Atmospheric Deposition

aquatic systems. These are termed dry deposition, wet deposition (through precipitation) and cloud droplet deposition (droplet impaction onto vegetation surfaces).

2.5.1 Dry deposition

The dry deposition of gases and particulates involves a transfer from the boundary layer to the vicinity of the surface, molecular diffusion and uptake at the surface by dissolution, sorption or chemical reactions (Reynolds and Ormerod, 1993). Untransformed oxides are deposited by adsorption and absorption while transformed gases fallout onto ground and

vegetation surfaces. These gases include HNO3, HCI and NHg. The rate of uptake is

governed by conditions at three levels. Above and within the forest canopy, deposition is dependent on windspeed and the aerodynamic roughness of the vegetation surface. Tilled soil, moorland and forestry are characterised by increasing surface roughness. At the vegetation/atmosphere interface and within the stomata, rates of uptake are controlled by molecular diffusion. At leaf surfaces uptake is governed either by chemical reactions occurring between the leaf surface and the gas or by entry into the leaf via stomata pores and subsequently by solution in the intercellular fluid. The dominant control over deposition

rates will depend on the reactivity of the gas deposited (Fowler et al., 1989). Other factors

can influence the uptake pathways of individual gases. At night uptake of SOg and NOg takes place via chemical reactions on the surface of the leaf whereas, during the day, when stomata pores are open, the uptake occurs through the pores and subsequently, in solution in intercellular fluids (Fowler and Cape, 1985). The control here is stomatal opening which is determined by changes in temperature and light. Surface wetness is also a factor which influences chemical reactions on the leaf surface although, for many vegetation types, dry deposition of SOg to wetted surfaces is not appreciably greater than uptake on dry surfaces (Fowler and Cape, 1985).

2.5.2 Wet deposition

Wet deposition comprises the atmospheric acids (and bases) which are deposited onto terrestrial and aquatic ecosystems via precipitation. This can occur, for example, when

H2SO4 is incorporated into water droplets or ice crystals and falls to the ground as

precipitation (rainout). Alternatively, if in particulate form, H^SO^ or HNO3 can be removed

from the atmosphere by raindrop impaction (washout). This includes the seeder feeder effect (Bader and Roach, 1977) which involves the scavenging of sulphuric and nitric acid held in mist or low lying orographic cloud formed as part of frontal weather systems by precipitation from overlying clouds (Bergeron, 1965). This tends to increase ionic deposition in cloud- capped upland areas. The feeder cap cloud is generally characterised by higher concentrations of acidic species than the precipitation from the seeder cloud above (Carruthers and Choularton, 1984). The lower mountain cloud caps incorporate the higher concentrations of ions in the atmospheric boundary layer whereas in the frontal clouds the processes of raindrop formation are initiated by the vapour growth of snowflakes. This does not efficiently incorporate the dissolved particulate cloud droplets and, as the snowflakes melt at lower altitudes, raindrops are formed which scavenge cloud droplets in the feeder cloud as a result of collision coalescence. Frontal weather systems are responsible for much

of the precipitation over upland areas (Fowler etal., 1995) as the moist boundary layer rises

over elevated terrain. Thus the seeder-feeder effect is primarily responsible for deposition of acidifying compounds in these areas, a supposition supported by experimental work

(Fowler at a!., 1988; Dore et a!., 1992; Inglis et a!., 1995).

2.5.3 Crowd droplet (occult) deposition

Exposed vegetation in upland areas can directly intercept water droplets held in wind driven cloud and fog. Concentrations of acidic species in cloudwater can be considerably greater

than in rainfall in the same area (Crossley etal., 1992) and it is suggested that cloud droplet

deposition in areas prone to low cloud could increase wet deposition estimates by up to 2 0 %

above that detected in rainfall gauges (Dollard et a!., 1993). Deposition loadings vary with

different vegetation communities (Ferrier etal., 1990). Estimates of wet and bulk deposition

should thus be modified to account for direct impaction in upland areas.

