Decomposition of aqueous solutions of phenol using high energy electron beam irradiation—A large scale study

~

Appl. Radiat. lsot. Vol.46, No. 12, pp. 1307-1316, 1995

Pergamon

096941043(95)00236-7

Copyright © 1995ElsevierScienceLtd Printed in Great Britain.All fights 0969-8043/95 $9.50+0.00

reserved

Decomposition of Aqueous Solutions of Phenol Using High Energy Electron Beam Irradiation A Large Scale Study KAIJUN

L I N I, W I L L I A M J. C O O P E R ~*, M I C H A E L G. N I C K E L S E N I, C H A R L E S N. K U R U C Z 2 a n d T H O M A S D. W A I T E 3

~Drinking Water Research Center, Florida International University, Miami, FL 33199, U.S.A. 2Department of Management Science and Industrial Engineering and 3Department of Civil and Architectural Engineering, University of Miami, Coral Gables, FL 33134, U.S.A. (Received I August 1994; in revised form 13 June 1995)

High-energy electron-beam irradiation was used to remove phenol from aqueous solution. The variables that affected phenol decomposition were solute concentration, absorbed dose and total alkalinity. Experiments were conducted at large scale (480L min-~), at solute concentrations of 10.6, 106 and 531/zmol L-~ (1, 10 and 50 mg L -~) over the pH range 5-9, and in the presence and absence of solids (3% w/w kaolin clay). Absorbed doses ranged from 0-7 kGy (0-700 krad). At low absorbed doses, catechol, hydroquinone and resorcinol were identified as the major reaction byproducts. These compounds are consistent with hydroxyl radical (OH.) addition to phenol. Subsequent ring cleavage of hydroxylated phenolic radicals and continued oxidative processes resulted in the formation of formaldehyde, acetaldehyde, glyoxal and formic acid. At high doses only trace amounts of the carbonyl derivatives were observed. Two recirculation experiments were conducted at higher phenol concentrations (g950 #mol L- ~) and it was shown that phenol was removed while the total organic carbon of the solution decreased only slightly. These results suggest that phenol was not mineralized but, rather, that irradiation resulted in the possible formation of higher molecular weight polymers.

Introduction Phenol and substituted phenols are widely distributed at low concentrations in natural waters. Trace amount of phenols may have detrimental effects on water quality, being toxic to fish and aquatic life in general. Natural sources of phenols in the aquatic environment include algal secretion (Seiberth and Jensen, 1969), hydrolyzable tannins and flavanoids (Swain, 1979), and humification processes (Mathur and Schnitzer, 1978; Freudenberg and Neish, 1969). However, at the high concentrations found in some industrial wastewater discharges and agricultural activities (Goerlitz et al., 1985; Wegman and Van Den Brook, 1983; Milner and Goulder, 1986) phenols may generally be toxic. U.S. Environmental Protection Agency (EPA) has included 11 phenols on the list of priority pollutants and these phenols have to be monitored at sub-ppb levels in surface water (Sitting, 1980). Phenols are also one of the most frequently found hazardous compounds at Superfund sites (McCoy and Associates, 1985).

*Author for correspondence.

Conventional biological methods of waste treatment are often inadequate for removing phenol, particularly if it is present at high concentrations. Physical processes such as activated carbon adsorption (Najm et al., 1991) may be effective, but they do not completely solve the problem because the spent carbon must either be regenerated or disposed of properly. A potential chemical treatment process is the oxidation of phenols by ozone (03). However ozonation has high selectivity to the reactants, for example, the rate of ozone reactions with phenol was three orders of magnitude slower than its reaction byproduct hydroquinone (Gurol and Nekouinaini, 1984). The rate of 03 reaction with phenol was directly proportional to the concentration of phenol and ozone, and increased with increasing pH in aqueous solution. At elevated pH the bimolecular rate constant approaches 9.0 x 102 M -j W -l (Konstantinova et al., 1991). However, under these conditions, longer processing times may preclude the use of 03 for many applications such as flow-through systems. To reduce the contact time and to increase removal eificiencies, 03 is often coupled with hydrogen peroxide or UV light (Gurol and Vatistas, 1987; McShea et al., 1987;

1307

Ka~un

1308

Takahashi, 1990). Even combined ozone treatment techniques have limitations when dealing with aqueous solutions high in total suspended solids. Additional methods such as photocatlysis (Richard and Boule, 1994) and peroxidase-catalyzed polymerization (Yu et al., 1994) may be alternative technologies for the control of phenol. Current studies using high-energy electron-beam irradiation have demonstrated that the process is efficient for the destruction of several classes of hazardous organic compounds (Nickelsen et aL, 1992; Cooper et al., 1992a--c; Kurucz et al., 1992). In this paper we extend this work to the destruction of phenol in aqueous solution in a large-scale flowthrough system. Studies have been conducted at several solute concentrations, in the presence and absence of clay, and at different pHs and alkalinity concentrations. Experimental

Electron Beam Research Facility (EBRF) All of the experiments discussed were conducted at the Electron Beam Research Facility (EBRF) located at the Miami-Dade Central District (Virginia Key) Treatment Plant, Miami, FL. The research facility has been described in detail elsewhere (Kurucz et al., 1992).

