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IV. Case Studies of Soil Change

IV. Case Studies of Soil Change

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ACID DEPOSITION ON FORESTED SOILS



43



histories (e.g., see Wolt and Lietzke, 1982). It embraces the inherent

assumption that all sites were similar at one stage, introducing the potential for misinterpreting initial site differences as a result of either vegetation occupancy or pollutant inputs. The second approach involves

measuring changes within individual stands or sites, determined by soil

remeasurements 10-60 years apart. When this sampling is done in conjunction with measurements of nutrient budgets, it offers by far the best

tool for measuring soil change and the causes of it. However, if soil sampling is conducted in isolation from either vegetation growth or nutrient

budget measurements, sorting out the causes of soil change becomes

nearly impossible.

Possibilities of capacity-type soil change were never in doubt, given

enough acid input over a significant amount of time. Well-known cases of

soil acidification near smelters clearly documented the validity of that

logical conclusion (e.g., Wolt and Lietzke, 1982). However, the amount of

time needed for such changes under typical field conditions was assumed to

be on the order of many decades to centuries, given the pool sizes of base

cations and acidity on exchange sites and the available information on

fluxes into and out of these pools (McFee, 1980; Johnson et al., 1982,

1985). However, several recent studies and literature reviews have conclusively demonstrated that forest soils in many parts of the world, both

polluted and unpolluted, are changing much more rapidly than previously

suspected (e.g., Hallbacken and Tamm, 1986; Berden et al., 1987; Binkley

et af., 1989b; Johnson et al., 1988a, 1991a). In some cases, these changes

are attributable primarily to nutrient uptake by vegetation (including uptake of Al; Turner and Kelly, 1977; Alban, 1982; Binkley et al., 1989a;

Johnson et al., 1988a), and in other cases, the changes are attributed to

both nutrient leaching and uptake (Van Miegroet and Cole, 1984; Hallbacken and Tamm, 1986; Johnson and Todd, 1990). In cases where soil

change is documented and leaching is known to be strongly impacted by

acid deposition, it is clear that acid deposition has played a role in causing

such changes (e.g., Tamm and Hallbacken, 1986; Billett et al., 1990b;

Johnson and Todd, 1990).

In contrast to a decade ago, the case studies of soil change are meanwhile too numerous to be reviewed in this article. The reader is referred to

Berd&n et al. (1987) and Johnson et al. (1991a) for a detailed review of

recent case studies of soil change. Here, we will only refer to a seiected

number of the these studies for the purposes of illustration. Specifically,

we will review case studies of intensity-type changes in which SO,’- and

NO, have caused marked increases in soil solution A1 concentration, and

capacity-type changes caused by uptake, leaching, and the combination

of the two.



44



WAYNE P. ROBARGE AND DALE W. JOHNSON



Figure 5. (a) Changes in soil exchangeable A13"at the Solling site, Germany, from 1966

to 1979. (b) Changes in soil exchangeable base cations (Ca2+, MgZ+,K + , and Na+) at the

Solling site, Germany, from 1966 to 1979. After Ulrich er al. (1980).



B. INTENSITY-TYPE

CHANGES

One of the earliest studies of soil change in an acid deposition-impacted

forest ecosystem (a beech forest at the Solling site in Germany) was the

work of Ulrich and his co-workers (1980). The Solling site is one of the

most intensively studied sites in the world with regard to element and Hf

budgets, and these element budgets are strongly impacted by acid deposi-



ACID DEPOSITION ON FORESTED SOILS



45



tion (Ulrich et al., 1980). However, inasmuch that soils at the site were well

within the aluminum buffering range before soil sampling took place, it is

not surprising that changes in exchangeable base cations over the 13-year

sampling period were minimal, at least in comparison to those noted by

Cole and Lamon (1981; as cited in Johnson et al. (1981), Johnson et al.

(1988a), and Binkley et al. (1989b)) (Fig. 5 ) . The authors did note increases

in A13+ in “equilibrium soil solution,” which is a water extract of fieldmoist soil, and, to a much lesser degree, increases in exchangeable A13+.

The authors interpreted the increase in exchangeable A13+ as a result of

dissolution of interlayer A1 polynuclear hydroxo complexes by H + .

