WATER POLLUTION CONTROL RESEARCH SERIES •
SAND AND GRAVEL
OVERLAY FOR CONTROL
OF MERCURY IN SEDIMENTS
U.S. ENVIRONMENTAL PROTECTION AGENCY
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WATER POLLUTION CONTROL RESEARCH SERIES
The Water Pollution Control Research Series describes the
results and progress in the control and abatement of pollution
in our Nation's waters. They provide a central source of
information on the research, development, and demonstration
activities in the water research program of the Environmental
Protection Agency, through inhouse research and grants and
contracts with Federal, State, and local agencies, research
institutions, and industrial organizations.
Inquiries pertaining to Water Pollution Control Research
Reports should be directed to the Chief, Publications Branch
(Water), Research Information Division, R&M, Environmental
Protection Agency, Washington, DC 20460.
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SAND AND GRAVEL OVERLAY FOR
CONTROL OF MERCURY IN SEDIMENTS
by
Leonard H. Bongers and
Mohammed N. Khattak
Research Insitute for Advanced Studies
(RIAS)
Martin Marietta Corporation
1450 South Rolling Road
Baltimore, Maryland 21227
for the
Office of Research and Monitoring
ENVIRONMENTAL PROTECTION AGENCY
Project Code #16080 HVA
Contract #68-01-0089
January 1972
For sale by the Superintendent of Documents, U.S. Government Printing Office, Washington, B.C. 20402 - Price 55 cents
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EPA Review Notice
This report has been reviewed by the Environmental Protection Agency
and approved for publication. Approval does not signify that the con-
tents necessarily reflect the views and policies of the Environmental
Protection Agency, nor does mention of trade names or commercial
products constitute endorsement or recommendation for use.
11
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ABSTRACT
The release of toxic mercurials by mercury-enriched river sediments
was examined in the laboratory. These tests indicated that about
1 \ig of methylmercury was released per m^ per day. The release of
such toxic mercurials could be prevented by a layer of sand, 6 cm in
thickness, applied over the mercury-enriched sediments. Layers of
fine or coarse gravel (6 cm deep) were as effective as sand. Thinner
layers of sand, 1.5 and 3 cm in thickness, appeared to be unsatisfactory.
The cost of applying 3-inch layers of sand or gravel over contaminated
river sediments is estimated to be about $3000 to $4000 per acre.
The formation of methylmercury occurred in sediments with low and
high organic content, in sediments with low and high cation exchange
capacity, and in aerobic and anaerobic sediments.
A convenient indicator of the potential toxicity of a contaminated sed-
iment is the presence of metallic mercury. The slow release of metallic
mercury occurred in aerobic sediments, but the release was much faster
in anaerobic sediments. Using ascorbate as an artificial electron donor,
metallic mercury could be released at high rates from aerobic sediments
as well. Ascorbate appeared to be a helpful indicator of the presence
of divalent biologically accessible mercury.
Although the laboratory investigations proved the soundness of the sand
blanket approach, its practical and economic feasibility must be deter-
mined in a combined field and laboratory analysis program.
This report is submitted in fulfillment of Project 16080 HVA, Contract
#68-01-0089, under the sponsorship of the Office of Research and
Monitoring, Environmental Protection Agency.
111
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CONTENTS
Section page
I Conclusions 1
II Recommendations 3
III Introduction 5
IV Materials and Methods 7
V Results and Discussion 17
VI Economic Considerations 37
VII Acknowledgments 41
VIII References 43
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FIGURES
Page
1 Schematic of closed system used for mercury analysis. 8
2 Calibration curve for absorption at 253. 7 nm for mercury 9
concentrations between 0 and 0. 7 )j,g/sample.
3 Calibration curve for absorption at 253. 7 nm for mercury 10
concentrations between 0 and 15 p,g/sample.
4 Calibration curve for absorption at 253. 7 nm for mercury 11
concentrations between 0 and 25 ng/sample.
5 Schematic of flow-through system used to measure 12
mercury vapor released directly from sediments.
6 Schematic of laboratory incubation procedure. 14
7 Time course of mercury assimilation by guppies. 20
8 Methylmercury uptake by guppies as a function of con- 21
centration.
9 Effect of a protective cover 6 cm in depth on mercury 28
accumulation by guppies.
10 Transformation pathways of mercury and its derivatives 30
in aquatic environment.
11 Rate of Hg° evolution from aerobic sediments spiked 32
with HgCl2.
12 Rate of Hg° evolution from anaerobic sediments spiked 34
•with 50 |j,g of
VI
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TABLES
No. Pag
1 Distribution of Hg between solid and liquid phase 17
in 10 g (dry weight) of sediment mixed with 1000 |j,g
of Hg as HgCl2.
2 Distribution of Hg between solid and liquid phase 18
in 10 g (dry weight) of sediment mixed with 200 |Xg
of Hg as CH3HgCl.
3 Effect of sand overburden on the mercury content 23
(Hg) of surviving guppies incubated for 3 weeks using
Sediment I enriched with HgCl2-
4 Effect of sand overburden on the mercury content 24
(ug) of surviving guppies incubated for 4 weeks using
Sediment II enriched with HgCl;?.
5 Mercury content of sand and sediment above and below 24
Hg-enriched layer.
6 Mercury content of the liquid phase at the end of a 25
3 -week incubation using Sediment I.
7 Effect of a 6 -cm overburden of sand or gravel on 26
the mercury content (|j,g) of surviving guppies
incubated for 3 weeks in the presence of Sediment
II enriched with
8 Estimate of mercury content (p,g) in the liquid phases 29
and acid extracts of mercury-enriched sediments in-
cubated aerobically or anaerobically.
9 Estimated fixed /variable costs of distributing sand 38
in an area south of Wyandotte.
10 Estimate of the cost involved in the application of 39
3 inches of sand to 2, 25, and 50 acres of sediment
contaminated with mercury.
VII
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SECTION I
CONCLUSIONS
Laboratory simulation studies dealing with the release of toxic mercury
from river sediments enriched with mercuric chloride showed that:
1. Approximately 1 p,g of methylmercury is released per m of
sediment per day.
2. Sand applied to a thickness of 6 cm will prevent the release of this
mercury during a 4-week laboratory incubation. Thinner layers of sand
were unsatisfactory under the experimental conditions.
3. Other aggregates, such as fine and coarse gravel, applied to the same
thicknesss were as effective as sand.
4. A field application would cost about $3000-$4000 per acre.
5. Ionic mercury introduced into a normal aerobic sediment to a con-
centration of about 200 ppm or less is rapidly complexed with sediment
entities or converted to mercuric sulfide. The complexed form might
give rise to a slow release of metallic mercury.
6. Significant quantities of metallic mercury might be released if
relatively small quantities of ionic mercury are introduced into an
anaerobic sediment.
7. Ascorbate might be used to identify sediments containing bio-
logically accessible mercury deposits.
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SECTION II
RECOMMENDATIONS
As a result of this research, it is recommended that:
1. A field feasibility study be conducted in an area known to release
substantial amounts of toxic mercury (e.g. , Detroit River).
2. The chemical and physical nature of the sediments be defined in
sufficient detail to allow a projection from the test area to another
contaminated area.