Solute preparation 90% Liquified phenol (Fisher Scientific) was used without any further purification. All experiments were conducted with 11,355 L (3000gal) batch solutions using either 4600 or 6000 gal tank trucks. The phenol was added directly to the tank truck and mixed by two pumps for 30 min to give the desired concentration (nominally 10.6, 106 and 531 #mol L -~, 1, 10 and 50 mg L -l, respectively). Potable water was used for all experiments. In addition to using the potable water as received (pH 8.5-9.0), hydrochloric acid (12N) was used to adjust the pH to 7 or 5. The pH adjustment was used either to alter the carbonate-bicarbonate ion ratio (pH 7) or to eliminate the total carbonate alkalinity (pH 5). To simulate sludge, sediments, or slurried solids, kaolin clay (EPK) (Miami Clay Co.) was added into tanks and mixed thoroughly by two diaphragm pumps prior to irradiation. The content of clay was approx. 3% in

Table 1. Typical water characteristic values for potable water used in studies to determine the removal et~ciency of phenol

Parameter pH Dissolved 0 2 (DO) nag L -I 0 2 Dissolved organic carbon (DOC) mg L - ~ C CaCO 3, mg L - 1 H C O f i a mg L-I CO~ - a , mg L -I NO~--N, mg L -I

JCalculated from equilibria.

Potable water 8.8 +_0.3 3.1 _+ 0.5 5 +_ 2 34 ± 4 20 ± 2 0.4-0.6 0.1-0.2

Lin et al. Table 2. Products formedas a result of the irradiation of aqueous

solutions and their concentrationsat several doses"

Concentration of products (raM) 0.50

1.00

5.00

G-value Species

(gmol J - ~)

efq H" CH' H30 +

0.27 0,062 0,28 0.27 0.047 0.073

H2

H202

Dose (kGy) 0.14 0.031 0.14 0.14 0.023 0.036

0.27 0.062 0.28 0.27 0,047 0.073

1.4 0.31 1.4 1.4 0.23 0.36

"Spinks and Woods (1990), Buxton et al. (1987).

the aqueous solutions and no obvious precipitation was observed during experiments. In addition, two recirculation experiments (pH 3.95 and 7.75) were conducted by plumbing the effluent line of the electron beam irradiator into the front end of a 4600 gal tank truck. The initial phenol concentration was approx. 950 #mol L-~ (90 mg L-~). During the recirculation experiment, various electron-beam doses were applied over a 2.5 h period.

Sample collection and analysis Influent (non-irradiated) and effluent (irradiated) samples were collected in 47 mL PTFE screw-cap glass vials. Triplicate vials were completely filled, immediately chilled on ice in the dark, and returned to the laboratory where phenol analyses were performed within 48 h. No post-irradiation loss of phenol was observed ( < 10%) in samples stored for 2 months with or without clay. General water characteristics that were monitored included dissolved oxygen (YSI electrode), pH, total alkalinity, chloride, bromide, sulfate, phosphate and nitrate ion (Dionex 4000i Ion Chromatograph), and total dissolved organic carbon (persulfate-UV oxidation to CO:), using standard methods (American Public Health Association, 1985). Table 1 summarizes typical water quality parameters for the experiments conducted in this study. Total alkalinity, in the waters used in this study, refers to the carbonate/bicarbonate ion concentration in the water.

Phenol and phenolic byproduct analysis A Hewlett-Packard model 1090 high performance liquid chromatograph (HPLC) with a 1040M

Table 3. Second-order reaction rate constants (L tool t s -~ ) of the reactive transient species formed in irradiated aqueous solutions with solutes of interest in this investigation'

Compounds Phenol

Phenoxide ion Bicarbonate Carbonate 02 H202

e~ 2.7 4.0 <1 3.9 1.9 1.1

x x x x x x

107 106 106 l0 s 107 104*

H.

OH.

1.7 x l0 g NR 2.1 x l0 s NR 2.1 x 107 9.0 x 104.

