I M P A C T OF A F F O R E S T A T I O N ON T H E S O I L S O L U T I O N

C H E M I S T R Y OF S T A G N O P O D Z O L S IN M I D - W A L E S

B. REYNOLDS 1, C. NEAL 2, M. HORNUNG x, S. HUGHES" and P. A. STEVENS a

(Received November 6, 1986; revised March 10, 1988)

Abstract. Chemical data for stagnopodzol soils and soil solutions are presented for grassland and plantation coniferous forest systems in upland mid-Wales. The results show that the introduction of the forest, with consequent increase in anion concentrations, has increased concentrations of A1 in the soil solution by 1.5 to 3 times. Hydrogen ion concentrations have only increased by a small amount. Simple cation exchange relationships have been proposed to explain some of these changes.

1. Introduction

Acidic deposition and conifer afforestation have both been implicated in the decline of fish and aquatic invertebrate populations in many upland streams of western and northern Britain (Harriman and Morrison, 1982; Stoner et al., 1984). These declines are associated with elevated streamwater concentrations of A1 and increased stream acidity (Howells, 1983). The majority of British upland soils are acidic, AI bearing and base cation depleted. They provide a source of A1 and H + to surface waters.

Predictive models are being developed to describe short- and long-term trends in streamwater acidification and dissolved A1 status to assess the importance of land management and atmospheric pollution on water quality (Christophersen et al., 1982, 1984; Cosby et al., 1985; Whitehead et aL, 1986; Neal et aL, 1986a). These models are dependent on a fundamental understanding of the processes controlling soil water chemistry which is currently incomplete. Almost exclusively, predictive models use the solubility of amorphous or crystalline AI(OH)3 to predict A1 chemistry within the soils and streams. However, this assertion has rarely been supported by field evidence in that; (1) AI(OH)3 has usually not been found in the soils modelled and (2) the implied cubic relationship between H + and Al3 + is not universally observed in soil waters and streamwater (Hooper and Shoemaker, 1985; Reynolds e taL , 1986; Sullivan et al.,

1986). Other solubility controls have been proposed including clay minerals (kaolins, micas, and smectites), low crystallinity Al-hydroxy silicates and Al-hydroxy sulphate (Jurbanite) (Prenzel, 1983). At Plynlimon, for example, thermodynamic calculations have shown that clay minerals are not in equilibrium with soil and streamwaters (Neal etal . , 1986b) and Al-hydroxy sulphates and Al-hydroxy silicates have not been identified in the soils. However, within the podzolic soils typical of the Plynlimon area,

1 Institute of Terrestrial Ecology, Bangor Research Station, Penrhos Road, Bangor, Gwynedd, North Wales, U.K. 2 Institute of Hydrology, Maclean Building, Crowmarsh Gifford, Wallingford, Oxon, U.K.

Water, Air, and Soil Pollution 38 (1988) 55-70. © 1988 by Kluwer Academic Publishers.

56 B. REYNOLDS ET AL.

TABLE I

Soil profile descriptions for grassland and forest stagnopodzol soils at Plynlimon, mid-wales

Grassland Altitude 435 m Slope 9 ° Land use: unimproved rough grazing

Horz. Depth (era)

Forest Altitude 380 m Slope 8 ° Sitka spruce plantation, planted 1949

Horz. Depth (era)

Ol 0-1 Mosses and grass

Of 1-2 Fibrous matted mosses and grass

Oh 2-12

Eag 12-17

Bf at 17 Bsl 17-30

Bs2 30-41

Bs/C 41-47

Black (5 YR 2/1) peat, Strong very coarse angular blocky structure; stoneless; firm; common fine fibrous and small woody roots. Pinkish grey (7.5 YR 6/2) silty clay loam with a few faint grey brown mottles; strong coarse angular blocky structure; stoneless; firm; few fine fibrous roots.

Weakly developed iron pan. Reddish brown (7.5 YR 6/8) silty clay loam; strong medium and coarse subangular structure; few small and gravel-size angular shale stones; few fine fibrous roots. Strong brown (7.5 YR 5/8) sandy loam; strong medium subangular structure; common gravel-size angular and few small angular shale fragments; friable; few fine fibrous roots. Yellowish brown (10 YR 5/6) sandy loam; weak fine subangular structure abundant gravel-size and small angular shale stones; loose; rare fine fibrous roots.