It is obvious from the discussion of cation exchange above that increases

or decreases in A1 in the equilibrium soil solution could occur over very

short intervals, without any change in exchangeable A13+ or base cations,

as mineral acid anions increase and decrease. In this connection, Khanna

et al. (1987) noted marked decreases in soil solution pH and marked

increases in soil solution A1 associated with increases in soil solution SO:at the Solling site over the period 1974-1978 (Fig. 6 ) . They interpreted

these results as indicative of the dissolution of jurbanite (A1OHSO4)rather

than an exchange process.

The preferential displacement of A13+ from the exchange complex in

acid soils in response to much shorter-term increases in mineral acid anion

concentrations was illustrated by Reuss (1989) for forest soils subject to

NO; pulses in a nitrogen-fixing red alder stand. It is also illustrated in

Fig. 7, which shows the response of soil solution A1 to pulses of both NO;

and SO$- in a red spruce stand (Johnson et al., 1991b). Aluminum concentrations in soil solution at this site ranged from 30 to over 200 pmol/

liter (600 pEq/liter in Fig. 7). These peak A1 values are well within the

range noted to inhibit Ca and Mg uptake in seedling solution culture studies in the laboratory (100 pmol/liter) (Raynal et al., 1990) and are at

the threshold for effects on root growth in solution culture studies

(200 pmol/liter) (Thornton et al., 1987; J o s h and Wolfe, 1988). Throughout the collection period in this study, total A1 concentrations (which were

80-90% monomeric Al) were positively correlated with solution NO, and

SO:- . Fluctuations in solution SO:- and NO; concentrations individually

could explain over 60% of the variation in A1 concentrations, and

NO; + SO:- combined could account for over 70% of the variation in

measured A1 concentrations from this site (Johnson et al., 1991b).

The presence of acid soils is a necessary but not sufficient condition for

the mobilization of A1 via exchange processes. The introduction of mobile,

mineral acid anions to an acid soil will cause increases in the concentration

of A1 in soil solution, but extremely acid soils in the absence of mineral acid

anions will not produce solutions that are high in Al. This is illustrated by



46



-8



WAYNE P. ROBARGE AND DALE W. JOHNSON



1



1



BEECH

0.8



.



SPRUCE



'.



'



.



0.3 '



YEAR



YWI



Figure 6. Changes in soil solution SO:- and A1 at the Solling site, Germany. After

Khanna ef al. (1987).



the results of the Integrated Forest Study for the relatively unpolluted

Findley Lake site as contrasted to the more poiluted Whiteface Mountain

and Great Smoky Mountains National Park (GSMNP) sites (Johnson

et al., 1989; Johnson and Taylor, 1989). Figure 8 shows the soil solution

anion and cation concentrations in B horizons from the pristine site at

Findley Lake, Washington, and the more polluted sites at Whiteface

Mountain, New York, and Clingman's Dome, GSMNP, North Carolina.

Though the Findley Lake site soil is as acid as those from the more

polluted eastern sites, concentrations of A1 in soil solutions at Findley Lake

are much lower than those at the latter sites. The differences in soil

solution A1 concentrations are almost totally a function of differences in

mineral acid anion (SO$- and NOT) concentration, both of which increase

from Findley to Whiteface to Clingman's Dome due to a combination of



ACID DEPOSITION ON FORESTED SOILS



47



700

600

500



f

.-c



400



s



300



-.

~



200

100

0 '



1000



800



-.-f



.

c



600



P



W



3.



400



200



0 .



-



.



.



.



,



~ .. .



- . . .,~. .



400



Sulfate



-.-z

.



300



-



200



-



100



-



c



w



3



0'



A



J O J A J O J A J O J A J O J

1985

1986

1987

1988

Figure 7. Temporal variations in A horizon soil solution Al, NO;, and SO:+ in a red

spruce forest in the Great Smoky Mountains National Park, North Carolina. After Johnson

etal. (1991b).



48



WAYNE P. ROBARGE AND DALE W. JOHNSON



Figure 8. Soil solution composition in the Findley Lake, Washington; Whiteface Mountain, New York; and Clingman's Dome, Great Smoky Mountains National Park (GSMNP),

North Carolina. After Johnson and Taylor (1989).



increases in atmospheric deposition and internal nitrification rate (Johnson

et al., 1991b).