6.6 x l0 9 9.6 x l09 3.2 x 106 3.9 x 108 NR 2.7 x 104

aSpinks and Woods (1990), Buxton et al. (1987). N R = not reported in the literature', * = measured in closed systems.

Electron beam treatment of phenol

1309

Table 4. Summaryof the doses required to remove99% of phenol(/)0.99in kGy) from aqueoussolutionusinghigh-energyelectronbeam irradiation, the percentage phenol removed and Go (/~molJ-') in the absence and presence of 3% w/w kaolin clay at three solute concentrations pH 5 pH 7 pH 9 pH 5a pH 7~ Dose (kGy) % Removed GD % Removed GD • Removed Go % Removed Go % Removed Go 0.25 0.50 1.00

99.2 0.050 99.6 0.025 99.6 0.012 Do 99 = 0.72

0.50 1.00 3.00

69.4 0.17 93.5 0.11 99.7 0.040 00,99 = 1,50

1.00 3.00 7.00

41.2 0.26 76.9 0.16 98.9 0.090 00.99 = 7.00

Initial concentration ~ 10.6/tmol L 93.8 0.062 81.0 0.040 94.8 0.045 99.0 0.033 99.9 0.024 99.9 0.024 99.0 0.018 99.9 0.013 99.9 0.011 00.99 = 0.66 00.99 = 0.60 Do.99= 0.63 Initial concentration ~ 106/zmol L 65.6 0.16 53.0 0.13 69.2 0.16 89.8 0.11 80.8 0.10 92.1 0.10 99.9 0.041 99.2 0.042 99.9 0.099 O009 = 1.90 D0,99 = 3.00 0099 = 1.50 Initial concentration := 531/~mol L J 38,3 0.24 25.0 0.16 43.8 0.25 79.4 0.17 50.5 0.11 79.5 0.15 98.2 0.090 83,7 0.075 99.0 0.079 00.99 = 8,10 00.99 = 15.40 D0.99 = 730 krad

94.8 0.037 99.9 0.020 99.9 0.010 Do 99 = 66 krad 64.7 0.14 91.4 0.099 99.9 0.035 D099 = 2.10 39.9 79.3 98.4

0.23 0.15 0.081 Do99 = 7.80

aPotable water with addition of 3% w/w kaolinclay.

diode-array UV/vis detection system (DAD) was used for the determination of phenol and phenolic byproducts (i.e. catechol, hydroquinone, resorcinol, etc.). The identification and quantitative analysis of phenols was based on the comparison of their retention times with authentic standards as well as the on-line confirmation that was the comparison of the UV spectrum of each separated compound with the standard files previously saved in the database. Separation was performed on a Hypersil 5 # m reversephase C~8 column (2.1 x 100 mm i.d.). A two-solvent gradient elution program was used. Solvent A consisted of 1% acetic acid/99% water (pH 2.8). Solvent B was 100% methanol. After an initial isocratic elution (flow 0.3 mL rain -j ) with 100% A for 4 min, a linear solvent gradient to 50% A in 11 rain was employed, followed by a second linear gradient to 100% B in 5 min. Elution with 100% B was maintained for 2 min and the column was then returned to 100% A over 2rain. The column temperature was maintained at 40°C throughout the chromatographic run. The lower limit of detection for all phenolic compounds was 0.1/~mol L -~ using an injection volume of 250/~L. In every case phenol determinations were duplicated. If the difference was less than 5%, the average was used, If it was greater, then a third analysis was run and the average of the three determinations used.

Low molecular weight aldehyde and ketone analysis Low molecular weight aldehydes and ketones were also analyzed by HPLC with retention time comparison and D A D on-line confirmation. The method was a pre-column derivatization reaction with 2,4-dinitrophenylhydrazine in acidic condition, followed by reverse phase HPLC analyses of the derivatized products (Puputti and Lehtonen, 1986; Kieber and Mopper, 1990). The pre-column derivatization was conducted automatically using HPI090 injection program. AR| 46/[2--B

Determination of organic acids A Dionex model 4000i ion chromatograph (IC) was used for analyses of low molecular weight carboxylic acids. The acids were separated using an lonPac ICE-AS5 exclusion column with an anionICE micro membrane suppressor (AMMS-ICE) and determined by their conductivity. The eluent was 0.8 mM heptaftuorobutyric acid maintained at 0.3 mL rain -~. The suppressor regenerant solution was 5 m M tetrabutylammonium hydroxide (2mL min -1 ).

