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2 STATE OF KNOWLEDGE ME:TAL-pH-ECOSYSTEM INTERACTIONS DI SCUSSTON OF FRAME:WORE QUESTIONS The following discussion is organized into four parts (atmosphere, terrestr ial ecosystems , aquatic ecosystems, and sediments) in accordance with the framework depicted in Figure 1.1. The questions listed in each part refer to the questions shown in the figure and addressed in the Appendix (Tables A.1 through A.18~. Atmospher e Question 1: Is Deposition Controlled by Human Processes? Whether human activity or natural emission processes control the current concentrations of metals in precipitation can be assessed in four ways: 1. by comparing the metal emission rates of human sources and natural sources, 2. by comparing the ratios of metal concentrations in the atmosphere to the ratios of metal concentrations in the natural sources, 3. by determining historical trends of metal concentrations in atmospheric deposition, 4. by comparing concentrations of metals in wet deposition in urban, rural, and remote areas. Using these four criteria, Galloway et al. (1982) divided the metals of concern in this report into three groups: · metals whose rates of atmospheric deposition in eastern North America are controlled by anthropogenic processes. · metals whose rates of atmospheric deposition in eastern North America are still controlled by natural processes. · metals for which there is insufficient evidence to determine the effect of anthropogenic activities on the rate of atmospheric deposition. 4

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5 A review of the literature suggests that deposition rates for AS, Cd, Cu. Pb, Mn, Hg, Ni, Se, Ag, V, and Zn are controlled by anthropo- genic activities (Galloway et al., 1982; Lantzy and MacKenzie, 1979~. There are insufficient data from which to draw conclusions on atmos- pheric deposition rates for Be, Co, Mo, Sn, Te, and T1. Of the 18 elements considered in this report, the sole one for which the atmospheric deposition rate is known to be controlled by natural sources is A1, due to its prevalence in geological materials (Galloway et al., 1982; Jeffries and Snyder, 1981~. The magnitude of the effect of anthropogenic activities is illustrated by a comparison of the concentrations of metals in wet deposition in urban, rural, and remote sites (Table 2.1~. Question 2: Do We Know the Speciation of the Metals in Atmopsheric Deposition? This question may be divided into three parts: (a) Do we know the physical speciation, i.e., dissolved versus particulate? (b) Do we know the chemical speciation? (c) Which is most important, wet or dry deposition? Physical Speciation. There have been few studies on the distribution of metals between the dissolved and particulate phases of wet deposition. For Ag, AS, Be, Co, Hg, Mo, Se, Sn, Te, T1, and V, we were unable to find any references. The few papers on the subject discussed A1, Cd, Cu. Me, Ni, Pb, and/or Zn. Lindberg et al. (1977) used 0.5-pm filters to separate dissolved from particulate material in summer rain from Walker Branch, Tennessee. They found that 90%, 75%, and 92% of the Ma, Pb, and Zn, respectively, were in the dissolved phase. Tanaka et al. {1981) used a 0.4-pm filter to differentiate between dissolved and particulate Ni, Zn, Cu. Pb, and A1 in winter rain from Tallahassee, Florida. Their results showed that A1 occurred exclusively in the particulate fraction, while Ni, Cu. and Zn were present principally as dissolved elements. Pb was more evenly distributed between the two fractions. Gatz et al. {1984) separated dissolved Zn , Cd, Cu. and Pb from particulate forms using a 0.4-pm filter on 49 weekly wet deposition samples collected in a Chicago suburb. They found that the medians of the percent dissolved distributions were 95%, 95%, 90%, and 83% for Zn, Cd, Cu. and Pb, respectively. - - Chemical Speciation. To be able to link metal deposition with ecological changes, it is necessary to know not only the rate of deposition but also the chemical Speciation of the metal. Its geochemical mobility and biological availability may both depend on the form of metal deposited. Dissolved metals in wet deposition can exist in a variety of ionic forms depending on the characteristics of the individual metals, pH of the precipitation, and kinetics of transformation from one chemical

