by either direct sampling or by modelling. r 2002 Elsevier Science Ltd. All rights reserved. Keywords: Radionuclides; Radioactive aerosols; Forest edge; Deposition; Wind tunnel; Field study
1. Introduction The process of dry deposition of airborne pollutants is influenced by various physical, chemical and biological *Corresponding author. Tel.: +44-207-276-8774; fax: +44207-276-8779. E-mail address: [email protected]. gov.uk (Z. Ould-Dada).
factors that can cause large variations in deposition on both small and large scales of time and space. Edges and other structural inhomogeneities of forests are important features that provide situations in which many factors controlling the deposition process change rapidly within small distances. The effect of forest edges on deposition of radionuclides to forest canopies was an important phenomenon observed in contaminated forests in the vicinity of the Kyshtym (southern Urals,
1352-2310/02/$ - see front matter r 2002 Elsevier Science Ltd. All rights reserved. PII: S 1 3 5 2 - 2 3 1 0 ( 0 2 ) 0 0 6 9 9 - 4
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1957) and Chernobyl accidents, 1986. After the Kyshtym accident of 1957, Tikhomirov and Shcheglov (1994) observed an enhancement in radionuclide deposition at the forest edge facing the source of release. Following the Chernobyl accident, however, the same authors reported gamma dose rate measurements, collected in an area situated at approximately 6 km to the west of the damaged reactor, which indicated a similar level of contamination in both forested plots and in adjacent pasture areas. This indicates a similar degree of deposition of radioactive materials in both forests and pasture, with none of the expected enhancement of deposition at the margin between the two. Several other authors, however, have found an increased deposition of airborne pollutants (not radionuclides) at forest edges (e.g. Raynor et al., 1974; Hasselrot and Grennfelt, 1987; Draaijers et al., 1988; Beier and Gundersen, 1989; Beier, 1991; Beier et al., 1992; Neal et al., 1994). This leads to the question of the likely role of the forest edge in controlling the deposition of radioactive aerosols. Forest edges are commonly exploited for food by both man and animals and this may lead to high exposure to contaminants given that enhanced deposition can occur within this area. The Forest Working Group of the Biosphere Modelling and Assessment programme (BIOMASS, organised by the International Atomic Energy Agency, Vienna), recognised the importance of forest edges in radioactive contamination of forests and stressed the need for the collection of relevant data and development of models (IAEA, 2001). The objective
of this paper is to examine the possible enhancement of aerosol deposition at the edge of forests using results from wind tunnel and field studies.
2. Wind tunnel study 2.1. Construction of composite grass and spruce canopy Swards of perennial rye grass mixture (Lolium perenne, Permogen long-term ley) were germinated in a peat-based compost in shallow plastic trays (4 cm 38 cm 38 cm). These swards were allowed to grow inside a green house to a height of approximately 5 cm. Ten trays of grass were then placed inside the wind tunnel covering the whole length of bays 1 and 2 (Fig. 1). Bay 3 of the tunnel was filled with 12 Norway Spruce saplings (Picea abies L.) positioned in six rows with an average height of 47 cm, i.e. about 10 times that of the grass. The height of the trunk space of the trees varied between 3 and 7 cm with an average value of 5.2 cm. A peat-based compost was used to provide a soil cover under the tree canopy; the soil was kept moist throughout the experiment to minimise any resuspension of particles from its surface. The soil surface of the tree stand was made level with that of grass. The soil level was adjusted by defining a height (i.e. the zero plane displacement) so that the distribution of drag over the full canopy depth is aerodynamically equivalent to the imposition of the entire stress at this height. The zero plane displacement was defined such that about 2/3 of
Fig. 1. Dimensions of grass and spruce canopies and positions of aerosol source and measurement points.
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the grass height was below the level of the airflow entering the working section. The positioning of grass and spruce canopies in the wind tunnel is shown in Fig. 1. 2.2. Meteorological measurements A free stream air velocity of 5 m s1 was established at about 30 cm above the tree canopy using hot wire anemometry. Wind speed measurements were made vertically at 10 mm intervals from a height of 0.87 cm above the top of the grass canopy and from 20 mm above the ground for the stand of trees. Measurements of vertical air speed profiles were made at four positions along the wind tunnel, i.e. 45, 5, +35 and +175 cm from the grass–tree margin (Fig. 1). These measurements allowed calculations of friction velocity, u ; (using the eddy correlation method) from the following relationship (see Kinnersley et al., 1994):
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grass–tree margin. The procedures for release and sampling of aerosol are described in detail in OuldDada (2002). 2.3.2. Sampling and analysis of plant material At the end of the experiment, each tree was divided into five horizontal layers of equal height and all trees, excluding the tree ‘trunk’, were sampled. Within each layer, new and old growth tissues were divided and 18 tissue categories obtained. Samples from the grass canopy were taken at four locations (15, 45, 90 and 120 cm) from the grass–tree margin and were oven dried at 701C for 24 h. The mass of uranium contained in each sample and each filter paper was determined by delayed neutron counting (DNC). The sampling procedure, the preparation of samples for irradiation and the analysis of samples are explained in detail in Ould-Dada (2002).
U* ¼ ðu0 v0 Þ0:5 ðm s1 Þ; where u0 and v0 are horizontal and vertical fluctuations on the mean wind velocity (m s1). The variation of the mean wind speed over time was measured as turbulence intensity, iu; at each sampling position. Assuming isotropic turbulence, iu was defined as the mean root square fluctuation on the mean velocity u (see Kinnersley et al., 1994): iu ¼ ðu02 Þ0:5 =u: Temperature and relative humidity were measured up to a height of 80 cm from ground level in the tree canopy and from the top of the grass. Measurements were made at 130, 20, +35 and +175 cm from the grass–tree margin. 2.3. Contamination and sampling 2.3.1. Release and sampling of aerosol The uranium aerosol used in this study consisted of 0.82 mm mass median aerodynamic diameter (MMAD) (0.45 mm mass median diameter, MMD) dry spherical particles produced by the aerosol generation technique described in Ould-Dada (2002). The aerosol generator was positioned adjacent to the wind tunnel throughout the experiment and particles were liberated into the tunnel via a short flexible hose and an injection manifold system for a period of 60 h. The injection manifold was adjusted so that the bottom row of injection nozzles was about 6 cm above the top of the grass canopy (Fig. 1). Fourteen isokinetic air samplers were used to sample aerosol particles at various levels within and above the tree canopy in order to determine their vertical air concentration profile (Fig. 1). The flow rate through each sampler was determined from the wind velocity profile measured at 75 cm distance from the
3. Field study Lady Wood (OSGR: NY 037045) is a coniferous woodland approximately 4.7 ha in size located 500 m northeast of BNFL Sellafield, Cumbria, UK. The woodland is dominated by a uniform stand of sitka spruce, Picea sitchensis, planted in the 1950s. It is surrounded by agricultural pasture and there are few grass and herbaceous species present beneath the dense canopy. The site has been characterised previously for radionuclide contamination derived primarily from the authorised discharges from BNFL Sellafield (Copplestone et al., 1999, 2000; Toal et al., 2001). The wood faces Sellafield with no major intervening natural or artificial features to intercept airborne particulates. Along the woodland edge facing the principal source of radionuclides, there is a steep slope varying between 5% and 33%, which presents a large surface area of canopy for the interception of radionuclides from the ambient atmospheric aerosol. For this work, two transects were set through the woodland at right angles to the front woodland edge. The transects were 230 and 260 m in length (transects 1 and 2, respectively) and were 50 m apart. The difference in length resulted from the non-uniform shape of Lady Wood. Soil samples were collected at intervals 20 m apart within the woodland and at 5, 10, 15, 20, 30, 40 and 50 m from the front and back woodland edges. The soil samples consisted of cores of 10 cm diameter, extracted to a depth of 9 cm. Each core was oven dried at 1051C, homogenised in a rotary mill and sub-sampled for determination of 137Cs and 241Am activity concentrations by gamma spectrometry. Details of sampling procedures and analytical techniques are presented elsewhere (Copplestone, 1996).
