United States
Environmental Protection
Agency
Office of Research and
Development
Washington DC 20460
EPA/625/R-97/004
July 1997
Technology Transfer
vvEPA Capsule Report
Aqueous Mercury Treatment
-------
EPA/625/R-97/004
July 1997
Capsule Report:
Aqueous Mercury Treatment
National Risk Management Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268
Printed on Recycled Paper
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Acknowledgements
This capsule report was prepared under contract number 68-C3-0315 by Eastern Research
Group, inc. (ERG) for the U.S. Environmental Protection Agency's (USEPA) Office of Research
and Development (ORD). Edwin Earth served as the work assignment manager and provided
technical direction. Linda Stein of ERG directed the editing and production of this report.
James Patterson, of Patterson and Associates, Inc., Chicago, IL was the primary author.
Technical reviewers of this report were:
Richard Osantowski, Radian international Corporation, Milwaukee, Wl
Thomas Sorg, National Risk Management Research Laboratory, USEPA, Cincinnati, OH
Donald Sanning, National Risk Management Research Laboratory, USEPA, Cincinnati, OH
Jerry Stober, Environmental Services Division, USEPA, Athens, GA
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Notice
This document has been reviewed in accordance with the U.S. Environmental Protection Agency's
peer and administrative review policies and approved for publication. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
in
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Contents
Page
Chapter 1
Chapter 2
Chapter 3
Chapter 4
Chapter 5
Executive Summary I -1
1.1 Purpose I -1
1.2 Summary I -1
Precipitation Treatment Processes 2-I
2.1 Sulfide Precipitation 2-I
2.2 Coagulation/co-precipitation 2-2
Adsorption Processes 3-I
3.1 Activated Carbon Adsorption 3-I
3.2 Xanthate Treatment 3-3
3.3 Other Adsorption Processes 3-4
3.3.1 BPHC Adsorption 3-5
3.3.2 MHBB Adsorption 3-5
3.3.3 Coal Fly Ash Adsorption 3-5
3.3.4 Forager Sponge Adsorption 3-5
Ion Exchange Treatment , , 4-I
Other Processes 5-I
5.1 Chemical Reduction 5-I
5.2 Membrane Separation 5-I
5.3 Emerging Technologies 5-5
5.3.1 Macrocycles Adsorption 5-5
5.3.2 Biological Detoxification 5-5
5.3.3 Membrane Extraction 5-5
IV
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Contents (Continued)
Page
Chapters Comparison of Treatment Processes , 6-1
6.1 Treatment Effectiveness 6-I
6.2 Residuals Management 6-I
6.2.1 Mercury Sludges 6-I
6.2.2 Spent Activated Carbon 6-I
6.2.3 Concentrated Brine Solutions 6-I
6.3 Economics 6-2
Chapter 7 Casestudy 7-I
7.1 Ion Exchange Removal of Mercury from Wastewater at
DOE's Savannah River Site Effluent Treatment Facility 7-I
7.1.1 TheETF 7-I
7.1.2 ETF Feed Streams 7-3
7.1.3 ETF Discharge Limits 7-3
7.1.4 Ion Exchange Role in the ETF 7-3
7.2 Effluent Treatment Facility Economics , 7-5
Chapter 8 References 8-I
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List of Tables
Table No.
Page
2-1
2-2
3-1
3-2
3-3
4-1
4-2
5-1
5-2
6-1
6-2
7-1
7-2
Sulfide Precipitation treatment for mercury 2-2
Coagulation/co-precipitation treatment results for mercury 2-3
Activated carbon mercury treatment results 3-3
Starch xanthate treatment for mercury 3-4
Freundlich isotherm parameters for mercury adsorption 3-5
Summary of mercury-selective chelate resins 4-2
Ion exchange treatment for mercury in drinking water 4-2
Performance of reduction processes for mercury treatment 5-2
Performance of membrane processes for mercury treatment 5-3
Summary of achievable effluent mercury concentrations 6-1
Operating costs (U.S. $/year 1987) of processes for mercury
removal from chlor-alkali wastewaters 6-3
Nonradioactive contaminants and discharge limits 7-4
Radioactive contaminants and discharge limits 7-4
Figure no.
List of figures
Page
2-I
3-I
3-2
5-1
5-2
5-3
5-4
6-I
7-I
7-2
Sulfide precipitation 2-I
Types of GAC column design 3-I
GAC process flowsheet 3-2
Principles of the crossflow microfilter 5-4
Filtrate polishing using microfiltration 5-4
Schematic representation of mercury extraction with an emulsion
liquid membrane 5-6
Schematic representation of mercury ion extraction with an
emulsion liquid membrane 5-6
Flow chart of unit operations utilized for technical-economical
analysis 6-2
F/H effluent treatment facility process flow sheet 7-2
ETF facility O&M cost breakdown 7-5
Vi
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Chapter 1
Executive Summary
1 .1 Purpose
This report describes established technologies and
identifies evolving methods for treating aqueous mercury.
The information provided encompasses full-, pilot- and
bench-scale treatment results as presented in the technical
literature. The report describes alternative technologies in
terms of (1) governing physical and chemical principles
(e.g., solubility, oxidation-reduction potential, volatility), (2)
key treatment parameters (e.g., speciation, pH,
precipitating agent type and dosage, or adsorbent type and
dosage), (3) pretreatment requirements, treatment
performance, advantages and disadvantages, design
considerations, and economics when available. This
information can be useful for evaluating mercury treatment
alternatives for industrial wastewater, groundwater, and soil
washing extract.
This document assumes that the reader is already well
versed with the technologies described and is using this
report to better understand each technology's applicability
for aqueous mercury removal. Thus, the report does not
provide basic descriptions of each technology; such
information can be found elsewhere in the literature. In
addition, the report does not present recommended values
for the common design parameters of technologies.
Values for such parameters as (1) contact time, (2)
volumetric loading rates, (3) dosages, (4) reaction times,
(5) breakthrough times, and (6) mixing requirements can
be determined by conducting treatability studies using the
wastewater to be treated.
1.2 Summary
A broad spectrum of mercury treatment technologies
has been described in the technical literature, ranging from
established full-scale applications to innovative approaches
investigated to date only at bench or pilotscale. The
literature, however, provides only limited information on
actual full-scale treatment technology performance and
almost no full-scale economic data or information on
mercury recovery.
Well-established and widely reported full-scale
technologies are precipitation, coagulation/co-precipitation,
and activated carbon adsorption. Representative data from
aqueous mercury treatment operations using these
methods are provided in this report.
Another technology is ion exchange treatment, which
has historically been limited to the use of anion resins to
process industrial wastewater that contains inorganic
mercury in the complex mercuric chloride form. Chapter 8
provides a case study illustrating the use of an ion
exchange system for mercury removal.
Other, less-established methods for treating aqueous
mercury that are discussed in this report include chemical
reduction, membrane separation, and emerging
technologies involving macrocycles adsorption, biological
treatment, and membrane extraction.
Each of the mercury treatment technologies described
in this report achieves different effluent mercury
concentrations. The effectiveness of treatment provided by
each type of technology depends on the chemical nature
and initial concentration of mercury as well as the presence
of other constituents in the wastewater that may interfere
with the process. As indicated by example data provided,
co-precipitation and ion exchange achieve the lowest
effluent mercury concentrations for many waste streams,
ranging from 0.5 to 5.0 /uglL. Membrane technology
typically achieves 80 to 90 percent rejection of mercury.
Other factors, however, such as residuals management
and costs, weigh heavily in selecting the appropriate
treatment approach.
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Chapter 2
Precipitation Treatment Process
This chapter presents information on precipitation and
coagulation/co-precipitation technologies, which are among
the most well-established approaches for removing
mercury from wastewater. The information provided
includes example data from aqueous mercury treatment
operations using these methods.
2.1 Sulfide Precipitation
One of the more commonly reported precipitation
methods for removal of inorganic mercury from wastewater
is sulfide precipitation. In this process, sulfide (e.g., as
sodium sulfide or another sulfide salt) is added to the
wastestream to convert the soluble mercury to the relatively
insoluble mercury sulfide form:
Hg2+ +
HgS(!
[2-11
As with other precipitation treatment, the process is
usually combined with pH adjustment and flocculation,
followed by solids separation (e.g., gravity settling, filtra-
tion). A typical process flow diagram for sulfide precipita-
tion is shown in Figure 2-1. The sulfide precipitant is
added to the wastewater in a stirred reaction vessel, where
the soluble mercury is precipitated as mercury sulfide. The
precipitated solids can then be removed by gravity settling
in a clarifier as shown in Figure 2-1. Flocculation, with or
without a chemical coagulant or settling aid, can be used to
enhance the removal of precipitated solids.
Table 2-I presents example sulfide treatment results.
For initial mercury levels in excess of 10 mg/L, sulfide
precipitation can achieve 99.9+% removal. Even with
polishing treatment such as filtration the minimum effluent
mercury achievable appears to be approximately 10 to
100 //g/L. The most effective precipitation, with regard to
minimizing sulfide dosage, is reported to occur in the near-
neutral pH range. Precipitation efficiency declines signifi-
cantly at pH above 9 (Patterson, 1985). Sulfide precipita-
tion appears to be the common practice for mercury control
in many chlor-alkali plants, removal efficiencies of 95 to
99.9 percent are reported for well designed and managed
Figure 2-1 Sulfide precipitation.