Determination of total organic carbon The total organic carbon (TOC) concentration was measured by a Shimadzu TOC-5000 total organic carbon analyzer. Absorbed dose. All of the experiments were conducted at a large scale (7.5-15 m 3 solutions) and a flow rate of ~0.45 m 3 min -~. To measure absorbed dose, five resistant temperature devices were installed, two before and three after irradiation. The temperature difference was used to estimate dose. An increase of I°C is approximately equal to 4.18 kGy (Kurucz et al., 1992).

Results and Discussion Table 2 summarizes the radiolytic radicals, ions and molecular species that exist 10 -7 S after irradiation of an aqueous solution, and their concentration with increasing doses. The three reactive species of greatest interest in the destruction of toxic organic compounds are the aqueous electron (e~q) hydrogen atom (H.) and hydroxyl radical (OH.) (Spinks and Woods, 1990). In oxygenated conditions, the superoxide formed through reactions of e~ and H" with oxygen may also be important in the removal of solutes and solute reaction byproducts. Table 3 lists the known bimolecular reaction rate constants of

Kaijun Lin et al.

1310

Table 5. Analysis of covarianee for the effect of pH and clay presence on D099 values adjusting for differences in initial concentration of phenol Source

DF

Seq. SS

Adj. SS

Adj. MS

F

P

Initial conch pH Clay Replicate Error Total

1 2 I 1 24 29

4,600,933 679,414 12,353 21,667 1,280,130 6,594,496

4,486,543 643,277 12,308 21,667 1,280,130

4,486,543 321,638 12,308 21,667 53,339

84,11 6.03 0.23 0.41

0.000" 0.008* 0.635 0.530

Term

Coeff.

SD

Constant Initial conch

93.28 15.649

59.59 1.706

Covariate properties, initial cohen t-value P 1.57 9.17

0.131 0.000"

*Significant at a = 0.05. DF = degrees of freedom; SS = sum of squares; MS = mean square; F = F-ratio; P = probability of obtaining the observed F-ratio, or greater, given the null hypothesis of no effect is true.

the three reactive species with phenol and phenoxide ion, and inorganic solutes commonly found in natural waters. From these data, the relative importance of each of the reactive species in the destruction of phenol can be estimated (i.e. 0,4% for e~, 5.4% for H. and 94.2% for OH.) (Nickelsen et al., I992). Based on these calculations the formation of hydroxylated compounds through OH. addition would be expected as the major reaction byproducts. Phenol decomposition

Table 4 summarizes the doses required to remove 99% of the phenol at each solute concentration, pH 25

•

(A) 2.0

.j

15

1.5 3

O z

"I" n

1.o 5

©

0.5 -

0.00

0.25

'

'

0.50

"=0.0

0.75

1.00

A B S O R B E D D O S E (kGy)

150

, 14 112

B-(~---o-------~ ~

100

10

°z

e ~

m 50 "1" ~ 0

0.0

4 2 0.5

1.0

1.5

2.0

2.5

0

3.0

80

60

~

60

500

50

Z

400

40 3

zoo 200

3° 20

100

10

700 6001 ~

Q Z 0.

100q

0

, 1

.

, 2

, 3

~ 4

= 5

. "x----.,-- 0 6 7

!

I

=

I

I

2.0

2,5

_z Z <

A B S O R B E D D O S E (kGy)

.-

and in the presence of 3% w/w kaolin clay. To calculate 99% destruction, the experimental dose constant (k), or, the slope of the line obtained by plotting the natural logarithm of diluent phenol concentration vs absorbed dose, was used (Cooper et aL, 1992a; Kurucz et aL, 1992). In radiation chemistry, G-values are defined as the number of molecules lost or formed for every 100 eV absorbed by the system and in SI units it is G x 0.104 #mol J-~. Therefore, another quantitative descriptor of removal efficiency is the radiation chemical yield Go, the G-value for the loss of phenol at any dose D (Nickelsen et al., 1992; Kurucz et al., 1992). The results of these calculations are also summarized in Table 4. The percentage phenol removed at each dose and concentration combination is also summarized in Table 4. The greatest percentage removal was observed at the lowest initial solute concentration (10.6#tool L -~) and in every case increased with increasing dose. The percentage removal decreased with increasing pH from 5 and 7 to pH 9, and appears unaffected by the addition of 3% solids as clay. These results are consistent with the statistical analysis in terms of D0.99 values and indicate that increased removal efficiency resulted from either a lower

U.I IZ

~ I1:

20

W 0

0.0

I

I

0.5

1.0

1.5

3.0

A B S O R B E D D O S E (kGy)

ABSORBED DOSE

Fig. l. R e m o v a l o f phenol: (A) 2 0 / ~ m o l L - t ; (B) 135 # m o l

Fig, 2. Comparison of phenol removal efflciencies for an initial concentration of 106 #mol L -I in potable water at pH 9 (O), at pH 5 (~7) and at pH 9 ([-I) with low total alkalinity.