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4. Results and discussion 4.1. Wind tunnel study 4.1.1. Wind velocity profiles The shape of the wind profile above the grass, at the 45 cm position, is simple with a boundary layer of less than 20 cm height whereas at the +175 cm position, a more complex profile with a ‘jet’ within the trunk space and a ‘deeper’ boundary layer typical of a tree canopy was established (Fig. 2). The upper part of the latter profile reflects the retarding effect of the wind tunnel ceiling. Because there was only about 0.70 m of space between the top of the trees and the ceiling of the wind tunnel, the wind speed profile at the +175 cm position was ‘squeezed’. Between these two extreme positions (i.e. 45 and +175 cm), the boundary layer is subject to
rapid alteration and deformation reflected in a scatter of points in the profile from the +35 cm position indicating an enhancement of turbulence. Friction velocity increased from about 0.5 m s1 over the grass canopy to about 1.5 m s1 over the stand of trees (Fig. 3). This reflects the considerably larger drag exerted by trees on the airflow than the grass and the air is decelerated as it moves from the smoother to the rougher surface. Similar results were obtained by Meroney (1970) who carried out wind tunnel studies, at a free stream velocity of 6 m s1, of the airflow at the leading edge and downstream of a model plastic tree canopy of about 18 cm height. The distribution of the foliage with height at the edge of the stand as well as inside the forest will affect the movement of the emerging wind. When the wind arrives at a forest edge, it penetrates and moves through the
Fig. 2. Wind velocity profiles over grass and spruce canopies.
Fig. 3. Friction velocity profiles over grass and spruce canopies.
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canopy forming a jet in the trunk space, which is usually more open compared to the canopy crowns. Even in the case of horizontal airflow over a uniform tree canopy, the efficiency of the airflow in penetrating downwards through the canopy will depend particularly on the foliage density. Measurements of turbulent fluxes of momentum in the extreme transition from a heathland of about 0.25 m height to a forest of approximately 10 m height were made by Gash (1986) who found that the flux reached an equilibrium at about 120 m from the heath–forest edge, i.e. a distance of approximately 12 tree canopy height. On this basis and on the evidence from Meroney’s (1970) study, therefore, the fetch downwind from the grass–tree margin in this study was insufficient to allow the reestablishment of a stable boundary layer above the tree canopy. However, within the constraints of the wind tunnel dimensions, no greater fetch could be achieved. Nevertheless, the configuration of the experimental canopy allowed measurement of particle deposition fluxes at and immediately behind the grass–tree margin. Since wind speed is one of the controlling parameters of the dry deposition process, the alteration in the airflow and the change in its profile at this edge can be expected to influence the patterns of aerosol deposition in this region.
both the trunk space and within the canopy reaching values of 0.7 and 2.2, respectively (Fig. 4). At this position, iu observed within the tree canopy was nearly 6 times higher than that observed over grass. This can be attributed to the rough surfaces and complex aerodynamic shapes of trees which create more turbulence than do aerodynamically smoother grass surfaces: the different values of iu probably reflect the size of the eddies generated by each surface. Similar profiles of iu were observed in wind tunnel over a plastic tree canopy (Meroney, 1970) and at a 10–13 m tall pine forest (Raynor et al., 1974). Moreover, the shape of iu profile at the +175 cm position was similar to those obtained for tree canopies (with no transition zone) used during deposition (Ould-Dada, 2002) and re-suspension (Ould-Dada and Baghini, 2001) experiments linked to the edge-effect study described here. The maximum value of iu in those profiles, which was also recorded at about the middle of the crown was, however, 5.5 times lower than that observed here. Turbulence is another factor controlling the dry deposition process and variations in iu observed here will affect the movement of aerosols and hence the deposition of airborne material to tree surfaces. Turbulence is likely to be influenced by both foliage density and the prevailing meteorological conditions.
4.1.2. Turbulence profiles Turbulence intensity (iu ) was measured at the same positions and height increments as wind velocity. Above the grass canopy, iu varied between 0.05 at a height of about 20 cm and 0.42 at the canopy top, i.e. 5 cm height (Fig. 4). At +35 cm from the grass–tree transition point, iu increased with height reaching a maximum value of 0.46 in the middle of the tree crown and 0.2 in the trunk space. Turbulence intensity remained constant above the tree canopy at this point. At +175 cm from the grass– tree edge, however, turbulence intensity increased in
4.1.3. Air temperature and humidity profiles Temperature and relative humidity were measured from ground level up to a height of 0.80 m at 130, 20, +35 and +175 cm from the grass–tree edge. Both temperature and humidity were constant with height above the grass canopy. Variations were confined within 21C (11C for the tree stand) and 5% for temperature and relative humidity, respectively. Temperature and relative humidity measured outside the working section of the tunnel were 10.61C and 45.8%, respectively. Variation in temperature is an additional factor to be taken into
Fig. 4. Turbulence intensity profiles over grass and spruce canopies.
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shown by measurements made at +175 cm from the edge. The enhancement in aerosol deposition flux observed at the edge of the tree stand can be explained by the alteration in the vertical wind profile resulting in an increase of turbulence as shown in Figs. 2 and 4. The two points of aerosol release were set up at a height within that of the tree canopy cross-section and the wind blew in a direction at right angles to the edge. Thus, the trajectory of aerosol particles was directly towards the tree stand edge where a considerable drag is imposed on the moving air resulting in an alteration of its vertical profile as described in Fig. 2. The current observations of enhanced deposition of aerosol particles at the model forest edge over that occurring on grass are in agreement with findings of several other authors. Draaijers et al. (1988) found the increase in dry deposition at forest edges to be strongly dependent on wind velocity and wind direction during ( dry deposition. Wiman and Agren (1985) showed from model studies that the higher wind speed at the forest edge increased the dry deposition of particles. In their study on particulate dispersion into and within a mixed
account when interpreting distribution patterns of deposition of air pollutants since the vertical temperature profile can either enhance or suppress vertical mixing of air and the materials contained within it. 4.1.4. Deposition rates 4.1.4.1. At canopy level. Aerosol deposition fluxes (F ) to both grass and tree canopies were calculated using the following relationship: F ¼ Qu =Ste ðmg cm2 s1 Þ; where Qu is the quantity of total uranium (mg), S is the projected area of plant material (cm2) and te is the time of exposure of canopy to aerosol (s). Deposition fluxes were then expressed in mg of uranium deposited per cm2 of ground area per second by multiplying the flux per unit area of plant material (i.e. F ) by the LAI. Deposition velocity (Vg ) was calculated as the flux divided by the mean air concentration (i.e. 3.1 107 mg U cm3) measured within the boundary layer above the tree canopy (see Ould-Dada, 2002). The results shown in Table 1 indicate a Vg of 1.28 cm s1 at the edge of the tree canopy, which is about 2.5 times compared with that measured inside the tree canopy (0.54 cm s1). The latter value is very consistent with that obtained for tree canopies (with no transition zone) during a wind tunnel deposition study (0.5 cm s1; see Ould-Dada, 2002). The results shown in Fig. 5 indicate a nearly constant deposition flux of about 1.32 107 mg U cm2 s1 over the grass canopy. This value was, however, three times higher at the edge of the tree canopy where a maximum flux of 3.97 107 mg U cm2 s1 was recorded. This flux decreased downwind of the edge showing a secondary ‘peak’ of 1.86 107 mg U cm2 s1 at +135 cm from the edge. This secondary ‘peak’ could be the result of the high turbulence developing inside the stand of trees as was
Fig. 5. Variation of uranium aerosol flux to grass and spruce with distance from grass–tree edge. Error bars indicate 7s.e.m. (standard error of the mean).
Table 1 Average LAI (cm2 cm2) at each tree layer and Vg values as a function of distance from grass–tree edge in the wind tunnel experiment (based on two trees at each distance) Distance from grass–tree edge (cm)
Vg to trees (cm s1)
LAI Tree layers
0 41 73 105 135 170 Mean a
Excludes Vg value at distance 0.