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6 TABLE 2.l Median Concentrations of Metals in Total Wet Deposition (pg/~) Metal Urban Rural Remote As 5.8 0.286 0.019 Cd 0.7 0.5 0.008 Co 1.8 0.75 -- Cu 41 5.4 0.060 Pb 44 12 0.69 Mn 23 5.7 0.194 Hg 0.745 0.09 0.079 Mo 0.20 —- - - Ni 12 2.4 -- Ag 3.2 0.54 0.007 V 42 9 0.163 Zn 34 36 0 .22 NOTE: If only one concentration was available, it was used as the median value. If the median fell between two numbers, the average of the two was used. If a range was reported, the midpoint of the range was used. Insufficient data exist for Be, Se, Sn, Te, and T1. No bulk data were used. SOURCE: Galloway et al. (1982~. species to another. There are two methods to estimate the chemical speciation of metals in precipitation: direct measurement and thermodynamic modeling. Both approaches have disadvantages. Existing analytical methods are often limited, either by their detection limits or because they involve perturbation of the sample that could alter the speciation of the metal. Thermodynamic modeling gives the speciatio~r assuming that all metals are at equilibrium. This ignores the fact that, at times, the kinetics of thermodynamically favorable reactions are too slow to result in equilibrium.

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7 TABLE 2 .2 Calculated Speciation of Dissolved Trace Metals in a Typical Rainwater (pa 4.14, pC1 4.88) Total Percent Concentration Aquo Ion Other Dissolved Species Metal ~lo-8M, MZ~ (H2O) n (>1%) Ag 0.50 98 AgCl+ (2%) Cd 0.45 100 Co 1.27 100 Cu 8.50 100 Hg 0.45 <1 - HgC12 (60%), HgCl+ (2%), HgOHCl (26%), Hg(OH)2 (118) Mn 10.4 100 Ni 4.09 100 Pb 5 . 7 9 10 0 Zn 55.1 100 As 0.38 Mo 0.21 V 17.7 H2As0~2 (100 % ) MMoO4 (3496), MOO42 (66%~ HVO4 (100%) We have been unable to find any references on measurements of the chemical speciation of metals in wet deposition and have found only one reference on the use of thermodynamic models to predict the speciation. Sposito et al. (1980) used the computer program GEOCHEM to calculate the equilibrium speciation of acid precipitation from New England for a few metals (Pb, Mn, Al, Cd, Cu. Zn, and Ni). In order to check their analyses and to include additional metals (Ag, Co, Hg, As, Mo, and V), we used the MINEQL thermodynamic model (Westall et al., 1976) to determine the chemical speciation of the metals in precipitation typical of central eastern North America. The data used are from the annual volume weighted concentrations of major inorganic ions in precipitation at the Hubbard Brook Exper imental Forest for the period 1963 to 1974 (Likens et al., 1977), and the median concentrations of trace metals measured in rural locations (from Table 2.1~. The results of our analyses (Table 2.2) agree very well with those of Sposito et al. (1980~. Specifically, based on the MINEQL and GEOCHEM models, Al, Ag, Cd, Co, Cu. Mn, Ni, Pb, and Zn are present as the aqua ion (M +tH2O)n). Mercury exists primarily as HgC12 (60~) and HgOHC1 (26%~. Arsenic exists as H2ASO4; Mo exists as MoO4 (66%) and HMoO4~34~; and V exists as HOOD. Changes in chloride concentration would influence the speciation of Ag and Hg.

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8 TABLE 2.3 Mean Values of Data Reported from All Seasons for Dry Fraction of Total Deposition Metal Mar ine Rural urban As -- -- 0.2 Cd 0.4 0.4 0.6 Cu 0.5 -- __ Pb 0.6 0.3 0.2 Mn 0.5 O.S 0.5 Ni 0.6 0.5 O.5 V 0.4 -- __ an 0.7 0.4 0.5 SOURCES: Duce, 1979 (marine); Feely and Larsen, 1979 (rural and urban); Lindberg and Harriss, 1981 (rural). Importance of Wet Versus Dry Deposition. Metals may be deposited from the atmosphere via wet or dry deposition. Dry deposition can occur in three ways: by gravitational settling (generally particles greater than 2 to 10 Am, depending on meteorological conditions), by aerosol diffusions and/or impaction (<10 ~m), and by gaseous diffusion. The rate of gravitational settling is typically estimated by what falls into an open container. To measure the rates of aerosol impaction and gaseous adsorption, estimates of atmospheric concentrations and deposition velocities are required. For the metals in this study, AS, Hg, Se, and perhaps Cd are the only ones with large vapor pressures. Therefore most metals will be in aerosol form and will be deposited by gravitational settling or aerosol impaction. Measurements of toxic metals in deposition suggest that the dry fraction is substantial. Few cases in the literature have indicated that the dry fraction is less than 0.1 of the total deposit. For the most part, the mean dry fraction lies between 0.3 and 0.6. The further the site is away from a source area, the less important dry deposition will be compared to wet deposition. Table 2.3 summarizes data for several toxic metals from deposition in marine, rural, and urban areas. No systematic trend between the different environments is evident. In summary, based on these initial studies, we are convinced that dry deposition of metals is important relative to wet deposition, but the degree of importance has yet to be determined primarily because of inadequate techniques to determine the rate of dry deposition. In addition, the deposition of fog, dew, frost, and cloud water has the potential to be a significant mechanism for transferring metals to forests. Lovett (1984) and others have shown that there is a signifi- cant amount of cloud water deposition to forests, but little work has