2.5.4 Monitoring and mapping deposition patterns

Measuring dry deposition presents particular difficulties as it tends to be governed by surface properties and a wide variety of measurement techniques have been developed for this purpose (Ross and Lindberg, 1994). Dry deposition rates of SOg onto vegetation and soil for different surfaces throughout Great Britain have been calculated (Garland, 1978; Fowler and Unsworth, 1979; Fowler and Cape, 1985). Rates are calculated from the product of near surface concentration and an appropriate deposition velocity. This is inversely related to the distance from the source. As a consequence, dry deposition tends to be greatest near major emission source areas and contributes more than 75% of total deposition in Southern

and Eastern England (Cottrill et al., 1986). Estimated annual inputs of acidity from dry

deposition of SOg have been mapped showing that dry deposition in the industrial Midlands and north of England is much greater than wet deposition while the converse is true in western Wales and north Scotland (Fowler and Cape, 1985).

Wet deposition can be measured relatively easily by collecting precipitation and multiplying the amount by solute concentrations. This precipitation weighting technique enables spatial and temporal patterns to be identified. In spatial terms two aspects of wet deposition require consideration, the concentration in precipitation of acidifying compounds and the amount of acidity actually deposited (Irwin and Williams, 1988). UK Maps showing the concentration of precipitation weighted non-marine sulphate and the amount deposited illustrate the

difference (Cottrill et al., 1987). The greatest concentrations are found in the east of Britain where rainfall levels are lower while deposition is much greater in areas of higher rainfall in

North West England and North Wales. The relative contributions of HgSO^ and HNO3 in the

UK have been estimated as 71% and 29% respectively (Fowler etal., 1982) although the

latter is becoming increasingly important both in absolute terms (Skeffington and Wilson, 1988) and relative to the former (Galloway and Likens 1981, Rodhe and Rood, 1986). The relative contribution of each to soil and water acidification is less easy to quantify due to the mitigating effects that ecosystem interactions have on N species (Sutton and Fowler, 1992).

The concentration of acidic species in precipitation also exhibits seasonal variation with non­ marine sulphate and nitrate maxima generally occurring in the spring or early summer (Irwin and Williams, 1981). Variations in composition can also occur between and within

precipitation events (Coscio etal., 1982). At one site in Eastern England 30% of the annual

sulphate deposition occurred in five days (UKAWRG, 1986) while in Wales 30% of deposited acidity falls on less than 5% of wet days (Reynolds, 1987). The implications for ecosystem response of pulsed deposition episodes are discussed below.

In general terms non marine sulphate (/.e that not derived from sea spray) and nitrate have fairly similar spatial patterns with lower concentrations in the north and west while those in

the East Midlands and East Anglia are up to a factor of 10 greater (Campbell et al., 1987).

For dry deposition, UK maps are not based on a monitoring network because of the difficulties involved in obtaining accurate measurements. Estimates are based on semi- empirical mathematical models which incorporate transport, transformation and removal processes (Barrett and InArin, 1983). Maps are produced on a 20km^ grid basis using the proportions of different land types in each square (UKRGAR, 1990). On a European scale

1995). National maps are based more on measurement networks. In the UK, wet deposition maps for S and N are based on a network of 38 monitoring sites together with the UK

Meteorological Office precipitation measurement network (Fowler et al., 1994). Maps have

been produced which incorporate the effects of orographic enhancement (UKRGAR, 1990;

Dore et a!., 1992). A modelling approach to deposition mapping has also been developed.

Initially, the Harwell Trajectory Model (HIM) coupled SOg, NO^, NHg and HCI with simple

meteorology data (DenA/ent et a!., 1988; Metcalfe et a!., 1989). The Hull Acid Rain Model

(HARM) refines this approach although it presently concentrates on modelling 8 deposition

(both at current emission levels and under future emission scenarios). Using data on emissions, rainfall, windspeed, trajectories, dry deposition and wet removal the HARM model produces similar deposition patterns as those using measured data (Metcalfe and Whyatt, 1994)

It is has been shown that there is considerable variation in deposition levels onto different landscape features and at different elevations. This variation is not fully incorporated into

maps at 2 0 km^ scale and precludes the use of these maps for identifying deposition at the

catchment scale. This has important implications for the application of the critical loads approach at this scale where it is necessary to compare the sensitivity of the surface waters for a specific catchment with the actual deposition loading to identify where critical load

exceedance may occur (Erisman, et a!., 1995). Acid loading onto individual catchments is

dependent on altitude, slope, aspect, vegetation cover and location, factors which can vary substantially from catchment to catchment, even at a local scale (Ross and Lindberg, 1994). The significance of these uncertainties on the development and application of a catchment

scale predictive model are discussed further in Chapters 3 and 8 .