Acid/Base
Sulfide precipitant
Chemical flocculants
and/or settling aids
Effluent
Sludge
clarifier
2-I
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Table 2-1. Sulfide Precipitation Treatment for Mercury (After Patterson, 1985)
Mercury Concentration (ug/L)
Treatment
Chemical
Initial
Final
Percent Mer-
cury Removal
Treatment
PH
Additional Treatment
Sodium sulfide
NA
300-6,000
1,000-50,000
<3
10-125
10
NA
58-99.8
99999.9
NA Vacuum filter
NA Pressure filter
NA Flocculation + activated car-
bon
Sodium hydrosulfide
Magnesium sulfide
"Sulfide" salt
131,50
5,000-10,000
300-6,000
NA
NA
NA
20
1 0-50
10-125
(50 avg)
100-300
100
1 0-20
>99.9
99-99.9
58-99.8
NA
NA
NA
3.0
10-11
5.1-8.2
NA
NA
NA
"Filter"
None
Filtration
None
None
Activated carbon
NA = Not available.
treatment Mercury systems (Perry, 1974; U.S. EPA, 1974).
A mercury effluent level of about 65 /wg/L has been re-
ported for sodium sulfide treatment of wastewaters from
the chlor-alkali industry; influent mercury concentration was
not reported (U.S. EPA, 1974). Costs of using the sulfide
process for the treatment of chlor-alkali wastewater were
reported to be $0.79/1,000 gal (1987 basis), exclusive of
sludge management. Capital cost (adjusted to 1995 basis)
for a chlor-alkali plant utilizing sodium sulfide addition plus
diatomaceous earth filtration for a 100-gpm flow was
$2,767.47 /1 ,000 gpd capacity (Perry, 1974). One conse-
quence of the application of sulfide precipitation technology
is stockpiles of mercury-laden process sludges, which must
be either disposed of in an environmentally acceptable
manner or processed for mercury recovery. Thus, the
sludge management approach chosen is a key factor in
evaluating the sulfide process for treating such wastewater.
In addition to its inability to reduce mercury below 10 to
100 ^g/L, other drawbacks of this method include: (1) the
formation of soluble mercury sulfide species at excess
dosage of sulfide, due to the common ion effect, (2) the
difficulty of real-time monitoring of reactor sulfide levels, (3)
the generation of toxic residual sulfide in the treated
effluent (a potential problem), (4) the difficulty of clarifica-
tion and sludge processing, and (5) the need to dispose of
sulfide sludges. Investigators have reported that mercury
can resolubilize from sulfide sludges under conditions that
can exist in landfills (Hansen and Stevens, 1992). This
could in mercury contamination of leachate and potential
result ground-water pollution.
2.2 Coagulation/co-precipitation
Information is available in the literature on the removal
of both inorganic and organic mercury by coagulation/co-
precipitation for a variety of mercury-containing waste-
waters (Patterson, 1985). Coagulants employed include
aluminum sulfate (alum), iron salts, and lime. For alum
and iron, the dominant mercury removal mechanism is
most likely by adsorptive co-precipitation (Patterson et al.,
1992). Here, one ion is adsorbed into another bulk solid,
formed, for example, by addition of alum and precipitation
of aluminum hydroxide or by addition of an iron (ferrous or
ferric) salt and precipitation of iron hydroxide. The adsorp-
tion process is isothermal, and treatment performance can
be enhanced by optimal bulk solids formation and by pH
manipulation to optimize bulk solid surface change and
soluble mercury speciation.
In studies on the treatment of inorganic mercury dosed
to domestic sewage, both iron and alum co-precipitation,
followed by filtration, reduced initial mercury levels of 50 to
60 /^g/L by 94% to 98%. Lime coagulation treatment,
applied at a higher mercury level of 500 //g/L, achieved 70
percent removal upon filtration (Patterson, 1985). Treat-
ment data for coagulation/co-precipitation are summarized
in Table 2-2. Effluent levels of mercury achieved by alum
treatment range from 1.5 to 102 i^g/L, with a typical 5 to 10
value, and by iron treatment from 0.5 to 12.8
2-2
-------
Table 2-2. Coagulation/co-precipitation Treatment Results for Mercury (After Patterson, 1985)
Mercury,
Coagulant
Coagulant Dosage
Salt (mg/L)
Alum 1,000
100
100
21-24
NA
220
20-30
20-30
Iron 34-72
NA
40
20-30
20-30
Lime 415
NA
"Organic mercury.
NA = Not available.
- = None
Percent Mer-
Initial
1 1 ,300
90
NA"
5.9-8.0
50
60
3-8
3-16"
4.0-5.0
50
50
1-17
2-17'
500
0.66
Final
102
11
10
5.3-7.4
26.5
3.6
1.5-6.4
2.3-21 .3
2.5
3.5
1.0
0.5-6.8
1.2-12.8
150
co.2
cut-y Re-
moval
99
88
NA
1 o-34
47
94
50-81
<23
38-50
93
98
50-97
40-93
70
>69
Treatment
PH
3
NA
NA
6.7-7.2
7.0
6.4
NA
NA
6.9-7.4
8.0
6.2
NA
NA
11.5
8.3
Additional
Treatment
Filtration
__
Filtration
Filtration
Filtration
Filtration
Filtration
Filtration
Filtration
2-3
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Chapter 3
Adsorption Processes
Adsorption processes have the potential to achieve
high efficiencies of mercury removal and/or low effluent
mercury levels. The predominant adsorption process
utilizes activated carbon, but the use of other adsorbents
also are reported in the literature. These include pro-
cessed vegetable or mineral materials such as
bicarbonate-treated peanut hull carbon (BPHC), modified
Hardwickia binata bark (MHBB), coal fly ash, and the
Forager sponge (Namasivayam and Periasamy, 1993; Sen
and De, 1987; Deshkar et al., 1990; U.S. EPA, 1994b).
Metal hydroxides are also used as adsorbents. When
metal hydroxides are employed for adsorptive treatment,,
the process is commonly termed coagulation or co-precipi-
tation. (This process is discussed in Chapter 2.) An
inherent advantage of adsorptive treatment, particularly
when the adsorbent displays isothermal or quasi -isother-
mal behavior, is that increased treatment efficiency results
from incremental adsorbent dosage. Isothermal behavior
is observed when, for a fixed initial pollutant concentration,
decreasing residual soluble concentrations are observed
as the dosage of adsorbing treatment material is added.
Unless adsorbent recovery is feasible, these incremental
dosages also result in production of increased wastewater
treatment residuals, requiring ultimate disposal. Variables
other than adsorbent type and'dosage can also affect
adsorption efficiency. Common variables include waste-
water pH and pollutant speciation.
4.1 Activated Carbon Adsorption
Granular activated carbon (GAC) is the most commonly
used adsorbent system for treating industrial waste (U.S.
DOE, 1994). This process is used in a variety of configura-
tions, as demonstrated in Figures 3-1 and 3-2. GAC
systems may be either pressure or gravity type. They may
Figure 3-1. Types of GAC column design (Calgon Carbon Corp.)
Influent
Granular
activated
carbon
Influent
Eff.
Granular
activated
carbon
Effluent
Downflow in series
(A)
Inf.
Moving-bed
(B)
,, Eff. Inf.
*> '
Downflow in parallel
(C)
Upflow expanded in series
(D)
3-1
-------
Figure 3-2. GAC process flowsheet (after Eckenfelder, 1989)
Backwash effluent
return
^o
Backwash
pump
Quench
tank
Regenerated carbon
High pressure
Eductor
High pressure
water
Eductor water
be upflow counter-current type with packed or expanded
carbon beds, or upflow or downflow fixed-bed units with
multiple columns in series (Figure 3-I). Contaminated
water is passed through the columns until the key contami-
nant is detected at a predetermined level in the effluent.
When multiple columns are placed in series, the first
column can be loaded to a greater capacity, while residual
levels of the contaminant are removed in the downstream
columns. When a column has been loaded to its design
capacity, it may be regenerated or the spent carbon can be
replaced while another column is brought online. An
alternative method of carbon treatment involves use of
powdered activated carbon (PAC). The PAC is typically
added as a slurry into a contact reactor, and the PAC solids
subsequently are removed in a solids separation stage.
The PAC is normally not regenerated for' reuse due to
unfavorable economics including poor recovery of the PAC.
Table 3-1 summarizes example activated carbon
mercury treatment data. The removal of mercury from
potable water using PAC was studied by Thiem and
colleagues (1976). Treating a spiked water solution
containing 10 /^g/L total mercury, they achieved approxi-
mately 80% removal at a pH of 7 and a PAC dosage of
100 mg/L. The study also demonstrated that the addition
of mercury chelating agents, such as ethylene diamine
triacetic acid (EDTA) or tannic acid, prior to contact with the
PAC increased mercury removal efficiency. Concentra-
tions as low as 0.02 mg/L EDTA and 1 mg/L tannic acid
increased mercury removal efficiencies by 10% to 20% .
The mercury removal efficiencies by concentrations of 50
to 200 mg/L also increased mercury removal efficiencies by
10% to 20% over those obtained by PAC alone.
The removal of mercury (II) from synthetic wastes by
11 different brands of commercial activated carbon was
studied by Huang and Blankenship (1984). Among the 11
different types of activated carbon, Nuchar SA and Nuchar
SN exhibited a high percent (>99.9) mercury (II) removal
over a wide pH range (2.5 to 11). The other activated
carbons studied displayed maximum total mercury (II)
removal at pH 4 to 5, and the percent mercury (II) removal
dropped markedly at pH values greater than and less than
4 to 5.
Pretreatment or modification of activated carbon with
carbon disulfide solution before use, has been shown to
enhance mercury removal. Humenick and co-investigators
(1974) utilized an activated carbon that was presoaked in
carbon disulfide and then dried and used as PAC. The
pretreated activated carbon removed mercury from an
initial concentration of 10 mg/L down to 0.2 Mg/L, versus
the 4 mg/L effluent value obtained with the untreated
carbon. The enhanced mercury removal was attributed to
chemisorption reactions. Sulfur atoms have a high affinity
for mercury, as evidenced by the Ksp of HgS (see Table 2-
2). The mercury removal mechanism proposed by Hum-
enick and colleagues (1974) involves transport and diffu-
sion to the carbon disulfide sites and subsequent formation
of a chemical bond between a carbon disulfide molecule
and the mercury ion.