Total
1
2
3
4
5
0.17 0.14 0.08 0.09 0.13 0.14
0.80 0.63 0.42 0.53 0.29 0.49
1.29 1.00 0.51 0.67 0.53 1.20
0.88 0.64 0.89 0.78 1.04 1.00
1.00 0.91 0.57 0.85 0.86 0.62
4.14 3.59 2.37 2.92 2.85 3.45
1.28 0.68 0.47 0.54 0.60 0.40
3.22
0.54a
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forest, Raynor et al. (1974) reported that deposition of stained ragweed pollen and other tracers ranging from 14 to 54 mm in diameter near the forest edge and in the tree tops took place by impaction whereas at the interior of the forest deposition occurred mainly by gravitational settling. Beier and Gundersen (1989) suggested that increased dry deposition at a Norway spruce forest edge could be attributed to a dramatic change in the surface roughness length, caused by the forest edge, which leads to turbulent air movements and lower canopy resistance. Beier (1991) stated that dry deposition at the forest edge may increase due to reduced boundary layer thickness which causes increases in interception and impaction of particles. In his study on throughfall deposition to a Norway spruce forest edge in Denmark, Beier found that deposition of SO+ 4 and Na was 11 and 21 times higher, respectively, at the front tree compared to the open field and 2.4 times and 8.3 times compared to the deposition inside the forest. In Beier’s study deposition decreased exponentially with distance from the forest edge reaching an almost constant flux at 15 m from the edge (B175 and 75 meq m2 yr1 for SO+ 4 and Na , respectively). Hasselrot and Grennfelt (1987) also measured throughfall deposition of SO4, NO3, NH4, Cl, Na, K, Mg and Ca at a pine forest on the Swedish west coast and observed a continuous decrease in the deposition of these elements from the edge downwind into the forest. A constant deposition flux was not reached until more than 50 m from the forest edge. Draaijers et al. (1988) measured atmospheric deposition in Douglas fir (Pseudotsuga menziesii) forest edges by monitoring canopy throughfall and assuming throughfall fluxes to be constant at 200 m from the forest edge. All these studies were conducted in coniferous forests (spruce, pine and Douglas fir) and different results could be due to differences in canopy structure which influence deposition (Beier, 1991). All these studies, however, showed an enhancement of deposition at forest edges with this effect decreasing with distance from the forest edge. In the present study, LAI was more or less constant with distance from the grass–tree edge (Table 1). In the field, however, LAI is often higher at the edge than inside the forest especially when spacing between trees is very small preventing the light from penetrating downward through the crown hence reducing the full development of the crowns inside the canopy. Model ( results from Wiman and Agren (1985) suggested an increase in aerosol deposition close to the edge of forests with dense structure. In a Norway spruce forest, Beier and Gundersen (1989) found LAI a factor of 2 higher at the edge than inside the forest. From 5 m downwind of the forest edge LAI remained nearly constant. According to these authors, higher LAI for trees at the edge could be one of the explanations for the enhancement of dry deposition at the forest edge, although the results of
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the current study suggest that alterations in airflow at the edge of a tree stand are sufficient to enhance the deposition flux of B1 mm aerosols. 4.1.4.2. Between tree layers. Figs. 6 and 7 show the variation of deposited aerosol at each canopy layer with distance from the grass–tree edge. Maximum deposition of material occurred at the top of trees (i.e. layer 1) whereas within the crown deposition varied only slightly but was about four times lower than that recorded at the tree tops (Fig. 6). If layer 1 is excluded, there is an indication that deposition rate decreased from the bottom to the canopy top (Fig. 7). In fact, tests carried out with a smoke generator set up at the same height as the aerosol source revealed that once in contact with the tree canopy edge most of the plume moved through the crowns and in the trunk space although part of the smoke flowed over the top of the trees. At the far edge of the tree canopy, smoke was observed on some occasions to be also moving upwards from inside the tree crowns. Furthermore, measurements of aerosol air concentration made within and above the tree canopy showed a decrease in air concentration with height, though, air concentration profile does probably vary with distance
Fig. 6. Variation of total amounts of uranium deposited to different layers of the spruce trees.
Fig. 7. Variation of total amounts of uranium deposited to different layers of the spruce trees (excluding top layer).
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intercept potentially large quantities of dry deposits. The enhancement of dry deposition at the edge of forests is expected to depend on forest edge characteristics such as tree species, stand density, tree height, open-trunk space and leaf surface area (i.e. LAI). This increase has also been found to vary with other factors. For example, Beier and Gundersen (1989) found the effect of Norway spruce forest edges on the deposition of atmospherically borne substances (Na, Ca, K, Mg, Cl, NO3, NH4 and SO4) to be far greater in the dormant season than in the growing season. The same authors also found that substances deposited as particles (mainly Na+ and Cl) showed the most pronounced edge effect compared to those deposited as gaseous substances. Beier (1991) found that dry deposition of particles increased by a factor of 13 at the front tree, whereas dry deposition of gaseous S increased by a maximum factor of 2.5. Within the vicinity of forest edges, total deposition will include a large proportion originating from edge effects. This has important implications when considering the likely spatial heterogeneity of ground surface contamination following release of radioactive aerosols to the near-surface atmosphere. In areas such as the Chernobyl 30 km zone, in which forests and fields form a complex mosaic, much of the observed variability of radionuclide deposition may be due to the effects of forest edges on dry deposition. Such effects should, therefore, be considered in assessments of fallout impact, whether these are to be made by either direct sampling exercises or modelling.
from the grass–tree edge. This is due to the increased drag exerted by trees on the oncoming flow resulting in a transfer of momentum, hence of aerosol material, from the air stream to the tree surfaces. The air concentration profile observed in this study (no data shown) is the reverse of that obtained in the wind tunnel deposition study (with no transition zone) which was part of the same project (Ould-Dada, 2002). As an explanation, in the deposition study aerosol material was released above a continuous tree canopy where aerosol penetration occurred from the boundary layer down through the canopy. In this case, the canopy layers offer more resistance to the moving air, hence allowing less material to pass from one layer to another. In the current study, however, the aerosol trajectory was directly towards the trees where the airflow moved through the tree crowns and the open trunk space. The latter, offering less resistance to the airflow, allowed more aerosol to circulate in this part of the canopy. Following a series of smoke puff releases at various heights from ground level to the canopy top, Raynor et al. (1974) observed that air penetrating the forest edge moved primarily into the more open trunk space where air speeds were greater than those in the crown for a downwind distance of about 60 m. Their results revealed that the rate of loss of material by deposition in both the canopy and the trunk space was much greater than above the forest or in the open. The authors concluded, however, that the rates of loss and partition of material between layers (i.e. trunk space, canopy, above canopy) were not sensitive to source distance or to height provided that much of the material reaches the forest below treetop height (canopy height was 10–13 m). From the results of the current study and from the literature reports discussed above, it can be seen that forest edges disturb the vertical wind profile and
4.2. Field study The measured 137Cs and 241Am levels (Figs. 8–11) are consistent with other reported work around Sellafield. For example, Jones et al. (1996) report values for Lady
35,000
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Fig. 8.
Cs (Bq m2) deposition in front of, through and behind Lady Wood. Transect 1.
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Fig. 9.
Cs (Bq m2) deposition in front of, through and behind Lady Wood. Transect 2.
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Distance from Front of Woodland (m)
Fig. 10.
Am (Bq m2) deposition in front of, through and behind Lady Wood. Transect 1.
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Wood of 1300 Bq m2 for 241Am and Cawse reports 137 Cs levels of 39,000 Bq m2 (unpublished data). There is generally a higher deposition of radioactivity along transect 2 than transect 1. This may reflect the different surface area of the canopy available for deposition that results from the trees growing on a variable slope along the front edge of the woodland. The slope varies between 5% and 30% along the front edge with the slope at transect 2 being approximately three times as steep as that at transect 1; consequently there is greater deposition of radionuclides along transect 2, e.g. 37.1 kBq m2 for 137Cs compared with 20.2 kBq m2 along transect 1. Airborne particulates that may have radionuclides associated with them can be deposited by two principal
mechanisms: precipitation scavenging (wet deposition) and dry deposition. Precipitation scavenging involves the removal of particulate matter and gases from the air by different forms of precipitation. In contrast, dry deposition occurs continuously within the surface boundary layer with particulate matter being deposited through the process of diffusion, impaction, interception and sedimentation under gravity. These processes have been described more fully elsewhere (e.g. Harrison et al., 1993; Nicholson, 1988). The deposition patterns observed within, and around, Lady Wood (Figs. 8–11) show clear elevation of the levels of 137Cs and 241Am within the woodland (0–100 m, transect 1 and 0–140 m, transect 2) compared with the surrounding agricultural land. From the data available
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0
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Distance from front of wood (m)
Fig. 11.
Am (Bq m2) deposition in front of, through and behind Lady Wood. Transect 2.