3. Caged bioaccumulators (e.g., carp, pike, or catfish) be used, prior
to and after the preventive sand or gravel overburdens are applied so
that their effectiveness can be determined.
4. Laboratory simulation tests be initiated to determine the long-term
(> 1 year) effectiveness of the abatement procedures.
5. The nature of mercury present in sediments be defined in much
greater detail and that the broader environmental hazard of the contam-
inated sediments be determined.
6. The fate of mercuric sulfide and the complexed mercuric ions in an
aerobic environment be determined.
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SECTION III
INTRODUCTION
Following the discovery in Sweden during the late sixties of high levels of
mercury in fish and seed-eating birds, fish from suspected waters in
Canada and the United States were examined for mercury contamination.
The findings ledto the banning of fishing at many sites (Lake St. Clair,
Ontario; Lake Erie, Detroit River) (1). Mercury levels in fish of 5 mg/
kg and more were observed—a level well in excess of the proposed
practical limit of 0. 05 mg Hg/kg for food (2).
Detailed studies by Westoo (3) showed that mercury in fish, and probably
in other biological systems as well (4), occurs predominantly as methyl-
mercury, a mercurial which was used as a fungicide in agriculture (seed
dressing) and industry (pulp preservation). Although contamination in
many cases was clearly related to the use of this fungicide, other organic
and inorganic compounds were implicated. Findings suggested that the
toxic products were formed from less toxic mercury compounds.
Over the years, industrial and other activities resulted in the release of
inorganic and organic mercurials in the aquatic environment. Here,
these compounds are easily absorbed by surfaces of organic particulate
matter, clays, silt particles, planktonic organisms, and hydrated ferric
oxides. Consequently, mercury compounds tend to accumulate in the
sediments of lakes and rivers, where the relatively harmless mercurials
can be transformed into highly toxic and soluble methylmercury or into
the harmless and insoluble mercuric sulfides (8).
Little is known concerning the biochemical and chemical processes
associated with mercury in these sediments. Intuitively, it is assumed
that mild anaerobic conditions would promote the formation of insoluble
sulfide. Strong anaerobic environments -would be undesirable because
these would promote the activity of methanogenic bacteria and, from the
work of Wood et al (5) and Jensen et al (6), this activity could enhance
methylation processes. Strong aerobic sediment conditions also would
be undesirable, because sulfides would be oxidized to sulfates, allowing
the methylation of the inorganic mercury. Methylation of mercury also
may proceed by a non-enzymatic process involving vitamin Bio(7). This
means that waters with high bacterial counts and those affected by
sewage discharges may promote these undesirable conversion processes.
Although much remains to be learned concerning the magnitude of these
processes, it appears that the release of toxic mercurials from contam-
inated sediments depends on biological processes occurring in these
sediments. In order to control the exchange of mercury bet-ween the
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sediments and the overlying water, the biological processes occurring
in the contaminated sediments must be controlled.
Clearly, in order to design efficient control methods, detailed knowledge
is needed concerning environmental conditions that enhance the formation
of mercuric sulfides and stimulate the release of toxic mercury. Much
of this knowledge is lacking, but programs are suggested (8) to test the
effectiveness of various control procedures. Proposed field methods
aim at (a) increasing the pH of the water-sediment interface, (b) intro-
ducing materials with stmng absorptive capacities, and (c) reducing
available oxygen and thereby promoting the development of hydrogen
sulfide and formation of insoluble mercuric sulfides.
The effect of an overburden of sand or gravel on the release of toxic
mercurials from contaminated sediments is examined in this report.
This overburden would reduce the availability of oxygen and thus promote
the conversion of mercury to mercuric sulfide. Natural sedimentation
could further bury the contaminated sediments reducing even more the
danger of exposure and subsequent release in the water. This abate-
ment procedure appears economically and ecologically appealing.
The sand overburden required to control the release of toxic mercury
from mercury-enriched sediments was investigated, and the effective-
ness of sand and fine and coarse gravel was compared. The cost in-
volved in a field application was estimated, and some preliminary
experiments were conducted to assess the immediate fate of inorganic
mercury added to aerobic and anaerobic sediments.
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SECTION IV
MATERIALS AND METHODS
Inorganic mercury was determined according to the flameless atomic
absorption technique (FAAS) described by Hatch and Ott (9) employing
the procedure and reagents described by Perkin-Elmer Corp. (10) and
Kolb (11), respectively. The mercury vapor was measured at 253. 7 nm. The
arrangement used is illustrated in Figure 1. To measure various con-
centration ranges, quartz-windowed absorption cells having a light path
of 1,25, Z. 50, or 20 cm were used. Figures 2, 3, and 4 show the
relationship between absorbancy and mercury concentration and the
concentration ranges for which the three cells were employed. Cali-
bration curves were prepared using a stock solution containing 0. 1354 g
of mercuric chloride (HgC^) dissolved in IN H2SO4- To prepare
standard curves, the stock solution was diluted to the required values.
Resolution of 0. 01 |j,g of Hg was obtained with the 20-cm cell. With 0. 1
M>g of Hg, a precision of about - 3% was obtained.
Metallic mercury vapor, present or formed in sediments was determined
with a flow-through system developed for this purpose (Figure 5). To
transport metallic mercury through the measuring cell, air or N£ was
used as a carrier gas; this gas was introduced into the sediment through
the fritted disk of medium porosity. A gas flow of 100 to 150 cc was
commonly used for 100 ml of a watery sediment containing some 30 g
dry weight of material. This allowed continuous monitoring of evolved
metallic mercury.
Organic mercury in fish or in sediment was determined, where possible,
by flameless atomic absorption spectroscopy (FAAS) after conversion
into inorganic mercury by acid digestion (Figure 1). Digestion was
carried out in 30-ml Kjeldahl flasks containing a 2-ml mixture of con-
centrated H2SO4 and HNO3, to which KMnO4 was added. Routinely,
digestion was complete after 20 to 30 min at about 100C. Using methyl-
mercuric chloride (CH3HgCl) and HgCl£ as test compounds, recoveries
of 85% to 100% were observed. Lower digestion temperatures (50C to
70C) gave incomplete recovery with CH3HgCl (12). This result was
attributed to incomplete digestion. Since methylmercury is the dominant
form of mercury reported in fish (13) and because its C-Hg bond is
relatively strong, digestion procedures developed for methylmercury
would apply to other mercurials as well. Thin layer chromatography and
two types of resin (13, 28) were used to separate inorganic mercury from
organic mercurials.
Reagents; Mercury compounds were obtained from the K and K Company,
New York. Organic mercurials contained 2 to 5% of inorganic mercury,
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PC
to aspirator
253.7nm
Figure 1. Schematic of closed system used for mercury analysis.
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900
I/)
§ 750
Z 600
O
O
CO
CO
450
300
150
Lp 20cm range 0-0.7 ftgHg
0.1 0.2 0.3 0.4 0.5
/igHg / SAMPLE
0.6 0.7
Figure 2. Calibration curve for absorption at Z53.7 nm for mercury
concentrations between 0 and 0.7 tig/sample.