01 0-2

Off 2-8

Of 2 8-10 Ah 10-16

Oh 16-19

Eag 19-24

Bf at 24 Bs 24-51

Bs/C 51-89

Sitka spruce needles and twigs Sitka spruce needles. Abundant medium woody roots. Molinia debris Very dark greyish brown (10 YR 3/2) organic silty clay loam; stoneless; weak fine subangular blocky structure; friable; common large and small woody roots; overturned plough furrow. Black (5 YR 2/1) peat, Weak fine subangular blocky structure; stoneless; greasy; common medium woody roots. Dark greyish brown (10 YR 4/2) clay loam with rare ochreous mottles; coarse moderate angular blocky structure; stoneless; firm few medium woody roots. Undulating iron pan Yellowish brown (10 YR 5/8) silty clay loam; fine weak crumb structure; common gravel-size shale stones; friable; common small woody roots.

Olive brown (2.5 Y 6/6) silty loam; structureless; common gravel-size shale stones; firm; few small woody roots.

IMPACT OF AFFORESTATION ON THE SOIL SOLUTION CHEMISTRY

Table I (congnued)

57

Grassland Altitude 435 m Slope 9 ° Land use: unimproved rough grazing

Forest Altitude 380 m Slope 8 ° Sitka spruce plantation, planted 1949

Horz. Depth Horz. Depth (cm) (cm)

C 47-87 Yellowish brown (10 YR 5/6) C 89-101 sand; structureless; abundant gravel size and small angular shale stones, loose; rare fine fibrous roots.

Pale olive (5 Y 6/4) silty loam; structureless; abundant gravel-size and medium shale stones; firm; roots absent.

especially in the Bs horizon, a large pool of readily mobilized A1 may be present in the form of amorphous oxides and hydroxides although the exact nature of this material is obscure.

To facilitate and encourage further development of predictive acidification models, results from comparative studies of grassland and forest systems in upland Wales are presented to highlight the importance of cation exchange, weathering and increased salt loading in determining the concentrations of A1 and H ÷ in soil waters.

2. Sites

The grassland and forest sites are located in the catchments of the Afon Cyff and Afon Hore, respectively, situated to the east of Plynlimon, approximately 24 km from the west coast of Wales. The area is underlain by base-poor lower Palaeozoic mudstones and shales. Ferric stagnopodzols (Avery, 1980), (Aquods, in the USDA soil classification) of the Hafren series (Lea, 1975), along with peats, comprise the largest proportion of the soil cover. Profile descriptions of the podzols at the grassland and forest sites are shown in Table I.

The vegetation at the grassland site comprises an Agrostis-Festuca dominated community. A similar vegetation was formerly present at the forest site which was planted with Sitka spruce (Picea sitchensis) in 1949. The area receives an annual average rainfall of 2480 mm and the mean annual air temperatures for February and August are 1.8 and 13.1 °C, respectively. The area has clean air with estimated mean annual SO2 concentrations of between 4 and 5 ppbv (Martin, 1982) and measured annual NO× concentrations of 4.4 ppbv (Ashenden, 1987). Detailed site descriptions are given in Newson (1976) and Reynolds et aL (1986).

58 B. REYNOLDS ET AL.

3. Methods

Soil water samples were collected fortnightly from four horizons (Oh, Eag, Bs, and C) within the stagnopodzols. Three replicate sets of samplers were used at each site and individual samples were bulked, by horizon, for analysis. A set of samplers comprised a plastic zero-tension tray lysimeter beneath the organic-rich Oh horizon and a 5 cm diameter porous ceramic cup soil water sampler (Stevens, 1981) in each of the three mineral horizons. Attempts to use porous cups in the Oh horizon failed due to physical instability. The mineral horizons proved too stony for installation of tray samplers. Although problems with ceramic cup samplers have been reported (e.g. Grover and Lamborn, 1970; Nagpal, 1982), our own laboratory and field studies (Hughes and Reynolds, 1988) suggest that difficulties are largely overcome by laboratory leaching prior to installation, and by a prolonged field stabilization period (cf. Hetsch et aL,

1979). Bulk precipitation was sampled using a continuously open collector in the Cyff

catchment. Earlier work indicated that samples from this site would be representative of solute concentrations in bulk precipitation across the catchments (Reynolds, 1984). Throughfall beneath the Sitka spruce was sampled with twelve open collectors and beneath the Nardus-Festuca grassland community six small plastic troughs were installed and connected, via PVC tubing, to polypropylene bottles located in the ground. Samples of precipitation and throughfall were collected fortnightly and individual samples were pooled for chemical analysis.

