Download - [Advances in Chemistry] Aquatic Humic Substances Volume 219 (Influence on Fate and Treatment of Pollutants) || The Fate and Removal of Radioactive Iodine in the Aquatic Environment

35 The Fate and Removal of Radioactive Iodine in the Aquatic Environment

R. Scott Summers1, Friedrich Fuchs, and Heinrich Sontheimer

Engler-Bunte-Institut, Universität Karlsruhe, 7500 Karlsruhe, Federal Republic of Germany

The reaction of iodine with aquatic humic substances (HS) and the subsequent removal of the products by typical drinking-water-treatment processes was investigated. Both iodine and iodide react completely with isolated HS in the concentration range below 0.03 mg of I per mg of HS and behave similarly with Rhine River water. The reaction is independent of pH, initial HS concentration, and HS molecular size. However, at higher I-HS ratios iodine reacts slightly more than iodide. Kinetic studies indicate that the reaction is complete within 10 min. No interaction was found between methyl iodide and HS. Flocculation and activated-carbon (AC) adsorption were effective for the removal of the I-HS complex, and the dissolved organic carbon measurement served as a good surrogate parameter. Volatilization and AC adsorption were effective for methyl iodide removal.

/JLFTER A NUCLEAR REACTOR ACCIDENT the release of radionuclides poses a problem for drinking-water-treatment facilities using surface waters as their raw water source. A recent nuclear power plant accident resulted in high levels of radioactivity in the environment throughout Europe, as shown in Table I for the Federal Republic of Germany. The highest activity levels occurred in the southern part of the country where, fortunately, 95% of the potable water originates from ground-water aquifers that were not directly contaminated. However, for communities that use surface waters, an un-

1Current address: Civil and Environmental Engineering Department, University of Cincinnati, Cincinnati, O H 45221-0071

0065-2393/89/0219-0623$06.00/0 © 1989 American Chemical Society

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5

In Aquatic Humic Substances; Suffet, I., et al.; Advances in Chemistry; American Chemical Society: Washington, DC, 1988.

624 AQUATIC HUMIC SUBSTANCES

Table I. Radioactivity Levels in the Federal Republic of Germany after the Chernobyl Accident

Maximum Reported Main Ref. Source Activity Component Ref.

Air (Bq/m3) 150 1-131 1 Ground (Bq/m2) 280,000 1-131 1 Aqueous (Bq/L)

Rain 35,000 1-132 or 1-131 2 River 370 1-131 2 Reservoir 570 1-131 2 Ground <1 — 2 Lake 360 1-131 2 Sewage 1,600 1-131 2

Sediments (Bq/kg) River 18,900 Cs-137 2 Lake 6,000 Cs-137 3 Drinking-water

treatment sludge 104,000 Ru-103 4 Sewage sludge 780,000 Ru-103 2

Biota (Bq/kg) Fish 1,640 Cs-137 3 Plankton 38,000 Cs-137 3

derstanding of the reduction in radioactivity during drinking-water treatment (DWT) is of great importance.

The high activity levels in both D W T and sewage sludges, shown in Table I, qualitatively indicate that the processes commonly employed are removing some of the radioactivity. Â survey of Swiss water-treatment plants showed that flocculation and filtration were able to reduce the radioactivity by 50-90% (5). Table I also shows that the main atmospheric and aqueous component was 1-131, while the activity accumulated in sediments and biota are mainly from Cs-137 and Ru-103. This situation is partially due to the relatively short half-life, 8 days, of 1-131 and the time periods before the activity of sediments and biota are measured. However, residence time in a water-treatment plant and distribution system is only a few days; this condition makes 1-131 a problem for communities using surface waters. The proposed U.S . Environmental Protection Agency maximum recommended level for M 3 1 in drinking water is 3 p C i / L , that is, 0.11 B q / L (6).