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I/)
c
900
750
"-~ 600
O
Q_
Qi
O
to
CO
450
300
150
Lp 2.5cm range 0-15
I
I
_L
I
0 2.5 5.0 7.5 10.0 12.5
/igHg / SAMPLE
15.0
Figure 3. Calibration curve for absorption at 253.7 nm for mercury
concentrations between 0 and 15 p,g/sample.
10
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900
'E 750
o
Z
o
O
co
CO
600
450
300
150
Lpl.25cm range 0-25/j.gHg
10
15
/ SAMPLE
20
25
Figure 4. Calibration curve for absorption at 253.7 nm for mercury
concentrations between 0 and 25 u,g/sample.
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253.7nm
Exh
oust f
Additions
mi
I
sediment
fritted disc
PC
Figure 5. Schematic of flow-through system used to measure mercury
vapor released directly from sediments.
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as indicated by thin layer chromatography and NMR spectroscopy.
Dithizone was Baker's analyzed; diphenyl thiocarbazone was Fisher's
certified; and tetrahydroxy-p-benzoquinone was an Eastman Kodak
Company product. Bismethylmercuric sulfide, (CH3Hg)2S, was pre-
pared according to a procedure of Dadii et al (14) from methylmercuric
bromide and sodium sulfide in ethanol. Male guppies were used as
accumulators of organic mercurials. As will be seen later, guppies
almost quantitatively removed organic mercurials of interest from
the water, but they did not accumulate inorganic mercury compounds,
even when such compounds were present at relatively high concentration.
Mercury observed in fish is assimilated as organic compounds.
Column preparation; Columns were prepared following the procedure
described by Jernelov (15) and as shown in Figure 6. For all experiments
reported here, the mercury-rich sediment -was prepared by adding 16 mg
Hg (as HgClz) to 160 ml of watery sediment (the amount of Hg-enriched
sediment used in each cylinder). This Hg-enriched sediment was spread
on a 10-cm sediment, relatively free of mercury,in a 10-cm diameter
column. The enriched layer contained about 200 M-g of Hg/g of dry sed-
iment and was about 1 cm in thickness. Subsequently, sand or coarse
or fine gravel was spread over the enriched layers to the required
thickness, and tap water was added for a total volume of about 4 liters.
The depth of water layer was approximately 12 cm. Six guppies were
added to the control and to each experimental series. The columns
were placed in an environmental chamber, which was kept at 20C and
subjected to 12-hr light/dark intervals. Illumination was provided by
fluorescent and incandescent light sources. The light intensity was
approximately 0.3 mW cm~2 (400 to 700 nm). The sand used for these
experiments was obtained from a local distributor and fractionated by
a standard sieve procedure before use. The -30 +70 fraction (particle
sizes between 210 and 595 microns) wasusedfor all experiments. The
fine gravel ranged from 3/16 to 1/4 inch in particle size. The size
distribution was between 3/4 and 1 inch for the coarse gravel.
Sediments: The sediments used for the experiments described in this
report were obtained from the fresh water section of the Patapsco River
in Maryland. Two sediment "types" were used. One sediment (collected
near Baltimore, Maryland, and referred to as Sediment I) was relatively
rich in sulfur as determined according to the methods of Vogel (16) and
Carius (17). Both methods suggested a sulfur content of about 1% in dry
sediment. The cation exchange capacity (CEC) of this sediment was about
10 milli-equivalents (me) per 100 g of sediment as determined by the
method of Toth and Ott (18). The organic content of Sediment I was
approximately 7%. This was determined by weight loss at 500C for
2 hr. The difference in weight before and after ignition was taken as
13
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I
water
Hg-rich layer
sediment
I
HI
sand
fine
gravel
coarse
gravel
Figure 6. Schematic of laboratory incubation procedure.
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the organic fraction and expressed as a percentage of the total weight.
Similar measurements were made on Sediment II, obtained from the
Patapsco River near Frederick, Maryland. The organic content of
Sediment II was about 10%, and the CEC was 19 me per 100 g of dry
sediment.
Organic and inorganic Hg in sediment: To determine inorganic Hg in
sediment, an aliquot of known weight (~ 1 g) •was mixed with 2 ml of con-
centrated acid (H2SC>4 + HNC>3) with KMnC>4 added and kept at room
temperature for 30 min. Subsequently, the sample was made up to 5 ml,
centrifuged at 5000 g, and the Hg content of the supernatant was analyzed.
To determine the total Hg content of a sediment, a similar procedure
was followed, except that the mixture was digested for 20 min at 100C
(19, 20). The organic content was taken as the difference between the
total Hg content of the sediment and the inorganic fraction.
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SECTION V
RESULTS AND DISCUSSION
Retention of mercurials by sediments; It is to be expected that both
HgCl^ and CH-jHgCl will, to a given extent, form complexes with sed-
iment entities. The extent to which this occurs depends on the affinity
of both salts for binding sites. We assumed that the binding ability of a
sediment is reflected by the cation exchange capacity and organic content
(21).
An assessment of the ability of sediments to retain inorganic mercury
was made by mixing 1 mg of Hg as HgCl2 with a sediment slurry con-
taining 10 g (dry weight) of Sediment I in about 40 ml volume. After a
2-hr incubation period at room temperature, the liquid fraction was
removed by vacuum filtration (Whatman 41), and the Hg content of the
leachate was determined. Subsequently, the solid fraction was mixed
with diluted HC1, to a final concentration of about IN. After 2 hr of
incubation, the leachate was removed, and the mercury content was
determined in the acid leachate. The mercury content of the solid
fraction also was measured. The observed results, recorded in Table 1,
show that a large majority of the added inorganic mercury was retained
by the solid sediment fraction.
Table 1
Distribution of Hg between solid and liquid phase in 10 g (dry weight) of
sediment mixed with 1000 M-g of Hg as HgCl2-
Fraction M*gHg % Total
Liquid (H2O) < 10 < 1
Liquid (IN HC1) < 50 < 5
Solid 800 90
% Recovery 90-95
Similar measurements were made with CH^HgCl. The results, recorded
in Table 2, show that methylmercury is not as readily absorbed as in-
organic mercury. For example, whereas water and acid extraction re-
moved only insignificant quantities of inorganic mercury from the
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absorbing complex, about two-thirds of the organic mercury was re-
moved by this procedure. The non-ionic nature of methylmercury
might be responsible for the difference in behavior. Thus, once formed
in the sediment layers, methylmercury could be released to the over-
lying water.
Table 2
Distribution of Hg between solid and liquid phase in 10 g (dry weight) of
sediment mixed with 200 M-g of Hg as CH^HgCl.
Fraction M*gHg
Liquid (H2O ) 58
Liquid (HC1) 80
Solid 50
% Recovery
% Total
29
40
25
94
Retention of inorganic mercury in the sediment also could occur through its
conversion to a sulfide (HgS). In contrast, the release of methylmercury
would be enhanced by complexing with sulfur. As will be discussed
later, the formation of bismethylmercuric sulfide, (CH3Hg)2S, anorganic
mercurial which is even more toxic than methylmercury, is likely to
occur under such conditions. Because of its non-polar nature, the
release.of this compound would occur more readily than methylmercury.