3-2
-------
Table 3-1. Activated carbon mercury treatment results
Mercury Concentra-
tion
Activated Car- Percent Additional Other
bon Type initial Final Removal Treatment Conditions Reference
PAC
PAC
PAC
10,000 4,000 60 None SW, BS
10,000 0.2 >99.9 5 ^m filtration, PAC SW, BS
presoaked in CS2 and dried
2,000 NA -100 Centrifugation or 0.45 ^m SW, BS
filtration
Humenick et al.,
1974
Humenick et al.,
1974
Huang and
Blankenship,
1984
PAC
PAC
GAC
GAC
10
1.0
0-100
1.7
1.5
NA
0.5
0.9
0.8
-80
50
47
47
0.45 ^m filtration
Settling
None
Filtration
SW, BS Thiem et al., 1976
PW, BS Guarino et al., 1988
SF, FS E.G. Jordan Co.,
1989
PW, BS Guarino et al., 1988
PAC
GAC
BS
SW
PW
SF
FS
NA
Powdered activated carbon.
Granular activated carbon.
Bench scale.
Synthetic wastewater.
Petrochemical wastewater.
Superfund wastewater.
Full scale.
Not available.
A study was conducted by Guarino and co-invest-
igators (1988) to establish the feasibility of using activated
carbon as an advanced treatment method for petrochemi-
cal wastewater. This study investigated petrochemical
wastewater at bench scale, utilizing GAC and PAC. Low
initial mercury levels of 1.5 and 1.7 /^g/L were reduced to
0.8 and 0.9 ^g/L, respectively, using GAC, while an initial
mercury concentration of 1.0 /ug/L was reduced to 0.5 fj.g!L
using PAC. The performance data reported in the literature
suggests that activated carbon treatment can achieve a
residual mercury level of 0.5 to 20 M9/L, dependent in part
on the initial wastewater mercury level (Patterson et al.).
Gates and colleagues (1995) conducted laboratory
work to investigate the feasibility of using inexpensive
sulfur-impregnated activated carbon beads, known as
Mersorb, for mercury removal from aqueous waste. These
studies were conducted to evaluate the treatability of
mercury-containing aqueous and solid mixed wastes stored
at DOE sites, such as the Oak Ridge Y-12 site. The from
aqueous solutions to below 0.2 mg/L. Mersorb worked
under acidic conditions (pH of 2), but its capacity at low pH
was reduced by 50% compared with neutral conditions.
Mersorb beads reportedly had favorable process econom-
ics compared with ion exchange.
3.2 Xanthate Treatment
An alternative adsorption material to activated carbon
is starch xanthate, yielding mercury-starch xanthate. One
modification is termed the Metals Extraction by Xanthate
Insolubilization and Chemical Oxidation (MEXICO) pro-
cess, also termed the Advanced MEXICO Precipitation
Process (Macchi et al., 1985; Tiravanti et al., 1987). Most
published data on this process appears to be from bench-
and pilot-scale studies. No published information was
available on full-scale application.
Example data for starch xanthate treatment are
presented in Table 3-2. Campanella and colleagues (1986)
were able to reduce the mercury concentration in a syn-
thetic wastewater at bench scale from 10 to 23
3-3
-------
Table 3-2. Starch Xanthate Treatment for Mercury
Mercury Concentration (mg/L.)
Initial
10
100
9.5
9.5
Final
0.023
0.001
0.01-0.1
0.005-0.02
Treatment
PH
1
5
5
5
Additional Treatment
Sedimentation
0.45 ^m filtration
Sedimentation
Sedimentation plus 0.45 ^m filtra-
Other
Conditions
SW, BS
SW, BS
cw, PS
cw, PS
Reference
Campanella et
Tiravanti et al.,
Tiravanti et al.,
Tiravanti et al.,
al., 1986
1987
1987
1987
tion
6.3
6.3
6.3
-0.2
0.01
0.001
11
11
NA
1 0 ^m filtration
Sodium hypochlorite addition
Activated carbon
cw, BS
CW, BS
CW, BS
Macchi et al., 1985
Macchi et al., 1985
Macchi et al., 1985
SW = Synthetic wastewater.
BS = Bench scale.
cw = Chlor-alkali wastewater.
PS = Pilot scale.
following sedimentation. Tiravanti and co-investigators
(1987) were able to reduce mercury at bench scale from
100 to 1 yug/L following 0.45 //m filtration. These research-
ers also conducted pilot-scale (15 m3/d) experiments on
chlor-alkali wastewater and were able to reduce the
mercury concentration from 9.5 mg/L to a range of 10 to
100 ^g/L following sedimentation, and to a range of 5 to 20
ng/L following sedimentation and 0.45 ^m laboratory
filtration (to estimate residual soluble mercury). Macchi
and colleagues (1985) conducted bench-scale experiments
on chlor-alkali wastewater and were able to reduce the
mercury concentration from 6.3 to 200 ^g/L following
10 /^m filtration, to 10 /^g/L following sodium hypochlorite
addition, and to 1 /^g/L following activated carbon treat-
ment. The process appears able to achieve an effluent
mercury level of 5 to 20 //g/L.
Macchi and colleagues (1985) also reported that
mercury can be recovered from the mercury-xanthate
sludges by treating the precipitate with 5 M hydrochloric
acid and sodium hypochlorite. The cost of sodium hypo-
chlorite is relatively insignificant for the chlor-alkali industry,
and the redissolved mercury reportedly could be recycled
to the head of the chlor-alkali plant.
3.3 Other Adsorption Processes
Various other adsorbent alternatives to activated
carbon have been reported to perform in comparable
fashion for mercury treatment. These adsorbents include
BPHC, MHBB, coal fly ash, and the Forager sponge. Each
of these adsorbents is described in the following sections.
Table 4-3 presents mercury adsorption Freundlich
parameter values for these adsorbents, except the Forager
sponge. The Freundlich adsorption equation is:
logx = log k + llogCe [3-1]
m n
Where:
x = the amount of solute (mercury) adsorbed
m = the amount of adsorbent required to adsorb x
k and 1= empirical constants (Freundlich parameters)
n
Ce = equilibrium concentration (mercury) The Freundlich
parameters k and are equal to the intercept and slope
of the line obtained by plotting log vs. log C,.
m
3-4
-------
Table 3-3. Freundlich Isotherm Parameters for Mercury Adsorption
Adsorbent
1
n
Reference
GAC
BPHC
Coal fly ash
(PH 2.2)
Coal fly ash
(PH3.1)
Coal fly ash
(pH 4.2)
MHBB
4.68 3.16 .Namasivayam and Periasamy, 1993
42.17 3.50 Namasivayam and Periasamy, 1993
1.014 0.053 Sen and De, 1987
1.094 0.333 Sen and De, 1987
1.230 0.361 Sen and De, 1987
1.07 0.324 Deshkaret al., 1990
GAC = Granular activated carbon.
BPHC = Bicarbonate-treated peanut hull carbon.
MHBB = Modified Hardwickia binata bark.
The value of k is roughly an indicator of sorption capacity,
and is an indicator of sorption intensity.
n
3.3.1 BPHC Adsorption
From bench-scale study, using a stock mercury
solution feed of 10 to 20 mg/L, Namasivayam and Peri-
asamy (1993) reported BPHC to be seven times more
effective than GAC for mercury (II) removal. This result
was attributed to the higher porosity plus moderate ion
exchange capacity of BPHC as compared to GAC. The
Freundlich parameters shown in Table 3-3 quantify the
sorption capabilities of BPHC. The desorption capabilities
of BPHC also were reported to be promising. Percent
recoveries of mercury from BPHC and GAC using 0.6 M
HCI were 47% and 13%, respectively, and 87% and 24%,
respectively, using 1.0% Kl (potassium iodide). No full-
scale data were available on this material.
3.3.2 MHBB Adsorption
A modified Hardwickia Binata bark was studied at
bench-scale for its adsorption of mercury (II) from water
(Deshkar et al., 1990). Although the media was shown to
be effective in removing mercury (II) from water, it is not as
effective as GAC, as indicated by the Freundlich parame-
ters listed in Table 3-3. No information was reported on the
desorptive properties of the Hardwickia binafa bark.
3.3.3 Coal Fly Ash Adsorption
Coal fly ash, an industrial waste solid, was shown to
adsorb mercury (II) (Sen and De, 1987). Coal fly ash did
not perform as well as GAC, however, as shown by the
Freundlich parameters listed in Table 3-3. Maximum
mercury adsorption by coal fly ash was observed in the pH
range 3.5 to 4.5 (Sen and De, 1987).
3.3.4 Forager Sponge Adsorption
The Forager sponge is an open-celled cellulose sponge
with an amine-containing polymer that reportedly has a
selective affinity for aqueous heavy metals in both cationic
and anionic states. The polymer is reported to form
complexes with ions of transition-group heavy metals,
providing ligand sites that surround the metal and form a
coordination complex. The polymers order of affinity for
metals is reportedly influenced by solution parameters such
as pH, temperature, and total ionic content. Mercury is one
of the metals that is claimed to be removed by the sponge.
In general, the following affinity sequence for representa-
tive ions is expected (U.S. EPA, 1994b):
Cd2t>Cu2+>Fe3+>Au3t>Mn2+>Zn2+>Ni2+>Co2+>Pb2+>Au(CN)2-
6>Se042->As043->Hg2+>CrO/->Ag+>AI3+>Ca2+>Mg2+
The sponge can be used in columns, fishnet-type
enclosures, or rotating drums. When used in a column,
flow rates of 3 bed volumes per minute are reported to be
obtained at hydrostatic pressure only 2 feet above the bed
and without additional pressurization. Therefore, sponge-
packed columns are claimed to be suitable for unattended
field use.