L-t; and (C) 600#mol L -I, from potable water at pH7 using high-energy electron beam irradiation (O--influent; Vl--effluent).

(kGy)

1311

Electron beam treatment of phenol Table 6. Comparison of phenol removal Go values (106 a m o l L -~ ) at pH 9, 5 and at 9 with N a O H adjustment

Table 8. The formation of formaldehyde, acetaldehyde, glyoxal and formic acid from the irradiation of phenol at 106 vmol L -)

Dose (kGy)

Products (/zmol L - I )

0.50 1.00 3.00

GD at pH 5

G o at pH 9

GD at pH 9 with N a O H

0.148 0.103 0.034

0.105 0.083 0.036

0.124 0.087 0.035

pH and/or a lower bicarbonate/carbonate ion concentration. The efficiency of phenol removal as determined using Go-values (Table 4) also decreases with increasing pH. However, at any given solute concentration and pH, the removal efficiency decreased with increasing dose. This suggests that as irradiation dose increases, radical-radical recombination reactions involving OH', e~ and H. also increase, thereby reducing removal efficiency (Trumbore et al., 1988; Goulet and Jay-Gerin, 1992). A three-factor analysis of covariance was performed on the phenol removal data of Table 4 using the General Linear Model (GLM) procedure of MINITAB. The GLM procedure was used to allow for the simultaneous consideration of both qualitative factors and quantitative variables. Replicate differences and the presence of clay were not significant at ~ = 0.05 (the probability of rejecting the null hypothesis of no effect, when indeed there is no effect) as shown by the P-values in Table 5. Both pH and the "covariate" (initial phenol concentration) were statistically significant at ~ = 0.05. The covariate coefficient value in Table 5 indicated that the required Do.99 increases by 0.157 kGy for each l mg L-J increase in initial phenol concentration. The pH effect was evident when the adjusted m e a n D0.99 values (not shown), which are 3.39, 3.31 and 7.04kGy at pH 5, 7, and 9, respectively, were compared. Parts A-C of Fig. 1 demonstrate the effective removal of phenol against various doses at three initial solute concentrations at pH 7. Two possibilities could account for the increased phenol removal with decreasing pH. First, the relative scavenging of the OH- at pH 9 was greater than at either of the two lower pHs. Second, it is possible that differences in the reaction rate of the phenol/phenoxide ion [pKa= 10 (Hoigne and Bader, 1976)] with the reactive species (Table 3) could lead to differences in removal efficiency at different pHs. (The effect of 02 concentration was not studied or controlled; however, it was measured before and after

pH5

Formaldehyde Acetaldehyde Glyoxal Formic acid Total yield

1.24 0,54 1.43 ND 3.21

Formaldehyde Acetaldehyde Glyoxal Formic acid Total yield

0.98 0.45 1.99 ND 3.42

Formaldehyde Acetaldehyde Glyoxal Formic acid Total yield

0.40 0.51 0.56 ND 1.47

pH7

pH9

pH5 ~

Absorbed 0.51 NDb 0.92 ND 1.43 Absorbed 0.42 0.39 0.78 ND 1.59

dose ==0..~0 kGy 0,62 1.18 0.29 0.58 0.39 0.,*4 9.6 12.4 10.9 14.6 dose --- 1.00 kGy 1.09 1.48 0.39 0.71 0.52 1.03 18.7 17.8 20.7 21.0 Absorbed dose = 3.00 kGy 0.17 0.20 0.57 0.34 0.44 0.81 0.31 0.72 0.26 ND 16. I 13.3 0.82 17.5 14.9

pH7 ~ 0.55 0,09 0.11 ND 0,75 0.66 0.32 0.33 ND 1.31 0.30 0.59 1.06 ND 1.95

~Potable water with 3% w/w kaolin clay. bND = not detected. Influent concn (amol L I): 0.04 formaldehyde; ND for acetaldehyde, glyoxal and formic acid. Detection limit (,umol L -~): 0.01 formaldehyde; 0.05 acetaldehyde; 0.01 glyoxal; 1.00 formic acid.

