Radiat. Phys. Chem. Vol. 47, No. 4, pp. 581-593, 1996
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RADIATION-INDUCED DEGRADATION OF WATER
POLLUTANTS--STATE OF THE ARTt
NIKOLA GETOFF
Institute for Theoretical Chemistry and Radiation Chemistry, University of Vienna
and Ludwig-Boltzmann-lnstitute for Radiation Chemistry and Radiation Biology,
Althanstr. 14, A-1090 Vienna, Austria
(Received 5 September 1994; accepted 20 February 1995)
Abstract--The radiation-induced decomposition of biological resistant pollutants in drinking as well as
in wastewater is briefly reviewed. First, some important units, definitions etc., radiation sources, as well
as dose-depth curves in water as functions of the electron energy and 6°Co-y-rays are mentioned. Following
is a schematical presentation of water radiolysis and of characteristics of primary free radicals. Then the
degradation of some aliphatic and aromatic chlorinated compounds in the presence of air is presented.
Some spectroscopic and kinetic data of transients resulting from chlorinated phenols are also quoted in
order to illustrate and to explain the rather complicated degradation mechanisms. In this respect the
synergistic effect of radiation and oxygen as well as that of ozone is also discussed. Finally, a scheme for
technical application of high energy elelctron beam is presented.
l. INTRODUCTION
2. UNITS AND CONVERSION FACTORS
As a result of the strong development of various
industries and the rapid growth of world population
a rather heavy overloading of water resources is
observed. In addition to this, the usage of various
chlorine-containing presticides and other chemicals,
as well as fertilizer in modern agriculture, contributes
to contamination of ground water. The killing of
microorganisms in drinking water containing humic
compounds by chlorination leads to the formation of
various halogenated hydrocarbons (Rook, 1974).
Such substances are carcinogenic and are usually
removed by filtration through activated charcoal
which is subsequently burned up. The resulting chlorine oxides contribute to the occurence of acid rain.
Due to the above mentioned pollution sources, environmental and toxicological effects have increased
considerably (Hutzinger et al, 1974, 1982, Fishbein,
1979; Strobel and Dieter, 1990). This is also reflected
in the health of the population in industrial areas.
Fundamental studies have shown that a complete
degradation of biologically resistant compounds, disinfection of sewage sludge as well as killing of
microorganisms can be achieved by ionizing radiation treatment. Abundant literature is available on
this topic. This is now the basis for technical electron
beam remediation in water.
In order to present a more complete picture of the
subject matter and as an introduction for newcomers
to this field, some basic knowledge in radiation
chemistry is briefly mentioned.
For convenience some radiation units, definitions
and conversion factors used in the radiation technology are given in Table 1.
2.1. Radiation sources
For radiation processing of polluted water high
energy electrons, such as 't-rays (e.g. from 6°Co,
zl/==5.26 years, Ey = 1.17 and 1.33MeV; ~37Cs,
~/2= 29 years, E~, = 0.66 MeV etc.) or X-rays can be
principally used. Based on the present state of technological achievements, preference is given to the electron-accelerator machines (EA-machines). Some
typical radiation sources are given in Table 2 for
comparison in respect to their output power. Obviously the EA-machines are far the best for this
purpose, because of their very high dose rate.
3. PENETRATION DEPTH OF RADIATION IN WATER
Only the absorved radiation energy can initiate
physical, chemical or biological effects. The energy
absorption in a medium, e.g. water, takes place in
10 -~5 s, in the course of which the dose distribution
is not uniform, because of electron scattering and the
"build-up" effects occuring during the interaction
between radiation and matter. As a result of this dose
distribution gradients are observed (Wiesner, 1979;
Seizer and Berger, 1987; Miller, 1990; Getoff, 1992a,
1993a). This effect is illustrated by curves, representing depth-dose (in %) for a set of electron energies
from 1 to 14 MeV, shown in Fig. 1.
Taking the curve e.g. for 10 MeV electrons the
extrapolated range in water, R¢~ = 5 cm (Fig. 1). This
tPresented at the IMRP-9 in Istanbul, Turkey, 11-16
September, 1994.
581
Nikola Getoff
582
Table I. Some useful definitions, radiation units and conversion
factors.
Output power of electron accelerator (in kW):
kW = mAx MeV
rnA = electron current, MeV = electron energy
Radiation yield: G-value = number of produced or decomposed
molecules per 100eV absorbed energy. For conversioninto Si-units:
multiply the G-value by 0.10364 to obtain G(x) in g mol J-I.
Absorbed dose:
1 rad = 100erg g -~ = 6•24 x 101JeVg -~
100 rad = 1Joule kg-t= 1 Gray (Gy)
1 krad = 10Joulckg t = 10Gy
1 Mrad = 104Joule kg i= 104Gy
1kW = 3.6 x I 0° J kg- ~= 360 Mrad kg- ~h
,0o
; . ,:f
.o
+,,
~it I
,,
I ,rvel
..-.
"'.- ",,
~+". xx
\".
,,
\'..
\
\
'.
~'
~" lOO
is approximately inverse proportional to the density
(0) of the absorbing material (for water 0 ~ lg/ml)
and roughly proportional to the electron energy (for
water R~ ~ 0.5 c m / l MeV penetration depth). Hence,
the shape of a depth-dose curve in a material is a
function o f electron energy, absorber density and
incidence angle o f the electron beam.
The depth-dose curves (D, %) in water for 6°Co-yrays and 2 MeV electrons are given for comparison in
Fig. 2. Obviously, the y-rays have a much deeper
penetration in water than the electrons (see also
Fig. 1).
4. WATER RADIOLYSIS
As a sequence of the interaction between ionizing
radiation (electrons, y-rays) and water electronically
excited and ionized molecules are formed. Subsequently this leads to the production of several very
reactive primary species (OH, e~q, H) and molecular
products (H2, H202). Their yields (G-values, related
to the absorbed dose) as well as the major reactions
taking place in irradiated water and the corresponding rate constants (k) are presented in Table 3.
In the presence of oxygen in water the reducing
species, H-atoms and the "solvated electrons" (e~q)
are converted into oxidizing species, perhydroxyl
radicals (HOt) and perhydroxyde radical anions
(O~-). The last ones together with the OH-radicals
can initiate degradation of water pollutants.
F o r information and practical interest the absorption spectra of the primary products of water radiolysis and their initial yields (Gt) as function of p H are
presented in Fig. 3. These data are also needed in the
case when reaction mechanisms are postulated in
order to get material and energy balance.
,
I
:iA
0
I
l "I
0
\c
I
1
'". F \X G
,
I
I
I
I
I
I
J
X%
I
I
I
1
2
3
5
6
7
Penetration depth in water (cm)
Fig. 1. Depth<lose (%) distribution in water for different
electron energies (see inserted Table)• Insert: discharge traces
of 3 MeV electrons in a plate of polymethylmethacrylate.