A third organic mercurial that is readily released from sediments is
the volatile dimethylmercury (Cr^HgCF^), which, according to Wood
and co-workers (5), is formed under alkaline conditions.
The nearly complete absorption of Hg to the solid fraction is not
surprising, considering the affinity of the divalent ions for absorption
sites and the fact that the amounts used are small (about 0. 112 me/100 g
sediment) relative to the cation exchange capacity of this sediment
(10 me/100 g).
Apparently, methylmercury is less readily absorbed on the solid sed-
iment fraction. The relatively low binding capacity of methylmercury is
indicative of the monovalent bond and the non-ionic nature of the salt.
Our results indicate that the interstitial liquid and the solid fraction of
the sediment should be sampled to determine the extent of the methylation
process.
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Uptake of organic and inorganic mercury by guppies; To determine
whether the mercury present in guppies is of an organic or inorganic
origin, measurements were made of the extent and the rate of Hg up-
take by male guppies exposed to known quantities of either organic or
inorganic mercury.
In these experiments, usually 8 to 10 guppies -were incubated at about
ZOC in 100 ml aged aquarium water spiked with about 25 [j,g Hg as
CH3HgCl, (CH3Hg)2S, or HgClz- At given time intervals, guppies were
removed from the incubator and analyzed for Hg content by the procedures
described earlier. Results of a typical series are presented in Figure 7.
A rapid increase in the Hg content was observed in guppies exposed to
methylmercury and bismethylmercuric sulfide. Guppies exposed to the
inorganic mercury also showed increased Hg content, but the total
amounts retained were much less and did not increase significantly
with time.
At initial mercury concentrations of 25p,g of methylmercury per 100 ml
of water, about 1 p,g of Hg accumulated per guppy (average weight was
100 mg). Occasionally, values of 2 to 3 |j,g per fish were observed. At
the lower values, most guppies survived.
Mercury concentrations of 50 ng of methylmercury per 100 ml volume
also were tested; but, at these levels, most of the animals died within
a 2-hr exposure.
To test whether guppies quantitatively remove methylmercury from the
medium, 7 guppies were exposed to 16. 5 |j,g Hg (as CH^HgCl) in 100 ml
water. After 77 hr the animals and the medium were analyzed for Hg,
and only 0. 1 p,g Hg was found in the medium, and a total of 12. 5 (ig of Hg
was found in the animals, indicating a recovery of about 75%.
The concentration of methylated mercurials used in the above experiments
is high relative to values most likely to be encountered in contaminated
environments. A more realistic approach would use less than one-
hundredth of that concentration and exposure times of weeks or months.
To simulate a natural situation more realistically and to obtain some
estimate of the lower threshold level at which guppies accumulate
organic mercurials, the rate of accumulation was determined as a
function of mercury concentration. To this end, guppies -were exposed
to concentrations of methylmercury varying between 3 |j,g to 15 |j,g per
liter. After 36 hr of exposure, the animals were sacrificed and
analyzed. The results, illustrated in Figure 8, show a near-linear
relationship between mercury accumulation by the guppies and the
mercury content of the liquid.
19
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20
15
E
o
O)
I
10
0
0
30 60 90 120
TIME (min)
Figure 7. Time course of mercury assimilation by guppies.
ZO
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12.O
9.6
»-•—
^o
O)
O)
7.2
4.8
2.4
0
0
3 6 9 12
CH3HgCI (/ig/liter)
15
Figure 8. Methylmercury uptake by guppies as a function of concentration.
21
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These findings suggest that guppies will tolerate prolonged exposure
to relatively low concentrations of methylmercury and that they are
capable of concentrating organic mercurials even if present at reason-
ably low concentrations. At the lower limit used in the above experiments,
a concentration factor of about 1000 was calculated, estimating a guppy's
volume at 0. 3 ml. On the basis of these results, the investigators are
confident that the methylmercury formed in the laboratory simulations
(described in the following sections) is accumulated by the fish.
Effect of sand on the release of organic mercury: The objective of this
series of experiments was to determine the extent of organic and inorganic
mercury release from mercury-enriched sediments covered by layers of
sand of varying thickness. The procedure is based on Swedish observations
(22) that mercury-rich layers become inactive rapidly when covered with
2 to 10 cm of mercury-poor sediment. To evaluate the efficacy of this
"burying procedure, " a laboratory simulation was carried out using sed-
iments enriched with mercuric chloride (200 M-g of Hg/g of dry sediment)
and covered with layers of sand 0, 1.5, 3, and 6 cm in thickness. (The
stratification is illustrated in Figure 6.) Incubators were supplied with
at least 6 guppies each, and, after a 4-week incubation, the mercury
content was determined in the fish, in the overlying water, in the enriched
sediment layers, in the sand.and in the sediments beneath the enriched
layers,. It was assumed that this type of analysis would provide some
indication of the extent of the mercury's vertical transport through the
sediment and sand layers.
The data on mercury accumulation by guppies are recorded in Tables 3
and 4. On the basis of the observed data, it is apparent that the for-
mation of methylmercury is low in those incubators supplied with a sand
cover of 6 cm over the mercury-enriched layer. Although the mercury
accumxilation by fish varied widely with sand covers of 1. 5 and 3 cm, it
appears that this overburden is only partially effective in controlling
the release of methylmercury.
Experiments reported in Tables 3 and 4 were conducted with Sediment I
and Sediment II differing in organic content (7% vs 10%) and cation ex-
change capacity (10 me vs 19 me). Considering prevailing experimental
conditions and the obtained results, it appears that these sediment
parameters have little effect on the release of toxic mercurials.
In all incubators with 1. 5 cm of sand, and in a few with 3 cm, a build-up
of black material on top of the sand layers was observed. This material
•was due to the activity of sludge worms (Tubificidae) present in large
quantities of Sediment I (obtained 2 miles below a sewage outfall). The
mercury content of the worms' rejects was measured; as expected, this
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content was quite similar to that in the mercury-enriched layer of the
experimental series. In the control series, no mercury was detected.
Although a similar build-up •was not observed in incubators with 6 cm
of sand, the results suggest that in areas where the sludge fauna carries
on its function actively, additional coverage might be required. Besides
the conveyance of mercury by sludge worms, vertical transfer apparently
•was limited, since only small amounts of mercury were detected in the
sand and sediment above and below the enriched layer. (See Table 5. )
Sediment II, collected far from sewage outfalls, did not exhibit this
sludge worm activity. Therefore, a shallower sand overburden would
control the release from a sediment typified by Sediment II.
Table 3
Effect of sand overburden on the mercury content (|j,g) of surviving
guppies incubated for 3 weeks using Sediment I enriched with
Sand Overburden
Additions 0 cm 1. 5 cm 3 cm 6 cm
f n 0.3 0.2 0.2 0.2
(Control)
HgCl2 No 2.5 2.5 0.5
Enriched survivors
HgCl2 3.5 3.3 1.0 0.4
Enriched
Note: Due to animal mortality, only a few fish
were available for analyses at the end of
the 3-week incubation. Results reported
here are based on the analysis of 2 fish
and are expressed as (j,g Hg/g fish.