241
on radionuclide concentrations within the woodland, it is believed that dryness is the most important factor in determining the pattern of distribution given that precipitation scavenging would be expected to produce a uniform distribution of deposition over the area receiving rainfall (Harrison et al., 1993). Copplestone et al. (2000) reported a uniform soil distribution of 134Cs throughout Lady Wood that varied between 250 Bq m2 at the front of the woodland nearest to Sellafield to 300 Bq m2 at the back of the woodland. Since Sellafield only releases minor quantities of airborne 134Cs, it is likely that the majority was deposited as a result of the passage of the Chernobyl plume (Jackson et al., 1987). Given that the majority of the radionuclide deposition from the Chernobyl plume occurred as a result of precipitation scavenging, a uniform distribution pattern would be expected. This is borne out by the 134Cs data observed. A similar pattern of deposition would have occurred for Chernobyl derived 137Cs. However, as the data in Figs. 8–11 clearly demonstrate, the deposition of 137Cs is not uniform and, based on the wind tunnel experiments reported here and by other workers, e.g. Beier and Gundersen (1989) and Neal et al. (1994), an elevation in the activity concentrations of the radionuclides within the woodland would be expected if dry deposition was occurring. The wind tunnel experiments have demonstrated the significance and mechanisms involved with dry deposition on the pattern of the uranium particle deposition to model spruce canopies. Consequently, Lady Wood provides field data as evidence of the underlying mechanisms described by the wind tunnel experiments. The elevated levels in Lady Wood compared with the surrounding pasture can be explained then by changes in
boundary layer dynamics as the air flows over the grassland in front of Lady Wood and into the forest edge. As mentioned previously, the canopy of a woodland acts as a porous barrier which can allow air to partially pass. This, however, is dependent upon the foliage density of the woodland. These factors will cause eddies to be produced within the canopy structure which will in turn give rise to increased turbulence and reduced velocity within the canopy. Within the time frame and scope of the Lady Wood study, it was not possible to measure these parameters directly within the canopy. However, the data provide good agreement with those from the wind tunnel experiments, thus demonstrating that the mechanisms involved in the deposition process are similar. In Lady Wood the foliage is very dense and the trees are narrowly spaced apart. It was therefore expected that the Lady Wood forest edge would have a significant impact on the airflow and consequent deposition of airborne particulates and any associated radionuclides within this field site. As demonstrated by the data presented in Figs. 8–11, there is a significant increase in the deposition of radionuclides within the woodland. This indicates the importance of the forest edge when determining the behaviour and distribution of radionuclides released during routine authorised discharges and accidental releases.
5. Uncertainties It has been demonstrated in previous studies that the use of model canopies for wind tunnel deposition experiments at Imperial College provides a realistic simulation of the factors affecting dry deposition of
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aerosol particles in real forest canopies (Kinnersley et al., 1994; Ould-Dada, 2002). However, it is generally difficult to simulate atmospheric wind with complete realism and wind tunnel conditions are expected to deviate from those of the field by some degree (OuldDada, 2002). Forest edges provide conditions in which many factors controlling the deposition process change rapidly within small distances. In this study, the fetch downwind from the grass–tree margin was insufficient to allow the re-establishment of a stable boundary layer. This is due to the constraints of the wind tunnel dimensions, which did not allow a greater fetch to be established. In addition, under field conditions, large turbulent eddies are expected to occur near the top of the canopy whereas in a wind tunnel, the scale of turbulence is restricted by the height and width of the tunnel. Nevertheless, the configuration of the experimental canopy allowed measurements of particle deposition fluxes to be made at and immediately behind the grass–tree margin. Differences between wind tunnel and field conditions will inevitably lead to some degree of uncertainty in extrapolating wind tunnel findings to ‘real’ forest canopies. Whilst it was not possible to quantify all uncertainties, efforts were made throughout the study to minimise uncertainties as much as possible.
findings from the wind tunnel experiments are of relevance to the improvement of our understanding of the effect of forest edges on radionuclide deposition and provide quantitative information on the magnitude of this effect for use in model development. The application of wind tunnel results in particular may be appropriate not only for radionuclides but also for pollutant aerosols such as SO4, NO3 and NH4, which are characterised by particle sizes in the range used in the present study.
6. Conclusion
References
The potential enhancement of aerosol deposition at forest edges compared with the surrounding grass/ pasture was investigated using a combination of wind tunnel and field studies. Results showed an increased deposition at the forest edge over that occurring in the surrounding grass/pasture. The wind tunnel study resulted in aerosol deposition rates at the ‘forest’ edge, which were three times higher than those on the grass and within the tree stand. The field study showed a clear elevation of 137Cs and 241Am levels within the woodland that can be attributed to changes in turbulence and other boundary layer factors as air flows over the forest edge. Within the vicinity of forest edges, total deposition of radionuclides will include a large proportion originating from edge effects. This has important implications when considering the likely spatial heterogeneity of ground surface contamination following the release of air pollutants to the near-surface atmosphere. Such effects should, therefore, be considered in assessments of fallout impact, whether these are to be made by either direct sampling or modelling. At present, data on the effect of forest edges on radionuclide deposition are lacking and such effects are generally not considered in current models. The field study results demonstrate the significant impact that the forest edge can have on the atmospheric deposition of radionuclides and thus the
Acknowledgements The financial support of the wind tunnel study by the Commission of the European Communities is gratefully acknowledged (contract number FI3PCT920016). The Imperial College environmental wind tunnel was constructed with partial funding from the Ministry of Agriculture, Fisheries and Food, UK. Special thanks are due to Rob Kinnersley and Margaret Minski for their advice and support, and to Nasser Baghini for his help with sample analysis. The financial support of BNFL for the Lady Wood study is gratefully acknowledged.
Beier, C., 1991. Atmospheric pollutants: separation of gaseous and particulate dry deposition of sulphur at a forest edge in Denmark. Journal of Environmental Quality 20, 460–466. Beier, C., Gundersen, P., 1989. Atmospheric deposition to the edge of a spruce forest in Denmark. Environmental Pollution 60, 257–271. Beier, C., Gundersen, P., Rasmussen, L., 1992. A new method for estimation of dry deposition of particles based on throughfall measurements in a forest edge. Atmospheric Environment 26A (9), 1553–1559. Copplestone, D., 1996. The food chain transfer of radionuclides through semi-natural habitats. Ph.D. Thesis, University of Liverpool, 369pp. Copplestone, D., Johnson, M.S., Jones, S.R., Toal, M.E., Jackson, D., 1999. Radionuclide behaviour and transport in a coniferous woodland ecosystem: vegetation, invertebrates and wood mice, Apodemus sylvaticus. Science of the Total Environment 239, 95–109. Copplestone, D., Johnson, M.S., Jones, S.R., 2000. Radionuclide behaviour and transport in a coniferous woodland ecosystem: the distribution of radionuclides in soil and leaf litter. Water, Air and Soil Pollution 122, 389–404. Draaijers, G.P.J., Ivens, W.P.M.F., Blenten, W., 1988. Atmospheric deposition in forest edges measured by monitoring canopy throughfall. Water, Air and Soil Pollution 42, 129–136. Gash, J.H.C., 1986. Observations of turbulence downwind of a forest–heath interface. Boundary Layer Meteorology 36, 227–237.
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Hasselrot, B., Grennfelt, P., 1987. Deposition of air pollutants in a wind-exposed forest edge. Water, Air and Soil Pollution 34, 135–143. Harrison, R.M., ApSimon, H.M., Beck, H., Devell, L., Dickerson, M., Garland, J.A., Graziani, J., Gudiksen, P., McKay, W.A., Mishra, U.C., Nicholson, K.W., Smith, F.B., Shapiro, C.S., 1993. Atmospheric pathways. In: Warner, F., Harrison, R.M. (Eds.), Radioecology after Chernobyl: Biogeochemical Pathways of Artificial Radionuclides. Scope 50. Wiley, New York, pp. 55–100. IAEA, 2001. Modelling the migration and accumulation of radionuclides in forest ecosystems. Draft Final Report of the Forest Working Group Activities 1998–2000. BIOMASS Theme 3, BIOMASS/T3FTM/WD02, International Atomic Energy Agency, Vienna. Jackson, D., Jones, S.R., Fulker, M.J., Coverdale, N.G.M., 1987. Environmental monitoring in the vicinity of Sellafield following deposition of radioactivity from the Chernobyl accident. Journal of the Society of Radiological Protection 7 (2), 75–87. Jones, S.R., Willans, S.M., Smith, A.D., Cawse, P.A., Baker, S.J., 1996. Deposition of actinides in the vicinity of Sellafield, Cumbria: accounting for historical discharges to atmosphere from the plant. Science of the Total Environment 183, 213–229. Kinnersley, R.P., Farrington-Smith, J.G., Shaw, G., Minski, M.J., 1994. Aerodynamic characterisation of model tree canopies in a wind tunnel. Science of the Total Environment 157, 29–33. Meroney, R.N., 1970. Wind tunnel studies of the air flow and gaseous plume diffusion in the leading edge and down-
stream regions of a model forest. Atmospheric Environment 4, 497–614. Neal, C., Ryland, G.P., Conway, T., Jeffery, H.A., Neal, M., Robson, A.J., Smith, C.J., Walls, J., Bhardwaj, C.L., 1994. Interception of chemicals at a forest edge for a rural lowlying site at Black Wood, Hampshire, Southern England. Science of the Total Environment 142, 127–141. Nicholson, K.W., 1988. The dry deposition of small particles: a review of experiments measurements. Atmospheric Environment 22 (12), 2653–2666. Ould-Dada, Z., 2002. Dry deposition profile of small particles within a model spruce canopy. Science of the Total Environment 286, 83–96. Ould-Dada, Z., Baghini, N.M., 2001. Resuspension of small particles from tree surfaces. Atmospheric Environment 35, 3799–3809. Raynor, G.S., Hayes, J.V., Ogden, E.C., 1974. Particulate dispersion into and within a forest. Boundary Layer Meteorology 7, 429–456. Tikhomirov, F.A., Shcheglov, A.I., 1994. Main investigation results on the forest radioecology in the Kyshtym and Chernobyl accident zones. Science of the Total Environment 157, 45–57. Toal, M.E., Copplestone, D., Johnson, M.S., Jackson, D., Jones, S.R., 2001. A dynamic compartmental food chain model of radiocaesium transfer to Apodemus sylvaticus in woodland ecosystems. Science of the Total Environment 267, 53–65. ( Wiman, B.L.B., Agren, G.L., 1985. Aerosol depletion and deposition in forests—a model analysis. Atmospheric Environment 19 (2), 335–347.