Adsorbed ions can be eluted from the sponge using
techniques typically employed to regenerate ion exchange
resins and activated carbons. Following elution, the
sponge can be reused in the next adsorption cycle. The
number of useful cycles is reported to depend on the
nature of the adsorbed ions and the elution technique
3-5
-------
used. Alternatively, the metal-saturated sponge could be sponge may be dried and reduced in volume to facilitate
incinerated. Metals volatilization would be of concern. The disposal (U.S. EPA, 1994b).
3-6
-------
Chapter 4
Ion Exchange Treatment
Resins containing the iminodiacetic acid group will
exchange for cationic mercury selectively over calcium and
magnesium, but copper and cobalt are also readily ex-
changed. Mercury in the form of anionic complexes, such
as HgCr3, can be treated by anion exchange resins. The
thiol resin, Duolite GT-73, is reported to be selective for
mercury in any of its three oxidation states (Ritter and
Bibler, 1992).
Ion exchange processes are typically operated as
packed columns. Usually four operations are carried out in
a complete ion exchange cycle: service,, backwash,
regeneration, and rinse. In the service step, the ion
exchange resin in the packed column is contacted with the
water containing the mercury to be removed. After a target
concentration of mercury in the column effluent is reached,
the resin is said to be spent. A backwash step is then
initiated to expand the bed and to remove fines that may be
clogging the packed bed. The spent resin is then regener-
ated by exposing it to a concentrated solution of the original
exchange ion, so that a reverse exchange process occurs.
The rinse step removes excess regeneration solution
before the column is brought back online for the next
service cycle.
Reported advantages and disadvantages of ion ex-
change include (Clifford et al., 1986):
Advantages
Operates on demand
Is relatively insensitive to variability
Can achieve essentially a zero level of effluent
contaminant
Is available in a large variety of specific resins
Can normally achieve beneficial selectivity reversal
upon regeneration
Disadvantages
Has potential for chromatographic effluent peaking
Results in spent regenerant brine that must be
disposed of
a Can yield variable effluent quality
Cannot typically be used for waters with a high
total dissolved solids content
Ion exchange technology for mercury removal has
historically been limited to the use of anion resins to treat
industrial wastewater that contains inorganic mercury in the
complex mercuric chloride form. For the process to be
effective, the chloride content of the wastewater must be
high, such as that generated by a chlor-alkali plant. This
will yield negatively charged mercury chloride complexes.
If the chloride content of the wastewater is low, either
chlorine or chloride salt could be added to improve removal
process efficiency (Sorg, 1979).
Cation exchange of mercury may be effective if the
anion content of the wastewater is low (Sorg, 1979).
Certain cation exchange resins (Amberlite IR-120 and
Dowex-50W-X8) are reported to be effective for ion
exchange treatment of mercury present in industrial
wastewater (Patterson, 1985). Also, Duolite GT-73, a
cationic resin, contains the thiol (-SH) group and reacts
with ionic mercury. The thiol functional group has a high
selectivity for mercury as well as a strong tendency to bind
certain other metal ions such as copper, silver, cadmium,
and lead.
A chelate resin is an insoluble polymer to which is
attached a complexing group or groups. This, in turn, can
bond metal cations within the structure so as to form a ring
(or chelate) into which the metal is incorporated. The
reaction involves both ion-exchange and chemical reac-
tions. Table 4-1 lists some chelate resins that are reported
to have a high selectivity for mercury; the table includes the
order of selectivity.
Example ion exchange treatment data for drinking
water are presented in Table 4-2. Mercury removal from
ground water was studied in point-of-entry treatment
(POET) systems installed on private water supply wells
(Sites and Obeholtzer, 1992). Table 4-2 indicates that
lonac SR-4, Purolite S-920, AFP-329, and ASB-2 were
able to remove mercury from the relatively low initial
ground-water concentrations to below 1 /^g/L, following
prefiltration.
A full-scale ion exchange process at a defense pro-
cesses facility has consistently removed mercury via ion
exchange from 0.2 to 70 mg/L down to levels of 1 to 5
,ug/L, following 0.2 /^m prefiltration (Ritter and Bibler, 1992).
This system utilizes a macroporous, weakly acidic, polysty-
rene/divinylbenzene cation resin, with thiol (SH) functional
groups. High levels of mercury in a synthetic wastewater
4-I
-------
Table 4-1. Summary of mercury-selective chelate resins (After Calmon, 1981)
Resin Order of Selectivity
Duolite ES-466
Dowex A-l
Nisso Alm-525
Diaion CR-I 0
Amberlite IRC-718
Unicellex UR-10
Sirorez-Cu
Sumichelate Q-l 0
Hg2*>Cu2*>Fe2+>Ni2+>Pb2+>Mn2*>Ca2+>Mg2*>Na+
Cu2+>Hg2VNi2*>Pb2Wn2*>Co2+>Cd2+>Fe2+>Mn2H>Ca2+>Na+
Hg2+>Cd2*>Zn2+>Pb2+>Cu2+>Ag+>Cr3+>Ni2+
Hg2*>Cu2+>Pb2+>Ni2+>Cd2+>Zn2+>Co2+>Mn2+>Ca2+>Mg2+>Ba2+>Sr2*>»Na+
Hg2+>Cu2+>Pb2t>Ni2*>Zn2t>Cd2+>Coz+>Fe2+>Mn2+>Ca2+
Hg2*>Cu2+>Fe3+>A13+>Fe2+>Ni2t>Pb2+>Cr3+>Zn2+>Cd2+>Ag2+>Mn2+>Ca2*>Mg2+>»Na2<
pH>5, Cu2+; pH>0, Hg2*
HgCI2>AuCr4>Ag+>Cr2072-
Table 4-2. ion Exchange treatment for mercury in drinking water
ion Ex-
change
Resin
lonac SR-4
Purolite
s-920
AFP-329
ASB-2
Duolite
GT-73
Amberlite
IRC 718
IRC 718
and
GT 73
Mercury
Concentration
(M9/L)
Resin Type Initial Final
Weak acid chelat- 14.88" 0.43"
ing resin
Hg-specific chelat- 10.67" 0.34"
ing resin
Weak base anion 12.21" 0.44a
resin
Strong base anion 14.31" 0.70"
resin
Weak acid cation 200-70,000 1-5
thiol
Iminodiacetic acid 11,800 15-35
resin
(See above) 14,000 15-1 ,200
Additional Other
Treatment Conditions Reference
Prefiltration GW, FSb
Prefiltration GW, FSb
Prefiltration GW, FSb
Prefiltration GW, FSb
0.2 p.m pretilter DFW, FS
None SW, BS
GT 73 used as SMW, BS
polishing
Sites and Oveholtzer, 1992
Sites and Obeholtzer, 1992
Sites and Oveholtzer, 1992
Sites and Oveholtzer, 1992
Ritterand Bibler, 1992
Becker and Eldrich, 1993
Becker and Eldrich, 1993
'Average value
b3 to 4 gpm
GW = Ground water.
FS = Full scale.
DFW = Defense facility wastewater.
SW = Synthetic wastewater.
BS = Bench scale.
SMW = Smelter wastewater.
4-2
-------
were removed to levels as low as 15 /^g/L after 77 bed
volumes of usage, and 35^g/L after 157 bed volumes of
usage (Becker and Eldrich, 1993). This system utilized
Amberlite IRC 718 in bench-scale testing. In further
bench-scale testing, smelter wastewater containing 14.0
mg/L of mercury at a flow rate of 6.7 mL/min was treated
with IRC 718 followed by a polishing ion exchange
column containing GT 73. This system removed mer-
cury to concentrations of 15 to 46 /^g/L after 289 bed
volumes, and 1,200 ^g/L after 325 bed volumes. This
study further showed that at pH 1.5, the iminodiacetic
acid resin (IRC 718) was highly selective for mercury (II)
over zinc, lead, and cadmium, and that mercury recov-
ery from wastewater on such a resin is feasible provided
strongly complexing anions such as chloride are absent.
Regeneration with 3 M NaCI or other complexant for
mercury at near neutral pH yields a solution for which
mercury can reportedly be recovered via reduction to an
insoluble and commercially valuable form (Becker and
Eldrich, 1993).
4-3
-------
Chapter 5
Other Processes
In addition to precipitation, adsorption, and ion ex-
change treatment technologies, the following processes
are also reported to be applicable to remove mercury from
wastewater: (1) chemical reduction, (2) membrane separa-
tion, and (3) various emerging technologies.
5.1 Chemical Reduction
The standard electrode potential of metals determines
their placement in the electromotive series, which is a
series of elements in descending order of their standard
potential. Ionic mercury can be displaced from solution via
reduction by another metal higher in the electromotive
series, and then separated by filtration or other solids
separation technique. Reducing agents include aluminum,
zinc, iron, hydrazine, stannous chloride, and sodium
borohydride. Example data on these reductants are
presented in Table 5-1.
Although the literature includes much discussion of
reduction processes, only limited actual treatment data are
presented. The main advantage claimed for reduction is
that mercury can be recovered in the metallic state (Pat-
terson, 1985). The data in Table 5-1, however, indicate
that most reduction processes cannot effectively achieve
mercury levels below 100 /J.Q/L, and their use would likely
require second-stage polishing.