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Table 4
Effect of sand overburden on the mercury content (M-g) of surviving
guppies incubated for 4 weeks using Sediment II enriched with
Additions
0 cm
Sand Overburden
1. 5 cm 3 cm
6 cm
None
(Control)
HgCl2
Enriched
HgCl2
Enriched
0. 1
0. 2
5. 0
2. 1
4. 3
4.9
0. 1
-
3.5
2.5
3.3
3.8
0. 1
0. 1
5.6
2. 1
0. 3
1.8
0. 1
-
2.6
0.4
0.3
0.8
0. 1
0.2
1. 0
0.5
0. 3
0.2
0. 1
-
0.8
0.2
0.3
0. 1
0.2
0.2
0.3
0. 1
0.6
0.2
0.2
-
0.3
0.2
0.4
0. 1
Note: Each value indicates an individual animal and is
expressed in M-g of Hg/g of fish.
Table 5
Mercury content of sand and sediment above and below Hg-enriched
layer.
Sand
0 cm 1. 5 cm 3 cm 6 cm
Sand
(M.g Hg/g sand)
Sediment
(5 cm beneath
enriched layer)
(M-g Hg/g sediment)
0.4
0. 5
0.7
0.2
0. 1
1. 5
Note: For details, see Table 3.
24
-------
The pH of the liquid phase also was measured and analyzed for mercury
content. Although the initial pH value of the -water in all incubators was
slightly acidic (6. 7 to 6. 9), alkaline values were observed in all incubators
at the end of the 3-week incubation. Considering the algal growth observed
in most incubators, the rise in pH might be due to a significant decline in
the carbon dioxide content of the water.
As recorded in Table 6, algal concentrations of incubators supplied with
0 and 1. 5 cm of sand were relatively high in comparison to values
observed with 3 and 6 cm of sand. However, these values might be mis-
leading since, in the latter cases, algal growth along the walls was
significant, and representative sampling -was virtually impossible.
Although no attempts were made to determine the specific distribution,
it appeared from the color that blue-green algae were predominant in in-
cubators with 3 and 6 cm; green varieties appeared to dominate in in-
cubators supplied with 0 and 1. 5 cm of sand. It is not known whether
sand constituents or differences in the rate of nutrient release from
the sediment is the principal cause of the phenomenon.
Table 6
Mercury content of the liquid phase at the end of a 3-week incubation
using Sediment I.
0 cm
Sand
1. 5 cm 3 cm
6 cm
Total Hg
Total Hg
Filtrate
Suspended
algae
(mg/liter )
.4
13
10
.4
23
4
. 5
.4
Note; 0.45 micron (Millipore) filter was used.
The algal content was derived from chlorophyll
estimation. To determine this content, an 80%
acetone extract was prepared and read at 663 nm.
To convert chlorophyll concentration to algal con-
centration, a content of 2% dry weight was assumed.
25
-------
Significant quantities of mercury were observed in the water (see
Table 6). This "suspended" mercury apparently was present on or
in the algal cells, since it could be removed almost completely from
the liquid phase by filtration over 0.45 micron filters (Millipore). As
shown in Table 6, at least 90% of the mercury remained on the filter.
This fraction was found to be inorganic in nature.
Comparison of sand and fine and coarse gravel; A layer of sand 6 cm
in thickness controlled the release of organic mercury to the overlying
water. To examine tne usefulness of other aggregates, the effectiveness
of fine gravel (5 to 6 mm) and coarse gravel (18 to 25 mm) was compared
with sand. (This type of coarse gravel is used routinely in the preparation
of "popcorn" concrete. Popcorn concrete was not tested because it was
felt that the addition of cement would significantly increase the pH and
thus invalidate the comparison with other aggregates. ) In this comparison,
6-cm layers of the two grades of gravel •were used instead of sand (see
Figure 6). This series employed two experimental (+Hg) and one control
(-Hg) incubator. Incubation time was 3 -weeks.
The results, recorded in Table 7, revealed that coarse and fine gravel
are as effective as sand in controlling the release of toxic mercury. In
laboratory experiments, the use of gravel required less precaution than
for sand, but it remains to be seen whether this also holds true for
field application.
Table 7
Effect of a 6-cm overburden of sand or gravel on the mercury content
(p,g) °f surviving guppies incubated for 3 weeks in the presence of
Sediment II enriched with
Aggregate Sand Fine gravel Coarse gravel
Size distribution (mm) 0. 2 to 0. 6 5 to 6 19 to 25
No Hg added 0. 1 0. 2 0. 2 0. 3 0. 2 0. 3
(Control)
0.2 0.2 0.3 0.2 0.3 0.3
HgC 2, 0.2 0.2 0.2 0.2 0.3 0.3
Enriched 0.2 0.2 0.2 0.2 0.3 0.3
Note: Content is based on values of individual
animals and expressed in y,g Hg/g fish.
26
-------
To demonstrate and summarize the effect of a 6-cm sand or gravel
overburden on mercury abatement, all data observed in the presence
of a 6-cm cover (irrespective of sediment type and the nature of the
cover material) and in the absence of a cover were pooled separately
and compared to the pooled control values. Results, illustrated in
Figure 9, show that a 6-cm cover surpresses the release of toxic
mercurials from contaminated sediments. Comparisons of the cal-
culated means and the standard deviations indicate the soundness of the
observed data.
Effect of anaerobioses on release of toxic mercury; Wood and co-workers
(5) reported that anaerobically incubated extracts of methanogenic bacteria
stimulate the conversion of inorganic mercury into methylmercury. If
a similar conversion occurs in sediments, one would expect tne production
of methylmercury to be relatively high under strictly anaerobic conditions.
This aspect -was examined by comparing methylmercury formation in an
anaerobic sediment (argon equilibrated) to an aerobic sediment (air
equilibrated). Containers similar in design to those used for aerobic
experiments were fitted with male and female ground joints and used as
anaerobic incubators.
After a one-month incubation, the liquid phases were removed from both
series, filtered over Whatman #40 filter paper, adjusted to pH 6. 5 with
HC1, and flushed with air to remove volatile gases (H2S, CH4) and to
raise the level of dissolved oxygen. Subsequently, eight guppies were
added to the liquid, and the mercury content of the guppies was deter-
mined after two and five days of exposure. To obtain a rough estimate of
the residual organic mercury in the sediment, part of the enriched layer
was removed from the incubators and extracted with IN HC1. It was
assumed that acid extraction would remove the loosely bound toxic
mercurials. After neutralization with NaOH, guppies were added to
the extract, incubated for a day, and, subsequently, analyzed as usual.
The results, recorded in Table 8, showed that the concentration of toxic
mercury (i. e. , guppy assimilable) is higher in the anaerobic incubator
than in the aerobic incubator. Considering the residual mercury content
of the sediment (as analyzed in the IN HC1 extract), it appears that the
toxic mercury is released to the liquid phase under anaerobic conditions,
but it is retained, at least in part, under aerobic conditions.