All water samples were filtered through 0.45 ~tm membrane filters in the laboratory on the day following sampling. Samples were analyzed for Na and K by flame emission and Ca, Mg, and A1 by flame atomic absorption spectophotometry (AAS). Silicon was determined colorimetrically using an auto-analyzer (Allen et aL, 1974) and SO n, C1, and NO 3 by ion chromatography. Dissolved organic carbon (DOC) was determined using a Carlo-Erba analyser and pH was measured on unfiltered subsamples using a Radiometer meter and dual electrodes. The precision for DOC and A1 determinations was + 10~o, pH + 0.05 units and all other solutes + 5~o.

Exchangeable cations were determined on soil samples taken from three pits at each site. Samples were air dried and sieved to < 2 mm grain size. For each horizon, samples from each of the three pits were bulked together prior to analysis. Exchangeable base cations were determined by flame AAS following a 1 hr extraction with 1M ammonium acetate at pH 7 and exchangeable A1 was determined, also by flame AAS, following extraction with 1M KCI. Exchangeable acidity was estimated on the same KC1 extracts by G-ran titration. All extractions were duplicated and the mean values are reported. The effective cation exchange capacity (ECEC) was estimated as the sum of the exchangea- ble base cations plus A1 and H + (Juo et aL, 1976; Lee et aL, 1985). Soil pH (1 : 2.5 w/v soil : water) and loss-on-ignition (at 375 °C for 16 hr in a muffle furnace (Ball, 1964)) were determined on duplicate samples of field moist soil and mean values are reported.

I M P A C T O F A F F O R E S T A T I O N O N T H E S O I L S O L U T I O N C H E M I S T R Y 59

O

~o

~0

o

= o ~

4-I

O

oo

o

o

<

"4-I + I + I -i-I

~I~I~I~I

+I ÷ I + I + I

+ I + I + I + I

+ l + t + l + l

+ I + I + I + I

+ I + I + I + I

+ I + I + I + I

+ I + I + ] + I

q-1 -I-I -I-I '4-I

-I-I -I-I -H "I-I

+l , d ,,:i "~ ~

m ~,'-1 ¢".1 ~-I ~

+1 +1 +1 +1 ~

I

,,....i

1 4-I -I-I - I - I

0 ~'~

+ I + I + I + I

~ °~

+I + I + I + I o,, ¢~i ¢r-a O o.i

. .o

+I + I + I + I I~

4-I + I + I + I '~

eel I '~ ',~" t"e/ .~0

'-I-I + I + I "+'I "~ -

60 B. REYNOLDS ET AL.

4. Results and Discussion

At Plynlimon, grassland and forest soil waters are acidic and A1 beating and show systematic variations in solute concentrations with depth (Tables II and III). Apart from H ÷ and organic anions (estimated as the difference between the mean of the cation sums and the mean of the inorganic anion sums), this variation corresponds to an increase in solute concentrations from the organic horizon to the mineral soil; within the mineral soil, patterns are less consistent. Concentrations of H ÷ decrease uniformly with depth (Table II) . Organic anions show an overall decrease in concentration from the Oh to the C horizon. This pattern contrasts, however, with the more uniform decrease in DOC concentrations with depth (Table III). The difference probably results from the varying

TABLE III

Arithmetic mean soil water anion and DOC concentrations ( + SE) for stagnopodzol soils under grassland and forest at Plynlimon, mid-Wales

Horizon S 0 4 C1 N 0 3 DOC Mean of inorganic Organic anions ~ - - laeq L - 1 mg L - 1 anion sums

~teq L -1

Grassland

Oh 68 + I0 124 + 12 29 + 6 9.5 + 1.1 221 + 21 50 Eag 87 + 5 209 + 18 39 + 5 6.0 + 0.7 335 + 22 30 Bs 71 + 5 184 + 15 28 + 4 3.7 + 0.4 283 + 17 35 C 77 + 5 202 + 19 44 + 6 2.3 + 0.4 323 + 19 27