Experiments were conducted by Gould and colleagues
(1984) at bench scale on Chemical Oxygen Demand
(COD) test wastewater using iron wire (nominal diameter
0.229 mm). Due to the high initial mercury levels (735 to
2,030 mg/L), high recovery efficiencies were observed
(96% to 99%); however, high residual mercury levels were
also observed (22 to 33 mg/L). Experiments were con-
ducted by Grau and Bisang (1995) on synthetic wastewater
with iron felt formed by compressing iron wool. As for other
studies, a high removal efficiency resulted at the high initial
mercury concentration, leaving 68 to 91 /^g/L residual
mercury.
As noted in Chapter 4, 'mercury removal from ground
water was studied in POET systems that were installed on
private water supply wells (Sites and Oberholtzer, 1992).
Table 5-I shows that a bimetallic oxidation/ reduction
compound, KDF, which consists of a finely ground alloy of
55% copper and 45% zinc, was able to remove low levels
of mercury down to a range of 0.4 to 1.08 //g/L, following
prefiltration. This process may be applicable only for
exceptionally clean solutions, however.
5.2 Membrane Separation
Several membrane processes have been applied for
water and wastewater mercury treatment. These include
ultrafiltration, charged filtration, crossflow microfiltration,
magnetic filtration, and reverse osmosis. Example treat-
ment data for these processes are shown in Table 5-2.
Ultrafiltration systems are pressure-driven membrane
operations that use porous membranes for the removal of
dissolved and colloidal material (Metcalf and Eddy, 1991).
These systems differ from reverse osmosis systems by the
relative driving pressures, usually under 150 psi (1034
kN/m2). Ultrafiltration is normally used to remove colloidal
material and large molecules with molecular weights in
excess of 5,000. Recent studies indicate that effluent from
ultrafiltration using spiral wound elements is suitable as a
feed source for reverse osmosis (Metcalf and Eddy, 1991).
Chelation in combination with ultrafiltration is a process
that has been described for the removal of heavy metals,
including mercury (Kosarek, 1981). This concept is based
on reacting ligands with cationic metallic constituents to
form a metal-containing complex (chelate), and then
removing these metal-containing complexes by ultrafiltra-
tion (Kosarek, 1981). The opposite charges of the ionized
ligand and metal attract each other to form a stable chelate
complex. The properties that facilitate ultrafiltration
membrane rejection of the metal-containing complex
(including mercury complexes) are thought to be (1) the
increased size of the metal chelate complexes, (2) alter-
ation in the ionic shape of the metal, (3) modified solubility,
and (4) reversal of charge from cationic metal to a function-
ally anionic or electroneutral chelate species (Kosarek,
1981).
Charged membrane ultrafiltration incorporates a
noncellulosic, high flux membrane that is negatively
charged as a result of dissociated subgroups within the
membrane structure. A beneficial aspect of the charged
ultrafiltration membrane is that the negative polarization
minimizes membrane fouling (Kosarek, 1981). Bhat-
tacharyya and colleagues (1979) conducted bench-scale
investigations to determine the feasibility of the simulta-
neous separation of various heavy metals from scrubber
blowdown wastewater generated in the primary copper
industry. They studied the application of low pressure
ultrafiltration with commercially available, negatively
charged noncellulostic membranes. Typical mercury values
5-1
-------
Table 5-1. Performance of Reduction Processes for Mercury Treatment.
Mercury
Reductant
Initial
Final
Treatment pH
Reference
Zinc
Iron
Iron felt
KDF'
Stannous chloride
Sodium borohydride
5,000-10,000
1,800
12,500
12,500
12,500
NA
734,000-
2,030,000
100,000
6.17-12.11
2,800
10,000
4,000
26,000
4,700
NA
5-10
140
830
750
470
600
22,000
-33,000
68-91
0.4-I ,08
500
220
420
820
200
<10
NA
11.5
10.0
6.0
2.5
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
Patterson, 1985
Patterson, 1985
Patterson, 1985
Patterson, 1985
Patterson, 1985
Patterson, 1985
Gould et al., 1984
Grau and Bisang, 1995
Sites and Oberholtzer, 1992
Patterson, 1985
Patterson, 1985
Patterson, 1985
Patterson, 1985
Patterson, 1985
Patterson, 1985
'Bimetallic copper-zinc oxidation/reduction compound.
NA = Not available.
in the scrubber blowdown were found to be 1.5 to 2.0 mg/L.
Mercury removals for a full-scale system were calculated
by computer simulation based on laboratory-scale data and
were reported to be 91% to 93% (AP = 5.6 x 105 N/m*,
channel velocity, U = 250 cm/s, and pH = 4.5). For a
3.6 x 10s kg/d (400 ton /day) copper production plant, the
net operating costs for a closed-loop scrubber blowdown
water recycle system was estimated to be $2,508 /day or
$0.0070/kg Cu production (1995 basis updated assuming
cost components follow change in skilled labor cost index).
The crossflow microfiltration system is based on the
concept of using a dynamic membrane to form a filtration
medium. This process, whose patented form is called
Exxflow, is a solid-liquid separation process in which the
feed suspension sweeps across the face of a filter mem-
brane while pressure differences cause the liquid phase to
pass through the membrane, leaving the solids to be
flushed away in the residual flow. By this means, the solids
are concentrated up in the suspension flow, which is
commonly recycled to the feed end. This contrasts with
"barrier" filtration systems in which the solids build up on
the filtering surface, gradually restricting the flow through
the filter (Squires, 1992). A schematic of the microfiltration
process is shown in Figure 5-I.
Mercury removal via crossflow microfiltration was
reported for a full-scale plant designed to process 200 m3
/day of mixed plating wastewater (Broom et al., 1994). A
process schematic of the plant is shown in Figure 5-2. The
filtrate from the rotary vacuum filter was combined with the
supernatant from a preclarification stage and stored in a
80 m3 balance tank, where the pH was adjusted to 11 to
12, primarily to precipitate cadmium. Sodium hydrogen
sulfide (NaHS) was also added to precipitate any soluble
metals remaining. This conditioned filtrate was then
pumped to the crossflow microfiltration unit, where it was
recycled at an average pressure of 150 kPa. The reject
flow was effectively a concentrate produced by the pas-
sage of clean permeate through the filter. With mercury
feed concentrations to the microfiltration plant of 1.27,
0.967, 0.15, and 2.28 mg/L, permeate concentrations of
0.015, 0.015, 0.088, and 0.03 mg/L were achieved, respec-
tively. This represents a removal efficiency of about 95 %.
Removal may have been enhanced by mercury co-precipi-
tation in the balance tank.
5-2
-------
Table 5-2. Performance of Membrane Processes for Mercury Treatment.
Mercury (M9/L)
Membrane
Process
Reverse osmosis
Initial Final
5,000 880*
9,000 1 5Q3.
8 1.5M.7*
Percent
Removal
82.4
83.3
79-81
Comments
BS.SW
BS, SW
PS
Reference
Sorg, 1979
Charged ultrafil-
tration
Crossflow micro-
filtration
Magnetic filtration
1,500-2,000
1,270
967
150
2,280
15,000
NA
15
15
88
30
3-117
91-93
99.8'
98.51
41.3*
98.7"
99.2-99.9
CS
FS, PW
FS, PW
FS, PW
FS, PW
BS, GSW
Bhattacharyya et al.,
1978
Broom et al., 1994
Terashima et al., 1986
'Calculated from removal efficiency data (% removal = (Initial - Final) x 100%
Initial
BS = Bench scale.
SW = Spiked wastewater (secondary effluent).
PS = Pilot scale.
CS = Computer simulation.
FS = Full scale.
PW = Plating wastewater.
GSW = Gas scrubber wastewater from solid waste incineration plant.
NA = Not available.
Magnetic filtration of soluble species such as ionic
metal is accomplished by forming a magnetic precipitate
through coagulation and magnetic seed addition, and then
passing the wastewater through a filter made with ferro-
magnetic wires, which represent a magnetic field. The
magnetic precipitate can be rapidly and efficiently removed
by magnetic filtration, even if the precipitate is quite fine
(Terashima et al., 1986). This occurs because of the
strong magnetic forces that act on the magnetic particles
as they move through the magnetic field. The magnetic
particles are captured on the filter matrix by the magnetic
force, which overcomes other competing forces of gravita-
tion, hydrodynamics, and inertia (Terashima etal., 1986).
Terashima and colleagues (1986) report mercury removals
in a bench-scale magnetic filtration unit fed gas scrubbing
wastewater from a municipal solid waste incineration plant.
For an influent mercury concentration of 15.0 mg/L, effluent
mercury concentrations of 0.003 to 0.117 mg/L were
achieved.
Reverse osmosis (RO) is a physical separation tech-
nique whereby an applied pressure in excess of the
inherent solution osmotic pressure forces water to perme-
ate a semipermeable membrane, which rejects the bulk of
the dissolved and suspended constituents. The pressures
applied to the membrane in RO processes range from 200
to 800 psi (Kosarek, 1981). The operation of an RO
membrane system is significantly affected by fouling,
scaling, pH-temperature-pressure-related hydrolysis, and
chemical or biochemical deterioration of the membranes.
The RO process has very strict feed water requirements,
particularly related to the concentration of suspended solids
and materials (e.g., oil or grease), which will foul the
membrane surface (Kosarek, 1981).
A literature review by Sorg (1979) described several
laboratory and pilot-plant RO studies for the removal of
metals from water or wastewater, but only one study
provided data on the removal of mercury. Investigators ran
a series of RO pilot-plant tests to evaluate the technique for
the removal of heavy metals, pesticides, and other toxic
chemicals from secondary waste water effluent. The
results of one-day batch tests with spiked concentrations
of 5 and 9 mg/L of inorganic mercury showed removals of
82.4% and 83.3%, respectively.