The investigators cannot offer a reasonable explanation for the observed
difference. Conceivably, the organic fraction constitutes a significant
part of the ion exchange capacity of this sediment, and anaerobiosis
could have adversely affected its capability to retain ions. From these
qualitative results one would conclude that aerobic and anaerobic
27
-------
4.0
3.0
O)
w 2.0
I
O)
1.0
.15 ±.01 (erg)
j*v
( DATA FROM
TABLES 3, 4 AND 7)
CONTROL
-Hg
NO COVER
+Hg
.25±.02(o-x)
6cm COVER
+Hg
Figure 9. Effect of protective cover 6 cm in depth on mercury
accumulation by guppies.
Z8
-------
conditions stimulate the formation of toxic mercury to about the same
extent, but that strong anaerobic conditions enhance the release of toxic
mercury from the sediment.
Table 8
Estimate of mercury content (|j,g) in the liquid phases and acid extracts
of mercury-enriched sediments incubated aerobically or anaerobically.
Aerobically Anaerobically
incubated incubated
Wa tor
, 0.2 to 0.3 5 to 6
phase
Acid 1.4 to 1.8 0.5 to 0.6
extract
Note: Values represent the content of mercury per
incubator (4 liters) and are rough estimates
only.
In a parallel experiment, mass spectrometric analysis was used to deter-
mine the formation of methane in the anaerobic incubator. A gradual
increase in the methane concentration was observed. The rate of in-
crease appeared to follow first-order kinetics, indicative of a logarithmic
cell reproduction. At the end of a one-month incubation, about 900 ^mole
CH^ was formed per 100 g sediment. In the same period and calculated
on an equimolar basis, approximately 0. 03 ^mole of methylmercury was
formed. Apparently, the rate of methylmercury formation is small
relative to the rate of methane formation, suggesting that both products
are formed independently.
Effect of redox state on release of mercury; Results discussed thus far
are concerned with transformations of mercuric chloride into organic
mercurials. From a toxicological point of view, the latter are of
principal concern. Transformations leading to their formation are
depicted in Figure 10.
In the general scheme, other inter conversions also must be considered,
both in connection with abatement procedures and the translocation of
mercury. Although little is known about the importance of reactions
leading to the formation of mercury complexes in a sediment, intuitively
one would assume that inorganic and humic entities of the sediment play
29
-------
INORGANIC
OR HUMIC
SEDIMENT
COMPLEXES
RCH2Hg
Hg
Ca
2K
^ Hg dimethylmercury
CI-LHg monomethylmercury
(CH,HgL S bismethylmercuric
* L sulfide
Figure 10. Transformation pathways of mercury and its derivatives in
aquatic environment.
30
-------
an important role in stabilizing mercury contaminated sediments. This
aspect was examined previously (see Tables 1 and 2). The formation
of the relatively insoluble mercuric sulfide (HgS) must be considered
as another stabilizing factor. Its formation and its ultimate fate is
determined, to a large extent, by the redox state of the sediment.
The measured redox state of sediments is largely a function of the micro-
bial activity, the presence of oxygen, and the availability of degradable
organic substances. Reducing conditions usually result if aeration is
relatively low, and these conditions promote the formation of insoluble
sulfides (25). The extent to which this occurs depends upon the avail-
ability of sulfur. If sulfur is limited and mercury is available in an
ionic form (Hg++), reduction to metallic mercury (Hg°) will occur.
Since the redox state of a sediment can be assessed by using simple
sensing probes, its value assumes significance principally because it
can be used in the field as a convenient indicator of the hazard posed
by a mercury-laden sediment. Therefore, the investigators assessed
the effect of the redox state on the release of metallic mercury by an
aerobic, i. e. , high-redox potential, sediment and by an anaerobic, or
low-redox potential, sediment.
The high-redox potential sediment was obtained by drying material in
air overnight at room temperature. Subsequently, 30 g of this material
was added to 100 ml distilled water, mixed thoroughly, and transferred
to the sediment chamber illustrated in Figure 5. The Eh of the pre-
pared sediment was about 450 mV (Orion 96-78-00). Various amounts
of HgCl2 were added through the entrance port (see Figure 5), and the
release of Hg° was measured as a function of time. The results are
illustrated in Figure 11.
The amount of metallic mercury evolved from an aerobic sediment was
dependent on the relative amount of mercuric chloride added to it. The
release pattern observed upon addition of different amounts of HgCl2 to
a constant amount of sediment (30 g), presented in Figure 11, also was
observed when the sediment was increased and the mercury content of
the liquid was held constant.
Figure 11 also shows that the rate of Hg° evolution increased gradually
upon addition of HgCl2 until, about an hour later, a steady state value
was attained. To determine the rate of Hg° formation, the efflux was
scrubbed by the solution of nitric acid and sulfuric acid, and the
mercury content of the solution was determined by the usual procedure.
It was found that approximately 1 ^g of Hg° was released per min by
30 g of sediment enriched with 100 mg of HgCl2 and flushed with about
100 cc of air per min.
31
-------
320 -
30g (dry weight)
Aerobic Sediment
Suspended in
100ml distilled
water
(20mg HgCI2
(lOmg HgCl2)
10 20 30 40
TIME (min)
60
600
Figure 11. Rate of Hg evolution from aerobic sediments spiked with
HgCl2.
32
-------
The release of metallic mercury from anaerobic sediments appeared to
follow different kinetics. For example, the addition of 50 p-g of HgCl-,
to an anaerobic sediment (Eh = ~ 50 mV) resulted in an immediate re-
lease of substantial amounts of metallic mercury (Figure 12, curve 1).
Subsequent addition of 50 pig each increased the rate of Hg° evolution
until, after 4 additions, the rate of release became constant at about
1 p,g of Hg° per min (Figure 12, curve 4).
A comparison between the two sediments reveals other differences as
well. For example, whereas Hg° evolution was at its maximum at
200 fig of HgCl2 per 30 g of anaerobic sediment, about 500 times as
much HgCl2 was required for the production of a similar amount of Hg°
from the aerobic sediment. Although no effect was observed if HgCl2
was added to an aerobic sediment, a 200 ug dose of HgCl2 added to an
anaerobic sediment produced a Hg° gush of a magnitude that was be-
yond the measuring range of the apparatus.
From the above, it appears that at least two processes control the re-
lease of metallic mercury in sediments. The slow process, as seen
with the aerobic sediment, could be a reflection of biological activity.
Tonomura et al (26, 27) reported the "bio-conversion" of organic mercury
into metallic mercury carried out by mercury-resistant microorganisms.
However, it is not clear whether this conversion was active or passive
in nature. In sediments, a physico-chemical process could be responsible
for the conversion of Hg++ into metallic mercury. A slow release of
electrons, due to a biological process, could lead to the reduction to
Hg°. As will be discussed later, once the divalent ion is complexed
•with a sediment entity or is converted into mercuric sulfide, it is much
less likely to be converted into the metallic form.
The fast process observed with anaerobic sediments undoubtedly is due
to a chemical reduction of HgCl2 according to Hg"1"1" + 2e—^Hg°. In an
anaerobic sediment, a pool of various reduced entities could provide
electrons for such a reduction. This point was investigated by using
ascorbate as an electron donor •with (a) aerobic sediment, (b) anaerobic
sediment, and (c) water. In all three cases, metallic mercury was pro-
duced from HgCl2 if ascorbate also was present.