The objective of this study was to determine the effectiveness of commonly employed treatment processes for the removal of radioactive iodine from drinking-water sources. In the first phase of the study, the reaction between iodine and aquatic humic-substances was investigated. In the second phase, the removal of radioactive iodine in the concentration range below 1000 B q / L (1.0 mBq/m 3 ) by bench-scale processes was examined. The dissolved organic carbon (DOC), U V absorption (λ = 254), and turbidity

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5


35. SUMMERS ET AL. Radioactive Iodine in the Aquatic Environment 625

were also measured in order to find an indicator parameter that could be measured more easily than radioactivity.

One problem encountered was the determination of the appropriate species of iodine to use in such an investigation. Metal iodide is the chemical form that escapes from the reactor core (7); however, the form of iodine predominating after exposure to the atmosphere is not completely understood. The form is thought to be dependent on the conditions in the containment building; the volatile elemental iodine and methyl iodide are important with respect to atmospheric release.

Experimental Details The humic substances used in this study were isolated by a strong basic anionic resin (Lewatit MP 500 A, Bayer Chemical Co.) used in the treatment of ground water with a high humic content (7 g of DOC/m3) in Fuhrberg, Federal Republic of Germany. The resins were regenerated with a solution containing 10% NaCl and 2% NaOH (8). The regenerate has a molecular size (MS) range of 200-4000, with an average of 1500 as estimated by gel-permeation chromatography (GPC). The experimental solution properties for the Fuhrberg humic-substances (FHS) are shown in Table II.

The GPC of the Rhine River sample indicates a MS range of 160-5000, with an average of about 1500. Samples of the Rhine River were taken at Karlsruhe, Federal Republic of Germany, approximately 360 km downstream of Lake Constance. Their properties are also shown in Table II. The radionuclide 1-131 was supplied in a carrier solution of Nal at a ratio of 4.72 Χ 104 Bq^g of I, and the solution activity was analyzed by 7-ray spectrometry. The concentrations of I" and I2 were analyzed by the leuco crystal violet method (6) with a detection limit of 5 mg/m3 in the Rhine River and FHS solutions. The DPD (N,N-diethyl-p-phenylenediamine) photometric method was used to measure chlorine (6). The concentration of CH3I was measured with an electron-capture detector-gas chromatograph with a detection limit of less than 0.1 mg/m3.

The reaction experiments between I~, I2, or CH3I and the FHS or Rhine River water were conducted in 0.1- or 0.25-L closed volumetric flasks. Flocculation was conducted in 1-L glass beakers with the addition of iron sulfate at high mixing intensities (250 rpm) for 5 min, followed by 10 min of flocculation at 50 rpm and 30 min of settling. The cationic polyelectrolyte, poly aery lam ide (PAA), when used, was added 2.5 min after the addition of iron sulfate. Filtration utilized glass-fiber filters or 0.45-μπι membrane filters. Adsorption experiments were conducted with closed 0.25-L bottles on a shaker-table (250 rpm) at a contact time of 2 days with pulverized activated carbon (F300, Chemviron Corp.). In the combined floccula-

Table II. Properties of Fuhrberg Humic Substances and Rhine River Sample DOC UV-254 Redox Temperature Turbidity

Sample (g/m3) (m-*) pH (mV) (°C) (FTUa) Rhine River 2.62 6.52 7.65 240 19.6 2.9 Fuhrberg humic substances 4.20 16.3 6.5 515 20 — "Formazine turbidity units.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



tion-activated-carbon adsorption experiments, the pulverized carbon was added to the beaker at high mixing intensities 10 min prior to the addition of iron sulfate; it was filtered out with the sludge; the result was a 1-h contact time.

Results and Discussion Reaction with Humic Substances. The reaction kinetics of ele

mental iodine, I 2 , and iodide, I", with F H S over a 2-week period are shown in Table III. A l l initial solutions of I 2 used throughout this study contained 27% iodide. For both iodine and iodide the reaction with F H S is very fast, with no additional reduction in solution concentration occurring after 10 min. The reaction may even be faster, but this possibility could not be assessed, as 10 min was required for analytical sample preparation. The reaction kinetics of iodine and chlorine with the F H S can be seen in Figure 1. The iodine reaction is much faster than that of chlorine, which displays continued formation of organic-bound chlorine over a 17-h period.