Sorg (1979) also reported that EPA's Drinking Water
Research Division (DWRD) conducted two one-day tests
for inorganic mercury (influent at 0.008 mg/L) removal from
Glendale, Ohio, well water, with two small-scale RO
systems. The first system used a spiral-wound (SW)
membrane and the second a hollow fiber (HF) membrane.
The SW system had a raw water flow of 2.2 Lpm and
operated at 1,400 to 1,600 kPa (200 to 230 psi) and 7% to
9% recovery. The HF system had a higher raw water flow
5-3
-------
Figure 5-1. Principles of the Crossflow Microfirter (Broom, et at, 1994).
Influent
, Concentrate
4444
4444
Filtrate
Membrane cross section
Dynamic membrane
Flexible textile tube
Filtrate Filtrate
Figure 5-2. Filtrate Polishing using Micro filtration.
Microfillration stage
Filter cake
Microfiltration feed tank
Permeate
5-4
-------
of 5 L/min and operated at 1,200-1,400 kPa (170 to
200 psi) and 40% to 50% recovery. The test results
showed 25% mercury removal with the SW system and
79% to 81% with the HF system.
The results for these two limited scope studies are not
as high as the 95% to 98% rejected range for organic
mercury that Sorg (1979) reports as equipment manufac-
turer estimates. Sorg also states, however, that full-scale
RO systems operated at high pressure and recovery
should achieve greater removals than pilot-scale units
performing under less than optimum conditions. A sum-
mary of membrane process performances by Kosarek
(1981) cites mercury removals of 95% to 97% by RO.
Based on these mercury treatment results, a broader range
of 80% to 99% membrane rejection is indicated.
5.3 Emerging Technologies
Various other mercury treatment processes are cited in
the technical literature; however, actual mercury treatment
data are extremely limited. These processes include
solvent extraction with high molecular weight amines, the
use of silicon alloys for reduction of elemental mercury, and
adsorption onto ground rubber and wool (Patterson, 1985).
Other processes include the use of bound macrocycles
(Izatt et al., 1991), biological detoxification (Hansen and
Stevens, 1992), and microemulsion liquid membrane
extraction (Larson and Wiencek, 1994). Selected pro-
cesses are briefly reviewed below.
5.3.1 Macrocycles Adsorption
The use of aza macrocycles bonded to silica gel
columns to selectively remove heavy metals from contami-
nated waters was reported (Izatt et al., 1 991 ). Bench-scale
tests using synthetic wastewater removed mercury from 4
to less than 1 //g/L.
5.3.2 Biological Treatment Systems
Biological treatment systems using mercury-resistant
bacteria in a completely mixed, aerobic treatment process
was studied by Hansen and Stevens (1992). Methods for
recovering mercury from the vapor phase of the bioreactor
are discussed by Hansen and Choudhury (1990). In
bench-scale studies, three influent mercury concentrations
(nominal levels of 2, 20, and 40 mg/L), three mean cell
residence times (12, 20, and 28 hours), and three tempera-
tures (15, 22.5, and 30°C) were employed in a 23 factorial
design at a nominal influent COD of 2,500 mg/L (Hansen
and Stevens, 1992). Effluent mercury concentrations
ranged form 0.010 to 18.6 mg/L, with mercury removal
efficiencies ranging form 6.8% to 99.5%.
Genetically engineered plants have been used in
laboratory studies to reduce mercury concentrations in soil.
Certain soil bacteria are known to reduce highly toxic
action of a mercuric ion reductase called MerA. Research-
ers modified the bacterial MerA gene and inserted it into
Arabidopsis thaliana plants. The results showed that the
genetically engineered plants not only absorbed and
concentrated Hg2+ ions from soil, but also reduced them to
less toxic elemental mercury (Anonymous, 1996).
5.3.3 Membrane Extraction
Microemulsion liquid membrane extraction of mercury
is an example of a liquid-liquid extraction technique.
Schematic representations of mercury extraction using an
emulsion liquid membrane are shown in Figures 5-3 and
5-4. A microemulsion forms spontaneously when oleic acid
tetradecane, DNP-8 surfactant, and 6N sulfuric acid are
mixed. The microemulsion is then dispersed in the
aqueous stream containing mercury. After extraction, the
emulsion and aqueous phase are separated. The micro-
emulsion is then demulsified to recover the internal aque-
ous phase, which is concentrated in mercury. During the
extraction process, mercury ion reacts with the oleic acid
at the surface of the emulsion droplet. The mercury/oleic
acid complex diffuses to the interior of the emulsion until it
encounters an internal droplet containing sulfuric acid. A
hydrogen ion is exchanged for the mercury ion on the oleic
acid molecule, which is then free to diffuse back to the
surface of the emulsion and extract another mercury ion
(Larson and Wiencek, 1994). As a result, mercury can be
pumped against its concentration gradient, with counter-
transport of hydrogen ions.
To demonstrate the efficiency of a microemulsion liquid
membrane compared to conventional solvent extraction,
Larson and Wiencek (1994) formulated a microemulsion
with 0.32 molar (M) oleic acid intetradecane, 10 weight %
DNP-8, and 6 normal (N) sulfuric acid. At equilibrium, a
clear microemulsion phase and excess aqueous phase
could be observed. The aqueous content of the micro-
emulsion phase was 11 weight %. The organic phase of
the control experiment consisted of 0.32 M oleic acid.
There was no internal phase. Extraction of the feed phase
with this formulation reduced the mercury content to 8.2
mg/L from 460 mg/L. After extraction with microemulsion
liquid membrane, however, the feed phase was reduced to
0.25 mg/L. A second-stage extraction may be required to
reduce mercury levels to a required limit via the micro-
emulsion liquid membrane process.
The demulsification and recovery of mercury via
electrostatic coalescence and butanol addition were eval-
uated by Larson and colleagues (1994). They concluded
that electrical demulsification with heat was not effective for
microemulsiohs due to the small size of the internal phase
droplets in microemulsions. They further reported that
5-5
-------
Figure 5-3. Shematic representation of mercury extraction with an emulsion liquid membrane.
Oil?
Water
ion
exchnger
surfactant
Emulsion
formulation
Aqueous waste stream
with dispersed emulsion
Mercury rich emulsion and
"clean" aqueous phase
microemulsions can be demulsified using butanol as an
an additive. The demulsification kinetics were found to
be proportional to the butanol concentration and the
temperature, and inversely proportional to the surfactant
concentration.
Figure 5-4. Schematic representation of mercury ion extraction with an
emulsion liquid membrane.
Receiving phase
(low pH to strip)
Microdrops
Macrodrops
Hg
Membrane phase (oil)
Feed phase
(high pH to extract)
5-6
-------
Chapter 6
Comparison of Treatment Processes
The preceding chapters describe a variety of mercury
treatment technologies. The most widely recognized full-
scale technologies are precipitation, coagulation/co-
precipitation, and activated carbon adsorption. Other
processes include starch xanthate adsorption, ion ex-
change, reduction, and membrane separation. Relatively
limited full-scale performance data are available for these
technologies and only general comparisons are possible.
This chapter presents brief comparisons of treatment
technologies based on (1) treatment effectiveness, (2)
residuals management, and (3) economics. An example of
a cost comparison from the literature is provided.
6.1 Treatment Effectiveness
Each of the mercury treatment technologies described
in the preceding chapters achieves different effluent
mercury levels. The effectiveness of treatment provided by
each type of technology depends on the chemical nature
and initial concentration of mercury as well as the presence
of other constituents in the wastewater that may interfere
with the process. Table 6-1 summarizes achievable
effluent mercury concentrations. Co-precipitation and ion
exchange achieve the lowest effluent mercury concentra-
tions, ranging from 0.5 to 5.0 /^g/L. Membrane technology
typically achieves 80% to 90% rejection of mercury.
Table 6-1. Summary of Achievable Effluent Mercury Concentrations
Achievable
Concentration (//g/L)
Treatment Process
Sulfide precipitation (+ filtration)
Co-precipitation
Activated carbon
Starch xanthate
Ion exchange
Reduction
Membrane separation
'Membrane technology typically achieves 80 to 90 percent rejection,
which is a measure of salt rejected into the brine.
10-100
0.5-5.0
0.5-20
5-20
0.5-5.0
10->100
a
6.2 Residuals Management
Each mercury treatment technology yields a waste
residual requiring further management.
Types of waste residuals include:
Sludge produced by chemical precipitation and co-
precipitation reactions.
Spent carbon from activated carbon adsorbers.
Concentrated ion exchange regenerant solutions.
Concentrated brine solutions generated from
membrane separation processes.
This section discusses potential options for mercury
recovery from these residuals.
6.2.1 Mercury Sludges
The amount of metal sludge produced from a precipita-
tion or co-precipitation process is typically estimated by
performing treatability studies. The amount of sludge
produced, the mass of mercury within the sludge, and the
physical handling characteristics must be examined for
each treatability method to effectively evaluate sludge
management options. Mercury sludges associated with
precipitation or co-precipitation are typically landfilled,
although thermal processing to volatilize mercury for
recovery would appear to be an option. The MEXICO
process generates voluminous sludge, but mercury
recovery from the sludge for reuse has been proposed
(Macchietal., 1985).
6.2.2 Spent Activated Carbon
Activated carbon columns may be regenerated either
chemically or thermally, but options to concurrently recover
mercury are not addressed in the literature. PAC is not
normally regenerated for reuse due to unfavorable eco-
nomics. GAC regenerated chemically would yield a
mercury-rich regenerant solution. If regenerated thermally,
a mercury-enriched vapor phase would result.
6.2.3 Concentrated Brine Solutions
Ion exchange and membrane processes both yield
concentrated residual mercury solutions requiring further
management. These solutions are typically high in total
6-1
-------
dissolved solids. Reduction or thermal processing for
mercury recovery may be applicable to these brines.