C. CAPACITY

CHANGES

DUETO LEACHING

In one of the most convincing case studies of soil change using the

space-for-time approach, Van Miegroet and Cole (1984, 1985) documented the effects of 55 years of excessive N fixation and NO; leaching

on soil properties in adjacent stands of red alder and Douglas fir [Pseudotsuga menziesii (Mirb Franco)] at the Thompson Research Center,

in the foothills of the Cascade mountains of Washington. The Douglas

fir stand was planted in 1931 after a series of wildfires following logging of

the original old-growth forest between 1910 and 1920. The adjacent red

alder forest established naturally a few years later where conifer planting

ceased in the burnt area. Nitrogen fixation in the red alder stand resulted

in the accumulation of approximately 5000 kg ha-' of soil N versus about

2000 kg ha-' in the adjacent Douglas fir stand. Nitrate concentrations in

the soil solution measured in 1981-1983 were on the order of 200300 pEq/liter in the red alder stand versus 0.5-2 pEq/liter in the Douglas

fir stand (Van Miegroet and Cole, 1988), indicating high rates of nitrification and considerable internal generation of nitric acid within the red alder

soil. It was estimated that nitrification resulted in an annual H+ input of

approximately 3.2 kmol ha-' yr-', a value not dissimilar from that attributable to wet and dry deposition reported for the Solling site (Matzner,



ACID DEPOSITION ON FORESTED SOILS



49



1989). This high rate of nitrate leaching decreased the surface soil pH

beneath the alder compared to the Douglas fir by over one-half pH unit

(pH 4.3 versus 5.0) and the base saturation by 50% (Fig. 9). The base

saturation in deeper soil horizons appeared to be higher, perhaps due to

displacement of base cations from the upper soil or perhaps due to increased weathering in these horizons. Such changes in soil acidity due to

occupancy by Ainus species have also been observed by Franklin et al.

(1968) and Binkley and Sollins (1990).

In another space-for-time study, Tamm and Hallbacken (1988),

through a process of elimination, attributed pH decreases in C horizons

of soils in southern Sweden over the period from 1926 to 1985 to acidic

deposition. In reaching this conclusion, they compared pH changes in all

horizons over this 60-year period at two Norway spruce (Picea abies) sites:

one in northern Sweden, where acidic deposition inputs were minimal, and

one in southern Sweden, where acidic deposition inputs were elevated. At

both sites, there was a clear trend toward decreasing pH in surface soils

with stand age, presumably due to accumulations of acidic humus. At the

northern site, no pH decreases were observed in the C horizon, whereas

substantial C horizon pH decreases were noted in the southern site

(Fig. 10). The possibilities that differences in forest history, climate, and

soil mineralogy could account for these differences in pH were discounted,

leaving leaching as the most likely cause. The authors make the tacit

assumption that natural leaching rates are similar at the two sites, and

therefore conclude that acidic deposition is the major cause of the C

horizon pH decreases at the southern site. There are no data to support the



Figure 9. Soil exchangeable cations beneath red alder (RA) and Douglas fir (DF) stands

at the Thompson Research Center, Washington. After Van Miegroet and Cole (1984).



50



WAYNE P. ROBARGE AND DALE W. JOHNSON



Figure 10. Changes in soil pH beneath spruce forests in northern (Kulbacksliden site,

left) and southern (right, Tonnersjoheden) Sweden. After Tamm and Hallbacken (1988).



ACID DEPOSITION ON FORESTED SOILS



51



assumption that natural leaching rates are similar at the two sites, and the

possible effect of previous occupancy by much deeper-rooted beech (Fagus

sylvatica) at the southern site (Tamm and Hallbacken, 1988) raises questions about the role of uptake in causing the C horizon pH decreases.

However, there is little doubt that acid deposition could either cause or

substantially contribute to increases in soil acidity over such a time period.

There have been several other studies of soil change in Sweden that are

of note. Falkengren-Grerup (1987) reported fairly consistent increases in

acidity and decreases in exchangeable bases in nine sites (both forested and

nonforested) in southern Sweden from 1949 to 1985. The increases in

acidity were much more substantial in soils that were initially less acid, as

predicted by Reuss (1983) and as also noted by Anderson (1988) (Fig. 11).