Effect of forest edges on deposition of radioactive aerosols Z. Ould-Dadaa,*, D. Copplestoneb, M. Toalc, G. Shawd a
Food Standards Agency, Radiological Protection and Research Management Division, Room 715B, Aviation House, 125 Kingsway, London WC2B 6NH, UK b ERC, University of Liverpool, Vanguard Way, Birkenhead, Wirral CH41 9HX, UK c Centre for Ecology and Hydrology, Pollution and Ecotoxicology Section, Monks Wood, Abbots Ripton, Huntingdon, Cambridgeshire PE17 2LS, UK d Department of Environmental Science & Technology, Faculty of Life Sciences, Imperial College at Silwood Park, Ascot, Berkshire SL5 7PY, UK Received 20 March 2002; received in revised form 1 August 2002; accepted 23 August 2002
Abstract The possible enhancement of aerosol deposition at forest edges was investigated in a wind tunnel and in the field. The wind tunnel study was carried out using 0.82 mm mass median aerodynamic diameter uranium particles and a composite canopy of rye grass and spruce saplings. The field study was undertaken at a coniferous woodland near to BNFL Sellafield, Cumbria, UK. Two transects were set through the woodland to determine the influence of the forest edge on atmospheric deposition of radionuclides released under authorisation from the Sellafield site. Results from the wind tunnel study showed that the deposition flux of uranium particles decreased with distance downwind from the grass– tree edge towards the interior of the canopy. The deposition flux at the edge was maximal at about 4 107 mg of U cm2 s1. This was 3 times higher than that observed over grass where a constant flux of about 1.32 107 mg of U cm2 s1 occurred. Results from the field study showed a clear influence of the forest edge on the atmospheric deposition of 241Am and 137Cs. Activity depositions of around 4750 and 230 Bqm2 for 137Cs and 241Am, respectively, were measured in front of the woodland. Activity deposition inside the forest edge, however, rose to levels of between 20,200 and 50,900 Bq m2 and 1100 and 3200 Bq m2 for 137Cs and 241Am, respectively, depending upon the transect. Similar activity concentrations were measured in the pasture to the front and behind Lady Wood. Results from these studies corroborate those obtained from various studies on air pollutants including radionuclides. This underlines the importance of deposition at the edge of forests and its contribution to the overall canopy deposition. The edge effect is therefore an important factor that should be considered in the assessment of fallout impact, whether this is to be made by either direct sampling or by modelling. r 2002 Elsevier Science Ltd. All rights reserved. Keywords: Radionuclides; Radioactive aerosols; Forest edge; Deposition; Wind tunnel; Field study
1. Introduction The process of dry deposition of airborne pollutants is influenced by various physical, chemical and biological *Corresponding author. Tel.: +44-207-276-8774; fax: +44207-276-8779. E-mail address: [email protected]. gov.uk (Z. Ould-Dada).
factors that can cause large variations in deposition on both small and large scales of time and space. Edges and other structural inhomogeneities of forests are important features that provide situations in which many factors controlling the deposition process change rapidly within small distances. The effect of forest edges on deposition of radionuclides to forest canopies was an important phenomenon observed in contaminated forests in the vicinity of the Kyshtym (southern Urals,
1352-2310/02/$ - see front matter r 2002 Elsevier Science Ltd. All rights reserved. PII: S 1 3 5 2 - 2 3 1 0 ( 0 2 ) 0 0 6 9 9 - 4
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1957) and Chernobyl accidents, 1986. After the Kyshtym accident of 1957, Tikhomirov and Shcheglov (1994) observed an enhancement in radionuclide deposition at the forest edge facing the source of release. Following the Chernobyl accident, however, the same authors reported gamma dose rate measurements, collected in an area situated at approximately 6 km to the west of the damaged reactor, which indicated a similar level of contamination in both forested plots and in adjacent pasture areas. This indicates a similar degree of deposition of radioactive materials in both forests and pasture, with none of the expected enhancement of deposition at the margin between the two. Several other authors, however, have found an increased deposition of airborne pollutants (not radionuclides) at forest edges (e.g. Raynor et al., 1974; Hasselrot and Grennfelt, 1987; Draaijers et al., 1988; Beier and Gundersen, 1989; Beier, 1991; Beier et al., 1992; Neal et al., 1994). This leads to the question of the likely role of the forest edge in controlling the deposition of radioactive aerosols. Forest edges are commonly exploited for food by both man and animals and this may lead to high exposure to contaminants given that enhanced deposition can occur within this area. The Forest Working Group of the Biosphere Modelling and Assessment programme (BIOMASS, organised by the International Atomic Energy Agency, Vienna), recognised the importance of forest edges in radioactive contamination of forests and stressed the need for the collection of relevant data and development of models (IAEA, 2001). The objective
of this paper is to examine the possible enhancement of aerosol deposition at the edge of forests using results from wind tunnel and field studies.
2. Wind tunnel study 2.1. Construction of composite grass and spruce canopy Swards of perennial rye grass mixture (Lolium perenne, Permogen long-term ley) were germinated in a peat-based compost in shallow plastic trays (4 cm 38 cm 38 cm). These swards were allowed to grow inside a green house to a height of approximately 5 cm. Ten trays of grass were then placed inside the wind tunnel covering the whole length of bays 1 and 2 (Fig. 1). Bay 3 of the tunnel was filled with 12 Norway Spruce saplings (Picea abies L.) positioned in six rows with an average height of 47 cm, i.e. about 10 times that of the grass. The height of the trunk space of the trees varied between 3 and 7 cm with an average value of 5.2 cm. A peat-based compost was used to provide a soil cover under the tree canopy; the soil was kept moist throughout the experiment to minimise any resuspension of particles from its surface. The soil surface of the tree stand was made level with that of grass. The soil level was adjusted by defining a height (i.e. the zero plane displacement) so that the distribution of drag over the full canopy depth is aerodynamically equivalent to the imposition of the entire stress at this height. The zero plane displacement was defined such that about 2/3 of
Fig. 1. Dimensions of grass and spruce canopies and positions of aerosol source and measurement points.
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the grass height was below the level of the airflow entering the working section. The positioning of grass and spruce canopies in the wind tunnel is shown in Fig. 1. 2.2. Meteorological measurements A free stream air velocity of 5 m s1 was established at about 30 cm above the tree canopy using hot wire anemometry. Wind speed measurements were made vertically at 10 mm intervals from a height of 0.87 cm above the top of the grass canopy and from 20 mm above the ground for the stand of trees. Measurements of vertical air speed profiles were made at four positions along the wind tunnel, i.e. 45, 5, +35 and +175 cm from the grass–tree margin (Fig. 1). These measurements allowed calculations of friction velocity, u ; (using the eddy correlation method) from the following relationship (see Kinnersley et al., 1994):
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grass–tree margin. The procedures for release and sampling of aerosol are described in detail in OuldDada (2002). 2.3.2. Sampling and analysis of plant material At the end of the experiment, each tree was divided into five horizontal layers of equal height and all trees, excluding the tree ‘trunk’, were sampled. Within each layer, new and old growth tissues were divided and 18 tissue categories obtained. Samples from the grass canopy were taken at four locations (15, 45, 90 and 120 cm) from the grass–tree margin and were oven dried at 701C for 24 h. The mass of uranium contained in each sample and each filter paper was determined by delayed neutron counting (DNC). The sampling procedure, the preparation of samples for irradiation and the analysis of samples are explained in detail in Ould-Dada (2002).