irradiation and for all experiments was at approximately the same concentration before irradiation as shown in Table 1. In every experiment the 02 concentration in the effluent sample, after irradiation, was lower with respect to the influent concentration. The amount of 02 removal was directly related to increased dose.) To examine these possibilities, the removal of phenol was determined in potable water as received at pH 9 (total alkalinity 55.7 mg L -~ as CaCO3) and at pH 5. A third experiment was conducted where the pH was lowered to 5, the water thoroughly mixed for 30 min to eliminate the CO2, and the pH adjusted back to 9 with the addition of NaOH. Phenol was added and the removal efficiency determined. Figure 2 shows the percentage phenol remaining at pH 9, 5 and 9 with the addition of NaOH (minimum carbonate alkalinity). The GDvalues calculated from the three experiments are listed in Table 6 for comparison. It appears from these results that alkalinity affected the removal efficiency of phenol while the effect of pH was small if even significant. The Go-values at the lower dose (50 krad) appear to be higher at pH 5 than either of the pH 9 solutions. However, as dose is increased the Go-values are nearly the same. An analysis of variance of this data showed that only dose was significant at ~ = 0.05, however, there was insufficient

Table 7. The formation of hydroquinone, catechol and resorcinol from the irradiation of phenol at 106t~mol L -L at an absorbed dose of 0.50 and 1.00 kgy

pH Hydroquinone (,umol L -)) Catechol (/~mol L I) Resorcinol (/~mol L ~) Total yield (pmol L - ) )

Detection limit (.umol L -I ) 0.10 0.05 1.00

Absorbed dose = 0.50 kGy

Absorbed dose = 1.00 kGy

5

7

9

5a

7a

5

7

9

5"

7"

16.3 20.1 ND 36.4

13.0 20.3 ND 33.3

2.8 18.8 ND 21.2

15.8 18.2 ND 34.0

4.4 12.5 3.8 20.7

6.0 9.5 ND 15.5

6.2 II.l 2.7 20.0

9.9 18.3 3.4 31.6

6.7 8.4 ND 15.1

4.8 7.3 ND 12.1

aPotable water with addition of 3% w/w kaolin clay, N D = not detected.

1312

Kaijun Linet al.

25 20

Z _0 15 In" I-Z 10 I./J 0 Z 0 0 5

0.0

0.5

1.0

1.5

2.0

2.5

3.0

ABSORBED DOSE (kGy)

Fig. 3. The f o r m a t i o n a n d d e c o m p o s i t i o n o f catechol ( V ) ,

hydroquinone (O) and resorcinol ([]) vs absorbed radiation dose as function of phenol removal at pH 5. The initial concentration of phenol was 106#mol L ~.

data to test for the effects of interaction between dose and pH/alkalinity. Formation o f reaction byproducts

The formation of several di-hydroxy substituted phenol derivatives was observed (yon Sonntag and Schuchmann, 1991). These compounds were identified as hydroquinone and catechol, with trace amounts of resorcinol. These results are consistent with those of others from the 6°Co ?-irradiation of benzene (Pan et al., 1993) and phenol (Ahrens, 1967; Mini6 et al., 1975; Hashimoto et al., 1992) and from the ozonation of phenol (Gurol and Vatistas, 1987; Takahashi, 1990; Legube et al., 1981). These dihydroxy phenol derivatives were not observed in the samples prior to irradiation (influent) or in the effluent samples at 0 krad, suggesting that they were indeed reaction byproducts from phenol decompositions. The concentrations of hydroquinone, catechol and resorcinol resulting from the irradiation of nominal 106 #mol L -~ phenol solutions are summarized in Table 7. Hydroquinone and catechol were formed in the highest concentration at low absorbed doses and their concentration decreased with increasing radiation to below detection limits at the highest dose of 300 krad. The distribution of these compounds was pH dependent. At pH 9 where the removal efficiency was most inefficient, the highest concentration of the di-hydroxy substituted phenols occurred at 100 krad. For pH 5 and 7 the maximum concentration of these products was found at 50 krad. This also suggests more efficient destruction of phenol at lower pH values. Figure 3 demonstrates the relative concentrations of the three di-hydroxy substituted phenols at pH 7 where the initial phenol concentration was 106 #mol L -z . The formation of di-hydroxy phenol derivatives is consistent with the -OH group on the phenyl ring acting as an ortho/para director for aromatic electrophilic addition of OH. (Land and Ebert, 1967). At high doses, tri-hydroxy phenol derivatives, such as