A l l nonorganic-bound iodine is reduced to iodide in the reaction between 1 2 and F H S , as shown in Table HI . This result can also be seen in Figure 2, where the initial concentrations of iodine were five times higher than in Table HI . In the system of I 2 and F H S , the total amount of iodine in solution decreases from 4.38 to 1.02 g / m 3 after 0.5 h; the I 2 component decreases from 3.20 to 0.23 g /m 3 . After 3.5 h the total amount of iodine does not change, but the I 2 component is completely reduced to iodide. In the system with both I" and C l 2 , all iodide is initially oxidized to I 2 . After reaction with F H S (0.5 h), most of the iodine left in solution has been reduced to iodide. After 3.5 h all solution iodine is in the form of iodide, but no additional organic-bound iodine was formed.

The relationship between organic-bound iodine and the added or total system iodine is shown in Figure 3 for both elemental iodine and iodide. Below a total iodine concentration, normalized for the D O C concentration of F H S , of 0.05 g of I per g of D O C , both forms of iodine react completely with the F H S . This reaction represents iodine concentrations as high as

Table ΙΠ. Reaction of Iodine and Iodide with Fuhrberg Humic Substances

Time Iodine (l2) Iodide (/)

Time h I 2 / h / - Σ / 0 650 240 890 ndf l 650 650 10 min nd 370 370 nd 280 280 8 h nd 400 400 nd 270 270 1 day nd 380 380 nd 285 285 4 day nd 400 400 nd 290 290 14 day nd 400 400 nd 290 290 NOTE: All values are concentration in milligrams per cubic meter. Solution conditions: 3.2 g of DOC/m3; pH 6.5. end, not detected.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



100

£S 80

X I

ο

c ο ο

" Ό

c D Ο m

60

40

20

Fuhrberg humic substances (FHS) / pH 6.5

C l 2

7* JPr~~"^ — . -o ^

(g /m 3 )

l l 2 0.89

f C l 2 2.42

1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1

DOC 3.20

0 15 5 10

Time, t (h)

Figure 1. Halogen and Fuhrberg humic substances reaction kinetics.

20

E en

c ο

D 3

c <D Ο C ο 2 ο φ c

~° 1 ο ι σ ο

Fuhrberg humic substance (FHS) / pH 6.5

r r+cu

0.0 M m. Τ777Λ

0.5 Time, t (h)

• r

C 0 (g/m 3 ) DOC 6.5 C l 2 3.0

3.5

Figure 2. Reaction kinetics and iodine species distribution for the reaction of I2, l'y or I~ + CU with Fuhrberg humic substances.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



0.4

\ cn

Ο Ο α

0.3

I <-> 0.2

CD C.

Ο

"D C 3 Ο

CD

0.1

Fuhrberg humic substances ( H S (

l

1 0 0 %

b o u n d

FHS C 0 (gDOC/m 3)

3.20 6.40

â

pH 3.5 6.5 6.5

Ο Δ Τ

0 0.2 0.4 0.6 C.8

Total iodine, l /DOC (g /g )

Figure 3. Organic-bound iodine as a function of the total iodine in the system with Fuhrberg humic substances.

160 m g / m 3 for a F H S concentration of 3.2 g of D O C / m 3 . At higher total iodine concentrations the percent of total iodine bound to the F H S decreases, but in the concentration region below 0.8 g of I per g of D O C , a saturation of the F H S by iodine is not found. Elemental iodine is slightly more reactive than iodide, but the reaction is independent of p H and F H S concentration, as shown in Figure 3. However, Skogerboe and Wilson (9) found that low p H values reduce the extent of reaction between a soil fulvic acid and both I 2 and I 3~.

F H S was separated by ultrafiltration into high (>1000) and low (<1000) molecular size (MS) fractions. The reaction of both I 2 and I" with these MS fractions was examined. For both high and low MS fractions at two levels of iodine addition, as shown in Table IV, both I 2 and Γ reacted to an extent similar to that of the unfractionated F H S .