5. RADIATION INDUCED DECOMPOSITION OF WATER
P O L L U T A N T S - - G E N E R A L REMARKS
As already mentioned above the radiation induced
degradation of pollutants in water is a very powerful
technique. Based on the very high radiation power of
the modern EA-machines this method is marked out
by a great product output, high efficiency and economics in comparison to other technologies (see e.g.
Kurucz et al., 1991; Cooper et al., 1992; Cooper et al.,
1993 and references therein).
The outstanding ability of electron beam irradiation in respect to water remediation is demon100
80
60
A
~
1
0
6°Co-~-source:
7.2 x 10~
7.2 x 107
1.8 x l0s
I 1
1 3
1 6
18
I 10
112
I
I G 114
';..
Table 2. Output power (kGy kg- 1h- i) of various radiation sources
Type of radiation source
Dose rate (Mrad/h)
X-ray source 10mA, 250 kV
0.18
65
130
". ~x
'.
,
i
A
B
C
D
E
F
50
~'~ 2 0 4 0 0
0.5 x 106 Curie
1.0 x 106 Curie
Electron accelerator:
-Van de Graaff (1 mA, 2 MeV)
-DYNAMITRON (40mA, 5 MeV)
-LINAC (50 mA, 10 MeV)
I
I
I
|
I
lo
0
0.5
1.0
I
I
I
I
5
10
15
20
Depth in water (cm)
Fig. 2. Depth<lose (D, %) curves in water for ~°Co-7-rays
(A) and 2 MeV electrons (B).
583
Radiation-induced degradation of water pollutants--state of the art
Table 3. Radiolysis of water and some primary reactions
Primary reactions:
H20,,.*HzO*---,H+ OH
H~O + + e-; e---,e~ + nH20~e~q
H20 + + H:O--*HaO + + OH
Gross reaction of water radiolysis (the G-values? at pH 7 are given ip
brackets):
H: O:,,-..e~,H. OH, H~, H2Oz, Ha~. OHaq
(2.7) (0.6) (2.8) (0.45) (0.7) (3.2) (0.5)
Major primary reactions:
(k= 1.4x 10"dm3mol-Js i)
H~ + OH/q~H20
(k = 1.0 x 10~°dm3mol t s a)
H + H~H 2
(k=2.5× 10t°dm3mol is i)
H + OH~H20
( k = 2 . 0 x 101°dm3mol Is L)
H + eaq-*H2+ OH,q
(k=6.0x109dm 3tool ~s ~)
OH + OH--*H202
(k =2.5 x 10~°dm3mol~ s ~)
OH + e~---*OH,q
( k = 3 . 0 x 109dm3mol ~s i)
eaq + e~q~H2 + 2OHaq
( k = 2 . 3 x 101°dm3mol's i)
e~+H~H
( k = 2 . 5 x 107dmrmol Is ~)
H + OH;q---,e,q
(pK = 11.9)
(pK = 11.65)
H202~H~ + HO2aq
In the presence of oxygen:
(k=2.1xl0t°dm3mol Js ~)
H + O2--*HO2
(k=l.9×101°dm3mol Is ~)
e~q+ O:-.O;
(pK = 4,8)
HO2.~-H+ + 02
In the presence of N20:
(k=0.91 × 10t°dm3mol Is ~)
%q + N20~OH + OH- + N~
in the presence of H~:
(k=3.5x107dm 3tool ~s ~)
OH + H2~H~O + H
tG-value = number of changed molecules per 100eV (1.60 x 10 ~Tj)
absorbed energy. For conversion into Sl-units: multiply the G-value by
0.10364 to obtain G(x) in #mol J 1.
5.1. Steady-state radiolysis of chlorinated hydrocarbons
s t r a t e d by s o m e examples. First, the r a d i a t i o n
induced oxidation of chlorinated hydrocarbon
m i x t u r e s in d r i n k i n g w a t e r a n d after t h a t this
o f t r i c h l o r e t h y l e n e ( T C E ) is discussed. It follows
the degradation of phenol and 4-chlorophenol
(4-C1PhOH).
As a l r e a d y m e n t i o n e d in the p r o c e s s o f d r i n k i n g
w a t e r purification by c h l o r i n a t i o n a n u m b e r o f chlor i n a t e d h y d r o c a r b o n s are f o r m e d . Hence, the effect o f
(B)
104
Iq62 ~
(A)
103 x 2.0 -
1.5
_
H+ + 6~
./-'\
/
/
pK = 4.8
\
,o
103
/°~
['
102
200
400
600
800
nlll
4 --
~o
1.0
-
0.5
I!OH~
o
200
i.I
"\
~5-
2
H~"~
o :,250
300
nm
350
3
eaq
i
i
5
7
9
I
13
pH
Fig. 3. (A) Absorption spectra of H, OH, H02, and O;- radicals as well as H202 (~ in dm 3 mol -~ cm-~).
Insert: pK-value of HO2. (B) Absorption spectrum of e~ (~ in dm3mol -~ era-I). (C) G:values of the
primary products of water radiolysls as a function of pH.
584
Nikola Getoff
Table 4. Radiation induced degradation of some chlorinated hydrocarbons (givenin #g. dm-3) in various mixtures in the presence of air
(0.25 x 10-3 tool' dm -3 02) at pH ~ 7. Applied dose: 1.9 kGy.
CH2CI2
CHCI3
CCI4
CI3C---CH3
C12C-------CHCI
CI2C-------CCI2
Sample no.
P
A
P
A
P
A
P
A
P
A
P
A
I
--32
nd
--0.7
nd
.
.
.
.
2
.
.
.
.
25
1
0.2
nd
.
.
.
.
3
48
nd
.
.
.
.
107
0.8
1.1
nd
--4
23
nd
2.3
nd
1.1
0.1
0.2
nd
.
.
.
.
5
--2.5
nd
0.2
nd
4.5
nd
361
nd
0.2
nd
tThe GC-analysis was performed with "Vista 6000", Varian Ltd., by Dr D. U. Bauer. Traceable limits: < 0.1 ggdm -3. P = prior and
A = after irradiation.
the radiation-induced degradation has been proven
by irradiating samples containing mixtures of chlorinated methanes, ethanes and olefines in various concentrations in the presence of air (dose: 1.9 kGy). The
analysis o f the samples were performed by GCmethod. It can be seen in Table 4 that the pollutants
are completely decomposed with exception o f CC14
and C13C--CH3 which show very small amounts left
(Getoff, 1986b, 1989).
It should be mentioned that in the case of natural
drinking water containing inorganic salts, humic
compounds etc. the required radiation dose is somewhat higher than for artificial samples (e.g. Proksch
et al. 1987).
In the following the radiation-induced degradation
of trichloroethylene (C12 = CHCI; TCE) is discussed
in some details. This pollutant can occur in various
kinds o f wastewaters and under certain conditions
also in drinking water. Its radiolysis has been previously studied (e.g. K r s t e r and Asmus, 1971;
Gehringer et al., 1985; 1986; Getoff 1986a; 1987,
1989).
