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OCR for page 4
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
OCR for page 5
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
OCR for page 6
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.
OCR for page 7
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.
OCR for page 8
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
OCR for page 9
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;
OCR for page 10
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:
OCR for page 11
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
OCR for page 12
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.
OCR for page 13
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
OCR for page 14
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
OCR for page 33
33
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OCR for page 34
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
OCR for page 35
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
OCR for page 36
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
OCR for page 37
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.
OCR for page 38
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.
. . . . . . —
OCR for page 39
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.
OCR for page 40
40
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