Two doses, 50 and 100 m g / m 3 each, of I 2 and I" were added to samples of the Rhine River water. One sample of the Rhine water had been glass-fiber filtered prior to the experiments. At the lower dose both I 2 and I" reacted completely with both the filtered and unfiltered Rhine water. At the higher dose both I 2 and I" reacted completely with the unfiltered Rhine water, but reacted slightly less, 93%, with the filtered sample. This decrease indicates that part of the reaction is with particulate organic matter found in the Rhine River, and that iodine reacts to a similar extent with both F H S and the organic matter in the Rhine River. Several mechanisms have been proposed for the reaction between iodine and organic compounds, including

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



electrophilic substitution, charge-transfer, and biochemical oxidation (10-12). However, the specific mechanism involved with humic substances is not clear because of the complex nature of this heterogeneous macro-molecular material.

The interaction of methyl iodide and F H S is shown in Table V. Although both sample concentrations, with and without F H S , decrease with time, no significant difference exists between the samples at a given time. This similarity indicates little interaction between C H 3 I and F H S . The decrease in solution concentration for both samples is an indication of the volatility of C H 3 I . It also appears from this data that F H S has no effect on the Henrys coefficient, as the liquid-to-gas volume ratio of both samples was the same.

Removal by DWT Processes. RHINE RIVER AND 1-131. The radionuclide 1-131 as Na l was added to a sample of the Rhine River water to yield an activity level of 1.17 m B q / m 3 (1170 Bq/L) , which is a factor of 3 larger than any river water value reported in the Federal Republic of Germany after Chernobyl (Table I). The sample was mixed for 0.5 h before chlorine was added at a concentration of 1.0 g of C l / m 3 , which raised the redox potential from 240 to 350 mV. Preoxidation is commonly practiced in D W T

Table IV. Influence of Fuhrberg Humic Substance Molecular Size on Organic-Bound Iodine

C û Bound Iodine, ClI (%) I (mg/m3) FHS High MSb Low MSC

I 2 274 88 81 85 890 58 62 58

i - 200 90 90 85 650 57 46 52

NOTE: Solution conditions: 3.2 g of DOC/m3; pH 6.5. "Initial C concentration. ^Molecular size >1000. 'Molecular size <1000.

Table V. Interaction Between Methyl Iodide and Fuhrberg Humic Substances

Time (h) Without FHS With FHS 0.5 42.0 41.1 1.0 — 39.0 5.5 39.2 38.4

22.0 34.9 34.6

NOTE: All values are CH 3I concentration in milligrams per cubic meter; solution conditions: 3.2 g of DOC/m3; 1.0 mM Na2HP04; pH 6.5.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



plants that directly utilize surface waters to control problems associated with biological growth.

The results of the flocculation experiments with iron sulfate are shown in Figure 4. The results with no addition of iron sulfate indicate the removal by volatilization and sedimentation. The maximum removal effectiveness by flocculation for 1-131 and D O C is about 30%. Effectiveness nearly doubles for UV254 and triples for turbidity. The addition of PAA at 0.1 g / m 3 with an iron sulfate dosage of 15 g / m 3 did not improve the removal effectiveness for any of the measured parameters. Similar removal values for 1-131 and D O C and higher removal values for turbidity indicate that most of the 1-131 has reacted with dissolved organic matter and not particulate organic matter. This finding is also supported by the filtration results shown in Table VI . Both glass-fiber and 0.45-μπι membrane filters are effective in the removal of turbidity, but significantly less so for 1-131, D O C , and U V ^ . With all

1 0 0

ο > ο Ε

-*-> c Φ ο ω

Rhein River

Turbidity 0 .90 FTU

U V 2 5 4 6.52 m " 1

1 3 1 l 1.17 M B q / m 3

DOC 2.62 g / m 3

5 10 15

Iron sulfate concentration ( g F e 3 + / m 20 25

3+ / _ 3 \ 30

Figure 4. Removal effectiveness of iron sulfate flocculation of the Rhine River sample.