Figure 4 shows the formation of C1- ions from
1 0 - 3 m o l d m -3 C I 2 C = C H C I as a function of the
absorbed dose in the presence of air (Getoff, 1989).
The initial yield Gi(CI') = 19.8 is rather high indicating the involvement of a short chain-reaction. As
further products were found a mixture of aldehydes
(G~ = 0.02) and simple carboxylic acids. They are
however, decomposed at higher doses. The required
radiation dose for the total degradation of
CI2C------CHC1 strongly depends on the substrate con-
, / . /
~
5 --
O f
1oo
,~o
I
I
I
0.5
1.0
1.5
(1 a)
(1 b)
k 1 = 1.9 × 101°dm3mol-I s )
(Koster and Asmus, 1971)
L~
100
?
L--*CI~------CHCI + C1-
C12C--~CHCI + O H - - p
~
r..)
0
CI 2C-------CHCI+ e~q --[-*CI 2C-------~H+ CI-
Simultaneously with the above processes O H radicals
are adding on the double bond of TCE-molecule
forming OH-adducts which are unstable and decompose, e.g.
pO~.---"
10 -4 X 10
centration. This fact is illustrated by the insert in
Fig. 4, where small amounts of trichloroethylene have
been irradiated under the same conditions. A complete degradation of the pollutant has been achieved
at 150 Gy, at which only traces of aldehydes and
carboxylic acids were observed. Similar results have
been reported by Gehringer et al. (1986).
In order to obtain information about the involvement of the indivuduai primary transients of water
radiolysis (OH, e~q, H), peroxide-radicals and ozone
in the degradation of trichlorethylene experiments
under the required conditions have been performed
(Getoff, 1989). In this case the yield of C! was taken
as an indicator for the decomposition process. The
obtained initial C1 yields Gi (CI-), from the various
experimental series are presented in Table 5.
The rather high Gi(CI-) = 11.2 observed in airfree
solutions seems to result partly by the specific e~q
attack on the Cl-atom of T C E and on the reaction of
O H forming predominantly OH-adducts which undergo a spontaneous decomposition. F o r the radiolysis of T C E in the absence of air the following major
reactions have been suggested (Getoff, 1989).
I
200
I Gy
2.0
kGy
Fig.
4. Dose-dependence
of
a-formation
from
10-6 tool dm -3 CIzCH==CHCI in the presence of air (pH
6.4). Insert: Radiation induced decomposition of
2.74 x 10 -6 tool dm -3 CI2CH~------CHCIin the presence of air
(pH ~ 6.5) as a function of dose (Getoff, 1989).
CI2C--CH(CI)OH
(2a)
C12(OH)C--(~HCI
(2b)
Table 5. Gi(CI ) obtained f r o m aqueous
1 x 10-4mol.dm 3 CI2~CHCI in the presence of
different O2-eoncentrations at pH ~ 6.5(Getoff,
1989).
Series
Gas-content
No.
( m o l d m 3)
GI(C I )
1
airfree
I 1.2
2
3
0.25x10 3Or
1.25x 10 3 02
1.25x
10
3 02
1.1 x 10'3 03
9.1
7.6
8.5
4
Radiation-induced degradation of water pollutants--state of the art
mean value: k2=3.3 x 109dm3mol-ls -l (K6ster)
and Asmus, 1971)
CI2C---~H(CI)OH-CI2(~---CHO + C1- + H + (3a)
585
The phosgen (C12CO) hydrolyses according to reaction (8). The formaldehyde is completely decomposed
at higher doses.
O ' - + H + ~ O H ( p K = 11.9)
C12(OH)C--t~HCI~CIOC--CHO + C1- + H + (3b)
k s > 7 x l0 s s -I (K6ster and Asmus, 1971)
The radicals (CI2C--CHO) are disappearing according to a second order reaction, e.g.
2C12C--CHO--* products
(4)
2k4 = 6.2 × 108 dm 3 mol -l s -1
(K6ster and Asmus, 1971)
The H-atoms can principally form also adducts or
abstract CI- or H-atom from TCE-molecule.
It has been reported earlier (Koltzenburg et al.,
1982; see also Lal et al., 1987; Getoff, 1993a) that in
general alkyl radicals with leaving groups on fl-position, resulting from the reaction of OH with a
halogenated compound, undergo a fast hydrolysis.
As a result of this OH-group is incorporated on the
place of the Cl-atom, e.g.
//~CI+H20~/~/OH+C1-
+H +
The produced O H / O ' - species [equations (10) and
(11)] are consumed according to reaction (2).
In addition to the above reaction steps also the
formation of tetraoxides and oxy-radicals can be
involved in the degradation process of TCE, e.g.
2C12(OH)C--CH(CI)O2--* R O O - - O O R
-r
'
(tetraoxide)
ROO"
~12)
ROO---OOR~O2 + 2C12(OH)C--CH(CI)O" (13)
(oxy-radical)
The oxy-radical can decompose after abstracting
an H-atom from a compound in the bulk of the
solution:
C12(OH)C--CH(C1)O + R H - , R "
+ CI2(OH)C--CH(CI)OH
(14)
C12(OH)C--CH(CI)OH--, C1OC - CHO
+2C1-+2H +
(5)
Under airfree conditions the obtained species can
disappear by dimerisation and/or disproportionation.
In the presence o f air (0.28 × 10 -3 moi-302 at 20°C
and 760 torr) or in water saturated with pure oxygen
(1.40× 10-3moldm 302 at 20°C, 760 torr) the
above mentioned radicals can add oxygen resulting in
peroxyl-radicals. Depending on the 02 as well as on
the TCE concentration in the aqueous solution a
competition for e~ and H-atoms between 02 and
TCE takes place. Both e~q and H-atoms, at low
pollutant concentration in aerated water, are converted into peroxyl-radicals (see Table 3). Hence, the
attacking.species under these conditions are OH as
well as HO2/O i which are leading to the corresponding TCE-adducts. They are unstable and can undergo
a multiple hydrolysis (Koltzenburg et al., 1982; Lal et
al., 1987; Getoff, 1993a) e.g.
(11)
(15)
It is also conceivable that the oxi-radical of TCE
can undergo a multiple hydrolysis according to the
following gross reactions:
C12(OH)C--CH(CI)O + H 20 ~ CICO
+HCOOH+2C1
+2H*
C1CO + H 2 0 ~ C O O H + C1 + H +
(16)
(17)
C O O H ~ H ÷ + CO~
(pK = 1.4; Bruxton and Sellers, 1973)
2CO2 ~(CO~ )2 (oxlate)
(18)
(19)
The observed G~(CI )-values in solutions saturated
with pure oxygen are somewhat lower compared to
those containing air (Table 5). In this case the
primary e~q attack on TCE is completely excluded,
but as shown above the produced O~- species are
involved in the process.
CI2(OH)C--CHCI + O:-~CI2(OH)C--CH(CI)O 2 (6)
5.2. Combined radiation-ozone
CI2(OH)C--CH(CI)O2 + H20-~H62 + HCHO
+C12CO -I- H + + CI
C12CO + H20--'2C1- + 2H + + CO2
(7)
(8)
The produced HO2/O 2 species [equation (7)] as
well as those originating from the conversion of e;q
and H-atoms (see Table 3) are involved in similar
reactions, e.g.