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9 been done on the metal composition of this cloud water. This is clearly an area for research. Terrestrial Ecosystems Question 3: Does the Metal Accumulate in the System? Question 4: Does the Metal Move in Solution to the Lake? Question 5: IS There an Interaction of the Metal with Acidification over the pH Range 7 to 4? In the general area of trace metal mobility in the terrestrial ecosystem (Questions 3 and 4) and possible interactions with acidification (Question 5), three types of investigation can be identified: 1. laboratory experiments with soils, or with individual soil components (e.g., clays, iron and manganese oxides, humic acids), in which interactions between a particular metal and the solid phasets) have been studied under different pH conditions (usually in stirred suspensions); 2. laboratory or field experiments in which soil columns have been subjected to simulated or real precipitation events, and the chemical composition of the percolate has been followed over time; 3. regional surveys of soils, vegetation, and/or surface waters in areas of relative geologic homogeneity where there exists a gradient in precipitation chemistry (increasing acidity). Each of these approaches is discussed below. Soil and Soil Components in Aqueous Suspension. In the first type of study, if the percent metal adsorbed is plotted as a function of pH, a relationship similar to that shown in Figure 2.1 is usually obtained; for those metals existing as cationic species, an abrupt increase in the amount of adsorbed metal occurs over a narrow pH range. In such a plot, pH50, the pH at which 50% of the original cation concentration is adsorbed, is a useful parameter for comparing the adsorption of several different cations on a given solid, or the adsorption of a particular metal on several different solid phases (Rinniburgh and Jackson, 1981~. (See Table 2.4 for an example with two different solid phases, Fe(OH)3 gel and A1(OH)3 gel.) These laboratory studies are useful in illustrating the potential effects of a relatively small change in pH on trace metal mobility and demonstrating the selectivity of certain solid phases for different trace metal cations. However, in a given metal-solid system, the position and shape of the pa adsorption sedges (i.e., the value of peso) will depend on the following: · the composition of the solid phase; · the concentration of the solid phase;

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10 100 c: UJ m a: o In a 50 Ct: _ l _ _— / / 1 pH FIGURE 2.1 Typical pH adsorption curve for divalent cations on hydrous metal oxides (pH50 values may range from about 3 to 8~. . the initial concentration of the metal, [M]; · the nature and concentration of any competing metals present; · the nature and concentration of any ligands present, [L] (Benjamin and Leckie, 1981a,b; Davis and Leckie, 1978~. Relatively few studies have been performed with natural soils (or even with two or more competing solid phases), and virtually none have been carried out at the high (solid/solution) ratios that prevail in a soil system. Accordingly, the quantitative transposition of the results of this type of laboratory study to f ield conditions is unwarranted. In a qualitative sense, however, it could be anticipated that the larger the fraction of a particular metal present on exchange sites in a soil, the more sensitive the metal will be to minor pH changes. Soil Columns. In the second type of study, involving the leaching of soil columns, it in not surpr ising that trace metals have been found to migrate at different rates and to exhibit different sensitivities to pa changes. Three representative studies are discussed here. Fuller et al. (1976) compared 11 soils representing 7 different soil types and reported that Al, Cu. Cd, Fe, Mn, Ni, and Zn were more highly solubilized by dilute acid (H2SO4, pH 3.0) than by deionized water; this increased mobility was particularly noticeable for Al' Fe, and En, whereas Co and Pb remained immobile even when subjected to the acid leach. The relative order of mobility was thus as follows:

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11 TABLE 2.4 Adsorption of Trace Metals on Colloidal Metal Hydroxides Fe (OH) ~ gela Al(OH)~. gel metalb: pE50 c metal: pH50 Pb < Cu < Zn < Ni < Cd - Co 3.0 4.3 5.3 5.7 5.9 Cu < Pb ~ Zn < Ni < Cd - Co 4.8 5.2 5.6 6.3 6.6 aGel concentration: 9.3 x 10-2M A1 or Fe. brace metal concentration: 1.25 x 10-4M. CpH50: pH at which 50% of original metal concentration is adsorbed. SOURCE: Kinniburgh and Jackson (19811. A1, Fe, Zn > Cd, Cu. Mh, Ni ~ Co, Pb. In a similar exper iment, Tyler (1978) leached two organic spruce forest soils with artificial rainwater (pH 4.2, 3.2, and 2.8) and established the following ranking (decreasing percent leached): Mn _ Zn > Cd ~ Ni > Cu > V > Cr > Pb. The residence times for all the above elements except V and Cr decreased with increasing acidity of the simulated precipitation. In a subsequent two-and-one-half-year field study, Tyler (1981) quantified the amounts of metal (A1, Fe, Mn, Cd, Cr. Cu. Ni, Pb, V, and Zn) leached from the A-horizon of a podzolic spruce forest soil in southern Sweden and identified two distinct seasonal patterns. One metal group (Cr. Pb, Ni, V, A1, and Fe) exhibited maximum concentrations In late summer and autumn and minimum values in winter and early spring; considerable losses of these elements occurred under conditions favoring the leaching of organic matter (high soil temperature and moisture content). Tyler notes that the Pb reprecipitates in the B-horizon and thus is not lost to the ground water. The second group (Cd, Mh, and Zn) attained maximum concentrations during the winter months. This latter group, comprising those metals having an appreciable fraction present in exchangeable form, was susceptible to minor pH changes in the soil percolate and the concentration variations were positively correlated with [H+~. Tyler speculated that a lowering of soil pH would reduce the residence times of these elements in particular. Regional Surveys. With reference to the possible accumulation of airborne metals in the terrestrial system, Steinnes (1983) sampled 500 natural soils in a nationwide survey of metals in Norway and found a

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12 distinct geographical distribution of As, Cd, and Pb, which closely resembled the pattern of measured atmospher ic deposition. Be proposed that the higher levels in southern Norway were due to long-range atmospheric transport. Data from peat profiles in 13 Norwegian ombrotrophic bogs (Hvatum et al., 1983) supported Steinnes's contention for the distribution of As, Cd, and Pb and showed similar patterns, also suggesting long-range transport and accumulation, for Co, Cr. Cu , Fe, Mn, and Ni. Local natural sources of airborne Se added to the levels of Se or iginating from long-range atmospher to transpor t of this element. The moss Hylocomium spl endens was also collected from the sites used for soil analysis, and its metal content showed very large geographical differences for As, Pb, and Sb, with much higher values in southern Norway. A similar but less pronounced pattern of elevation was shown for Ag, Cd, Se, V, and Zn (Rambaek and Steinnes, 19801. In general, the net amount of metal added to soil from the atmosphere as a result of long-range transport is not considered a major source of metal contamination for terrestrial systems. Although there is good evidence of increased inputs, any environmental effects are likely to be a result of changes in mobility, because absolute increases in metal levels are relatively small in comparison with background levels. In regional surveys designed to evaluate the mobility of metals in the terrestrial system, changes in soil, water, or sediment chemistry noted along a precipitation pH gradient have been attributed to the interaction of the acid precipitation with the terrestrial system. Several examples are presented below. In the course of a regional soil survey in Scotland, McLaren and Crawford (1973a) determined the relative distr ibution of Cu among the forms: Cu-exchangeable ~ Cu-adsorbed · Cu-adsorbed - _ (inorganic) ~ (organic) As the natural pH of the soil decreased, this equilibrium was observed to shift to the left; i.e., a greater proportion of Cu was found in the easily exchangeable form. In their general review of the ecological effects of acid precipitation, Almer et al. (1978) present data for waters from different lakes in Sweden showing an increase in the concentrations of Al, Mn, Cd, and Hg along a spatial gradient toward lower pH (see Figure 2.2 for the Al data). More recently, Dickson (1980) has added Pb to this group of metals, noting that these trends apply to differently acidified lakes in Sweden subjected to similar atmospheric loadings. This Similarity of atmospheric loadings," a key point for Cd, Hg, and Pb, could not be verified as the relevant data were neither presented nor cited. Henriksen and Wright (1978) reported similar results for Pb and Zn in a series of small lakes in southern Norway, but they noted that they could not distinguish between Acidification + mobilization" or Acidification + concomitant increased atmospheric loading. as pass ible mechanisms.