Forest

Oh 108 + 12 135 + 10 13 + 2 13.1 + 1.5 256 + 20 88 Eag 159 + 10 233 + 16 11 + 3 11.8 + 1.0 399 + 23 139 Bs 206 + 7 327 + 22 1 + 1 6.1 + 0.3 535 + 22 56 C 186 + 5 281 + 14 16 + 2 4.1 + 0.3 483 + 14 65

Sampling period: = O Horizon 9 October, 1984-30 December, 1985; mineral soil 3 January, 1984-30 December, 1985. a Organic anions: estimated as difference between mean of cation sums and mean of inorganic anion sums.

degree of dissociation of the organic acids from which the anions are derived and because organic anions only comprise a part of the DOC measurement. Bicarbonate is only a trace component in the soil water and makes an insignificant contribution to the total anion concentration.

Systematic differences are observed between the grassland and the forest soil waters in that concentrations of SO4, C1, A1, Na, Ca, and Si are greater in the forest compared with the grassland soils (Tables II and III). The relative enrichment of cations is in the order A13 + > sum of divalent cations (M 2 + ) > sum of the monovalent cations (M ÷ ) (Figure 1). The largest difference in H ÷ is observed in the Oh horizon whereas the difference in A1 concentrations between the two sites becomes greater at depth. The differences in M 2 ÷ between the forest and grassland soil waters is largely due to the

I M P A C T O F A F F O R E S T A T I O N O N T H E S O I L S O L U T I O N C H E M I S T R Y 61

Ratio 1 2 3 4 I l ! I

Oh.

Eag-

Horzn.

Bs-

C-

• aO

Ratio 1 2 3 I I I

Oh-

Eag-

Horzn.

Bs-

C-

/ / KeYM . • • • ~ + +

• • ^, []

Fig. 1. Forest to grassland ratios of cation and Si concentrations in soil waters from Plynlimon.

higher concentrations of Ca in the forest. Concentrations of DOC are uniformly greater down the profile in the forest soil compared with the grassland soil.

The difference in anion concentrations between the forest and the grassland soil waters are partly explained by the increased production of organic anions with forest development. Higher concentrations of DOC and organic anions in forest soil waters have probably resulted from breakdown of the spruce litter although the degree of dissociation of organic acids will be controlled by soil water pH. The main increase in organic and total anion concentration (Table III) is attributable to higher SO4 and C1 concentrations in the forest soil waters.

Anion concentrations in throughfall beneath both vegetation canopies are generally greater than bulk precipitation'with the exception of N O 3 (Table IV). Concentrations of SO 4 and C1 are greater in spruce throughfaU compared with Nardus throughfall resulting from greater scavenging of fog and cloud water (occult deposition) by the forest canopy and increased capture of aerosol particles by the trees (Lovett et al., 1982).

The water balances for the 2yr period (1984-1985) for the forested part of the Hore

62 B. R E Y N O L D S ET A L

0

@

=

0

I i

I I

I I

I I I

I I i

...,.,

i I I

I I I

i I I

.~ ~

¢q [

oo I

"7

O

o

- - , .g

~ " N . f f ' ~

m ~

IMPACT OF AFFORESTATION ON THE SOIL SOLUTION CHEMISTRY

TABLE V

Water balances for the grassland Cyff catchment, the forested portion of the Hore catchment and the forest site at Plynlimon, mid-Wales

P Q ET ET/P (mm) (mm) (mm) (~o)

Cyff 4631 3789 842 18.2 Hore 4599 3466 1133 24.6

P T & S IL IL/P (mm) (mm) (mm) (~o)

Forest site 4424 3468 956 21,6

Measurement period: 1 January, 1984-31 December, 1985. P = Rainfall; Q = Discharge; ET = Evapotranspiration loss; T = Throughfall; S = Stemflow; IL = Interception Loss.