6.3 Economics
As for the performance-related aspects of mercury
control technologies, relative economics are not well
defined. As an example, a cost comparison of the xan-
thate treatment process in comparison with the sulfide
precipitation process is presented below.
A technical-economic evaluation of the xanthate
treatment process in comparison with sulfide precipitation
processes was conducted by Tiravanti, et. al. (1987) for
the treatment of chlor-alkali wastewater.
Investigators assumed a continuous treatment of 50 m3/d
of wastewater containing mercury at 3.0 mg/L, according
to the flow diagram depicted in Figure 6-1. An investment
capital cost of $280,000 (1987 basis) was assumed for all
plants. Cost data derived for the treatment processes are
presented in Table 6-2. The results indicate that total
annual costs, assuming the same disposal route for the
residual sludge, are comparable. The amount of sludge
generated via xanthate treatment, however, may be greater
than by sulfide treatment and this could affect sludge
disposal costs. It should also be taken into account that
the sludge produced by the xanthate process can report-
edly be treated for mercury recovery and reuse.
Figure 6-1. Flow chart of unit operations utilized for technical-economical analysis (Travanti, et. Al. 1987).
Filtration |+\ Effluent |
6-2
-------
Table 6-2. Operating Costs (U.S. $/year, 1987) of Processes for Mercury Removal from Chlor-alkali Wastewaters
(after Tiravanti et al., 1987).
Cost Item
Mexico
Sodium Sulfide
Chemicals
Precipitating reagent
Polyelectrolyte
Ferric sulfate
Sludge Treatment
Lime
Ferric chloride
Transportation*
Sludge disposalb"
Maintenance"
Manpower*
Electric energy'
Total
$ 7,267
10,200
267
333
933
5,533
8,400
20,000
24,533
$77,466
$ 466
2,067
733
1,134
1,667
4,600
27,400
8,400
20,000
24,533
$91 ,000
'Average distance = 30 km; unit cost = 16.7 #/kg.
'Including inertization and disposal of sludge unit cost =
C3 percent of the total investment costs.
dPlant is operated by the chloride production team; manpower has been increased by one operator.
'Average power consumption = 0.6 KWh/m3; unit cost = 9.33 0/kWh.
6-3
-------
Chapter 7
Case Study
7.1 Ion Exchange Removal of Mercury from
Wastewater at DOE's Savannah River Site
Effluent Treatment Facility
DOE's Savannah River Site (SRS) is a facility that
produces nuclear materials for national defense. The SRS
houses two separation areas where uranium and plutonium
are separated from fission products. Mercury is used at
SRS as a catalyst in the dissolution of fuel elements
composed of uranium-aluminum alloys. As a result,
mercury is present in varying concentrations in some SRS
wastestreams. Mercury is but one of several constituents
being addressed in a system involving various treatment
processes. Mercury is primarily removed by a cation
exchange treatment system.
Depending on the stream, mercury may be present in
solution as Hg', Hg^+. The site's Effluent Treatment
Facility (ETF) is designed to treat several dilute waste-
streams associated with the nuclear materials operation.
The ETF uses cation exchange resins to selectively
remove mercury in each oxidation state. Hg2* and Hg22*
are removed by typical ion exchange mechanisms. Hg' is
believed to be removed either by oxidation via dissolved
oxygen and subsequent ion exchange, or by van der Waals
attraction between Hg° and the matrix of the resin.
The discussion in this section focuses on (1) the ETF,
(2) ETF feed streams, (3) ETF discharge limits, and (4) the
role of ion exchange at the ETF. Actual ETF treatment
data are not presented in this report because they were
unavailable; however, the effluent limitations on which the
ETF system was designed are described.
7.1.1 The ETF
Treatment at the ETF involves a four-stage process of
filtration, organics removal, RO, and ion exchange. A
process flow diagram for the ETF is shown in Figure 8-1.
Process influent feed is collected in two 450,000-gal feed
tanks, which normally operate at 40% to 50% capacity. No
agitation is provided in the tanks. Mixing is accomplished
by recirculation, using a sump pump. The pH of the
influent wastewater typically ranges from 2.5 to 5.0;
however, pH values in the range of 1.5 to 13 are observed.
Nitric acid is added to the wastewater in the feed tank to
reduce the pH to the range of 2.0 to 2.5. Aluminum nitrate
[Al (NO3)3] (available on site) is added in liquid form to yield
a final concentration of Al3* of 35 mg/L. Al (NO3)3 is added
as a coagulant and, reportedly, to reduce biofouling in the
treatment system.
The wastewater is pumped from the feed tank into two
pH-adjustment tanks that operate in series. The volume of
the first tank is 1,500 gallons and the volume of the second
tank is 2,500 gallons. The pH is adjusted to 3.5 in the first
tank and to 7.5 in the second tank to precipitate ions
present in the feed, such as Al3* and Fe3*. These metal
precipitates are removed by the filtration system, which is
located upstream of the organic removal, RO, and ion
exchange systems. Some mercury removal may occur at
this point via co-precipitation. The wastewater flows from
the second pH-adjustment tank into a 2,500-gal filtration
feed tank that operates 50% full. The wastewater is then
pumped from the feed tank to the filtration system, which
includes a-alumina membrane filters operating in the
crossflow mode. The filtration system operates at a
wastewater flux ranging from 200 to 500 gpd/ft2. A feed
pump provides the required filtration pressure. The
pressure differential is approximately 30 to 40 psi. The
filtration system includes three treatment trains, each with
three stages in series. Each stage has four parallel
modules. Each module has two bundles of 10 filters. The
filtration system has recirculation on each stage and
blowdown on the third stage. During each filtration run, the
filter is backpulsed for about 30 seconds every 10 minutes.
As headloss increases and flux decays, the filtration
system is taken offline to be cleaned. The filtration tubes
are cleaned first by oxalic acid and then by caustic and
household bleach before they are put back online. In the
past, irreversible biofouling of the tubes was observed and
the tubes were replaced after being online for 3 years.
The concentrate (brine) exiting the filtration system
goes into the filter concentration tank and then to the
evaporator feed tank. The filtrate flows out of the filtration
system radially to the organic removal feed tank, which
also receives overheads from the evaporator. The organic
removal system was not originally included in the ETF. RO
was the main component of the process, intended to
remove cesium. It was subsequently determined, however,
that the waste water contained organics (e.g., tributyl
phosphate) that interfered with the performance of the RO
system. To remove the organics, the wastewater is now
pretreated by activated carbon columns. However, acti-
vated carbon accumulated mercury, which was present in
the wastewater, so ion exchange columns were added to
remove mercury before activated carbon treatment. The or-
ganics removal system includes three Duolite GT-73 ion e-
xchange columns operating in parallel, and three activated
carbon columns operating in series. Two ion exchange
7-1
-------
Figure 7-1. F/H effluent treatment facility process flow sheet.
Process Waste Collection
HN03 AI(N03)3
Treated water
Storage
Receives untreated
influent from F/ H
separations and
waste management
facilities
pH adjust- Filter Organic
Tient tank feed tank Filtration removal
k -H r
Concentrate
ed
> '
snt
+. Hg Carbon
columns columns
Reverse
osmosis
4_
Concentrate
_>
k
Evaporation
i u
Ion
exchange
Hg
columns
Cation
columns
Regenerant
i
r
To Upper
Three Runs
creek
Overheads
(vapor phase)
Waste concentrate
Bottoms
(brine)
To discharge
columns are in service and one is on standby (flooded
status). Each column is 4 feet in diameter, 5 feet high, and
contains 40 ft3 of resin. The columns operate in a down-
flow mode. The wastewater is fed at 135 psi and the
pressure drop is 15 psi. The ion exchange columns
operate on the sodium cycle instead of the hydrogen cycle
to avoid the need for downstream pH adjustment. The ion
exchange columns are prewashed with NaOH in order to
convert them from the hydrogen to the sodium cycle.
Spent NaOH is directed to the evaporator. Regeneration
of the ion exchange columns has not been required for
about 4 years, since the system became operational.
Headloss usually occurs because of biofouling. The head-
loss problem has diminished significantly, however, since
the use of AI(NO3)3 was initiated.
The ion exchange capacity of the resins is reported to
last for about 2 to 3 years. The effluent of the columns is
monitored weekly for mercury. When the ion exchange
capacity of a column is exhausted, it is taken out of service
and buried on site, since the resin exceeds Toxicity
Characteristic Leaching Procedure (TCLP) requirements.
Effluent from the ion exchange columns is directed to
activated carbon columns. The granular activated carbon
has a size 16 mesh and contains fines of 0.1% by volume.
Each column is 16 feet tall, has a diameter of 10 feet, and
contains 23,000 pounds of carbon. The activated carbon
columns operate in the downflow mode. When there is a
high pressure drop, the columns are pulsed with air and
then are backwashed with water. The backwash water
goes back to the head of the system. The activated carbon
columns are backwashed every 2 weeks because of high
headless, believed to be due to biofouling.
The wastewater exiting the organic removal system
enters a spiral wound RO system. The pH is reduced to a
range between 5.5 and 6.5 ahead of the RO system, using
dilute HNO3. Each RO membrane represents an 8-in.
diameter module and is 40 in. long. The RO membranes
used are seawater-type, high-rejection membranes.
Reportedly, they have the following decontamination (i.e.,
concentration) factors: (1) 20 for monovalent ions, such as
Na+, (2) 40 for divalent ions, such as Ca2* and SO42', and
(3) 60 for trivalent ions, such as PO43~. The decontamina-
tion factor is defined as the ratio of concentration of ions in
the feed over the concentration of ions in the product.
The RO system has two modules operating in parallel.