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13 700 600 - o 500 - ~r CC 400 z o i) 300 i 200 J 100 o . . · ·:: ·. i. :-. · — . · . - - . _ . . . I I · · Hi · ; ~ · 1 4 5 6 7 8 pH FIGURE 2.2 Relationship between pH and total (unfiltered) Al concentrations in Swedish lakes (redrawn from Dickson, 1980~. For well-drained organic soils in North America, Hanson et al (1982) studied a number of metals (Al, Ca, Cd, Mg, Mn, Na, Pb, and Zn) and reported that they are differentially leached at an accelerated rate as precipitation pH decreases. The results of the study showed the following: Mn > Ca > Mg ~ Zn > Cd; Na > Al. In this system, nearly 100% of the incoming Pb in precipitation is retained in the organic soil layer (see also Smith and Siccama, 1981~. The authors suggest that Zn is also accumulated in soils where the pH values for precipitation and soil solution are generally Greater than 5.0-5.5~; at lower pH values Zn and other Chemically similar elements are leached at an accelerated rate. Recent studies of trace metal profiles in forest floors in remote regions showed similar vertical profiles for Pb, Cu. Zn, Ni, and percent organic matter in three different forest types, with highest concentrations of each of these metals 2 to 4 cm below the surface. Metals concentrated in the He horizon (Friedland et al., 1984a). Studies of the metal cycling by forest vegetation in the Solling

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14 project in Germany Showed annual inputs of Cr. Mn, and Ni from the atmosphere to be low (<30%) compared with the amount stored in the annual increment of biomass. For Fe the percentage was higher (40 to 60%) , while for Cu. inputs were 100% from the atmosphere (Heinrichs and Mayer, 1980~. Time trends of metal accumulation in soils are rarely available except for high loadings, e.g., smelter studies. However, a recent study by Friedland et al. (1984b) on Camels Hump Mountain indicated an increase of Pb, Cu. and Zn in the forest floor over the period of 14 years since 1966; this increase was consistent with annual atmospheric deposition rates reported in the literature for these metals. Johnson et al. (1982} estimated a 5- to 10-fold increase in Pb in forest floors of the northeastern United States over the last century. Summary. Based on an analysis of the available experimental results (types 1, 2, and 3), the 18 trace metals of current concern are tentatively classified according to their potential mobility in a terrestrial environment subject to acid precipitation: high mobility: A1, Cd, Me, Zn moderate mobility: Cu. Ni low mobility: Co, Pb, V Insufficient data are available for Ag, As, Be, Hg, Ma, Se, Sn, Te, and T1. Question 6a: Is the Metal Bioavailable? Question fib: Does the Metal Bioaccumulate? There is no strong evidence for accumulation by terrestrial plants of metals derived from long-range atmospheric transport, although metals do accumulate in litter and organic layers of forests in areas remote from point sources (Johnson et al., 1982~. However, if soils acidify, there is some potential for plant uptake to increase with the mobilization and increased availability of the following metals in soils: A1, Mn, Fe, Zn, Cu. and Ni (Hutchinson and Collins, 1978~. These metals may occur naturally in soils, or result from atmospheric deposition. In naturally acid soils from California, A1 and Mn in soil solution ranged from O. . 1 to 108 mg/L and 2.6 to 200 mg/L r respectively (Straughan et al., 19811. It has long been known that one of the beneficial effects of liming is decreased solubility of Al and Mn (Vlamis, 1953~. Foy et al. (1978) found that the limitation to the growth of certain calcicolous plants on acid soils is directly related to A1 toxicity. In the absence of measurements of metal uptake and accumulation by plants from soils exposed to acid precipitation, the information cited in the Tables A.1 through A.18 has been taken from recent articles and reviews that deal with relatively heavy loadings; for the most part these have addressed either sewage sludge application or local point

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33 ~2 0 U] ~4 As o A Sit EN o o o A s o to o U] 0 to . ED 0 c 0 0 fir: to $ U] - o _' a: 0 A: . ~ V] · - 5 to O o ~ a' ~ ~ e ~ ~ ~ ~ o ~ ~ ~ ~ ~ co ~ e ~ ~ ~ ~ IS, C ' ~ ~ e.^ ~ ~ en ~ ~ ~ - ~ · 0 0 . - ~ ' 3 ~5 ~ O ~ - - - ~ O ~ ~ ~ ~ O ~ O S c) t) 0 c; ~ U] ~ U) ~r ~ ~ c~ · · · ~ ~r ~ ~ ~r 1 1 1 1 ~r ~ ~ c~ ~ ~ O · · · e · · · O . · · 1 O O O O O ~ ~ ~ O ~) In 11~) o o x: :1E: x: :lE~ U~ kD X 1 1 1 1 1 O O O O O I ~ t~ 1 X X X X X o 1 cn · ~ ~ X O Q ~ ~ Q O C) C) ~ ~ C) C~ ~ C) . ~ · U) · ' U) ~ ~ ~ ~ _I _t ~ c) ' ·— ' ~ s ~ · S ~ e~ ~ . - ~ O P~ C .' - ~1 ~q ~ 0 ~ · Z Z 0 C ~Q ~ ~ O ~ O ~ C ~ U) ~1 ~ $¢ O 0 o IO U] o erl 0 o o · - - 11~ e~ · - o S o dP o U, - S ut o eO _~.