63

catchment and the grassland Cyff catchment are given in Table V and show that evapotranspiration losses from the forest were 6.4~0 greater than from the grassland. Evapotranspiration (ET) is the sum of the interception loss (evaporation from tree canopy, IL) and the transpiration loss (TF). During 1984-1985 measurements of rainfall, throughfall and stemflow were made at the forest site. From these data (Table V) IL accounted for 21.6~o of the incoming rainfall. This compared with 22.8 ~o estimated from data collected over a 4 yr period at six sites distributed throughout the forest (Scanlan, 1981). Assuming that ET accounted for 24.6 ~o of the incoming rainfall to the forest site (Table V) then the transpiration loss (TF = ET-IL) was 132 mm or 3~ of the incoming rainfall and compares with an annual average of 4.2~o for the 4 yr period 1977 to 1980 (Hudson, 1988). Thus, the major zone for evaporative concen- tration of solutes in the forest is the tree canopy where it directly influences the composition of throughfall. Soil solution chemistry will, to some extent reflect throughfall chemistry as throughfall (and stemflow) comprise the inputs to the forest floor. Additional evaporative concentration of solutes in the soil, represented by TF, will be small but will vary seasonally through the year. Accordingly, arithmetic averaging of the soil water concentrations may exaggerate the differences between forest and grass- land soils as high concentrations will occur when the soil moisture content is low.

Simple concentration processes cannot fully account for differences in NO 3 and SO4 concentration between forest and grassland soils. The lower NO 3 levels in the forest probably reflect greater utilization of N by the trees compared with the grass sward and differences in the rates of nitrification for the two systems. Increased mineralization of organic S compounds, leading to higher soil water SO 4 concentrations, may also result from the drying and oxidation of the organic horizons as a consequence of site prepa- ration and increased evapotranspiration.

Many of the processes already described influencing anion chemistry will also affect the concentrations of cations and Si. Atmospheric deposition and canopy processes are

64 B. REYNOLDS ET AL.

clearly important for base cations and H ÷, as concentrations are increased in throughfall compared with bulk precipitation. For K and Ca, biological uptake and cycling are significant as soil solution concentrations in the Oh horizon, the major rooting zone in both systems, are much depleted compared with throughfall. The Oh horizon is also a major source of H +. The organic anion data indicate that organic acid production occurs in both systems, although it is greatest in the forest. NH 4 is present in bulk precipitation and throughfall in both systems but is near or below detection limit (<0.1 mgL -1 NH4) in grassland and forest soil waters. Several processes may account for this, e.g., nitrification, direct uptake by vegetation, and adsorption onto the soils. The rates of these processes have not been investigated in this study.

For A1 and Si, which are not derived from atmospheric sources, the differences in soil water concentrations will be ultimately related to different rates of release from primary aluminosilicates (chlorite, vermiculite, and mica; see Chapman, 1986) in the soil and bedrock by weathering reactions, a process which also affects major cations. However, Si concentrations may be buffered by solubility controls involving crystalline or partially crystalline SiO 2 (equilibrium Si concentrations vary between 36 and 107 ixmol L-1 depending on temperature and crystalinity; Casey and Neal, 1986). Thus, Si cannot be used as a primary indicator of aluminosilicate weathering rates. Amorphous A1 oxides will provide a readily available source of A1 to solution, especially in the Bs horizon.

4 .1 . T H E ROLE OF CATION E X C H A N G E

The quantities of individual exchangeable cations in each mineral horizon (Eag, Bs, and C) are similar for the grassland and forest soils. The cation exchange sites are dominated by A1, although Na, K, Ca, Mg, and H ÷ are also present (Table VI). The pool of exchangeable cations in the soil is large compared with the amount of cations in the associated soil waters. For example, assuming typical values for porosity (55 ~o), cation

TABLE VI

Exchangeable cations and pH of bulked samples of stagnopodzol soils under grassland and forest at Plynlimon, mid-Wales

Location Na K Ca Mg H A1 ECEC ~ pH LOI b meq 100 g - 1 (~o)

Grassland

Forest

Oh 0.49 0.78 0.86 1.44 0.15 5.23 8.95 3.67 71 Eag 0.13 0.03 0.16 0.12 0.07 8.25 8.76 3.75 9 Bs 0.13 0.06 0.11 0.06 0.04 4.11 4.51 4.20 7 C 0.12 0.06 0.15 0.05 0.01 0.90 1.29 4.48 11

Oh 0.41 0.31 0.71 0.22 0.20 12.05 13.90 3.43 38 Eag 0.12 0.07 0.17 0.03 0.07 9.34 9.80 3.81 8 Bs 0.10 0.03 0.15 0.02 0.03 4.64 4.97 4.01 7 C 0.07 0.01 0.14 0.01 0.02 2.03 2.28 4.31 2

a Effective cation exchange capacity. b Loss-on-ignition.