C12C-----CHC1+ 6 2 --* - O: (el 2)C--OHCI
(9)
++CI-+O'-
It is well known that a great number of organic
pollutants in water can be decomposed just by reaction of ozone (03). In the combined radiation-ozone
water treatment HO2/O2- as well as OH radicals
act as carriers for the chain-reaction of ozone degradation (e.g. Hoigne, 1982; Sehested et al.,
1983, 1984, 1991, 1992; Bfihler et al., 1984; Staehelin
et al., 1984). In the absence of pollutants in water the
chain-mechanisms by Weiss (1935) is operative. The
major reaction steps are:
03 + H62--'OH + 202
-O2(C12)C - C'HCI + H20--~C12CO
+HCHO+H
processing o f water
(10)
(20)
k20 ~ 104 dm 3 tool ~s -l (Sehested et al., 1984)
586
Nikola Getoff
0 3 + 0 2 - - - . 0 ;- + 02
Based on the higher efficiency of the synergistic effect
of the radiation-ozone treatment of polluted water
the author of this review is proposing the utilization
of the ozone formed as a byproduct in the air-layer
between the water surface and the accelerator-window. The gas-mixture (air and ozone) can be collected and pumped at some distance in front of the
irradiation sector. Due to the water-flow 03 will be
well mixed with the water before the irradiation step.
By means of this procedure an enhanced degradation
of the water-pollutants can be achieved. The irradiation of air will result in formation of small
amounts of nitrogen oxides which will be subsequently decomposed.
(21)
k2t = 1.5 x 109dm3mol -j s -1 (Sehested et aL, 1983)
O3 + O H ~ H 6 2 + 02
k22~--. 1.1 x
(22)
10Sdm3mol -I s -1 (Sehested et al., 1984)
O3 + efq---,O3-
(23)
k23 = 3.6 x 101°dm3 mol -l s -l
(Sehested et aL, 1982)
02 + O ' - - - , 0 3 -
(24)
k24 = 2.5 x 109 dm3 mol -l s -1
(Adams et aL, 1966)
03- + OH ----~ O 3 + O H - (30%)
(25a)
5.3. Degradation o f phenol
/
L - ~ 2 0 ~ - + H + (70%)
a
(25b)
Phenols, chlorinated phenols, anisoles, biphenyles
and many other aromatic compounds which appear
as frequent pollutants in waste waters and in some
cases even in drinking water, have been subject of
investigations. Besides, phenol has been used as a
typical representative of these compounds for studies
of radiation-induced decompositioned under different conditions (e.g. Gilbert and Giisten, 1973; Micic
et al., 1975; Hashimoto et al., 1979, 1980; Takehisa
and Sakumoto, 1982; Getoff and Lutz, 1985; Getoff,
1986a, 1987, 1992b, 1993b; Waite et al., 1992; Lin
et al., 1993).
Micic et al., (1975) reported that the decomposition
of phenol in oxygenated water solutions strongly
depends on its concentration and on the applied
dose rate. The degradation yield, G ( - P h O H ) of
2 x 10-2moldm -3 phenol increases from 2.6 to
250 by decreasing the dose rate from 100 to
13 Gy h - X.This effect is explained by a chain reaction
initiated by OH and O~- radicals and propagated
mainly by the O~- transients. Similar observations
•
03- + H + ~ HO3
(26)
b
k2~= 5.2 × 101°dm3 mol-I s -I
(Staehelin et al., 1984)
k26b= 3.3 × 102s -l (Staehelin et aL, 1984)
The ozone intermediates HO'~ (2 = 350 nm,
~350= 300 dm 3 mol-I cm-l) and 03- (2 = 430 nm,
E430= 2000 dm 3 mol-~ c m - l (B/ihler et al., 1984) are
strongly oxidizing agents and hence they can decompose organic pollutants. The simultaneous application of radiation and ozone for water treatment
leads to a very efficient degradation of pollutants (e.g.
Takehisa et al., 1985; Arai et aL, 1986 Gehringer
et aL, 1988, 1990a, 1990b, 1992; Getoff and Solar,
1988; Getoff, 1989, 1990; Haberl et al., 1991).
The synergistic degradation effect of radiation
and ozone is illustrated in the case of TCE (see also
Table 5). The following major reactions are invloved
in the process:
0
/\
CH2"-CHCi
+ o, - - - - - c h . c - . c n o
I
0
-----
I
0
cl c
CaCl
(27)
0~0
0
0
/~
ch :
0~0
OH OH
cno
I
+a3o
I
chc
0~0
The phosgen (C12CO) resulting by reaction (28) decompose by hydrolysis [equation (17)]. The chlorinated aldehyde is also not stable under these
conditions:
CIHCO + H20--.HCOOH + C1- + H ÷
(29)
I
cno
I
a o2
+
CIHCO + C!2CO
(28)
have been made also by Pikaev and Shubin (1984 and
reference therein). At low dose-rates in aerated phenol solutions they reported G ( - P h O H ) = 250-500.
As a main process is suggested the formation of
polymers. The increase of the temperature to 80°C is
leading to a two fold higher decomposition of phenol.
Takehisa and Sakumoto (1982) showed that the
587
Radiation-induced degradation of water pollutants--state of the art
D
Air &
Polluted A~L"
~
EBGOB
--"
j
Purified
~" i i ~ i ~ + r , A i r
Woter
__.
water
__.
Irrodiotion
Zone
Fig. 5. Utilization of the ozone formed in the air-layer between accelerator window and water surface.
EA = Electron accelerator; P = Gas pump; GCB = Gas collecting bell.
applied dose can be reduced 20-fold by combination
of radiation and ozonation. In this case preferentially
carboxylic acids (mainly oxalic acid) are produced
which are subsequently decomposed to H20 and CO,,.
As final produces of oxygenated 10 3 mol/dm 3
phenol solutions Hashimoto et al., (1979) have determined by means of HPLC-method: G(pyrocatechol) = 1.42, G(hydroquinone)= 0.94 and traces of
resorcinol and
phloroglucinol as
well
as
G ( - P h O H ) =2.66. The same authors Hashimoto
et al., 1980) studied the degradation of 10-3mol
dm 3 phenol by electron beam processing in airfree
as well as in solutions containing 02 or saturated with
pure oxygen in flow system. The phenol-content was
reduced upto 90% at a dose of 5 kGy and a flow rate
of 1.5-5dm3/min -1 and 2mA current beam. The
formation of the same products was observed as
described above.