pyrogallol, hydroxyhydroquinone and phloroglucinol, might also be formed as reaction byproducts of phenol (Chrostowski et al., 1983), The tri-hydroxy phenol derivatives were not detected in these experiments using our HPLC method. The yellow-brown coloration of irradiated samples at the highest initial solute concentration (106 ~mol L-~ ) suggests that the decomposition pathway of phenol may also result in polymerization through an oxidative coupling mechanism (Chrostowski et al., 1983). The polymerization may lead to the formation of "humic-like" substances as reported by Li and coworkers (Li et al., 1979); however, there was no precipitate. In addition to these hydroxy phenol reaction byproducts, carbonyl compounds were also found in all irradiated samples of the nominal 106/~mol L -~ phenol concentrations. These carbonyl compounds were identified as formaldehyde, acetaldehyde and glyoxal and their concentrations are summarized in Table 8. The formation of aldehydes, particularly glyoxal, indicate that phenol decomposition undergoes aromatic ring cleavage similar to the ozonation of phenol in aqueous solution (Baiely, 1972; Takahashi, 1990). These carbonyl compounds were found at all radiation doses and accounted for less than 1% of the initial phenol concentrations. Formaldehyde and glyoxal were observed at lower radiation doses, i.e. 50 and 100krad, but not at the higher doses. Another byproduct identified was formic acid, and its concentration at various doses is also summarized in Table 8. The concentration of formic acid increased to more than 10#tool L -~ with increased radiation dose. Formic acid has been shown to be formed as a result of the ozone oxidation of low molecular weight aldehydes (Caprio et al., 1989). Therefore, it is assumed that a majority of its formation results from the oxidation of acetaldehyde, formaldehyde, glyoxal, and other unidentified aldehydes and low molecular weight carboxylic acids. However, the formation of formic acid only accounts for <10% of the initial phenol concentration, indicating that the irradiation process probably

Table 9. Analysis of variance of the yield of reaction byproduct as a percentage of initial phenol concentration Hydroquinone Source

P-value

Dose pH Clay

0.276 0.452 0.501

Catechol Source Dose pH Clay

Formaldehyde Dose pH Clay

0.001 * 0.004* 0.088**

Acetaldehyde Dose pH Clay

Glyoxal Dose pH Clay

0.551 0.252 0.139

P-value 0.023* 0.873 0.178 0.045* 0.018 * 0.066**

Formic acid Dose pH Clay

• Significant at a = 0.05. **Significant at ~t = 0.10.

0.631 0.003* 0.026*

Electron beam treatment of phenol results in some mineralization at these lower phenol concentrations. An analysis of variance of reaction byproducts data was performed using the G L M procedure to identify which factors affected byproduct yield and their P-values are listed in Table 9. In order to eliminate the possible effect of differences in the initial concentration of phenol, the yields of byproduct were expressed as a percentage of the initial concentration. For catechol and hydroquinone, however, the results showed that the only significant effect (at ~t = 0.05) is that of dose for the yield of catechol, as shown in Table 9. It seems that the variability of the catechol and hydroquinone data is quite large, probably because the formation of catechol and hydroquinone are dependent on both dose and solution pH. For example, at pH 9 and 0.50kGy, a relatively small amount of phenol was destroyed and therefore a small amount of hydroquinone was produced. Further, because the concentration of hydroquinone was relatively low, the dose of 0.50kGy effectively removed it to approx. 2.8 #mol L - L Alternatively at pH 9 and a 1.00 kGy dose, more phenol was converted to hydroquinone and because of its relatively high concentration, the dose of 1.00kGy was not 1200

(A)

1000

' ~

1313

sufficient to destroy the hydroquinone below 9.9 #tool L -~. Thus the effect of high pH appears to be inconsistent between doses. It should be noted that the interactions between dose and pH, and dose and the presence of clay were not significant. The interaction between clay and pH could not be tested because of the unbalanced data. Carbonyl byproduct and formic acid formations were also analyzed, again expressing yields as a percentage of initial phenol concentration. The results in term of P-values, as presented in Table 9, are generally more consistent as compared to the catechol and hydroquinone. Both dose and pH effects were significant for the formation of formaldehyde. The highest yield occurred at 1.00 kGy and relatively low yields occurred at 0.50 and 3.00 kGy. Yield was minimal at pH 7 with the maximum occurring at pH 9. There was a marginally significant increase of yield in the presence of clay. The results are basically the same for acetaldehyde. However, the variability in the formation of glyoxal prevented any effect from becoming statistically significant. Concentrations of formic acid below detection limits prevented dose from becoming significant. The pH and clay effects were highly significant. Yield increased in the

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Fig. 4. Removal of phenol at an initial concentration of 950/Jtool L - i (A) and total organic carbon (inset)

(©--influent: O--effluent) and the formation and removal of irradiation-producedeatechol (D--influent; O--etttuent) and hydroquinone (V--influent; D--etttuent) (B), during the recirculation experiment at pH 3.95. The duration of the experiment was 2.5 h at a flow rate of 380 L min-L Different doses were applied during the experiment as indicated on the x-axis.