U* ¼ ðu0 v0 Þ0:5 ðm s1 Þ; where u0 and v0 are horizontal and vertical fluctuations on the mean wind velocity (m s1). The variation of the mean wind speed over time was measured as turbulence intensity, iu; at each sampling position. Assuming isotropic turbulence, iu was defined as the mean root square fluctuation on the mean velocity u (see Kinnersley et al., 1994): iu ¼ ðu02 Þ0:5 =u: Temperature and relative humidity were measured up to a height of 80 cm from ground level in the tree canopy and from the top of the grass. Measurements were made at 130, 20, +35 and +175 cm from the grass–tree margin. 2.3. Contamination and sampling 2.3.1. Release and sampling of aerosol The uranium aerosol used in this study consisted of 0.82 mm mass median aerodynamic diameter (MMAD) (0.45 mm mass median diameter, MMD) dry spherical particles produced by the aerosol generation technique described in Ould-Dada (2002). The aerosol generator was positioned adjacent to the wind tunnel throughout the experiment and particles were liberated into the tunnel via a short flexible hose and an injection manifold system for a period of 60 h. The injection manifold was adjusted so that the bottom row of injection nozzles was about 6 cm above the top of the grass canopy (Fig. 1). Fourteen isokinetic air samplers were used to sample aerosol particles at various levels within and above the tree canopy in order to determine their vertical air concentration profile (Fig. 1). The flow rate through each sampler was determined from the wind velocity profile measured at 75 cm distance from the
3. Field study Lady Wood (OSGR: NY 037045) is a coniferous woodland approximately 4.7 ha in size located 500 m northeast of BNFL Sellafield, Cumbria, UK. The woodland is dominated by a uniform stand of sitka spruce, Picea sitchensis, planted in the 1950s. It is surrounded by agricultural pasture and there are few grass and herbaceous species present beneath the dense canopy. The site has been characterised previously for radionuclide contamination derived primarily from the authorised discharges from BNFL Sellafield (Copplestone et al., 1999, 2000; Toal et al., 2001). The wood faces Sellafield with no major intervening natural or artificial features to intercept airborne particulates. Along the woodland edge facing the principal source of radionuclides, there is a steep slope varying between 5% and 33%, which presents a large surface area of canopy for the interception of radionuclides from the ambient atmospheric aerosol. For this work, two transects were set through the woodland at right angles to the front woodland edge. The transects were 230 and 260 m in length (transects 1 and 2, respectively) and were 50 m apart. The difference in length resulted from the non-uniform shape of Lady Wood. Soil samples were collected at intervals 20 m apart within the woodland and at 5, 10, 15, 20, 30, 40 and 50 m from the front and back woodland edges. The soil samples consisted of cores of 10 cm diameter, extracted to a depth of 9 cm. Each core was oven dried at 1051C, homogenised in a rotary mill and sub-sampled for determination of 137Cs and 241Am activity concentrations by gamma spectrometry. Details of sampling procedures and analytical techniques are presented elsewhere (Copplestone, 1996).
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4. Results and discussion 4.1. Wind tunnel study 4.1.1. Wind velocity profiles The shape of the wind profile above the grass, at the 45 cm position, is simple with a boundary layer of less than 20 cm height whereas at the +175 cm position, a more complex profile with a ‘jet’ within the trunk space and a ‘deeper’ boundary layer typical of a tree canopy was established (Fig. 2). The upper part of the latter profile reflects the retarding effect of the wind tunnel ceiling. Because there was only about 0.70 m of space between the top of the trees and the ceiling of the wind tunnel, the wind speed profile at the +175 cm position was ‘squeezed’. Between these two extreme positions (i.e. 45 and +175 cm), the boundary layer is subject to
rapid alteration and deformation reflected in a scatter of points in the profile from the +35 cm position indicating an enhancement of turbulence. Friction velocity increased from about 0.5 m s1 over the grass canopy to about 1.5 m s1 over the stand of trees (Fig. 3). This reflects the considerably larger drag exerted by trees on the airflow than the grass and the air is decelerated as it moves from the smoother to the rougher surface. Similar results were obtained by Meroney (1970) who carried out wind tunnel studies, at a free stream velocity of 6 m s1, of the airflow at the leading edge and downstream of a model plastic tree canopy of about 18 cm height. The distribution of the foliage with height at the edge of the stand as well as inside the forest will affect the movement of the emerging wind. When the wind arrives at a forest edge, it penetrates and moves through the
Fig. 2. Wind velocity profiles over grass and spruce canopies.
Fig. 3. Friction velocity profiles over grass and spruce canopies.
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canopy forming a jet in the trunk space, which is usually more open compared to the canopy crowns. Even in the case of horizontal airflow over a uniform tree canopy, the efficiency of the airflow in penetrating downwards through the canopy will depend particularly on the foliage density. Measurements of turbulent fluxes of momentum in the extreme transition from a heathland of about 0.25 m height to a forest of approximately 10 m height were made by Gash (1986) who found that the flux reached an equilibrium at about 120 m from the heath–forest edge, i.e. a distance of approximately 12 tree canopy height. On this basis and on the evidence from Meroney’s (1970) study, therefore, the fetch downwind from the grass–tree margin in this study was insufficient to allow the reestablishment of a stable boundary layer above the tree canopy. However, within the constraints of the wind tunnel dimensions, no greater fetch could be achieved. Nevertheless, the configuration of the experimental canopy allowed measurement of particle deposition fluxes at and immediately behind the grass–tree margin. Since wind speed is one of the controlling parameters of the dry deposition process, the alteration in the airflow and the change in its profile at this edge can be expected to influence the patterns of aerosol deposition in this region.
both the trunk space and within the canopy reaching values of 0.7 and 2.2, respectively (Fig. 4). At this position, iu observed within the tree canopy was nearly 6 times higher than that observed over grass. This can be attributed to the rough surfaces and complex aerodynamic shapes of trees which create more turbulence than do aerodynamically smoother grass surfaces: the different values of iu probably reflect the size of the eddies generated by each surface. Similar profiles of iu were observed in wind tunnel over a plastic tree canopy (Meroney, 1970) and at a 10–13 m tall pine forest (Raynor et al., 1974). Moreover, the shape of iu profile at the +175 cm position was similar to those obtained for tree canopies (with no transition zone) used during deposition (Ould-Dada, 2002) and re-suspension (Ould-Dada and Baghini, 2001) experiments linked to the edge-effect study described here. The maximum value of iu in those profiles, which was also recorded at about the middle of the crown was, however, 5.5 times lower than that observed here. Turbulence is another factor controlling the dry deposition process and variations in iu observed here will affect the movement of aerosols and hence the deposition of airborne material to tree surfaces. Turbulence is likely to be influenced by both foliage density and the prevailing meteorological conditions.
4.1.2. Turbulence profiles Turbulence intensity (iu ) was measured at the same positions and height increments as wind velocity. Above the grass canopy, iu varied between 0.05 at a height of about 20 cm and 0.42 at the canopy top, i.e. 5 cm height (Fig. 4). At +35 cm from the grass–tree transition point, iu increased with height reaching a maximum value of 0.46 in the middle of the tree crown and 0.2 in the trunk space. Turbulence intensity remained constant above the tree canopy at this point. At +175 cm from the grass– tree edge, however, turbulence intensity increased in
4.1.3. Air temperature and humidity profiles Temperature and relative humidity were measured from ground level up to a height of 0.80 m at 130, 20, +35 and +175 cm from the grass–tree edge. Both temperature and humidity were constant with height above the grass canopy. Variations were confined within 21C (11C for the tree stand) and 5% for temperature and relative humidity, respectively. Temperature and relative humidity measured outside the working section of the tunnel were 10.61C and 45.8%, respectively. Variation in temperature is an additional factor to be taken into
Fig. 4. Turbulence intensity profiles over grass and spruce canopies.