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34 TABLE 2.11 Metals Influence of pa on the Speciation of Sedbment-8Ound Trace Metal ApH Effect Reference Cd 8 ~ 5 Relatively mobile; organic forms Rhalid decreased as pH lowered; et al., 1981 accompanying increase in dissolved and exchangeable forms; no significant changes in Cd - associated with reducible fraction. Pb 8 ~ 5 Immobile; little or no dissolved Pb Gambrell detected at any pa; exchangeable et al., 1980 forms increased under moderately acid conditions (pa 5.0~; no Significant changes in Pb associated with reducible fraction. Zn 8 ~ 5 Relatively mobile; dissolved and Ga~hrell exchangeable Zn increased et al., 1980 markedly as pH decreased; concomitant decrease in Zn associated with reducible fraction. Hg 8 ~ 5 Immobile, little or no dissolved fig Gambrell detected at any pa; exchangeable et al., 1980 Hg increased slightly under moderately acid, reduced (pa 5.0, -150 mV) and weakly alkaline, oxidized (pE 8.0, +500 mV) conditions. A different laboratory approach has been to determine the effect of a change in pH on trace metal partitioning in the sediment, an deter- mined experimentally by sequential selective extractions. Several such studies on river sediments have been carried out by Gambrell et al. (1980) and Khalid et al. (1981) at the Laboratory for Wetland Soils and Sediments, Louisiana State University {Table 2.113. In their closed systems, a decrease in pa from 8 to 5 led to an increase in the levels of dissolved and exchangeable Cd and Zn, i.e., to an increase in the geochemical mobility of these metals. In a natural (i.e., open) system, Such an increase in mobility would result in a net loss of the metal from the sediment. An interesting field corroboration of this suggestion is provided by the recent work of Reuther et al. {1981), who determined the partitioning of several trace metals in sediment cores from lakes Hov~atn (pH 4.4) and Langtjern (pa 4.95) in Norway. They

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35 TABLE 2.12 Trace Metals Released from Sediments in Response to Acidification Metal Type of Evidence Reference Al Experimental acidification Schindler et al., 1980a,b of lakes Experimental acidification Hall et al., 1980 of streams Paleolimnological study of Schofield, 1980 lake sediments Mn, Fe Experimental acidification Schindler et al., 1980a,b of lakes Zn Experimental acidification Schindler et al., 1980a,b of lakes Regional survey of lake Hanson et al., 1982 sediments Paleolimnological study Hanson et al., 1982 of lake sediments Norton et al., 1981 Reuther et al., 1981 Cd, Co, Ni Paleolimnological study Reuther et al., 1981 of lake sediments found that Cd and Zn had been remobilized from the recent sediment strata in the lake of lower pH, mainly from the organic and the easily reducible fractions, respectively. Co and Ni also showed evidence of remobilization, but to a lesser degree, whereas Pb and Cu were little affected. Regional Sediment Surveys/Experimental Acidification Experiments. Both regional sediment surveys in areas affected by acid precipitation and the results from the experimental acidification of lakes and streams suggest that the acidification of an overlying water column will increase the geochemical mobility in the sediments of the following trace metals: Al, Cd, Co, Fe, Mn, Ni, and Zn. AS indicated in Table 2.12, the evidence for Cd, Co, and Ni is somewhat less persuasive than for the remaining elements because it is based on a single observation (Reuther et al., 1981~. This latter report suggests that remobilization of Cd, Co, Ni, and an from the sediment does not occur until the pH is below 4.95. The concept of a critical or threshold pH value has also been suggested for Zn remobilization (Norton et al., 1981~. Paleo- 1 Prolog ical data from sediment cores collected from lakes with relatively undisturbed watersheds revealed concentration versus depth profiles for Pb that increased monotonically toward the sediment-water