Table VI. Removal by Filtration of the Rhine River Sample

Parameter Glass Fiber Membrane 1-131 7.6 11 DOC 3.0 4.6 UV-254 8.0 11 Turbidity 69 80 NOTE: All values are percents.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



parameters, membrane filtration is slightly more effective. The removal as measured by U V ^ most closely matches that of 1-131, while D O C is slightly less removed.

Adsorption by activated carbon (AC) in the dosage range of 5-1000 g of A C per m 3 is shown in Figure 5 for the Rhine River sample prior to pretreatment. At nearly all dosages, the adsorption as measured by UV254

and D O C is greater than that of 1-131, although for D O C the difference is normally less than 15%. The maximum removal at 1000 g of A C per m 3 is 70%, 83%, and 97% for 1-131, D O C , and U V ^ , respectively. However, when A C is applied after flocculation, as shown in Figure 6, removal by adsorption is increased. This result can be seen by comparing the removal at a dosage of 100 g of A C per m 3 (Table VII). Prior to flocculation this dosage of activated carbon results in a removal of 57%, 71%, and 87% for 1-131, D O C , and U V 2 5 4 , respectively. After flocculation the respective removals by adsorption increased to 79%, 76%, and 91%. The removal of 1-131 after flocculation is paralleled by that of D O C , for both adsorption and the combined results of flocculation followed by adsorption, as shown in Table VII.

Another method of applying activated carbon is the addition of powdered activated carbon (PAC) during the flocculation process. In this alternative approach the PAC is added to the rapid-mixing tank prior to the addition of coagulant aids and settled out with the sludge in the sedimentation basin. The results of this simultaneous process are shown in Figure 7. Contact

D > Ο

c ο

Q_

100

80

60

20

0 (

10

Rhein River

• AC: F300 uv254

t c= 2 days

s A — ^ _ 1311

/ // Co

1 3 1 l 1.17 M B q / m 3

— i — 1 i 1 i 1 1 1 1 1 1 I

DOC 2.62 g / m 3

U V 2 5 4 6.52 m " 1

1 1 1 11 1 1 1 1 1 1 1 1 1 3 4 5 6 7 8 9 1

10 ' 3 4 5 6 7 8 9 1

1 0 2

3 4 5 6 7 8 9 1

1 0 3

Activated carbon dosage, D (gAC/m 3 )

Figure 5. Activated-carbon adsorption with the Rhine River sample.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



Table V H . Removal by Activated Carbon with the Rhine River Sample

Pretreatment 1-131 DOC None 57 71 87 Flocculation

Adsorption 79 76 91 Combined 85 83 m Simultaneous adsorption and flocculation 82 75 91

NOTE: All values are percents; activated carbon dosage: 100 g/m3.

time for the PAC process is short, 1 h in the case of Figure 7, compared to the adsorption residence time in a typical granular-activated-carbon column, modeled by the 2-day contact time in Figures 5 and 6. Even at this short contact time, the PAC-flocculation results in comparable removal, as can be seen in Table VII. The results for all three parameters after a 1-h contact time are nearly the same as those of the combined removal of flocculation followed by adsorption at a 2-day contact time.

The good removal achieved with the simultaneous PAC-flocculation process has significant implications for D W T plants that do not utilize granular-activated-carbon adsorbers, but do include sedimentation. In emergency situations of high levels of radioactivity or other contaminants, PAC could be added to the flocculation system and settled out in the sedimentation

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



100

80

60

40

Rhein River ° U V 2 5 4 " AC: F300 ° U V 2 5 4

t c= 1 h ^ _£) 131J

" DOC

Co 131J 1.17 M B q / m 3

DOC 2.62 g / m 3

uv254 6.52 r rT 1

15 g F e 3 + / m 3 and 0.1 g P A A / m 3

« 1 ι I ι I ι 20 40 60 80 100

Activated carbon dosage, D (gAC/m 3 ) 120

Figure 7. Simultaneous activated-carbon adsorption and flocculation of the Rhine River sample.

basin with the sludge. This procedure would require the ability to add the PAC and to handle increased amounts of a now-hazardous sludge. This approach could also be applied to wastewater-treatment plants.