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shown by measurements made at +175 cm from the edge. The enhancement in aerosol deposition flux observed at the edge of the tree stand can be explained by the alteration in the vertical wind profile resulting in an increase of turbulence as shown in Figs. 2 and 4. The two points of aerosol release were set up at a height within that of the tree canopy cross-section and the wind blew in a direction at right angles to the edge. Thus, the trajectory of aerosol particles was directly towards the tree stand edge where a considerable drag is imposed on the moving air resulting in an alteration of its vertical profile as described in Fig. 2. The current observations of enhanced deposition of aerosol particles at the model forest edge over that occurring on grass are in agreement with findings of several other authors. Draaijers et al. (1988) found the increase in dry deposition at forest edges to be strongly dependent on wind velocity and wind direction during ( dry deposition. Wiman and Agren (1985) showed from model studies that the higher wind speed at the forest edge increased the dry deposition of particles. In their study on particulate dispersion into and within a mixed
account when interpreting distribution patterns of deposition of air pollutants since the vertical temperature profile can either enhance or suppress vertical mixing of air and the materials contained within it. 4.1.4. Deposition rates 4.1.4.1. At canopy level. Aerosol deposition fluxes (F ) to both grass and tree canopies were calculated using the following relationship: F ¼ Qu =Ste ðmg cm2 s1 Þ; where Qu is the quantity of total uranium (mg), S is the projected area of plant material (cm2) and te is the time of exposure of canopy to aerosol (s). Deposition fluxes were then expressed in mg of uranium deposited per cm2 of ground area per second by multiplying the flux per unit area of plant material (i.e. F ) by the LAI. Deposition velocity (Vg ) was calculated as the flux divided by the mean air concentration (i.e. 3.1 107 mg U cm3) measured within the boundary layer above the tree canopy (see Ould-Dada, 2002). The results shown in Table 1 indicate a Vg of 1.28 cm s1 at the edge of the tree canopy, which is about 2.5 times compared with that measured inside the tree canopy (0.54 cm s1). The latter value is very consistent with that obtained for tree canopies (with no transition zone) during a wind tunnel deposition study (0.5 cm s1; see Ould-Dada, 2002). The results shown in Fig. 5 indicate a nearly constant deposition flux of about 1.32 107 mg U cm2 s1 over the grass canopy. This value was, however, three times higher at the edge of the tree canopy where a maximum flux of 3.97 107 mg U cm2 s1 was recorded. This flux decreased downwind of the edge showing a secondary ‘peak’ of 1.86 107 mg U cm2 s1 at +135 cm from the edge. This secondary ‘peak’ could be the result of the high turbulence developing inside the stand of trees as was
Fig. 5. Variation of uranium aerosol flux to grass and spruce with distance from grass–tree edge. Error bars indicate 7s.e.m. (standard error of the mean).
Table 1 Average LAI (cm2 cm2) at each tree layer and Vg values as a function of distance from grass–tree edge in the wind tunnel experiment (based on two trees at each distance) Distance from grass–tree edge (cm)
Vg to trees (cm s1)
LAI Tree layers
0 41 73 105 135 170 Mean a
Excludes Vg value at distance 0.
Total
1
2
3
4
5
0.17 0.14 0.08 0.09 0.13 0.14
0.80 0.63 0.42 0.53 0.29 0.49
1.29 1.00 0.51 0.67 0.53 1.20
0.88 0.64 0.89 0.78 1.04 1.00
1.00 0.91 0.57 0.85 0.86 0.62
4.14 3.59 2.37 2.92 2.85 3.45
1.28 0.68 0.47 0.54 0.60 0.40
3.22
0.54a
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forest, Raynor et al. (1974) reported that deposition of stained ragweed pollen and other tracers ranging from 14 to 54 mm in diameter near the forest edge and in the tree tops took place by impaction whereas at the interior of the forest deposition occurred mainly by gravitational settling. Beier and Gundersen (1989) suggested that increased dry deposition at a Norway spruce forest edge could be attributed to a dramatic change in the surface roughness length, caused by the forest edge, which leads to turbulent air movements and lower canopy resistance. Beier (1991) stated that dry deposition at the forest edge may increase due to reduced boundary layer thickness which causes increases in interception and impaction of particles. In his study on throughfall deposition to a Norway spruce forest edge in Denmark, Beier found that deposition of SO+ 4 and Na was 11 and 21 times higher, respectively, at the front tree compared to the open field and 2.4 times and 8.3 times compared to the deposition inside the forest. In Beier’s study deposition decreased exponentially with distance from the forest edge reaching an almost constant flux at 15 m from the edge (B175 and 75 meq m2 yr1 for SO+ 4 and Na , respectively). Hasselrot and Grennfelt (1987) also measured throughfall deposition of SO4, NO3, NH4, Cl, Na, K, Mg and Ca at a pine forest on the Swedish west coast and observed a continuous decrease in the deposition of these elements from the edge downwind into the forest. A constant deposition flux was not reached until more than 50 m from the forest edge. Draaijers et al. (1988) measured atmospheric deposition in Douglas fir (Pseudotsuga menziesii) forest edges by monitoring canopy throughfall and assuming throughfall fluxes to be constant at 200 m from the forest edge. All these studies were conducted in coniferous forests (spruce, pine and Douglas fir) and different results could be due to differences in canopy structure which influence deposition (Beier, 1991). All these studies, however, showed an enhancement of deposition at forest edges with this effect decreasing with distance from the forest edge. In the present study, LAI was more or less constant with distance from the grass–tree edge (Table 1). In the field, however, LAI is often higher at the edge than inside the forest especially when spacing between trees is very small preventing the light from penetrating downward through the crown hence reducing the full development of the crowns inside the canopy. Model ( results from Wiman and Agren (1985) suggested an increase in aerosol deposition close to the edge of forests with dense structure. In a Norway spruce forest, Beier and Gundersen (1989) found LAI a factor of 2 higher at the edge than inside the forest. From 5 m downwind of the forest edge LAI remained nearly constant. According to these authors, higher LAI for trees at the edge could be one of the explanations for the enhancement of dry deposition at the forest edge, although the results of
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the current study suggest that alterations in airflow at the edge of a tree stand are sufficient to enhance the deposition flux of B1 mm aerosols. 4.1.4.2. Between tree layers. Figs. 6 and 7 show the variation of deposited aerosol at each canopy layer with distance from the grass–tree edge. Maximum deposition of material occurred at the top of trees (i.e. layer 1) whereas within the crown deposition varied only slightly but was about four times lower than that recorded at the tree tops (Fig. 6). If layer 1 is excluded, there is an indication that deposition rate decreased from the bottom to the canopy top (Fig. 7). In fact, tests carried out with a smoke generator set up at the same height as the aerosol source revealed that once in contact with the tree canopy edge most of the plume moved through the crowns and in the trunk space although part of the smoke flowed over the top of the trees. At the far edge of the tree canopy, smoke was observed on some occasions to be also moving upwards from inside the tree crowns. Furthermore, measurements of aerosol air concentration made within and above the tree canopy showed a decrease in air concentration with height, though, air concentration profile does probably vary with distance
Fig. 6. Variation of total amounts of uranium deposited to different layers of the spruce trees.
Fig. 7. Variation of total amounts of uranium deposited to different layers of the spruce trees (excluding top layer).
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intercept potentially large quantities of dry deposits. The enhancement of dry deposition at the edge of forests is expected to depend on forest edge characteristics such as tree species, stand density, tree height, open-trunk space and leaf surface area (i.e. LAI). This increase has also been found to vary with other factors. For example, Beier and Gundersen (1989) found the effect of Norway spruce forest edges on the deposition of atmospherically borne substances (Na, Ca, K, Mg, Cl, NO3, NH4 and SO4) to be far greater in the dormant season than in the growing season. The same authors also found that substances deposited as particles (mainly Na+ and Cl) showed the most pronounced edge effect compared to those deposited as gaseous substances. Beier (1991) found that dry deposition of particles increased by a factor of 13 at the front tree, whereas dry deposition of gaseous S increased by a maximum factor of 2.5. Within the vicinity of forest edges, total deposition will include a large proportion originating from edge effects. This has important implications when considering the likely spatial heterogeneity of ground surface contamination following release of radioactive aerosols to the near-surface atmosphere. In areas such as the Chernobyl 30 km zone, in which forests and fields form a complex mosaic, much of the observed variability of radionuclide deposition may be due to the effects of forest edges on dry deposition. Such effects should, therefore, be considered in assessments of fallout impact, whether these are to be made by either direct sampling exercises or modelling.
from the grass–tree edge. This is due to the increased drag exerted by trees on the oncoming flow resulting in a transfer of momentum, hence of aerosol material, from the air stream to the tree surfaces. The air concentration profile observed in this study (no data shown) is the reverse of that obtained in the wind tunnel deposition study (with no transition zone) which was part of the same project (Ould-Dada, 2002). As an explanation, in the deposition study aerosol material was released above a continuous tree canopy where aerosol penetration occurred from the boundary layer down through the canopy. In this case, the canopy layers offer more resistance to the moving air, hence allowing less material to pass from one layer to another. In the current study, however, the aerosol trajectory was directly towards the trees where the airflow moved through the tree crowns and the open trunk space. The latter, offering less resistance to the airflow, allowed more aerosol to circulate in this part of the canopy. Following a series of smoke puff releases at various heights from ground level to the canopy top, Raynor et al. (1974) observed that air penetrating the forest edge moved primarily into the more open trunk space where air speeds were greater than those in the crown for a downwind distance of about 60 m. Their results revealed that the rate of loss of material by deposition in both the canopy and the trunk space was much greater than above the forest or in the open. The authors concluded, however, that the rates of loss and partition of material between layers (i.e. trunk space, canopy, above canopy) were not sensitive to source distance or to height provided that much of the material reaches the forest below treetop height (canopy height was 10–13 m). From the results of the current study and from the literature reports discussed above, it can be seen that forest edges disturb the vertical wind profile and
4.2. Field study The measured 137Cs and 241Am levels (Figs. 8–11) are consistent with other reported work around Sellafield. For example, Jones et al. (1996) report values for Lady
35,000
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Fig. 8.