In effect, a stochiometric relationship was observed between the mercury
gush and ascorbate addition il HgCl2 was added immediately after the
addition of ascorbate. However, if HgCl2 was thoroughly mixed with
the aerobic sediment and ascorbate added later, no mercury gush
occurred,and normal release kinetics were observed. It made no
difference whether aerobic sediment or an artificial resin was used.
The latter (Chelex 100) retained Hg~l~+ (28) as firmly as the sediment, and
33
-------
240
0)
±. 160
c
o
3
o
0)
o
O)
X
80
30g (dry weight)
Anaerobic Sediment
Suspended in 100ml
distilled water
468
TIME (min)
10
12
14
Figure 12. Rate of Hg° evolution from anaerobic sediments spiked
with 50 ug of HgCl2.
34
-------
ascorbate was unable to reduce the divalent ion and release it in the
metallic form.
These observations strongly suggest that once the divalent ion is con-
verted into the sulfide form or complexed with a sediment entity
(probably most of them are organic in nature), the ion no longer is
readily reduced by natural electron donors or artificial ones with a re-
ducing potential equal to that of ascorbate. One is tempted to draw a
similar conclusion for enzymatic and non-enzymatic reductions.
UBBARX ua. ifJ»
-------
SECTION VI
ECONOMIC CONSIDERATIONS
In the laboratory simulation program, the use of sand and gravel was
tested as a means of preventing the release of toxic mercury from
mercury-enriched bottom sediments. This laboratory simulation led
to the conclusion that common sand and fine or coarse gravel, applied
to a thickness of about 2 to 3 inches, can virtually eliminate the release
of toxic mercurials from contaminated sediments. In this section, the
approximate cost of applying this abatement procedure in the field is
evaluated.
Trenton Channel of the Detroit River near Wyandotte was selected as a
"representative" location. The channel varies from approximately 1/8
mile to 1/4 mile in •width. The landside depth is about 13 ft, the channel
depth is 28 ft, and the depth on the island side is 4 ft. The area of
interest is a 200-ft-wide and one-mile-long band on the Michigan side
immediately below Wyandotte Chemical Company. This strip is known
to be heavily contaminated with mercury (1).
These economic projections assume the use of the cheapest grade of sand
selling at about $. 95 a yard (approximately 1. 5 ton). (Gravel is not con-
sidered separately since the cost of application appears very similar to
that of sand. ) Considering typical trucking, loading, and unloading
charges, the price increases to $2. 25 a yard in a scow at dockside.
Since the overland transportation costs are significant, the possibility
of a waterborne source of supply--possibly from a nearby dredging
operation—was explored. The nearest large-scale sand mining oper-
ation is in Lake Erie and, although materials of various grades are
shipped to the Detroit area, this source appeared not to be competitive
with pit sources in Michigan--at least in the cheap, low-grade, sand
market.
A number of application methods were considered, but a barge-mounted
system appeared to be the most adaptable and economical. This system
consists of a fixed-boom clam shell, feeding a conveyor belt that
delivers the material to a swivel-type sand piler. A clam shell -with a
2-1/2 to 3-yard bucket can load a conveyor at the rate of about 200 tons
per hour. The swivel piler would be a 16-inch unit, handling about
200 tons per hour, allowing maximum use of the piler. A swivel piler
can spread sand or other small-lump aggregate material over a 270°
arc. Depending upon the speed of the barge, swivel motion, and sedi-
mentation characteristics at the locale, a fairly even layer of material
could be applied over the mercury-laden sediments.
37
-------
The distribution equipment would be assembled on a small equipment
barge at Wayndotte, towed to the site, and anchored there. A tug would
deliver sand-filled 800- to 1000-yard-capacity scows alongside the
barge-mounted equipment as a supply source. Ideally, the tug would
tow enough material to the site during the early morning to allow
continual operation of the spreading machinery during the daylight hours.
Table 9 presents estimates of the fixed and variable costs associated
with this hypothetical case. The following assumptions were made:
(a) the site is near enough to port for the tug to return empty sand scows
at night and to deliver loaded ones by early morning, (b) weather conditions
are ideal, (c) traffic at the site is negligible, and (d) no machinery down-
time is experienced.
Table 9
Estimated fixed/variable costs of distributing sand in an area south of
Wyandotte.
Fixed Costs:
Spreading Equipment System $20, 000. 00
(i. e. , swivel piler, conveyor, clam
shell, fixtures, hopper, etc. )
Variable Costs;
Sand, dockside, per yard 2.25
Tug boat and crew, per 12-hour day 1, 900. 00
Deck scow, 800-to 1000-yard capacity, per day 100.00
Equipment barge, per day 30. 00
Labor, per day (2) 80.00
Equipment maintenance, per day 10. 00
Based on these considerations, the cost of applying 3 inches of sand to
2-, 25™, and 50-acre areas was estimated. The results, shown in
Table 10, suggest that the overall cost of application •would increase
linearly as a function of the size of the area treated.
38
-------
Table 10
Estimate of the cost involved in the application of 3 inches of sand to
2, 25, and 50 acres of sediment contaminated with mercury.
Number of Acres
2 25*50
Fixed Costs ($):
Variable Cost($):
20, 000
20,000 20,000
Sand
Tug rental
Scow rental
Barge
Labor
Maintenance
S/Total
Number of days
Yards of sand
1, 670
1, 900
100
30
80
10
3,790
1
740
20, 800
24, 700
1, 300
390
1, 040
130
48, 360
13
9,250
41,600
47, 500
2, 500
750
2, 000
250
94,600
25
18,500
* An area of this size is used as an example in the discussion.
The preliminary nature of this cost evaluation is recognized. Such
factors as the local conditions of transportation, the actual sediment
characteristics, the accessibility and topography of the site, water
currents and depth, prevailing weather conditions, and the availability
of labor, materials,and hardware have an impact on the cost. Such
factors, which are site-dependent, could not be fully evaluated at
this time.
39
-------
SECTION VII
ACKNOWLEDGMENTS
The authors wish to thank Dr. Ed H. Parkison and Messrs. Donald R.
Talbot, Werner F. Furth, Otto J. Ollinger, and Thomas J. Quinn of
RIAS for their contributions to the work presented in this report. The
many discussions of the complexities associated with the mercury abate-
ment problem during the course of this work with Drs. George M. Cheniae,
Bessel Kok.and Kenneth L. Zankel, also of RIAS, have been very helpful.
The discussions with Dr. Curtis C. Harlin, Jr. , Project Officer, Robert S.
Kerr Water Research Center, EPA, Ada, Oklahoma, have been partic-
ularly stimulating.
41
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SECTION VIII
REFERENCES
1. Turney, G.W., "Mercury Pollution: Michigan's Action Program, "
J. Water Pollution Control Federation, 43, 7, pp 1427-1439 (1971).
2. Ackefores, H. , "Effects of Particular Pollutants; Mercury Pollution
in Sweden with Special Reference to Conditions in the Water Habitat, "
Proc. Roy. Soc. London B, 177, pp 365-387 (1971).
