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

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                                      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

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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

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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

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 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

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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

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                                              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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-------
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                                                    8-3

                                                        &U.S. GOVERNMENT PRINTING OFFICE: 1997 - 650-001/80154

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