Cs (Bq m2) deposition in front of, through and behind Lady Wood. Transect 1.
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Fig. 9.
Cs (Bq m2) deposition in front of, through and behind Lady Wood. Transect 2.
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Fig. 10.
Am (Bq m2) deposition in front of, through and behind Lady Wood. Transect 1.
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Wood of 1300 Bq m2 for 241Am and Cawse reports 137 Cs levels of 39,000 Bq m2 (unpublished data). There is generally a higher deposition of radioactivity along transect 2 than transect 1. This may reflect the different surface area of the canopy available for deposition that results from the trees growing on a variable slope along the front edge of the woodland. The slope varies between 5% and 30% along the front edge with the slope at transect 2 being approximately three times as steep as that at transect 1; consequently there is greater deposition of radionuclides along transect 2, e.g. 37.1 kBq m2 for 137Cs compared with 20.2 kBq m2 along transect 1. Airborne particulates that may have radionuclides associated with them can be deposited by two principal
mechanisms: precipitation scavenging (wet deposition) and dry deposition. Precipitation scavenging involves the removal of particulate matter and gases from the air by different forms of precipitation. In contrast, dry deposition occurs continuously within the surface boundary layer with particulate matter being deposited through the process of diffusion, impaction, interception and sedimentation under gravity. These processes have been described more fully elsewhere (e.g. Harrison et al., 1993; Nicholson, 1988). The deposition patterns observed within, and around, Lady Wood (Figs. 8–11) show clear elevation of the levels of 137Cs and 241Am within the woodland (0–100 m, transect 1 and 0–140 m, transect 2) compared with the surrounding agricultural land. From the data available
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Am (Bq m2) deposition in front of, through and behind Lady Wood. Transect 2.
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on radionuclide concentrations within the woodland, it is believed that dryness is the most important factor in determining the pattern of distribution given that precipitation scavenging would be expected to produce a uniform distribution of deposition over the area receiving rainfall (Harrison et al., 1993). Copplestone et al. (2000) reported a uniform soil distribution of 134Cs throughout Lady Wood that varied between 250 Bq m2 at the front of the woodland nearest to Sellafield to 300 Bq m2 at the back of the woodland. Since Sellafield only releases minor quantities of airborne 134Cs, it is likely that the majority was deposited as a result of the passage of the Chernobyl plume (Jackson et al., 1987). Given that the majority of the radionuclide deposition from the Chernobyl plume occurred as a result of precipitation scavenging, a uniform distribution pattern would be expected. This is borne out by the 134Cs data observed. A similar pattern of deposition would have occurred for Chernobyl derived 137Cs. However, as the data in Figs. 8–11 clearly demonstrate, the deposition of 137Cs is not uniform and, based on the wind tunnel experiments reported here and by other workers, e.g. Beier and Gundersen (1989) and Neal et al. (1994), an elevation in the activity concentrations of the radionuclides within the woodland would be expected if dry deposition was occurring. The wind tunnel experiments have demonstrated the significance and mechanisms involved with dry deposition on the pattern of the uranium particle deposition to model spruce canopies. Consequently, Lady Wood provides field data as evidence of the underlying mechanisms described by the wind tunnel experiments. The elevated levels in Lady Wood compared with the surrounding pasture can be explained then by changes in
boundary layer dynamics as the air flows over the grassland in front of Lady Wood and into the forest edge. As mentioned previously, the canopy of a woodland acts as a porous barrier which can allow air to partially pass. This, however, is dependent upon the foliage density of the woodland. These factors will cause eddies to be produced within the canopy structure which will in turn give rise to increased turbulence and reduced velocity within the canopy. Within the time frame and scope of the Lady Wood study, it was not possible to measure these parameters directly within the canopy. However, the data provide good agreement with those from the wind tunnel experiments, thus demonstrating that the mechanisms involved in the deposition process are similar. In Lady Wood the foliage is very dense and the trees are narrowly spaced apart. It was therefore expected that the Lady Wood forest edge would have a significant impact on the airflow and consequent deposition of airborne particulates and any associated radionuclides within this field site. As demonstrated by the data presented in Figs. 8–11, there is a significant increase in the deposition of radionuclides within the woodland. This indicates the importance of the forest edge when determining the behaviour and distribution of radionuclides released during routine authorised discharges and accidental releases.
5. Uncertainties It has been demonstrated in previous studies that the use of model canopies for wind tunnel deposition experiments at Imperial College provides a realistic simulation of the factors affecting dry deposition of
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aerosol particles in real forest canopies (Kinnersley et al., 1994; Ould-Dada, 2002). However, it is generally difficult to simulate atmospheric wind with complete realism and wind tunnel conditions are expected to deviate from those of the field by some degree (OuldDada, 2002). Forest edges provide conditions in which many factors controlling the deposition process change rapidly within small distances. In this study, the fetch downwind from the grass–tree margin was insufficient to allow the re-establishment of a stable boundary layer. This is due to the constraints of the wind tunnel dimensions, which did not allow a greater fetch to be established. In addition, under field conditions, large turbulent eddies are expected to occur near the top of the canopy whereas in a wind tunnel, the scale of turbulence is restricted by the height and width of the tunnel. Nevertheless, the configuration of the experimental canopy allowed measurements of particle deposition fluxes to be made at and immediately behind the grass–tree margin. Differences between wind tunnel and field conditions will inevitably lead to some degree of uncertainty in extrapolating wind tunnel findings to ‘real’ forest canopies. Whilst it was not possible to quantify all uncertainties, efforts were made throughout the study to minimise uncertainties as much as possible.
findings from the wind tunnel experiments are of relevance to the improvement of our understanding of the effect of forest edges on radionuclide deposition and provide quantitative information on the magnitude of this effect for use in model development. The application of wind tunnel results in particular may be appropriate not only for radionuclides but also for pollutant aerosols such as SO4, NO3 and NH4, which are characterised by particle sizes in the range used in the present study.
6. Conclusion
References
The potential enhancement of aerosol deposition at forest edges compared with the surrounding grass/ pasture was investigated using a combination of wind tunnel and field studies. Results showed an increased deposition at the forest edge over that occurring in the surrounding grass/pasture. The wind tunnel study resulted in aerosol deposition rates at the ‘forest’ edge, which were three times higher than those on the grass and within the tree stand. The field study showed a clear elevation of 137Cs and 241Am levels within the woodland that can be attributed to changes in turbulence and other boundary layer factors as air flows over the forest edge. Within the vicinity of forest edges, total deposition of radionuclides will include a large proportion originating from edge effects. This has important implications when considering the likely spatial heterogeneity of ground surface contamination following the release of air pollutants to the near-surface atmosphere. Such effects should, therefore, be considered in assessments of fallout impact, whether these are to be made by either direct sampling or modelling. At present, data on the effect of forest edges on radionuclide deposition are lacking and such effects are generally not considered in current models. The field study results demonstrate the significant impact that the forest edge can have on the atmospheric deposition of radionuclides and thus the
Acknowledgements The financial support of the wind tunnel study by the Commission of the European Communities is gratefully acknowledged (contract number FI3PCT920016). The Imperial College environmental wind tunnel was constructed with partial funding from the Ministry of Agriculture, Fisheries and Food, UK. Special thanks are due to Rob Kinnersley and Margaret Minski for their advice and support, and to Nasser Baghini for his help with sample analysis. The financial support of BNFL for the Lady Wood study is gratefully acknowledged.
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