FUHRBERG HUMIC SUBSTANCES AND 1-131. 1-131 was also added to a solution of F H S with the properties shown in Table II, to yield an activity level of 1.17 m B q / m 3 . The bulk solution was not chlorinated. Membrane filtration of the solution after mixing yielded no removal of the three parameters measured: 1-131, D O C , and U V ^ . Aeration in a 0.5-L bubble column for 20 min yielded an insignificant removal of 5% and 4% for 1-131 and D O C , respectively, a result that indicates that the I - H S complex is not any more volatile than the original F H S .

Removal effectiveness by activated-carbon adsorption in the dosage range 5-1000 g of A C per m 3 is shown in Figure 8. Adsorption was found to be less effective for the F H S in the low dosage, <100 g of A C per m 3 , than for the Rhine River sample.

However, activated carbon was more effective in the high-dosage range, >250 g of A C per m 3 , with removals of 94%, 86%, and 99% for 1-131, D O C , and UV254, respectively, at an adsorbent dosage of 1000 g of A C per m 3 . Again D O C seems to be an adequate indicator of the filtration, volatilization, and adsorption behavior of 1-131.

FUHRBERG HUMIC SUBSTANCES AND METHYL IODIDE. As shown in Table V, methyl iodide does not react with the F H S . However, the data in

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



this table indicate that methyl iodide is volatile and has an estimated Henry's constant, f i c , of 0.2. To estimate the volatilization of methyl iodide during DWT, a solution of 95.4 m g / m 3 in a background of 3.2 g of D O C per m 3

of F H S was tested with the flocculation apparatus without coagulant aid addition. Membrane filtration of this solution yielded no removal of C H 3 I . The mixing conditions were the same as for the other flocculation tests. A control solution was placed into the mixing vessel and exposed to the atmosphere for the same time but without mixing. The results, shown in Table VIII, indicate a 23% removal by volatilization during mixing; standing in an open vessel yielded an 8.2% removal. No significant change in D O C or UV254 was found. The removal of C H 3 I could probably be improved if the process was optimized for C H 3 I removal. Significant additional removals

ο > ο Ε

100

80

60

- 40 ω ο

CL 20

ot

10

Humic substances (FHS)

C 0 1311

131J

DOC uv 2 5 4

1.17 MBq /m 3

4.20 g / m 3

16.3 m" 1

-

I 1 i 1 > 1 U 1 1 1 1 I I I I I

AC: F300 t c = 2 d a y s

3 4 S 6 7 8 9 1

10' 3 4 S 6 7 8 9 1

10 2

3 4 5 6 7 8 9 1

10*

Activated carbon dosage, D (gAC/m 3)

Figure 8. Activated-carbon adsorption of Fuhrberg humic substances and 1-131.

Table V i n . Volatilization of Methyl Iodide in the Flocculation Apparatus

Time Concentration CHJ Percent Condition (min) C (mg/m3) Removal Initial 0 95.4 After mixing 45 72.1 23 Open vessel 45 87.2 8.2 NOTE: Fuhrberg humic substances in solution with 3.2 g of DOC/m 3 and 1.0 mM Na2HP04 at pH 6.5.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



would be expected if a packed column or other processes designed for the removal of volatile substances were used.

The activated-carbon adsorption C H 3 I - F H S solution in the dosage range 9-150 g of A C per m 3 is shown in Figure 9. The adsorption process is much more effective for the removal of C H 3 I than for D O C and U V 2 5 4 , and for 1-131, as shown in Figure 8. At a carbon dosage of 100 g of A C per m 3 , 92% of the C H 3 I is removed. Only 67% of the 1-131 is removed in a similar F H S solution. This result seems to be due to the solution interaction between 1-131 and the F H S . Because this interaction is missing, C H 3 I can independently adsorb.