Waite et al. (1992) used high energy electron
(1.5 MeV beam) and a current varied upto 50mA
(doses up to 8.5 kGy) for decomposition of phenol
(0.75 x 10-3 moldm 3) at pH 5 and 7 in addition to
studies conducted to the radiolysis of trichloroethylene ( 6 x l 0 - S m o l d m - 3 p H 7) and chloroform
(6 x 10 ~to 1 x 10 4 mol dm -3, pH 5) in absence and
presence of clay. In the case of phenol (no clay) a
complete decomposition was observed at a dose of
about 7 kGy. About the same dose was needed for the
degradation of chloroform, but only 0.5 kGy were
sufficient for the radiolysis of trichloroethylene. Lin
et al. (1993) conducted electron beam (I.5MeV,
50mA) experiments for removing phenol (0.01 to
0.95 x 10-3moldm -3) over the pH-range of 5-9
using doses upto 7 kGy under flow stream (480) dm 3
min-t) in absence and presence of 3% kaolin clay. At
low doses as major products were found: catechol,
hydroquinone and resorcinol. However, at continued
oxidative procedure the formation of formaldehyde,
acetaldehyde, glyoxal and formic acid were identified.
Although phenol (0.95 x 10 -3 mol dm -3) was decomposed by a recirculation experiments the total organic
carbon in the solution decreased only slightly, indicating the formation of polymers.
From the above reported data it is obvious that the
radiation induced-phenol degradation proceeds by
rather complicated reaction mechanisms depending
on many factors: substrate concentration, applied
dose and dose-rate, availability of oxygen etc. The
degradation of 10-~ to 10-3 mol dm -3 phenol saturated with oxygen (1.25 x 10 3moldm -3 02,
pH =6.5) as well as the formation of the major
products resulting from 10-4moldm -3 phenol as a
function of dose (dose-rate: 70 Gy min -1) is presented
in Fig. 6 for illustration. Obviously, at a dose of
about 1,2 k Gy the 10 -4 mol dm -3 phenol is practically completely degradated (Fig. 6, curve B). How-
100
(°I0)
1
<:..e
~Xo~i
80
A
li
%,~, I I ' ~ l
60
"..x
'~
no
40
-
C
liOl.,i=i+,IPli°lllC_~-PtOI41~
1o -s
10-+
o.~
+.l
'O-'
12.'
C
x~°
B
A
:'z~ A ~°x-,
20
""-Z~.
~*O~
.......n.
- 7A~- -'"~ .........-z~'~'"7"-" :
0
I
'
I%)
....
20
o
"~----o-- ~~_ b
,°, 9
,
0
0.2
0.4
0.6
0.8
,
1.0 kGy
Fig. 6. (I) Radiation-induced decomposition of various
concentrations of aqueous PHENOL as a function of dose
in the presence of 1.25 x 10-3moldm-302 (pH~6.5 at
25°C). Insert: G(-PhOH) observed at 10-5-10-3 mol dm -3
(PhOH), (II) Major products resulting from solution (B) as
a function of dose: (a) pyrocatechol (G+= 0.9),
(b) hydroquinone ( G i = 0.6) and (c) hydroxyhydroquinone
( G i = 0 . 0 8 ) (Getoff, 1986a, 1992b).
588
Nikola Getoff
ever, at the same time pyrocatechol, hydroquinone
and hydroxyhydroquinone are formed (Fig. 6, II).
The decomposition of phenol in aerated solutions
is initiated by OH-attack (k = 1.4 x 10J°dm3 mol -j
s-~) resulting in the formation of OH-adducts (hydroxycyclohexadienyl radicals) on o-, m-, p- and
ipso-positions (e.g Getoff and Lutz, 1985; Getoff,
1986a, 1987, 1992b, 1993a). These transients can add
02 forming the corresponding peroxyl-radicals which
are leading to ring-opening and splitting of HO:/02species. The last ones are then reacting in the same
manner as OH radicals. All these reaction steps are
favored at low substrate concentrations and low
dose-rates. At higher concentrations and dose-rates
the phenol-transients react preferentially with the
phenol leading to the formation of oligomeres. The
major reactions explaining the formation of the
above mentioned products are given for illustration
of some reaction steps [see equations (30--37)] on the
opposite page.
The HO:/O2- and the OH radicals resulting from
the above processes are initiating chain-reactions.
The observed products are attacked further by OH,
HO:/O 2- species in a prolonged irradiation procedure and can be decomposed upto CO: and H: O.
5.4. Degradation of chlorinated aromatic compounds
Chlorophenols have been often chosen as model
substances for studies in respect of the irradiation-
induced degradation of chlorinated aromatic compounds, e.g. pesticides used in agriculture, or basic
substances applied in the chemical industry
(e.g. Getoffand Solar, 1986, 1988; Draper et al., 1989
etc.).
It is of special interest to show the synergistic
effect of radiation and ozone, taking 10-4 mol dm -3
4-Cl-phenol (4-C1PhOH) as an example (Getoff and
Solar, 1988). The course of its degradation has been
followed as a function of dose (dose rate:
50 Gy min -I) in aerated (0.25 x 10-3 mol dm-302) as
well as in solutions saturated with oxygen
(1.25 x 10 -3 mol dm -3 02)
and
containing
1.1 x 10-5 mol dm -3 03) which are presented in Fig.
7. Obviously, in the first case (Fig. 7, A/I, curve a) a
dose of above 12 kGy is needed for degradation of the
pollutant, compared to a dose less than 1.5 kGy for
the solutions containing even a small amount of
ozone (Fig. 7, A II, curve a). The course of the CIyields in respect to dose in both cases are rather
similar, however the aldehyde yield is strongly
effected by ozone (compare curves c in I and II
inserts, Fig. 7).
The yield of the substrate degradation, G (---4C1PhOH), as well as the formation of CI- and
aldehyde for the same substrate concentration under
different conditions are presented in Fig. 7, Table B.
Here again is illustrated that the highest G~-values are
observed in the presence of ozone.
(A)
I. Air
100
--
II. Oxygen & 03
100
b ~'~
I/
I
I
50
10-6 x I0 ~ : c """ *" "*'" "'8
"'""
mo___ll 6 ~ /
dm3 4
~
2
I
--
1"" 10-6x8 -'*
"'*.
mol 4 .
dm3
c
50
I
""
•
2~
kGy
a
I -10
5
6
t b
I
'~0
kGy
"'......
0
1
[
5
kGYl
10
2
kGy
(8)
In the presence of:
Gi(-4-CIPhOH)
Gi(CI-)
Gi (Aldeh.)*
Argon
2.6
2.4
0.28
0.25 mM 02
2.0
1.50
0.23
1.25 mM 02
2.8
1.30
0.22
1.25 mM 02
0.011 mM 03
3.7
8.30
1.00
Fig. 7. (A) Radiation induced decomposition of 10-4 mol dm -3 4-Cl-phenol (pH ~ 6.5) as a function of
dose. (I) In the presence of air (0.25 x 10-3 mol dm-302). (II) In the presence of 1.25 x 10-3 mol dm -3
02 and 1.I x 10-~moldm-303. (a) decomposition of 4-Cl-phenol; (b) formation of CI- and
(c) aldehyde. (b) Initial yields (G~)of the decomposition of 10-4 mol dm -3 4-C1phenol as well as of the
formation of Cl- ions and of aldehydes observed under various irradiation conditions at pH ~ 6.5 (Getoff
and Solar, 1988).