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36 interface. In contrast, the concentration of Zn first increased and then decreased in the most recent strata near the top of the core. This general relationship holds for some 20 lakes in the northeastern United States and in Norway where the pH is less than 5.5 (Norton et al., 1981~. The authors suggest that at pa values of >5.5 lake sediments either act as a sink for incoming particulate En or chemically scavenge dissolved Zn from the water column. In acidified lakes (pH of <5.5), not only do the sediments not scavenge Zn from the water column, but they also apparently yield dissolved Zn to the overlying waters. It has been suggested id. C. Kennedy, USGS, Menlo Park, California, personal communication, 1984) that the Remobilization threshold. may mark the boundary between desorption of exchangeable ions as the major process contributing to metal exchange, and dissolution of carrier phases containing occluding trace metals (e.g., Fe or Mn oxides). The paleolimnological data available for lake sediments in Norway, the northeastern United States, Ontario (Dillon and Evans, 1982), and Quebec (Ouellet and Jones, 1983) all suggest that Pb is strongly bound in lake sediments. The critical pH value for release of Pb lies some- where below the lowest observed water column pa values. Experimental confirmation of this immobility was recently provided by Davis et al. (1982), who subjected the upper strata from two sediment cores to acidification in the laboratory; significant lead release {>5% of total concentration) only occurred at pH < 3.0 in the sediment from Woods Lake and pH < 2.0 in that from Sagamore Lake. Question 16a: Is the Metal Bioavailable? Question lab: Does the Metal Bioaccumulate? Question 17: Does the Metal Have Biological Effects on the Benthic System? It is commonly accepted that most of the metals transported into an aquatic system are scavenged or precipitated and accumulated in sediments. This fact is the basis for using downcore profiles of metals for estimating historical patterns of deposition. Yet for the biota, which can only assimilate soluble material, only a fraction of the sediment metal is potentially available, even to sediment dwelling (benthic) organisms. For example, Tessier et al. (1984) showed that the Cu. Pb, and Zn levels in various tissues of the Ellintio comulanata a suspension-feeding freshwater mollusk, were best related not to total metal concentrations in the adjacent sediment, but rather to one or more of the relatively easily extracted fractions. Body burdens of Cu. Pb, and Zn were also influenced by the protection or competitive effect of other sediment constituents, notably amorphous iron oxyhydroxides and, to a lesser extent, organic matter. Benthic organisms can assimilate metals from dissolved forms in pore water or from particles ingested into the gut. Studies with tubificids in contaminated sediments indicated uptake mainly through

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37 the integument, but also showed deposition of ingested metals in the foregut (Back, 1983; Prosi, 1983~. The metal, released from Sediment by chemical or biochemical processes, is bioavailable for direct uptake, and metals in ingested sediment, if weakly bound, may be assimilated in the gut. Not surprisingly, it is technically difficult to distinguish between metals obtained by the biota directly from the water column or pore water, and those obtained via sedimented materials. If the sediment should be a major source for metals in the water column, this distinction for the biota may be academic. Studies on the geochemical characteristics of metals in sediments and on the relationship between pa and metal release from sediments suggest that there is a reasonable basis for expecting some metals in sediments to show increased bioavailability as pH declines. The literature is, however, quite deficient in definitive experiments on biological uptake made in conjunction with geochemical studies; those that exist tend to relate to heavily polluted sediments . and Anew ial lv ~ ~ = ~ , ~——_ _ = _ ~ ~ ~ ~ to marine systems (e.g., Luoma and Bryan, 1979). In responding to the questions on bioavailability and bioaccumulation, we found that the most frequently encountered studies used correlation analysis on field collections to determine relationships between metals in sediment (total or identified fractions) and metals in biota; and only one of these studies dealt specifically with acidified systems {Schindler et al., 1980a}. Significant correlations would provide circumstantial evidence for the sediment as a source of metal. Copper in benthic blots exhibits a relationship with Cu in sediments, but this has only been shown for relatively contaminated water bodies such as sites affected by mining activities (e.g., Tessier et al., 1982), or systems spiked with Cu (e.g., DikS and Allen, 1983~. Stokes et al. (1983a) found significant accumulation of Cu in benthic algae related to sediment contamination, but there was no relationship for softwater lakes undergoing acidification unless they were close to _ point sources ot CU. zn also chows sediment-related accumulation in biota (Forstner and Wittmann, 1981; Tessier et al., 1982), and a release of Zn from sediments of relatively low contamination occurred in the presence of an active infauna (Rrantzberg and Stokes, 1982~. Mercury mobilization from sediments has not been reported from experimental studies on acid-stressed systems. However, Beijer and Jernelov (1979) suggested that methyl mercury was formed in sediments. At pH below 5.0, the monomethyl form of mercury is expected to predominate (Fagerstrom and Jernelov, 1972~. Transplant experiments with freshwater clams (Karbe en al., 1975) showed Hg uptake to be related to a number of parameters including organic content and oxygen in sediment, but not to total Hg in sediments. Hakanson (1980) produced an empirical model that related Hg in fish to Hg in sediment, pH of the overlying water, and the nutrient status of the sediment (referred to as the bioproduction index). Stokes et al. (1983b) found no significant correlation between Hg in benthic algae and total Hg in sediments in a series of acidic and neutral lakes lacking any point source of Hg. Hg levels in algae were, however, as high as or higher than those in sediments and up to 10,000 times higher than in water. Algal Hg showed a significant correlation with Hg in yearling perch from the same lakes.