These results indicate that both volatilization and activated-carbon adsorption are effective removal pathways if the radioactive iodine occurs as C H 3 I . However, D O C and U V ^ are not good surrogate parameters for monitoring C H 3 I removal.

Summary and Conclusions Both elemental iodine and iodide react completely with humic substances in the concentration range expected to occur after a nuclear reactor accident. Methyl iodide did not react with the humic substances. The results of bench-scale processes typical of drinking-water treatment indicate that flocculation and activated-carbon adsorption are effective in the removal of radioactive

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5



iodine when it is complexed with organic matter. D O C was found to be a good surrogate parameter to more easily assess the removal of complexed radioactive iodine in both humic substances and the Rhine River. Volatilization and activated-carbon adsorption were found to be effective for the removal of methyl iodide in the presence of humic substances. However, D O C and UV254 are not good indicators of C H 3 I removal.

Additional characterization of the removal effectiveness of treatment processes is needed at a scale larger than the bench scale investigated herein for 1-131 and other radionuclides likely to contaminate water supply sources, such as Cs-137 and Ru-103. Pilot-plant investigations should include the flocculation process with the addition of powdered activated carbon and should include fixed-bed columns of granular activated carbon, as this is the mode in which carbon is most commonly used in European drinking-water treatment. The effectiveness of aeration by packed columns and the impact of processes involving biological degradation should also be examined.

References 1. Haberer, K. GWF Gas Wasserfach: Wasser/Abwasser 1986, 127, 597-603. 2. Friedman, L.; Amann, W.; Lux, D. GWF Gas Wasserfach: Wasser/Abwasser

1986, 127, 604-613. 3. Laschka, D.; Herrmann, H.; Hübel, K.; Lunsmann, W. GWF Gas Wasserfach:

Wasser/Abwasser 1987, 128, 128-135. 4. Eberle, S. H.; Fuchs, F.; Haberer, K.; Summers, R. S.; Sontheimer, H. Ra

dioaktiv kontaminierte Rohwässer bei der Trinkwassergewinnung. Agrar- und Umweltforschung in Baden-Württemberg, Band 17, Ministerium fur Ernährung, Landwirtschaft, Umwelt und Forsten, 1986.

5. Massarotti, A. Gas Wasser Abwasser 1986, 66, 827-832. 6. Standard Methods for the Examination of Water and Waste Water; American

Public Health Association: Washington, DC, 1985. 7. Campbell, D. O.; Malinauskas, A. P.; Stratton, W. R. Nucl. Technol. 1981, 53,

111-119. 8. Kölle, W. In Adsorption Techniques in Drinking Water Treatment; Roberts, P.

V.; Summers, R. S.; Regli, S., Eds.; EPA 570/9-84-005; U.S. Environmental Protection Agency, Office of Drinking Water: Washington, DC, 1984; pp 805-811.

9. Skogerboe, R. K.; Wilson, S. A. Anal. Chem. 1981, 53, 228-232. 10. Dore, M.; Merlet, Ν.; De Laat, J.; Goichon, J. J. Am. Water Works Assoc.

1982, 74, 103-107. 11. Foster, R. Organic Charge-Transfer Complexes; Academic Press: London, 1969. 12. Behrens, H. In Environmental Migration of Long-Lived Radionuclides;

IAEA-SM-257/36, International Atomic Energy Agency: Vienna, Austria, 1981.

RECEIVED for review July 24, 1987. ACCEPTED for publication December 21, 1987.

Dow

nloa

ded

by U

CSF

LIB

CK

M R

SCS

MG

MT

on

Sept

embe

r 4,

201

4 | h

ttp://

pubs

.acs

.org

P

ublic

atio

n D

ate:

Dec

embe

r 15

, 198

8 | d

oi: 1

0.10

21/b

a-19

88-0

219.

ch03

5


Download - [Advances in Chemistry] Aquatic Humic Substances Volume 219 (Influence on Fate and Treatment of Pollutants) || The Fate and Removal of Radioactive Iodine in the Aquatic Environment

Top Related