Radiation-induced degradation of water pollutants--state of the art
589
OH
c,.,o. ÷ o . - -
~of"
c,.:
(3O)
+ H20 (-70%)
(OH - adductson o-. m.p - and ipso - positions)
2 C6Hs(OH) ~
C6HsOH + ~ , . O H
(31)
( - 30%)
OH
~
OH
OH
+
02 ~
~
~~o.
02
OH
H
nol
(32a)
:-c~ °
"" ~¢...c.~O + sol
(32b)
+
OH
(pyrocatechol)
H
(mucondialdehyde)
OH
OH
OH
---- ~
uo~
+
(33)
OH
Cnydroquinone)
OH
+ OH ~
+ 02
OH
I ~
OH
:.:o?
OH
OH
N . ~ °,
coo.
+ 02
[ ~
+
"~..coou
OH
OH
~,,~jOH
~
i,~'~" COOH
0
L--f'~'~c~O
CCOOn
cncooall cOOHI ncoon ~ cth
•% . . c ~ . °
-~..coou
cncoon
coon
Mar'oak
add
Maide
acid
Onlk
acid
Mueou
aldehyde
OH
(35)
OH
+
H
(34)
OH
~J'COOH
-_. :-c~o.
>coo.
OH
HO~
OH
(hydroxybydlro*
qulnone)
OH
"H
OH
+
Scheme 2
(36)
(37)
Form~
acid
590
Nikola Getoff
The very efficient synergistic effect of radiation and
ozone is illustrated by the following reaction steps
given in equation (38).
OH
By means of pulse radiolysis technique it has been
possible to gain a deeper insight in the complicated
degradation mechanisms of chlorinated phenols (e.g.
OH
+ 03 --~
OH
.03
°
----
+ H20 ---
H
(Complex)
OH
(Muconic aldehyde)
~'~COOH
~t~/COOH
0 2 4
6 ps
0 4 8 12p.s
•
0 40 80120Ill
0 1.63,2 4.8m~
0.005
300
350
+ H2
(38)
(2-Hydroxy muconic acid)
Naturally, there are a number of further reactions
in which 03 is involved, but they are not further
discussed here.
In addition to aldehyde also hydroxybenzenes (pyrocatechol, hydroquinone, hydroxyhydroquinone,
small amounts of resorcinol and traces of pyrogallol)
were observed. Their yields depend on the experimental conditions, e.g. on dose, dose-rate, substrate concentration, pH of the solution etc. At higher doses all
these products can be completely decomposed.
0
(Hydrolysis)
OH
~ C ~ H O + H202
0
(Ozonide)
400
nm
Fig. 8. Transient absorption spectra resulting from pulse
radiolysis of 5 x l0 -4 tool dm -3 4-Cl-phenol (pH 8) saturated with oxygen at: (A) 2/~s and (B) 100 ~s after 0.4/~s
pulse (OD-values normalized to 10 Gy). Insert: Transient
kinetic at 315 nm using 5 x 10-4 tool dm -~ 4-Cl-phenol (pH
7.6) in the presence of 1.25 x 10-3moldm -3 02 .
(A) Formation and decay of OH-adducts, (B) Begin of
O2-addition on the OH-adducts, competing with water
splitting process, (C) Decay of the peroxylradical,
(D) transient decay and reaction of 0 I- species with the
substrate (Getoff and Solar, 1988).
Getoff and Solar, 1986, 1988). The transient absorption spectra obtained at 2/~s and 100 #s after a 0.4 ps
electron pulse in the case of 4-CIPhOH solution
saturated with oxygen is presented in Fig. 8 (Getoff
and Solar, 1988).
First, the OH-attack on the substrate takes place
resulting in the formation of OH-adduct, where
k(OH + 4-CIPhOH) = 1.5 x 101°dm3 mo1-1 s -1. The
resulting transients are represented by spectrum A,
Fig. 8. Simultaneously, e~q and H-atoms are converted in the slow reacting O2- species (see also Table
3). The next step is addition of oxygen to the OHadduct (k ,~ 2 x l0 s dm 3 mol ) s -1) under formation
of peroxyl-radicals, followed by splitting off of water
and HO2/O ~- species. The last ones attack the substrate as well as the radicals present in the bulk of the
solution. Hence, the measured absorption spectrum
(B) at 160ps after pulse expresses a mixture of
transients. This fact is illustrated by the kinetic traces
given as insert in Fig. 8. Finally, the reactions of
HO2/O~,- species with substrate and with the radicals
present in the bulk in the solution take place as well
as opening of the aromatic ring. All these reaction
steps are competing with each other. The reaction
mechanisms are similar to that of phenol [see
equation (30-37)] leading to the degradation of 4CIPhOH.
A special environmental problem is the degradation of the extremely toxic polychlorinated biphenyls (PCBs). It has been found by several authors
(e.g. Singh et al., 1985 and references therein) that in
alkaline isopropanol solution a radiolytic dechlorination by chain-reaction takes place (G i > 1000). Further investigations in this respect are urgently needed.
5.5. Combination o f radiation and conventional techniques
Depending upon the kind and concentration of
given pollutants in water the efficiency and economy
591
Radiation-induced degradation of water pollutants--state of the art
(B)
(A)
Reservoir of
raw drinking
water
Equilibration
~iwlth air and
" 0 s addition
[ ~
I
I water
I [
Storage of ~
purified
~I
water
/
[
1
To consumer
I Pretreatment
I
I Ct
emica, ~ ~fe~er~°zrk._dsteps:
Neutralization, I
phLnt
Radiation
p.roc?ssing:
dismtection
and
decomposition
OT pollutants
~
waste
water
I
I
I
Isedimentation,
IcoaQuiat on etc.
Radiation
L _ ~ Microp,roc.esslng: ~
biological
o.ecompos]tionl
I purification
oT pollutants I
I (sewage)
dis'infection'
I~
To river/sea
Fig. 9. Schemes for purification of drinking (A) and of industrial wastewater by electron processing (B).
of their decomposition can be essentially increased
by combination of conventional methods and irradiation. Sakumoto and Miyata (1984) reported the
purification of polluted water by combined application of radiation and conventional methods, such
as biological oxidation, coagulation with Fe2(SO4)3
and ozonation. The choice of combination of
methods depends on the biological, chemical and
physical properties of the pollutants and their behaviour against radiation. Water containing organic
compounds having functional groups, such as hydroxyl, carbonyl, carboxyl etc. will be first irradiated
and then introduced to a biological degration procedure. Soluble monomers and polymers in water can
be converted under irradiation into higher molecular
weight substances. Hence, their removal can be successfully achieved by a subsequent coagulation in the
presence of air. Although ozone is a powerful oxidizing agent it cannot lead in some cases to a complete
destruction of pollutants without double bonds in
their molecule. As shown above 0 3 is a rather selective reagent to an unsaturated part of a pollutant
molecule, however, as mentioned before, the ozone
radiolysis in water results in the formation of OH and
H02/O ~- radicals (e.g. Weiss, 1935; Taube and Bray,
1940; Alder and Hill, 1950 etc.) which are able to
decompose the organic substances. This is one of the
major reasons for the efficient synergistic effect of the
O3-radiation combination. The electron beam processing of drinking water as well as of wastewaters is
schematically presented in Fig. 9 (see also Getoff,
1992a, 1993a, 1993b).