1314

Ka~un Linet al.

presence of clay, particularly at pH 5. The minimum yield occurred at pH 7 and the highest at pH 9 in the absence of clay, and pH 5 in the presence of clay. Although the unbalanced data did not allow a test of pH and clay interaction, it appears that this effect is clearly present. Recirculation experiments

Experiments were designed to test the capability of the high-energy electron-beam irradiation process to treat solutions containing relatively high concentrations of phenol (Kurucz et al., 1991). A solution of approx. 950/~mol L -~ phenol was prepared in potable water and continuously recirculated throughout the experiment at a flow rate of 380 L min- ~. The absorbed dose was varied, 0, 1.5, 4.5 and 6 kGy over a 2.5 h experiment period. The experimental design, that is random changes in applied dose, resulted in a non-uniform removal of phenol. The no-irradiation control (zero dose) was used to test the absolute extent of phenol removal from the entire system or non-irradiation loss of phenol in the system. Parts A-B of Figs 4 and 5 demonstrate the data for the two experiments with initial pHs of 3.95 and 7.75, respectively. It appears that only a 14% decrease of

total organic carbon occurred at pH 3.95, however, after the third recirculation (by volume) the phenol concentration was reduced to >99% of its initial concentration. When zero dose was applied during the recirculation experiment, the phenol concentration in the effluent was the same as in the influent, indicating that there was no non-irradiation phenol loss. As the phenol was removed, the major reaction byproducts observed were hydroquinone, catechol and resorcinol. As can be seen in Part B of Fig. 4, these reaction byproducts increased as irradiation began and were subsequently removed to below their detection limits. In the recirculation experiment conducted at pH 7.75 (part A insert of Fig. 5) it was observed that the total organic carbon decreased by only 2% during the course of the experiment. Phenol removal was less efficient at pH 7.75 and it was reduced to >95% of its initial concentration. Similarly, the concentrations of the reaction byproducts were higher than in the experiment conducted at pH 3.95 (part B of Fig. 5). The results of the recirculation experiments agreed with the flow-through experiments, i.e. higher removal efficiency at low pH and low total alkalinity. At the high concentration of phenol, the stable

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Fig. 5. Removal of phenol at an initial concentration o£950 #too! L-t (A) and total organic carbon (inset)

(O--influent; V'l--effluent)and the formation and removal of irradiation-produced catechol ([-'l--influent; Q--effluent) and hydroquinone (~7--influent; m--effluent) (B), during the recirculation experiment at pH 7.75. The duration of the experiment was 2.5 h at a flow rate of 380 L rain -I. Different doses were applied during the experiment as indicated on the x-axis.

Electron beam treatment of phenol reaction products have not been identified. The process appears to result in the formation of higher molecular weight c o m p o u n d s as major products. The T O C values indicate that the phenol was not mineralized as expected, because the total organic carbon did not significantly decrease at either pH. This indicates that during the main removal pathway of high concentrations of phenol, during the recirculation experiment over a period of 2.5 h, was not through mineralization, but more likely through polymerization processes. However, since each recirculation experiment demonstrated a reduction in T O C concentration, we speculate that continued recirculation o f the waste stream through the electron beam would lead to the destruction of any polymeric intermediates.

Conclusion It has been shown that high-energy electron-beam irradiation effectively removed phenol from aqueous solutions at large scale. Phenol was affected by solute concentration, absorbed dose, and total alkalinity. Other water quality parameters such as the presence of suspended solids and solution p H were not important factors influencing c o m p o u n d removal. At high phenol concentration ( ~ 9 5 0 / ~ m o l L-~), recirculation of the waste stream was necessary to destroy all of the phenol and its reaction byproducts. It was shown that high-energy electron-beam irradiation leads to the formation of oxidized reaction byproducts. At low doses and low phenol concentrations, more than 5 0 o o f the decomposition of phenol was attributed to the formation of hydroquinone, catechol and trace amounts resorcinol. The formation of these di-hydroxy phenol derivatives corresponds closely to those reported as reaction byproducts of phenol with OH. radicals. In addition, aldehydes and formic acid were also identified as reaction byproducts, indicating ring structure rupturing of phenol and further oxidation and mineralization in some cases. Acknowledgements--The cooperation of the Miami-Dade Water and Sewer Authority was essential in the completion of this work. This research was supported by the National Science Foundation under Grants CES-8714640 and BCS9108033, the Environmental Protection Agency under Cooperative Agreement CR-816815-01-0 and Grant R816932-01-0. The work described in this paper has not been reviewed by the U.S. EPA and therefore the contents do not necessarily reflect the views of the Agency and no official endorsement should be inferred.

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