The concentrate from these modules is directed to a third
module, and the concentrate exiting the third module
enters a fourth module. Oxalic acid and sodium metabisul-
fite are used to clean the RO membranes. The RO
membranes are typically cleaned monthly, after they have
treated 1.0 to 1.5 million gallons of wastewater. The
wastewater from the RO system goes to a pH-adjustment
system, where NaOH is added to adjust the pH in the
range of 6 to 9.
Cation exchange columns are used at the end of the
treatment train as a final polishing step. The wastewater
first enters two cation exchange columns operating in
parallel for additional mercury removal and then enters two
7-2
-------
cation exchange columns intended for the removal of
cesium and strontium. These columns operate on the
sodium cycle and are regenerated by NaNO3 solution. The
regenerant is directed to the evaporator feed tank.
Treated wastewater exiting the ion exchange polishing
step goes into three 120,000-gal storage tanks. The
following parameters are monitored: pH, specific conduc-
tivity, beta and gamma radiation, oil and grease, Pb, Cu,
Zn, and Hg. Treated wastewater that does not meet the
discharge limitations is retreated. Effluent meeting the
discharge criteria is discharged to a surface receiving
water. The evaporator bottoms go to a tank farm where
they are solidified and buried.
7.1.2 ETF Feed Streams
Feed to the ETF is effluent from the SRS separation
areas, evaporator condensate, and other waste streams,
including contaminated cooling water, surface water runoff,
and miscellaneous laboratory wastes. The normal sources
of wastewater are evaporator overheads and effluent from
the separation areas. The ETF is also designed to treat
cooling water that may have accidentally become contami-
nated in the separation process. In addition, the ETF is
designed to treat contaminated water that might result if a
leak in a waste storage tank or transfer system should
occur simultaneously with a rainstorm. Although the ETF
is designed to handle an average flow of 288,000 gallons
per day, it routinely handles a flow of 90,000 gallons per
day because the volume of wastewater generated has
decreased over the years.
Table 7-1 lists major nonradioactive constituents and
Table 7-2 presents radioactive contaminants present in the
wastewater.
7.1.3 ETF Discharge Limits
There are two sets of applicable discharge limits for the
ETF. The first set of discharge limits is set by the state of
South Carolina. These limits are generally related to water
quality guidelines for conventional nonradioactive waste
(see Table 7-1). The second set of discharge limits
involves guidelines established by DOE for the discharge
of radioactivity from the ETF (see Table 7-2).
7.1.4 Ion Exchange Role in the ETF
Two types of ion exchange columns are used in the
ETF system: mercury removal ion exchange columns, and
cesium and strontium removal ion exchange columns. The
mercury removal ion exchange resin is Duolite GT-73
cation exchange resin, which has thiol functional groups.
The cesium and strontium ion exchange resin is a macro-
porous sulfonic acid cation exchange resin. Each of these
applications is discussed below.
Duolite GT-73 Resins
Three columns containing Duolite GT-73 resins
operating in parallel are located between the filtration
system and carbon columns. Two identical columns
operating in parallel are located between the RO and the
sulfonic acid resin columns. Duolite GT-73 resin has been
reported to be very efficient in the removal of heavy metals,
especially mercury (Ritter and Bibler, 1992). Duolite GT-73
is a macroporous, weakly-acidic polystyrene/ divinyl
benzene cation resin with thiol functional groups that have
a pronounced selectivity for mercury in any of its three
common oxidation states. The following ion exchange
reactions are postulated for the removal of Hg2" and Hg22+
(Bibler etal., 1986):
R-S2H2 + Hg
,2+
R-S2H2
Hg22+
R-S2Hg
R-S2Hg2
2H+
2H+
[7-1]
[7-2]
Two plausible mechanisms that address the sorption of
Hg° have been proposed. The basis of the first mecha-
nism is as follows: The solubilities of Hg° and HgO in air-
free water are 60 ppb and 52 ppm, respectively. The
solubility of Hg° in an aqueous solution with dissolved air
(oxygen), however, increases by a factor of 700 compared
to its solubility in air-free water, and the final solubility is
nearly the same as that for HgO (Ritter and Bibler, 1992).
These facts suggest that Hgc can be partially ionized by O2
in an aqueous solution and thus be converted into a form
that can be removed by the Duolite GT-73 resin. Accord-
ing to the second mechanism, sorption of Hg° by the
Duolite GT-73 resin may also be explained by the van der
Waals attraction between metallic mercury (Hg°) and the
matrix of the resin, thereby interacting with the resin by
physical absorption rather than ion exchange (Ritter and
Bibler, 1992).
The Duolite GT-73 resin reportedly operates over a pH
range of 1 to 13, a much broader range than for other
commercially available mercury-selective resins. The
Duolite GT-73 resin remains physically and chemically
stable when exposed to ionizing radiation and is insoluble
in most common solvents, but decomposes slowly in HNO3
solution of greater than one molar concentration. The
manufacturer's reported ion exchange capacity of the resin
for Hg2* is 1.4 meq/mL (Bibler et al., 1986).
Three alternatives have been studied at the ETF for
handling the exhausted Duolite GT-73 ion exchange resins.
These alternatives are (1) storage of mercury-containing
resins as such, (2) storage of mercury containing resins
incorporated into grout, and (3) recovery of mercury
desorbed from the resin. Samples of Duolite GT-73 resin
were saturated with mercury and analyzed by the EP
toxicity test. The spent resin passed the EP toxicity test,
7-3
-------
Table 7-1. Nonradioactive Contaminants and Discharge Limits (After Bibler and Wallace, 1987).
Influent to ETF (ppm)
Influent to Ion Exchange (ppm)
Proposed Limits (ppm)
Ion
NH',
Hg2*
Zn2*
Cr3
Cu2*
Pb2*
Mn2*
NO,
Average
16
0.053
1.1
0.031
0.14
0.15
0.18
1,015
Maximum
110
10
100
240
18
38
21
22,400
Average
4.0 x 10''
1.3 x 10"
2.8 x10'2
7.8x10-"
3.5 x10'3
3.8 x10-3
4.5 xlO'3
25
Maximum
2.8
2.4x10"'
2.5
6.0
5.0 x 10-'
9.5 x IO"
5.3 x 10''
560
Average
20
4.5 x 10"'
1.48
1.71
1.30
2.2 x10'1
Maximum
1.75x ID'1
2.61
2.77
1.89
4.5 x 10-'
Table 7-2. Radioactive Contaminants and Discharge Limits (After Bibler and Wallace, 1987).
Influent to Ion
Radionuclide
cs- 134,137
Sr - 89, 90
Co-60
Other p-y
Total a
mCi/yr
70
9
5
3.0
3.5 x IO"
Exchange
mCi/mL
1
2.
1.
,7x
2 x
2x
7.5 x
8
.7x
10-'°
10'"
10-'°
io-'2
10 3
Release to Streams
mCi/yr
0.7
0.09
0.5
3.0
3.5 X10'3
mCi/mL
1
2
1
7
8
.7
.2
.2
.5
.7
x
x
x
x
x
io-'2
io-'3
io-'2
io-'2
io-'5
DOE Guide
mCi/mL
2.9 x 10'5
3.3X10-6
3. Ox 10'5
indicating that simple storage was a viable option. Similar
resin samples were incorporated in Portland Type II grout
and subjected to structural integrity tests and the Extraction
Procedure (EP) toxicity test. These samples passed the
structural integrity test but not the EP toxicity test. Because
the EP toxicity test involved maintaining the sample at pH
5 with acetic acid and given the high concentration of
calcium ions in grout, exchange of calcium ions for mercu-
ric ions may have taken place. Thus, the storage of spent
resin without incorporation into grout was preferred.
Although regeneration of the spent resin is possible, it
is not deemed desirable at ETF. Mercury can be eluted
from the resin using 3 M HCI or 2 M NaSCN, neither of
which is chemically compatible with materials of construc-
tion or processes at SRS. Dissolution of the resin and
reclamation of mercury by chemical means such as
precipitation as the sulfide or reduction to the metal is an
attractive alternative to storage, should recovery and
removal become desirable.
Sulfonic Acid Cation Resins
A macroporous, strong acid cation exchange resin was
chosen for removal of cesium and strontium. Several
commercially available resins have demonstrated cesium
and strontium removal capabilities coupled with ease of
regeneration. Cesium and strontium in the regenerant can
by a relatively small volume of neutral reagent be concen-
trated further and incorporated in concrete for final dis-
posal. Spent resin can be decontaminated and discarded
in an approved manner.
Several test runs were conducted to determine the
performance of the sulfonic acid resin throughout several
simulated feed, wash, and regeneration cycles. Results
indicated that the effectiveness of the process was less
7-4
-------
than desirable after several cycles had been completed. A
chromatographic effect was observed where concentrated
bands of all metals present were detected in the effluent at
unpredictable times during feed cycles. The frequency and
concentrations of such eluted bands cannot be accurately
predicted in the ETF due to the varying daily concentrations
of influent to the facility. To prevent this behavior the ion
exchange feed was first processed through the Duolite GT-
73 columns for mercury removal, allowing the sulfonic acid
columns to operate as designed for the removal of cesium
and strontium.
7.2 Effluent Treatment Facility Economics
The treatment plant operates Friday through Sunday of
each week. There are 5 operators during each shift, which
lasts 12 hours, with an additional 0.5 hour turnaround.
During the 1994 fiscal year, 22 million gallons of waste
water were treated at a total cost of $18.8 million. This
results in a unit cost of about $1/gallon. Figure 8-2 pres-
ents the ETF cost breakdown components.
Figure 7-2. ETF facility O&M cost breakdown.
Facility support
$6.5 M
(34.7%)
A = Health protection - $0.85 M - (4.5%)
B = Materials/chemicals - $0.4 M - (2.1%)
C = Central services - $0.13 M - (0.7%)
7-5
-------
Chapter 8
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