IMPACT OF AFFORESTATION ON THE SOIL SOLUTION CHEMISTRY 65

exchange capacity (10 meq 100 g- 1) and soil (solid) density of 2.5 g cm -3, then 99.89 and 99.82~o of the cations in the soil-soil water system are on cation exchange sites in the grassland and forest soils, respectively; from forest planting to the present time (~ 40 yr) the total ionic flux to the forest floor is only of the order of 20~o of the pool of exchangeable cations. The size of this pool dictates that ion exchange is a major process controlling solution chemistry in these soils.

In order to test this hypothesis, simple cation exchange relationships have been used to predict the changes in soil solution cation concentrations resulting from increased anion loading consequent on afforestation. The approach is based on the 'mobile anion' concept originally described by Nye and Greenland (1965) and applied to forest systems by Cole and Gessel (1965). The ion exchange relationships developed here build on this concept in two ways: (1)A1 and H ÷ relationships are defined by ion exchange rather than the more commonly used AI(OH)3 solubility (see, e.g., Christophersen et aL, 1984) and (2) the regulation of soil water pH by the inorganic CO 2 system has been excluded, as for these acid soils its effect is negligible. Note that the ion exchange relationships used here are very simple. They predict short-term exchange processes. Long-term consequences of afforestation on soil solution chemistry cannot be represented using ion exchange relationships but rather depend on site specific factors such as nutrient turnover, atmospheric inputs and weathering rates.

Two basic principles are used to describe the ion exchange behaviour of these acid soils. Firstly, the pool of cations on the soil exchange surfaces is large compared with that in solution. Secondly, the cation exchange reactions are described by simple mass action thermodynamic relationships of the type:

2Ads(M +) + M 2+ = Ads2(M 2+) + 2M + ,

with

and

with

K' = (Ads2(M2+)) [M+ ]2

(Ads(M +)) 2 [M2+ l

3Ads(M +) + M 3+ = Ads3(M 3+) + 3M +

K" = (Ads3(M3 +)) [M+ ]3

(Ads(M +))3 [M 3 + l '

where 'Ads' refers to the adsorbed/exchangeable cations and K' and K" to the cation exchange equilibria constants.

From the first principle, it follows that changes in the concentration of cations in the soil solution have an insignificant effect on the pool of exchangeable cations and thus the relative proportions of exchangeable cations remains constant. Combining this with the second principle gives constant values for (Ads2(M2+))/(Ads(M+)) 2,

66 B. REYNOLDS ET AL.

Cation Conc'n ~eq / l

(a)

400"

300-

200-

100-

0 0

G F

200 400 600 800

Sum of Anions jueq/I

Cation Conc'n }Jeq/I

(b)

300 G F

+

200- I l

100- H ÷

0 0 200 400 600 800

Sum of Anions }Jeq/I

Fig. 2. Predicted variations in coneentrations of ( a ) M +, M z +, and A13 ÷ and (b )H + and A13 + in grassland soil waters with total (inorganic plus estimated organic) anion concentrations. Measured forest soil water cation concentrations marked O. Mean total anion concentrations for mineral horizons in

grassland and forest soils denoted by lines G and F, respectively.

IMPACT OF AFFORESTATION ON THE SOIL SOLUTION CHEMISTRY 67

(Ads3(M3+))/(Ads(M+)) 3, and the ratios [M + ]2/[M2+ ] and [M + ]3/[M3+ ]. The

concentrations of individual cations in solution can be determined from knowledge of these ratios and the total anion charge in solution ~ ( n A n - ) using the principle of electrochemical neutrality: ~ m [ A m- ] = ~ n [ M n+ ].

These equations have been applied to data from the mineral horizons (Eag, Bs, and C) of the grassland soil to: (1) show how the concentrations of individual cations vary with total (inorganic plus estimated organic) anion concentration and (2) to predict the soil water chemistry response to afforestation as an effect of increased anion concen- tration on the exchange equilibria. The Oh horizon has been excluded from the analysis as this is the zone of most intense biological activity. The behavior of the predominantly organic exchange surfaces in this horizon will also be very different from the mineral material of the lower horizons due to pH dependence of the CEC and differences in the selectivity for individual cations. The ratios of [M + ]2/[M2 + ] and [M + ]3/[M3 + ] were calculated using the mean soil water chemistry for the three mineral horizons of the grassland soils. Further, as A1 is the dominant trivalent cation this was considered as M 3 +. Preliminary data on A1 speciation using the method of Driscoll (1980) indicated that inorganic A1 predominated and A1 has been assumed to be in solution primarily as A13 + (French, 1985). Thus, for example, using units of rteq L - 1