The authors hypothesize that acid deposition was a major factor in causing

the increases in acidity, but have no data from element budgets to compare

the relative effects of acid deposition, plant uptake, and natural leaching

on these systems. Falkengren-Grerup and Eriksson (1990) noted decreases

in pH and exchangeable cations between 1947 and 1988 in the C horizons

of F. sylvatica and Quercus robur sites in southernmost Sweden. An

unexpectedly large increase in the growth of the beech stands noted during



0

00



ce



I



0



0

5,



I



.



8 O

5.

9 6



00



.9



Figure 11. Changes in soil pH as a function of original pH in A horizon samples from a

variety of sites in Sweden between 1949 and 1970 and 1984 and 1985. After FalkengrenGrerup et al. (1987).



52



WAYNE P. ROBARGE AND DALE W. JOHNSON



this period was attributed to increases in atmospheric N deposition. Once

again, no element budget data were available to calculate the relative

importance of acidic deposition, uptake, and natural leaching, any one of

which might have caused a substantial amount of soil acidification over the

four decades involved in this comparison.

Subsequent to the initial studies at Solling (Ulrich et af., 1980), decreases

in exchangeable Ca2' and Mg2+ have been found in many sites in Germany (Hildebrand, 1986a; Grimm and Rehfuess, 1986; Hauhs, 1989;

Ulrich, 1989; see also review by Johnson et af., 1991a). The study by

Hauhs (1989) at Lange Bramke was accompanied by a partial budget

indicating that leaching was of major importance in causing substantial

reductions in soil exchangeable base cations over a 10-year period (19741984). Biomass uptake and increment values are not presented, but the

high rates of S (46 kg ha-' yr-') and N (19.4 kg ha-' yr-') deposition,

the high soil solution SO$- concentrations (230-500 pEq/liter), and the

trend of decreasing soil solution Ca2+ concentrations from 1977 to 1985

collectively present a very convincing case for acid deposition-caused soil

acidification.

The importance of leaching in causing soil change can vary considerably

among individual cations, as shown by Johnson et at. (1988a) and Johnson

and Todd (1990) in their studies of soil change and the reasons for it on

Walker Branch Watershed, Tennessee. Johnson et al. (1988a) noted

marked reductions in exchangeable Ca2' and Mg2+ in subsurface horizons

over the period 1971-1982 (Fig. 12). The decreases in exchangeable Ca2+,

which occurred in about half of the eight sites sampled, were expected due

to large accumulations of Ca in forest biomass. The reductions in exchangeable Mg2+, which occurred in nearly every site sampled, could not

be accounted for by Mg accumulation in forest biomass, however, leaving

leaching as the most likely cause. Previous studies on Walker Branch had

shown that ieaching was dominated by SO:- from atmospheric deposition

(Richter et af., 1983; Johnson et af., 1985), and thus it was hypothesized

that the reductions in exchangeable Mg2+ were due primarily to acidic

deposition. In a follow-up study on four of the eight sites, Johnson and

Todd (1990) constructed element budgets (using tension lysimeters to

measure leaching output) to test the hypotheses posed in the initial study.

They found support for the hypothesis that the Ca2+ decreases were due

primarily to tree Ca accumulation, which greatly exceeded leaching in

those sites where exchangeable Ca2+ decreased (Table 111).However, the

authors could neither support nor reject the hypothesis that leaching was

the primary cause of the Mg2+ decreases across all the sites from the data

available. Leaching was clearly the dominant mechanism of Mg export in



53



ACID DEPOSITION ON FORESTED SOILS



a

0.5



w



3 0.4

\



m

5 0.3



-



0



5



OS2

0.1

0.0



26



98



281



PINE POPLAR



42



179



91



107



237



CH. OAK



OAK-HICKORY



42



91



b

1.0



w



1



0.8



x



\



Q



0



0.6



-



0



Eu



OA

0.2



0.0

26



98



281



PINE POPLAR



179



107



237



CH. OAK OAK-HICKORY



Figure 12. (a) Changes in Bt horizon exchangeable Mg2+ at selected sites in the Walker

Branch Watershed, Tennessee, from 1971 to 1982. Numbers on the x axis represent individual plots. Asterisks indicate differences detected at the 5% level of significance.

(b) Changes in Bt horizon exchangeable CaZ+at selected sites in the Walker Branch Watershed, Tennessee, from 1971 to 1982. Numbers on the x axis represent individual plots. Asterisks indicate differences detected at the 5% level of significance. After Johnson et al. (1988a).



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