3. Westoo, G. , "Methylmercury Compounds in Animal Foods, " Chemical
Fallout, Charles C. Thomas, Springfield, 111., pp 75-93 (1969).
4. Berglund, F. , and Berlin, M. , "Risk of Methylmercury Cumulation
in Man and Mammals, " Chemical Fallout, pp 258-273 (1969).
5. Wood, J. M. , Kennedy, F.S., and Rosen, C.G. , "Synthesis of
Methylmercury Compounds by Extracts of a Methanogenic Bacterium, "
Nature, (London) 220, pp 173, 174 (1968).
6. Jensen, S. , and Jernelov, A. , "Biological Methylation of Mercury
in Aquatic Organisms," Nature, (London), 223, pp 753, 754 (1969).
7. Bertilsson, L. , and Neujahr, H. Y. , "Methylation of Mercury Com-
pounds by Methylcobalamin, " Biochemistry, 10, 14, pp 2805-2808
(1971).
8. Nelson, N. , et al, "Hazards of Mercury, " Environmental Research,
4, pp 29-31 (1970).
9. Hatch, R. , and Ott, W. L. , "Determination of Sub-Microgram
Quantities of Mercury by Atomic Absorption Spectrophotometry, "
Anal. Chem. 40, pp 2085-2087 (1968).
10. Perkin-Elmer Corporation, "Mercury Analysis System, " Operating
Directions 303-2119, Norwalk, Connecticut, March (1971).
11. Kalb, G. W. , "The Determination of Mercury in Water and Sediment
Samples by Flameless Atomic Absorption, " Atomic Absorption
Newsletter, 9, 4 pp 84-87 (1970).
12. Uthe, J. F. , Armstrong, F. A. J. , and Stainton, M. P. , "Mercury
Determination in Fish Samples by Wet Digestion and Flameless
Atomic Absorption Spectrophotometry, " J. Fisheries Res. Board
Can., 27, pp 805 (1970).
43
-------
13. Westoo, G. , "Determination of Methylmercury Compounds in Food
Stuffs, " Acta. Chem. Scand. , 20, pp 2131-2137 (1966).
14. Dadii, M. , Grdenic, D. , "Preparation of Bismethylmercury Sulfide, "
Croat. Chem. Acta. , 32, pp 39 (1969).
15. Jernelov, A. , "Release of Methylmercury from Sediments with Layers
Containing Inorganic Mercury at Different Depths," Limol. Oceanog. ,
15, pp 958-960 (1970).
16. Vogel, A. I. , "Determination of Sulfur, " Quantitative Inorganic
Analysis, Wiley and Sons, N. Y. , pp 408-721 (1961).
17. Steyermark, A. , "Determination of Sulfur, " Quantitative Organic
Micro Analysis, Acad. Press. N. Y. , pp 277-290 (1968).
18. Toth, S. J. , and Ott, A. N. , "Characterization of Bottom Sediments :
Cation Exchange Capacity and Exchangeable Cation Status, " Environ-
mental Sci. and Technology, 4, pp 945-939 (1970).
19. Gorgia, A. , and Monnier, D. , "Determination of Mercury (II) in the
Presence of Organic Mercurials, " Anal. Chem. Acta. , 54, pp 505-
570 (1971).
20. Magos, I. , "Selective Atomic-Absorption Determination of Inorganic
Mercury and Methylmercury in Undigested Biological Samples, "
Analyst. 96, pp 847-858 (1971).
21. Hugget, R. J. , Bender, M. E. , and Slone, H. D. , "Mercury in
Sediments from Three Virginia Estuaries, " Chesapeake Science,
III, pp 280-282 (1971).
22. Jernelov, A. , "Conversion of Mercury Compounds, " Chemical
Fallout, Charles C. Thomas, Springfield, 111., pp 68-74 (1969).
23. Tatton, J. O'G. , and Wagstaffe, P.J., "Identification and Deter -
mination of Organic Mercurial Fungicide Residues by Thin-Layer
and Gas Chromatography, " J. Chromatog. , 44, pp 284-289 (1969).
24. Chau, Y. K. , and Saitoh, H. , "Determination of Submicrogram
Quantities of Mercury in Lake Water, " Environmental Science and
Technology, 4, 10, pp 839-841 (1970).
25. Stumm, W. , and Morgan, J. J. , "Oxidation and Reduction, " Aquatic
Chemistry, Wiley-Intersc. , N. Y. , pp 300-379 (1970).
44
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26. Tonomura, K. , Maeda, K. , Futai, F., Nakagami, T. , and Yamada,
M. , "Stimulative Vaporization of Phenyl Mercuric Acetate by
Mercury-Resistant Bacteria, " Nature, 27, pp 644-646 (1968).
27. Tonomura, K. , and Kanzaki, F. , "The Reductive Decomposition
of Organic Mercurials by Cell Free Extract of Mercury Resistant
Pseudomonad, " Biochim. Biophys. Acta, 184, pp 227-229 (1969).
28. Law, S. L. , "Methylmercury and Inorganic Mercury Collection by
a Selective Chelating Resin, " Science, 174, pp 285-286 (1971).
45
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'1
Ac cess; on Number
w
5
2
Subject field
Organization _ , _ ,•,,/•
Research Institute for
& Group
SELECTED WATER RESOURCES ABSTRACTS
INPUT TRANSACTION FORM
Advanced Studies (RIAS)
Martin Marietta Corporation
1450 South Rolling Road, Baltimore, Maryland 21227
Title
Sand and Gravel Overlay for
Control of Mercury in Sediments
10
Authors)
Leonard H.
Mohammed
Bongers
N. Khattak
16
21
Project Designad'on
EPA, Project Code #16080 HVA,
r.nmtrari- #68-01-0089
Note
22
Citation
23
Descriptors (Starred First)
Mercury Abatement; Organic Mercury, Methylmercury, Release of Mercury;
Mercuric Sulfide Formation; Fate of Mercury in Sediment Environment
25
Identifiers (Starred First)
Mercury pollution control, methylated mercurials
27
Abstract
^nsiri
The release of toxic mercurials by mercury-enriched river sediments was examined in
the laboratory. Tests showed a release of 1 ^g of methylmercury per m^, per day.
Methylmercury occurred in sediments with low and with high organic content, in sed-
iments with low and high cation exchange capacity, and in aerobic and anaerobic
sediments. The release of toxic mercury could be prevented by a layer of sand, 6 cm
in thickness, applied over the mercury-enriched sediments. Layers of fine or coarse
gravel (6 cm deep) were as effective as sand. Thinner layers of sand, (1.5 and 3 cm)
were unsatisfactory. The cost of applying 3-inch layers of sand or gravel was about
$3000 to $4000 per acre. A slow release of metallic mercury occurred in aerobic
sediments. The release was much faster in anaerobic sediments. Using ascorbate as
an artificial electron donor, metallic mercury could be released at high rates from
aerobic sediments as well. Ascorbate appeared to be a useful indicator of divalent and
biologically accessible mercury. The laboratory investigations proved the soundness
of the sand blanket approach. Its practical and economic feasibility must be determined
in a combined field and laboratory analysis program.
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