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38 These studies for Hg illustrate the complexity of the task of resolving the interacting factors determining biological uptake of sedimented metals. It is clear that bioaccumulation of a metal by a sediment-dwelling organism does not necessarily relate in any direct or simple way to the geochemically mobile forms of metals in sediments. The dilemma faced by the researcher is quite obvious: biota can only assimilate metals from solution, which they can either obtain by ingesting solid material and absorbing weakly bound metals from the solid, or by absorbing metal directly from the water or from the water that is ingested with the sediment. Since there is a flux of metals between sediment and water, it may be in fact Impossible to determine whether sediment-dwelling organisms are accumulating metal from water or sediment. Furthermore, it becomes impossible to determine whether observed effects of acidification on benthic biota result from metals in sediments, or directly from H+. The question on effects therefore has to be given the response ~unknown. for all the metals of concern. Benthic faunal communities show certain broad responses to acidification, but data for blots are available for the most part only from studies of acidification of the water column. and not from studies of the conditions in the ~:~ imprint. · . ~ ~ * ~ e itself {Krantzberg, 1982~. Faunal biomass and species composition of benthos may show quite wide variations among lakes of apparently comparable pa and alkalinity in the water column. The obvious question to address concerns conditions in the sediment. Sediment pH and metal availability may have been measured in geochemical studies, but these have not been linked to faunal or floral response=. Baker and Schofield (1982) were able to look at the effects of pH and A1 on fish and found, for example, that the presence of A1 at low pa was beneficial to certain life stages. They stated that Simple generalizations concerning the effect of increasing A1 concentrations with acidification are not possible. ~ For benthic biota, where our knowledge of effects is much more rudimentary, it has to be concluded that while the potential exists for metal toxicity in sediments of acid-stressed systems, there are as yet no available studies that separate the respective influence of low pa, low nutrients, and increased metal availability. The implications for increased availability of metals to the benthic biota are not trivial. Increased body burden may lead to transfer of metals to the terrestrial food chain for birds that feed on benthos or emerging insects (Erikason et al. , 1980), while toxicity to benthos may result in depletion of food supplies for these animals at subsequent trophic levels. Bioaccumulation of metals in benthic algae may be an important factor in the cycling of metals in acid-stressed systems. Finally, the long-term fate of sedimented metals in post- acidification (neutralization or recovery) scenarios is likely to be not only chemically but also biologically determined. Local effects of bioturbation (Rrantzberg and Stokes, 1981) may provide for release over extended periods of time, reinforcing the concept of the sediment as a reservoir rather than as a sink for metals. . . . . . . —

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39 SCARY This section contains a summary of the information presented in the preceding sections. Table 2.13 summarizes the materials presented in the Appendix (Tables A.1 through A.18~. The reader can either look down the table to determine quickly how much is known about individual metals, or across the table to determine how much Is known about specific questions indicated in Figure 1.1. To synthesize the results of our analysis and to extract an answer to the general question--Which metals should be of primary concern in the context of acid deposition?--we condensed the original framework of seventeen questions (Figure 1.1) to five factors: 1. Are metal concentrations in the atmosphere controlled by anthropogenic activities? 2. Do metal concentrations increase in response to acidification, in either the soil or the aquatic environment? 3. Does a pH change in the critical range 7 to 4 cause significant changes in metal speciation? 4. What is the inherent toxicity of the metal, and has toxicity to biota been observed in relation to acidification? 5. Does the metal have a tendency to bioconcentrate? The results of this summary analysis are presented in Table 2.14.

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