[M + ] 2 [M + ]3 - 8 7 5 - - - 2 1 . 9 × 104 ,

[ M 2+ ] [A13+ ]

where M + = 50 ~teqL - I , M 2+ = (50)2/875 = 2.9 ~teqL -1 and A13+ =

= (50)3/21.9 x 104 = 0.6 Beq L - 1 and the total anion charge

~,(nA " - ) = M + + M 2+ + A13+ = 35.5 ~teqL -1 .

The results of this exercise (Figure 2a) show that M +, M 2 +, and A13 + increase with

TABLE VII

Measured and predicted (from grassland data) cation concentrations for stagnopodzol soil waters from the Plynlimon forest site

Measured Predicted ~o difference a

H 53.3 54.5 2.3 A1 154 137 - 11.0 Na 245 252 2.9 K 5.3 6.4 20.8 Ca 46.0 30.5 - 33.7 Mg 53.0 81.0 52.8

Units: ~eq L -1

measured - predicted a Difference =

measured x 100.

increasing anion loading. For individual ions the predicted results compared well with observations for Na, H +, and A1 in the forest soils (Table VII) and point to the

68 B. REYNOLDS ET AL.

importance of ion exchange in controlling their behavior. In particular, the results show why in Table II, A1 concentrations are between 1.5 and 3 times higher in the forest compared with the grassland soils (Figure 2b) as power relationships hold between AI concentration and the concentrations of the other cations. There are, however, larger differences between predicted and observed concentrations for the other cations. This would be expected given such a simple predictive relationship and the complex and interactive nature of the processes already discussed. The cation exchange relationships assume a constant exchange capacity, however, the CEC in the forest is slightly greater and the pool of exchangeable A1 larger than in the grassland soils.

This probably reflects differences in the exchange complex between the two systems such as an increased proportion of organic exchange surfaces in the forest soil with a higher affinity for H ÷ and A1. In the longer term, ion exchange reactions should lead to a depletion of the exchangeable A1 rather than the observed enrichment on afforestation. Thus a source of A1 is required to replenish this pool. This suggests an increased rate of breakdown of A1 bearing minerals, such as amorphous A1 oxides, in the forest compared with the grassland soils.

While the cation exchange relationships predict the difference in M a ÷ concentrations between grassland and forest, poorer predictions are obtained for the individual cations Ca and Mg (Table VII). The prediction for K is also poorer than for the other monovalent ions. Calcium, Mg and K will be involved in biologically mediated processes. The poorer predictions for these ions may reflect differences in the rates of nutrient turnover and biomass production between the forest and grassland systems, although detailed nutrient budgets are required to prove this. For Ca additional sources are available at the forest site in the form of small CaCO3 mineral veins in the bedrock (Reynolds et aL, 1986). Greater Ca weathering rates are to be expected, therefore, so that the exchange store of Ca in the forest is not depleted and soil water Ca concentra- tions are higher than those predicted by simple ion exchange.

5. Conclusions

The results from Plynlimon show that for similar stagnopodzol soils, soil water concen- trations are different under grassland and forest vegetation. Concentrations of A1 are between 1.5 and 3 times higher in the forest soil waters compared with the grassland and anion concentrations are 1.5 times higher. The results may also indicate that afforestation of a grassland catchment increases the rate of breakdown of A1 bearing minerals, such as poorly crystalline to amorphous A1 oxides and aluminosilicates. More work is required to characterise these materials and to define their contribution to the chemistry of A1 in acid soils.

The features demonstrated at Plynlimon for soil solution chemistry and exchangeable cation chemistry are now being identified at other localities in Wales. It is suggested that H ÷ and A1 relationships in the soil solution can be assessed by cation exchange processes.

IMPACT OF AFFORESTATION ON THE SOIL SOLUTION CHEMISTRY 69

Acknowledgments

The authors are grateful to J. D. Roberts and colleagues of the ITE Chemical Service, Merlewood for performing anion, DOC, and Si analyses. The study was part-funded by the UK Department of the Environment and the Welsh Office.

References

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