In the case of drinking water no chlorination
procedure is needed before the electron beam processing (Fig. 9A). The optimal dose (and dose-rate) have
to be elaborated by preliminary experiments. The
scheme for electron treatment of wastewater is given
in Fig. 9B. It represents a combination of radiation
and conventional methods. The most efficient procedure should be developed by experiments designed
for the individual purpose.
6. ECONOMICS
Concerning the economics for degradation of pollutants in water by electron beam processing in the
prsence of ozone or by combined methods of radiation and conventional techniques, it is difficult to
give precise data in advance. This is so because of
many factors involved in the process, such as kind
and amount of pollutants in water, their properties
(chemical, biological, physical etc.) as well as the
dose, dose-rate needed and their decomposition in the
presence of air, ozone etc. These parameters can be
determined by laboratory and pilot-plant experiments. As an example in this respect the studies by
Hashimoto et al. (1988) can be useful.
7. CONCLUSION
After a brief introduction concerning some basic
data of radiation units, sources and primary products
of water radiolysis, the radiation induced degradation
of chlorinated pollutants in water are reviewed. The
decomposition process of aliphatic, olefinic and aromatic substances under various conditions are
discussed and probable reaction mechanisms are
presented. The synergistic effect of radiation and
ozone is also reviewed. Finally, schemes for purification of drinking water as well as of industrial
wastewater are proposed.
REFERENCES
Adams G. E., Boag J. W. and Michael B. D. (1966)
Transient species produced in irradiated water and
aqueous solutions containing oxygen. Prec. R. Chem.
Soc. Ser. A 289, 321.
Alder D. A. and Hill G. R. (1950) The kinetics and
mechanisms of hydroxide ion catalyzed ozone dgcomposition in aqueous solution. J. Am. Chem. See. 72, 1984.
Arai H., Arai M. and Sakumoto (1986) Exhaustive degradation of humic acid in water by simultaneous application
of radiation and ozone. Water Res. 20, 885.
592
Nikola G-etoff
B/ihler R. E., Staehelin J. and Hoigne J. (1984) Ozone
d~omposition of water studied by pulse radiolysis.
HO2/O~- and H()3/O~- as intermediates. J. Phys. Chem.
88, 2560.
Buxton G. V. and Sellers R. M. (1973) Acid dissociation
constant of the carboxyl radical. J. Chem. Soc. Faraday
Trans. L 69, 555.
Cooper W. J., Waite I. D., Kurucz C. N., Nickelsen M. G.
and Lin K. (1993) High energy electron beam irradiation
for the destruction of toxic organic chemicals. Proc.
Radtech. Europe 93, 385.
Cooper W. J., Nickelsen M. G., Meachan D. E., Cadavid
E.-M., Waite T. D. and Kurucz C. N. (1992) High energy
electron beam irradiation. An innovative process for the
treatment of aqueous based organic hazardous wastes.
J. Environ. Sci. Health A27, 219.
Draper R. B., Fox M. A., Pelzetti E. and Serpone N. (1989)
Pulse radiolysis of 2,4,5-trichlorophenol: formation, kinetics and properties of hydroxytrichloro-cyclohexadienyl,
trichlorophenoxyl and dihydroxytrichlorocyclohexadienyl radicals. J. Phys. Chem. 95, 1938.
Fishbein L. (1979) Potential Industrial Carcinogens and
Mutagens. Elsevier, Amsterdam.
Gehringer P., Proksch E. and Szinowatz W. (1985) Radiation-induced degradation of trichloroethylene and tetrachloroethylene in drinking water Int. J. Appl. Radiat. Isot.
36, 313.
Gehringer P., Proksch E., Szinowatz W. and Eschweiler H.
(1986) Radiation-chemical degradation of traces of trichloroethylene and perchloroethylene in drinking water.
Z. Wasser Abwasser Forsch. 19, 196.
Gehringer P., Proksch E., Szinowatz W. and Eschweiler H.
(1988) Decomposition of trichloroethylene and tetrachloroethylene in drinking water by a combined radiation-ozone treatment. Water Res. 22, 645.
Gehringer P., Proksch R., Eschweiler H. and Szinowatz W.
(1990a) Removal of chlorinated ethylenes from drinking
water by radiation treatments. Radiat. Phys. Chem. 35,
456.
Grehringer P., Proksch E., Eschweiler H. and Szinowatz W.
(I 990b) Remediation of groundwater polluted from chlorinated ethylenes by ozone-electron beam irradiation
treatment. Report OEFZS-4558, Seibersdorf, Austria.
Gehringer P., Proksch E., Eschweiler H. and Szinowatz W.
(1992) Remediation of groundwater polluted with chlorinated ethylenes by ozone-electron beam irradiation treatment Int. J. Appl. Radiat. Isot. 43, 1107.
Getoff N. (1986a) Radiation-induced decomposition of biologically resistant pollutants in water. Int. J. Appl. Radiat. Isot. 37, 1103.
Getoff N. (1986b) Radiation-induced decomposition of
some chlorinated methanes. Water Res. 20, 1261.
Getoff N. and Lutz W. (1985) Radiation induced decomposition of hydrocarbons in water resources. Radiat. Phys.
Chem. 25, 21.
Getoff N. (1990) Decomposition of biological resistant
pollutants in water by irradiation. Radiat. Phys. Chem.
35, 432.
Getoff N. (1987) Radiation-induced decomposition of chlorinated pollutants in water. Proc. Prospect for Intense
Radiation Sources. AECL, Accelerator Systems, Deep
River, Ontario, Canada.
Getoff N. (1989) Advancements of radiation-induced degradation of pollutants in drinking and waste water Int. J.
Appl. Radiat. Isot. 40, 585.
Getoff N. (1992a) Radiation processing if liquid and solid
industrial wastes. In Application of Isotopes and Radiation in Conservation of the Environment, pp. 153-169.
IAEA, Vienna, Austria.
Getoff N. (1992b) Radiation-induced decomposition of
phenol and chlorinated phenols in water. In Application
of Isotopes and Radiation in Conservation of the Environment, pp. 198-200. IAEA, Vienna, Austria.
Getoff N. (1993a) Radiation-induced decomposition of pollutants in water. A short review. In Proc. Radtech Europe
93, pp. 371-383. Italy.
Getoff N. (1993b) Using high-energy electrons to solve
problems. Waste Magazine 1, 45.
Getoff N. and Solar S. (1986) Radiolysis and pulse radiolysis
of chlorinated phenols in aqueous solution. Radiat. Phys.
Chem. 28, 443.
Getoff N. and Solar S. (1988) Radiation-induced decomposition of chlorinated phenols in water. Radiat. Phys.
Chem. 31, 121.
Gilbert E. and Giisten H. (1973) Radiation chemical degradation biological resistant organic pollutants in aqueous
solution. Vom Wasser 41, 359.
Haberl R., Urban W., Gehringer P. and Szinowatz W.
(1991). Treatment of pulse-bleaching effluents by activated sludge precipitation, ozonation and irradiation.
Water Sci. Technol. 24, 229.
Hashimoto S., Miyata T. and Kawakami W. (1980) Radiation-induced decomposition of phenol in flow system.
Radiat. Phys. Chem. 16, 59.
Hashimoto S., Nishimura K. and Machi S. (1988) Economic
feasibility of radiation-composing plant of sewage sludge.
Radiat. Phys. Chem. 31, 109.
Hashimoto S., Miyata T., Washimo M. and Kawakami W.
(1979) A liquid chromatographic study on the radiolysis
of phenol in aqueous solution. Environ. Sci. Technol. 13,
71.
Hoigne J. (1982) In Handbook of Ozone Technology and
Applications (Edited by Rice R. G. and Netzer A.), Vol. 1,
pp. 341-379. Ann Arbor, Ann Arbor, MI, U.S.A.
Hutzinger O., Safe S. and Zitko V. (1974) The Chemistry of
Polychlorinated Triphenyls. CRC Press, Boca Raton, Fla.,
U.S.A.
Hutzinger O., Frei R. W., Merian E. and Pocchiari F.
(1982). Chlorinated Dioxines and Related Compounds-Impact on the Environment. Pergamon Press, Oxford.
Koltzenburg G., Behrens G. and Schulte-Frohlinde D.
(1982) Fast hydrolysis of alkyl radicals with
leaving groups in the fl-position. J. Am. Chem. Soc. 104,
7311.
Ktster R. and Asmus K. D. (1971) Die reaktionen chlorierter ethylene mit e~ und OH radikalen in w~isseriger
16sung. Z. Naturforsch. 26b, 1108.
Kurucz C. N., Waite I. D., Cooper W. J. and Nickelsen
M. J. (1991) High energy electron beam irrdiation of
water, wastewater and sludge. Adv. Nucl. Sci. Technol.
22, 1.
Lal M., Ktnig J. and Asmus K.-D. (1987) Acid formation
in the radical mediated degradation of chlorinated
ethanes in aqueous environmen. A radiation chemical
study. J. Chem. Soc. Perkin Trans. II, 1639.
Lin K., Cooper W. J., Nickelsen M. G., Kurucz C. N. and
Waite T. D. (1993) In Proc. Radtech Europe 93,
pp. 403-413. Italy.
Micic O., Nenadovic and Markovic (1975) Radiation
chemical destruction of phenol in oxygenated solutions.
In Radiation for a Clean Environment, pp. 233-239.
IAEA, Vienna, Austria.
Miller A. (1990) Maximum and minimum doses in gamma
and electron irradiated products. Beta-Gamma 3/4, 6.
Pikaev A. K. and Shubin V. M. (1984) Radiation treatment
of liquid wastes. Radiat. Phys. Chem. 24, 77.
Proksch E., Gehringer P., Szinowatz W. and Eschweiler H.
(1987) Radiation-induced decomposition of small
amounts of perchloroethylene in water. Int. J. Appl.
Radiat. lsot. 38, 911.
Rook J. J. (1974) Formation of haloforms during
chlorination of natural waters. Water Treat. Exam. 23(2),
234.
Sakumoto A. and Miyata T. (1984) Treatment of waste
water by a combined technique of radiation and conventional method. Radiat. Phys. Chem. 24, 99.
Radiation-induced degradation of water pollutants--state of the art
Sehested K., Holcman I. and Hart E. J. (1983) Rate
constants and products of the reactions of e~, Oj- and H
with ozone in aqueous solutions. J. Phys. Chem. 87, 1951.
Sehested K., Holcman J., Bjergbakke E. and Hart E. J.
(1982) Ultraviolet spectrum and decay of the ozonide ion
radical, O ; in strong alkaline solution. J. Phys. Chem. 86,
2066.
Sehested K., Corfitzen H., Holcman J. and Hart E. J. (1992)
Decomposition of ozone in aqueous acetic acid solutions.
J. Phys. Chem. 96, 1005.
Sehested K., Holcman H., Bjergbakke and Hart E. J. (1984)
A pulse radiolytic study of the reaction OH + 03 in
aqueous medium. J. Phys. Chem. 88, 4144.
Sehested K., Corfitzen H., Holcman J., Fischer C.H. and
Hart E. J. (1991) The primary reaction in the decomposition of ozone in acidic aqueous solutions. Environ. Sci.
Technol. 25, 1589.
Selzer S. M. and Berger M. J. (1987) Energy deposition
by electron, bremstrahlung and 6°Co-gamma-ray beams
in multilayer. Int. J. media Appl. Radiat. Isot. 38,
349.
Singh A., Kremers W. and Bennet G. S. (1985) Radiolytic
dechlorination of polychlorinated biphenyls. Proc. Int.
Conf. New Frontiers for Hazardous Management,
pp. 489-493. Pittsburgh, U.S.A.
Staehelin J., Bfihler R. E. and Hoigne J. (1984) Ozone
decomposition in water studied by pulse radiolysis.
RPC 47~4~F
593
OH and H04 as chain intermediates. J. Phys. Chem. 88,
5999.
Strobel K. and Dieter H. J. (1990) Toxikological pickbenefit aspects of drinking water chlorination and of
alternative desinfection procedures. Z. Wasser Abwasser
Forsch. 23, 152.
Takehisa M. and Sakumoto A. (1982) In Industrial applications of radioisotopes and radiation technology, pp.
217-233. Proc. Conf. Grenoble, IAEA, Vienna, Austria.
Takehisa M., Arai H., Arai M., Miyata T., Sakumoto A.,
Hashimoto S., Nashimura K., Watanaba H., Kawakami
W. and Kuryama I. (1985) Inhibition of trihalomethane
formation in city water by radiation-ozone treatment.
Radiat. Phys. Chem. 25, 63.
Taube H. and Bray W. C. (1940) Chain reactions in aqueous
solutions containing ozone, hydrogen perioxide in acid.
J. Am. Chem. Soc. 62, 3357.
Waite T. D., Kurucz C. N., Cooper W. J. and Nickelsen M.
(1992) High energy electrons. Innovative treatment for
detoxifying waste streams and contaminated industrial
sites. In. Applications of lsotopes and Radiation in Conservation of the Environment, p. 171. IAEA, Vienna~ Austria.
Weiss J. (1935) Investigations on the radical HO2 in solution. Trans. Faraday Soc. 31, 668.
Wiesner L. (1979) Penetration of electron beams through
unevently distributed absorbers. Radiat. Phys. Chem. 14,
457.